ChapterPDF Available

Nutrient Management Strategies for Coping with Climate Change in Irrigated Smallholder Cropping Systems in Southern Africa

March 2017

March 2017

DOI:10.1007/978-3-319-52381-1_16

In book: Phytoremediation (pp.423-437)

Authors:

Davie Kadyampakeni

University of Florida

Isaac Fandika

Department of Agricultural Research Services, Malawi

Lawrent Pungulani

Ministry of Agriculture, Irrigation and Water Development

Sound management of soil nutrients is critical for optimizing crop vegetative and reproductive development and realizing high yields in irrigated cropping systems. This paper discusses the work done in Africa and presents lessons from other parts of the world for improved nutrient management under irrigation. Considering the rising temperatures and erratic rainfall as a consequence of climatic change and depleted soil nutrients as a result of continuous cropping, this review offers remedial options for managing soil fertility while optimizing water use and crop yields. The paper intends to inform agricultural policy makers and help farmers and organizations in Africa to manage soil nutrient and water resources efficiently and achieve high yields. Importantly, this discussion should stimulate further research in nutrient and water management under varying ecological scenarios of southern Africa to provide a cogent basis for climate change adaptation interventions.

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AbidA.Ansari· SarvajeetSinghGill

RituGill· GuyR.Lanza

LeeNewman Editors

Phytoremediation

Management of Environmental

Contaminants, Volume 5

Phytoremediation

guarino@unisannio.it

Abid A. Ansari • Sarvajeet Singh Gill • Ritu Gill

Guy R. Lanza • Lee Newman

Editors

Phytoremediation

Management of Environmental

Contaminants, Volume 5

guarino@unisannio.it

ISBN 978-3-319-52379-8 ISBN 978-3-319-52381-1 (eBook)

DOI 10.1007/978-3-319-52381-1

Library of Congress Control Number: 2015931077

This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of

the material is concerned, specically the rights of translation, reprinting, reuse of illustrations, recitation,

broadcasting, reproduction on microlms or in any other physical way, and transmission or information

storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology

now known or hereafter developed.

The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication

does not imply, even in the absence of a specic statement, that such names are exempt from the relevant

protective laws and regulations and therefore free for general use.

The publisher, the authors and the editors are safe to assume that the advice and information in this book

are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the

editors give a warranty, express or implied, with respect to the material contained herein or for any errors

or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims

in published maps and institutional afliations.

Printed on acid-free paper

This Springer imprint is published by Springer Nature

The registered company is Springer International Publishing AG

The registered company address is: Gewerbestrasse 11, 6330 Cham, Switzerland

Editors

Abid A. Ansari

Department of Biology

Faculty of Science

University of Tabuk

Tabuk, Saudi Arabia

Ritu Gill

Centre for Biotechnology

Maharshi Dayanand University

Rohtak, Haryana, India

Lee Newman

College of Environmental

Science and Forestry

State University of NewYork

Syracuse, NY, USA

Sarvajeet Singh Gill

Centre for Biotechnology

Maharshi Dayanand University

Rohtak, Haryana, India

Guy R. Lanza

College of Environmental

Science and Forestry

State University of NewYork

Syracuse, NY, USA

guarino@unisannio.it

Preface

“Obscurity knows Nature will light the lamps”

Dahomean Proverb

The editors of Phytoremediation: Management of Environmental Contaminants

originally planned a two-volume book to provide a broad global perspective on the

development and use of phytoremediation to repair and restore contaminated ter-

restrial and aquatic habitats. The success and acceptance of Volumes 1 and 2 led to

the production of three additional volumes that provide a wide diversity of phytore-

mediation laboratory studies and case histories completed in many parts of the

world. Volume 5 contains the nal chapter contributions in the series and adds new

information on the application of soil microorganisms as inoculants or enhancement

agents in contaminated terrestrial habitats including petroleum-contaminated sites.

Other chapters describe the use of both woody and herbaceous plants for the bio-

monitoring and treatment of contaminants and provide new information on the trace

element and toxic metals present in medicinal plants.

In the area of aquatic ecosystems, Volume 5 offers chapters that describe impor-

tant new approaches to applying phytoremediation to increase the efciency of

aquaculture systems and the management of pharmaceutical and personal care

products using constructed wetlands. Other chapters describe the general use of

aquatic plants and oating wetlands to treat polluted water.

Several chapters in Volume 5 offer special applications of phytoremediation in

terrestrial and aquatic habitats and include information on the genetic control of

metal sequestration in hyperaccumulating plants, the use of engineered nanomateri-

als to remove metals/metalloids and their implications on plant physiology, apply-

ing plant biosorbents to extract metals from soils and water, and the phytomining of

rare and valuable metals. Nutrient management strategies for coping with climate

change in irrigated smallholder cropping systems and the phytoremediation of land-

ll leachates are covered in two chapters, and a chapter on the modeling of phytore-

mediation and another on the phytoremediation of contaminated air complete

Volume 5.

guarino@unisannio.it

The complete ve-volume series of Phytoremediation: Management of

Environmental Contaminants is designed to share a diversied sample of the current

laboratory research and eld applications of phytoremediation in a global context.

As editors, we hope that the series will be both useful and informative to academics,

government ofcials, and private sector managers and consultants interested in the

potential for cost-effective and sustainable approaches to improving the environ-

mental quality of terrestrial and aquatic ecosystems.

Tabuk, Saudi Arabia AbidA.Ansari

Rohtak, Haryana, India SarvajeetSinghGill

Rohtak, Haryana, India RituGill

Syracuse, NY, USA GuyR.Lanza

Syracuse, NY, USA LeeNewman

Preface

guarino@unisannio.it

vii

Part I Phytoremediation Using Soil Microorganisms

1 Microbial Inoculants-Assisted Phytoremediation

forSustainable Soil Management ............................... 3

Elizabeth Temitope Alori and Oluyemisi Bolajoko Fawole

2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant

Growth-Promoting Rhizobacteria (PGPR) toEnhance

Phytoremediation ........................................... 19

Karen E. Gerhardt, Gregory J. MacNeill, Perry D. Gerwing,

and Bruce M. Greenberg

3 Successful Integrated Bioremediation System

ofHydrocarbon-Contaminated Soil ataFormer Oil Refinery

Using Autochthonous Bacteria andRhizo-Microbiota ............. 53

Valentina Spada, Pietro Iavazzo, Rosaria Sciarrillo,

and Carmine Guarino

4 Phytoremediation ofPetroleum-Contaminated Soil

inAssociation withSoil Bacteria ............................... 77

Prayad Pokethitiyook

Part II Higher Plants in Biomonitoring and Environmental

Bioremediation

5 The Use ofHigher Plants inBiomonitoring andEnvironmental

Bioremediation .............................................103

Svetlana Vladimirovna Gorelova

and Marina Vladimirovna Frontasyeva

6 Phytoremediation Applications forMetal- Contaminated

Soils Using Terrestrial Plants inVietnam ........................ 157

Bui Thi Kim Anh, Ngyuen Thi Hoang Ha, Luu Thai Danh,

Vo VanMinh, and Dang Dinh Kim

Contents

guarino@unisannio.it

viii

7 Essential Elements and Toxic Metals in Some Crops,

Medicinal Plants, and Trees ...................................183

Elena Masarovičová and Katarína Kráľová

Part III Phytoremediation of Aquatic Ecosystems

8 Phytoremediation Using Aquatic Macrophytes ...................259

Amtul Bari Tabinda Akhtar, Abdullah Yasar, Rabia Ali,

and Rabia Irfan

9 Remediation ofPharmaceutical andPersonal Care

Products (PPCPs) inConstructed Wetlands: Applicability

andNew Perspectives ........................................277

Ana Rita Ferreira, Alexandra Ribeiro, and Nazaré Couto

10 Floating Wetlands fortheImprovement ofWater Quality

andProvision ofEcosystem Services inUrban Eutrophic Lakes .....293

Eugenia J. Olguín and Gloria Sánchez-Galván

11 Green Aquaculture: Designing andDeveloping

Aquaculture Systems Integrated withPhytoremediation

Treatment Options .......................................... 307

Guy R. Lanza, Keith M. Wilda, Sushera Bunluesin,

and Thanawan Panich-Pat

Part IV Special Applications of Phytoremediation

12 Modelling Phytoremediation: Concepts, Models,

andApproaches .............................................327

Edita Baltrėnaitė, Pranas Baltrėnas, and Arvydas Lietuvninkas

13 Genetic Control ofMetal Sequestration inHyper-Accumulator

Plants .....................................................343

Shahida Shaheen, Qaisar Mahmood, Mahnoor Asif, and Raq Ahmad

14 Engineered Nanomaterials forPhytoremediation

ofMetal/Metalloid- Contaminated Soils: Implications

forPlant Physiology .........................................369

Domingo Martínez-Fernández, Martina Vítková, Zuzana Michálková,

and Michael Komárek

15 Phytoremediation Application: Plants asBiosorbent forMetal

Removal inSoil andWater .................................... 405

Rasha H. Mahmoud and Amal Hassanein Mohammed Hamza

16 Nutrient Management Strategies forCoping withClimate

Change inIrrigated Smallholder Cropping Systems

inSouthern Africa ........................................... 423

Davie M. Kadyampakeni, Isaac R. Fandika,

and Lawrent L.M. Pungulani

Contents

guarino@unisannio.it

17 Phytoremediation ofLandfill Leachates ......................... 439

Prasanna Kumarathilaka, Hasintha Wijesekara, Nanthi Bolan,

Anitha Kunhikrishnan, and Meththika Vithanage

18 Phytomining ofRare andValuable Metals .......................469

Luís A.B. Novo, Paula M.L. Castro, Paula Alvarenga,

and Eduardo Ferreira da Silva

19 Air Phytoremediation ........................................487

Stanislaw W. Gawronski and Helena Gawronska

Index .........................................................505

Contents

guarino@unisannio.it

Contributors

Raq Ahmad Department of Environmental Sciences, COMSATS Institute of

Information Technology, Abbottabad, Pakistan

Amtul Bari Tabinda Akhtar Sustainable Developmental Study Centre, GC

University, Lahore, Pakistan

Rabia Ali Sustainable Developmental Study Centre, GC University, Lahore,

Pakistan

Elizabeth Temitope Alori Department of Crop and Soil Science, Landmark

University, Omuaran, Kwara, Nigeria

PaulaAlvarenga Department of Technologies and Applied Sciences, School of

Agriculture, Polytechnic Institute of Beja, Beja, Portugal

BuiThiKimAnh Institute of Environmental Technology, Vietnam Academy of

Science and Technology, Hanoi, Vietnam

Mahnoor Asif Department of Environmental Sciences, COMSATS Institute of

Information Technology, Abbottabad, Pakistan

EditaBaltrėnaitė Vilnius Gediminas Technical University, Vilnius, Lithuania

PranasBaltrėnas Vilnius Gediminas Technical University, Vilnius, Lithuania

NanthiBolan Global Centre for Environmental Remediation, Faculty of Science

and Information Technology, University of Newcastle, Newcastle, Callaghan, NSW,

Australia

SusheraBunluesin WHO Country Ofce for Thailand, Ministry of Public Health,

Nonthaburi, Thailand

PaulaM.L.Castro Faculty of Biotechnology, Centre of Biotechnology and Fine

Chemistry, Catholic University of Portugal, Porto, Portugal

guarino@unisannio.it

xii

Nazaré Couto CENSE, Departamento de Ciências e Engenharia do Ambiente,

Faculdade de Ciências e Tecnologia, Universidade Nova de Lisboa, Lisbon, Portugal

LuuThaiDanh College of Agriculture and Applied Biology, University of Can

Tho, Can Tho, Vietnam

Isaac R. Fandika Department of Agricultural Research Services, Kasinthula

Agricultural Research Station, Chikwawa, Malawi

Oluyemisi Bolajoko Fawole Department of Agronomy, University of Ilorin,

Ilorin, Kwara, Nigeria

AnaRitaFerreira CENSE, Departamento de Ciências e Engenharia do Ambiente,

Faculdade de Ciências e Tecnologia, Universidade Nova de Lisboa, Lisbon, Portugal

Marina VladimirovnaFrontasyeva Joint Institute for Nuclear Research, Frank

Laboratory of Neutron Physics, Sector of Neutron Activation Analysis and Applied

Research, Moscow, Russia

Helena Gawronska Laboratory of Basic Research in Horticulture, Faculty of

Horticulture, Biotechnology and Landscape Architecture, Warsaw University of

Life Sciences, Warsaw, Poland

StanislawW.Gawronski Laboratory of Basic Research in Horticulture, Faculty

of Horticulture, Biotechnology and Landscape Architecture, Warsaw University of

Life Sciences, Warsaw, Poland

Karen E. Gerhardt Department of Biology, University of Waterloo, Waterloo,

ON, Canada

Perry D. Gerwing Earthmaster Environmental Strategies Inc., Calgary, AB,

Canada

Svetlana VladimirovnaGorelova Institute of Advanced Training and Professional

Retraining of Education Employees of Tula Region, Tula, Russia

BruceM. Greenberg Department of Biology, University of Waterloo, Waterloo,

ON, Canada

CarmineGuarino Department of Science and Technologies, University of Sannio,

Benevento, Italy

NgyuenThiHoangHa VNU University of Science, Vietnam National University,

Hanoi, Vietnam

Amal Hassanein Mohammed Hamza Biochemistry Department, Faculty of

Science, King Abulaziz University, Jeddah, Saudi Arabia

Biochemistry and Nutrition Department, Faculty of Women, Ain Shams University,

Cairo, Egypt

PietroIavazzo Lande S.p.A.Environment\Heritage\Archaeology, Naples, Italy

Contributors

guarino@unisannio.it

xiii

Rabia Irfan Sustainable Developmental Study Centre, GC University, Lahore,

Pakistan

Davie M. Kadyampakeni Soil and Water Sciences Department, University of

Florida, Citrus Research and Education Center, Lake Alfred, FL, USA

International Water Management Institute, Cantonment, Accra, Ghana

Dang Dinh Kim Institute of Environmental Technology, Vietnam Academy of

Science and Technology, Hanoi, Vietnam

Michael Komárek Faculty of Environmental Sciences, Department of

Environmental Geosciences, Czech University of Life Sciences Prague, Prague,

Czech Republic

KatarínaKráľová Faculty of Natural Sciences, Institute of Chemistry, Comenius

University in Bratislava, Bratislava, Slovakia

Prasanna Kumarathilaka Environmental Chemodynamics Project, National

Institute of Fundamental Studies, Sri Lanka

Anitha Kunhikrishnan Chemical Safety Division, Department of Agro-Food

Safety, National Academy of Agricultural Science, Wanju-gun, Republic of Korea

GuyR.Lanza College of Environmental Science and Forestry, State University of

New York, Syracuse, NY, USA

ArvydasLietuvninkas Vilnius Gediminas Technical University, Vilnius, Lithuania

GregoryJ. MacNeill Department of Biology, University of Waterloo, Waterloo,

ON, Canada

Qaisar Mahmood Department of Environmental Sciences, COMSATS Institute

of Information Technology, Abbottabad, Pakistan

RashaH.Mahmoud Biochemistry Department, Faculty of Science, King Abulaziz

University, Jeddah, Saudi Arabia

Biochemistry and Nutrition Department, Faculty of Women, Ain Shams University,

Cairo, Egypt

DomingoMartínez-Fernández Faculty of Environmental Sciences, Department

of Environmental Geosciences, Czech University of Life Sciences Prague, Prague,

Czech Republic

Elena Masarovicˇ ová Faculty of Natural Sciences, Department of Soil Science,

Comenius University in Bratislava, Bratislava, Slovakia

Zuzana Michálková Faculty of Environmental Sciences, Department of

Environmental Geosciences, Czech University of Life Sciences Prague, Prague,

Czech Republic

VoVan Minh University of Education, University of Da Nang, Danang, Vietnam

Contributors

guarino@unisannio.it

xiv

Luís A.B. Novo Department of Geosciences, GeoBioTec Research Center,

University of Aveiro, Aveiro, Portugal

Eugenia J. Olguín Institute of Ecology, Environmental Biotechnology Group,

Veracruz, Mexico

Thanawan Panich-Pat Faculty of Liberal Arts and Science, Department of

Science, Cluster of Environmental Science and Technology, Kasetsart University,

Nakorn Pathom, Thailand

Prayad Pokethitiyook Department of Biology, Faculty of Science, Mahidol

University, Bangkok, Thailand

LawrentL.M.Pungulani Department of Agricultural Research Services, Chitedze

Agricultural Research Station, Lilongwe, Malawi

AlexandraRibeiro CENSE, Departamento de Ciências e Engenharia do Ambiente,

Faculdade de Ciências e Tecnologia, Universidade Nova de Lisboa, Lisbon, Portugal

Gloria Sánchez-Galván Institute of Ecology, Environmental Biotechnology

Group, Veracruz, Mexico

RosariaSciarrillo Department of Science and Technologies, University of Sannio,

Benevento, Italy

ShahidaShaheen Department of Environmental Sciences, COMSATS Institute of

Information Technology, Abbottabad, Pakistan

Eduardo Ferreira da Silva Department of Geosciences, GeoBioTec Research

Center, University of Aveiro, Aveiro, Portugal

ValentinaSpada Department of Science and Technologies, University of Sannio,

Benevento, Italy

MeththikaVithanage Environmental Chemodynamics Project, National Institute

of Fundamental Studies, Kandy, Sri Lanka

Martina Vítková Faculty of Environmental Sciences, Department of

Environmental Geosciences, Czech University of Life Sciences Prague, Prague,

Czech Republic

HasinthaWijesekara Global Centre for Environmental Remediation, Faculty of

Science and Information Technology, University of Newcastle, Callaghan, NSW,

Australia

Keith M. Wilda Blue Stream Aquaculture, Island Grown Initiative, WE

Aquaculture, Edgartown, MA, USA

AbdullahYasar Sustainable Developmental Study Centre, GC University, Lahore,

Pakistan

Contributors

guarino@unisannio.it

Part I

Phytoremediation Using Soil

Microorganisms

guarino@unisannio.it

3© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_1

Chapter 1

Microbial Inoculants-Assisted

Phytoremediation forSustainable Soil

Management

ElizabethTemitopeAlori andOluyemisiBolajokoFawole

Abstract Agricultural soil pollution refers to its accumulation of heavy metals

and related compounds which could be from natural or anthropogenic sources. This

threatens food quality, food security, and environmental health. The traditional

physico-chemical technologies soil washing used for soil remediation render the

land useless as a medium for plant growth, as they remove all biological activities.

Others are labor-intensive and have high maintenance cost. Phytoremediation, sus-

tainable and cheaper in situ remediation techniques was therefore considered.

However, plants do not have the capability to degrade many soil pollutants especially

the organic pollutant. It is therefore imperative to take advantage of the degrading

ability of soil microorganisms. This chapter therefore focuses on phytoremediation

techniques augmented by microbial inoculants.

Keywords Inoculants • Microbes • Phytodegradation • Phytoremediation • Soil

pollution • Soil management • Sustainable

1.1 Introduction

Pollution of agricultural soils refers to its accumulation of heavy metals and related

compounds which could be from natural or anthropogenic sources. This threatens

food quality, food security, and environmental health [1]. Soil pollution produces

change in the diversity and abundance of biological soil populations [2]. This is

critical because of the role of soil organisms in plant establishment and survival.

Such elimination of soil organisms can lead to problems with plant establishment

and survival. Crops raised on polluted soil may contain harmful levels of pollutants

that can be passed on to the animals and human that eat them [3]. Inhaling dust

E.T. Alori, B.Agric., M.Sc., Ph.D. (*)

Department of Crop and Soil Science, Landmark University, Omuaran, Kwara, Nigeria

e-mail: aloritope@yahoo.com

O.B. Fawole, B.Sc. M.Sc., Ph.D.

Department of Agronomy, University of Ilorin, Ilorin, Kwara, Nigeria

e-mail: fawoleob@unilorin.edu.ng

guarino@unisannio.it

blown from polluted soil can be injurious to one that inhales it. More also, polluted

soil cannot be used for commercial development, parks or recreation [4]. Soil pol-

lutants alter plant physiology. It can cause cell membrane disruption, damage to

photosynthetic apparatus, and can also alter the physical and chemical properties of

the soil where plants are growing [5].

Cleaning of polluted soil may be very difcult because both soil pollutants and

soil minerals carry small electric charges that cause each to bond with each other. It

is well known that heavy metals cannot be chemically degraded and need to be

physically removed or be immobilized [6]. Traditionally, remediation of heavy

metal-contaminated soils is either on-site management or excavation, and subse-

quent disposal to a landll site [7]. However, this method of disposal merely shifts

the contamination problem elsewhere. Soil washing for removing contaminated soil

is an alternative to excavation and disposal to landll. This method is however

costly and produces a residue rich in heavy metals, which will require further treat-

ment or burial. Moreover, these physico-chemical technologies used for soil reme-

diation render the land useless as a medium for plant growth, as they remove all

biological activities. Other technologies such as vitrication, leaching, electrokinet-

ics soil vapor extraction, thermal desorption, chemical processing, etc., are labor-

intensive and have high maintenance cost [8, 9]. It is therefore imperative to develop

a sustainable on-site technique for remediation of heavy metal contaminated sites.

For better soil management, an increase in use of biological potential is impor-

tant. Phytoremediation is one of the sustainable and cheaper in situ remediation

techniques to be considered. Phytoremediation is a novel green technology that uses

specialized plants and associated soil microbes to remove, destroy, sequester, or

reduce the concentrations or toxic effects of contaminant in polluted soil and water

[4]. The plant root-colonizing microbes or the plants themselves absorb, accumu-

late, translocate, sequester, and detoxify toxic compounds to non-toxic metabolites.

Five important approaches can be considered in the use of plants to clean up pol-

luted soil. (1) Phytostabilization, a process in which pollutants are immobilized by

plant activity resulting in attenuation of the wind and soil erosion and runoff

processes into the ground water or air. (2) Hydraulic control, plants act like a pump,

draws the groundwater up through their roots to keep it from moving. This reduces

the movement of contaminated groundwater toward clean areas off-site. (3) Phyto-

volatization involves use of plants to take up certain contaminants and then converts

them into gaseous forms that vaporize into the atmosphere. (4) Phytoltration refers

torhizoltration where contaminants such as metals are precipitated within the

rhizosphere. (5) Phytoextraction (Phytoaccumulation) which involves metal hyper-

accumulating plants which can contain more than 1% of metals in harvestable

tissues [10, 11] (Fig. 1.1).

However, plants do not have the capability to degrade many soil pollutants. It is

therefore imperative to take advantage of the degrading ability of soil organisms.

Organic toxins containing carbon such as the hydrocarbons found in gasoline and

other fuels can only be broken down by microbial processes [12]. Symbiotic root

E.T. Alori and O.B. Fawole

guarino@unisannio.it

colonizing microorganism through metal sequestration increases metal tolerance in

plants. The remediation by plant using the degrading ability of soil organisms is

called phytodegradation. This helps us to understand integrated activity patterns

between plants and microbes [13]. Some soil microbes such as the arbuscular

mycorrhizal fungi (AMF) secret glycoprotein called glomalin. This can form com-

plexes with metals. Microbial organisms within the rhizoplane can take part in phy-

toremediation by protecting the plants from the toxic effect of the contaminants

while the plants in return provide the microbial processes the boost they need to

remove organic pollution from the soil more quickly. Plants excrete organic materi-

als that serve as food for microbes thus playing a key role in determining the size

and health of soil microbial population. Bioaugmentation enables an increase of

biodegradation of contaminated sites by the introduction of single strains or assem-

blages of microorganisms with the desired catalytic capabilities [14]. Microbial

assemblages are found to be efcient since each partner can accomplish different

parts of the catabolic degradation [15]. In this chapter, our focus is mainly on phy-

toremediation augmented by microbial inoculants. We begin with the contribution

of plants and microbial inoculants in phytoremediation process. Then the methods

of inoculating plants with microbial inoculants, the various mechanisms used by the

microbial inoculants to assist plant in remediation, and the limitations of microbial

inoculants-assisted phytoremediation are summarized and discussed.

PHYTOEXTRACTION

PHYTOVOLATIZATION

PHYTODEGRADATION

PHYTOSTIMULATION

Microbial

inoculants

PHYTODEGRADATION

HYDROLIC

CONTROL

PHYTOEXTRACTION

PHYTOSTABILIZATION

Fig. 1.1 Mechanisms of microbial-assisted phytoremediation

1 Microbial Inoculants-Assisted Phytoremediation forSustainable Soil Management

guarino@unisannio.it

1.2 Sources ofSoil Pollution

Soil pollutants get introduced to the soil from various sources ranging from natural

(Lithogenic) to anthropogenic activities (Fig. 1.2). Heavy metals commonly get

introduced via human activities that are related to energy and mineral consumption

[5], while petroleum hydrocarbons usually come from accidental spills of petroleum-

based products commonly used. Various industrial processes and anthropogenic

activities in urban areas induce the release of metals and metalloids (MM) (toxic

and genotoxic compounds) in natural environments.

Agricultural inputs such as chemical fertilizers, herbicides, and pesticides leaves

the soil polluted with heavy metals [16]. According to Pietrzak and Uren [17],

excessive use of fungicides and herbicides that are rich in heavy metal results in soil

pollution. Copper for instance is used as a broad-spectrum bacterial and fungicidal

agricultural pesticide and as fertilizer component because of its antimicrobial prop-

erties, but Cu is a common soil pollutant that persists in the soil providing a chronic,

long-term stress on the soil microbial community [18]. Industrial activities such as

chemical works, service stations, metal fabrication shops, paper mills, tanneries,

textile plants, waste disposal sites, and intensive agriculture equally brings about

the appearance of serious environmental problems such as soil pollution [19].

Indiscriminate waste disposal practices have led to signicant build upon a wide

range of metal(loid)s, such as arsenic (As), cadmium (Cd), chromium (Cr), copper

(Cu), mercury (Hg), lead (Pb), selenium (Se), and zinc (Zn) in soils [20]. Kierczak

etal. [21] found that soils in the areas around historic smelters are highly polluted

Sources of

Soil

Pollution

Natural

Processes

Volcanic

eruption

Mining Combustion

of fossil fuel

Military

activities

Industrial

discharge

Sewage

effluents

Air

Pollution

fall out

Agricultural

inputs e.g.

pesticides,

Fertilizers etc

Continental

dust

Weathering

processes of

earth crust

Soil erosion Urban

runoff

Anthropogenic

Source

Fig. 1.2 Sources of soil pollutants

E.T. Alori and O.B. Fawole

guarino@unisannio.it

with metal(loids)s (up to 4000 mg/kg Cu, 1500 mg/kg Zn, 300 mg/kg As, and

200mg/kg Pb). Fossil fuel combustion is another source of soil pollution reported

by Krgović etal. [22]. Vehicle emissions, industrial processes, or waste incineration

plants were revealed to introduce some pollutant such as heavy to what should

have been valuable soil [23]. Soil pollutants could originate from the mining and

smelting of metal ores [24], runoff of urban soils, fertilizer application, or efuents

discharged [25].

1.3 Contributions ofPlants andMicrobial Inoculants

inPhytoremediation

Microbial-assisted phytoextraction optimizes the synergistic effect of plants and

microorganisms and has been used for the cleaning-up of soils contaminated by

metals [2].

Plant translocates and sequesters pollutions such as heavy metals while microbes

degrade organic contaminants. Plants can store many contaminants in biomass that

can later be harvested, while microbial assemblages can also convert contaminants

such as heavy metals to stable and/or less toxic form. They can facilitate the uptake

of pollutants such as heavy metals by plant roots. Microorganisms that reside on or

within aerial plants tissue can help to stabilize and/or transform contaminants that

have been translated which may limit the extent of volatization [13]. Plant root

exudates such as enzymes, amino acids, aromatics, simple sugars, and aliphatics

stimulate the growth of root-associated microorganisms; on the other hand,

microbes can reduce the phytotoxicity of the contaminants in the soil or augments

the capacity of the plant to degrade contaminant [3]. Ability of plant root to extend

deeper into soil, allowing access to water and air and therefore changing the con-

centration of carbon dioxide, the pH, osmotic potential, redox potential, oxygen

concentration, and moisture content of the soil, could lead to an environment that

will better able to support high micro-biomass [26]. This enhanced trace element

uptake by plants can be ascribed to an increase in root absorption ability and/or

to an enhancement of trace metal bioavailability in the rhizosphere, mediated by

microorganisms.

Plants can increase biodegradation through the transfer of oxygen to the rhizo-

sphere and the release of soluble exudates that provide nutrient sources for micro-

organisms [27]. Thus, plants enhance microbial growth and hence the associated

contaminant-degradation processes. Microorganism contribution in immobilizing

elements or facilitating plant absorption plants may signicantly contribute to MM

removal through uptake in biomass [28]. Microbial assemblages improve plant

health and growth, suppress disease-causing microbes, and increase nutrient avail-

ability and assimilation [29].

1 Microbial Inoculants-Assisted Phytoremediation forSustainable Soil Management

guarino@unisannio.it

1.4 Methods ofInoculating Plants withMicrobial Inoculants

Plants to be used as phytoremediator to clean polluted soils could be inoculated with

microbial assemblages via quite a number of techniques. These methods could

include: (1) Seedinoculation, (2) Soaking plant roots with microbial suspension,

when the root of ryegrass was soaked with a suspension of an endophytic Massilia

sp. (Pn2) the same was found to have been translocated to the plant shoots [30].

(3) Painting plant leaves with microbial suspension [31–33]. Afzal etal. [34] dis-

covered the cells of BurkholderiaphytormansPsJN in the internal tissue of the

shoot and root when the plant was inoculated via leaf painting. Root colonization

strategy was found to be the optimal colonization method for circumventing the risk

of plant organic contamination [32].

1.5 Types ofSoil Pollutants

Soil pollutant could be organic or inorganic present in the hydrosoluble fraction

(complexed, adsorbed onto particles or dissolved). The most common inorganic

contaminants are heavy metals and mineral oils such as Cd, Cr, Pb, Cu, Hg, NiSe,

As, and Zn [35]. Industrial efuents release organic pollutants like hydrocarbons,

polycyclic aromatic hydrocarbons, and anionic detergent. Other soil pollutants

include plant organic materials, petroleum hydrocarbons, and organochlorines [36].

Table 1.1 reveals some examples of soil pollutants that could be removed from soil

via a microbial-assisted phytoremediation technique.

1.6 Mechanisms ofMicrobial Inoculants

inPhytoremediation ofPolluted Soil

Microbial inoculants can improve pollutant removal through various mechanisms.

Some has the potential to produce metal chelating siderophores, which could

improve metal bioavailability [37]. Moreover, they produce biosurfactants (rhamno-

lipids) that can enhance the solubility of poor water-soluble organic compounds and

the mobility of heavy metals [38]. Formation of biolm is another mechanism by

which microbial inoculants assist plants in remediation of polluted soils [39]. In

addition, these microbes can transform metals into bioavailable and soluble forms

through the action of organic acids, biomethylation, and redox processes [39].

Diverse soil microbes have the ability to secrete plant hormones such as indole-

3- acetic acid (IAA), cytokinins, gibberellins (GAs), and certain volatiles which

promote plant growth by altering root architecture [16]. The microbial plant growth

stimulatory actions result from the manipulation of the complex and balanced net-

work of plant hormones that directly are responsible for growth and root formation.

For example, IAA produced by soil microbes has been demonstrated to enhance

E.T. Alori and O.B. Fawole

guarino@unisannio.it

Table 1.1 Some examples of soil pollutants that could be removed from soil via microbial- assisted

phytoremediation technique

Plant Microorganism Pollutants References

Helianthus annus Micrococcus

sp. MU1 and

Klebsiella sp.

BAM1

Cd Prapagdee etal. [50]

Polygonum

pubescens

Enterobacter

sp. JYX7 and

Klebsiella sp.

JYX10

Cd Jing etal. [51]

Zea mays L Azotobactor

chroococum

and Rhizobium

leguminosarum

Pb Hadi and Bano [52]

Solanum melongena Pseudomonas sp.NaCl Fu etal. [53]

Vigna unguiculata Scutelospore

reticulate,

Glomus phaseous

Al, Mn Alori and Fawole [2]

Solanum nigrum Pseudomonas

sp. LK9

Cd Chen etal. [54]

Brassica napus Pantoea

agglomerans

Jp3–3, and

Pseudomonas

thivervalensis

Y1–3-9

Cu Zhang etal. [55]

Brassica juncea Paenibacillus

macerans

NBRFT5, Bacillus

endophyticus

NBRFT4, B.

pumilus NBRFT9

Cu Tiwari etal. [56]

Loliummultiorum

Lam

Staphylococcus

sp. strain BJ06

Pyrene Sun etal. [57]

Brassica oxyrrhina Pseudomonas

sp. SRI2,

Psychrobacter

sp. SRS8 and

Bacillus sp. SN9

Ni Ma etal. [58]

Brassica napus Acinetobacter

sp. Q2BJ2 and

Bacillus sp.

Q2BG1

Pb Zhang etal. [55]

Cytisus striatus Rhodococcus

erythropolis

ET54b

Sphingomonas

sp. D4

Hexachlorocyclohexane

(HCH)-

Becerra- Castro

etal. [59]

Cichorium intybus Rhizophagus

irregularis

Diesel Driai etal. [60]

Medicago sativa Pseudomonas

aeruginosa

(Cu, Pb and Zn and

petroleum hydrocarbons)

Agnello etal. [35]

(continued)

guarino@unisannio.it

Table 1.1 (continued)

Plant Microorganism Pollutants References

Orychophragmus

violaceus

Bacillus subtilis,

B. cereus, B.

megaterium, and

Pseudomonas

aeruginosa

Cd Liang etal. [61]

Cytisusstriatus

(Hill) Rothm

Rhodococcus

erythropolis E T

54b and

Sphingomonas

sp. D4

Becerra- Castro

etal. [62]

Arabidopsis

thaliana

Achromobacter

xylosoxidans

Phenolic Ho etal. [63]

Solanum

lycopersicum

Penicillium

janthinellum LK5

Al Khan etal. [64]

Brassica napus Rahnella sp. JN6 Cd He etal. [65]

Triticum aestivum Pseudomonas

putida KT2440

Cd, Hg, Ag Yong etal. [66]

Brassica juncea Bacillus subtilis

SJ-101

Ni Zaidi etal. [67]

Sedum

plumbizincicola

Bacillus pumilus

E2S2 and

Bacillus sp. E1S2

Cd Ma etal. [68]

Brassica napus Pseudomonas

uorescens G10

and Microbacterium

sp. G16

Pb Sheng etal. [69]

Trifolium repens Arbuscular

mycorrhizal fungi

and Bacillus

cereus

Heavy metals Azcón etal. [70]

Iris pseudacorus Arbuscular

mycorrhiza fungi

Pb, Fe, Zn, and Cd Wężowicz etal. [71]

Brassica juncea Rhizobium

leguminozarum

Zn Adediran etal. [72]

Rahnella sp. Amaranthus

hypochondriacus,

A.Mangostanus

and S. nigrum

Cd Yuan etal. [73]

Brassica juncea Staphylococcus

arlettae

NBRIEAG-6

As Srivastava etal. [74]

Orycoprhagmus

violaceus

Bacilus subtilis,

B. cereus,

Flavobacterium

sp. and

Pseudomonas

aeroginosa

(Zhang etal. [55])

Zn He etal. [75]

Lupinus luteus Burkholderia

cepacia VM1468

Ni and trichloroethylene

(TCE)

Weyens etal. [76]

Alnus rma Bacillus

thuringiensis

GDB-1

As Babu etal. [77]

guarino@unisannio.it

root proliferation [40]. In addition, soil microbes possess growth- promoting traits,

including phosphorus solubilization, nitrogen xation, iron sequestration, and phy-

tohormone, which improve plant growth and increase plant biomass [39].

In addition to degrading soil pollutants microbial assemblages, also partake in

phytoremediation by producing hormones, xing atmospheric nitrogen, or solubi-

lizing P [41]. One of the most important mechanisms by which microbial assem-

blages respond to stress condition such as from soil pollutant is by increasing

ethylene levels that result to an increase in cell and plant damage [42]. Many

microbes that augment phytoremediation destroy a precursor of the ethylene (1- am

inocyclopropane- 1-carboxylate (ACC)) that by producing the enzyme ACC deami-

nase, that in turn facilitates plant growth and development by decreasing plant

ethylene levels [39]. Figure 1.3 depicts strategies of phytoremediation through

microbial assemblages.

1.7 Challenges ofMicrobial Inoculants-Assisted

Phytoremediation

The success of microbial inoculation-assisted phytoremediation encounters some

set back due to the following reasons: (1) The number of degrading microbes

available regarding the pollutant to be degraded may be low or non-detectable, (2).

Production of

organic acids Formation of

biofilm

Biomethy

lation Produce metal

chelating

siderophores

Produce

biosurfactants

(rhamnolipids)

Secretion of

plant growth

hormones

Production of the

enzyme ACC

deaminase

Organic

acids

Microbial

assemblages

Redox

processes

Fig. 1.3 Strategies of phytoremediation through microbial assemblages

1 Microbial Inoculants-Assisted Phytoremediation forSustainable Soil Management

guarino@unisannio.it

The physical and chemical properties of pollutants. The various types of soil

pollutants vary in their mobility, solubility, degradability, and bioavailability. These

properties play very important role in the removal of the pollutants from the soil.

Pollutant or mixtures of pollutants sometimes require several metabolic pathways

operates simultaneously with sometimes metabolic intermediates whose toxicity

toward indigenous microbes may be high, and (3) Some polluted areas requiring

long microbial adaptation period of time justifying soil bioaugmentation [14, 43].

Other abiotic factors that also affect the success of microbial inoculation-assisted

phytoremediation include; temperature, aeration, soil pH, cation exchange capacity

(CEC), soil organic matter content, sorptive capacity of soil, and redox potential.

According to Diels and Lookman [44], microbial inoculation-assisted phytoreme-

diation is inuenced by temperature in the range 5–30°C.It therefore means that

the success of microbial inoculation-assisted phytoremediation will depend largely

on season as this will be ineffective during winter in temperate countries. Grundmann

etal. [45] reported that the efciency of microbial inoculation-assisted phytoreme-

diation depends on pH in the range 5–8. Many metal cations like Cd, Cu, Hg, Pb,

and Zn are reported to be more soluble and available in the soil solution at low pH

(below 5.5) [46]. However, Phytoremediation of atrazine by two microbial consortia

was seriously affected by pH and soil organic matter content. At pH6.1 only one

consortium degraded atrazine but at pH >7 atrazine was effectively degraded by the

consortia, the microbial inoculants were ineffective at pH5.7 because of their inter-

action with organic matter [47]. pH for the degradation of phenol and TCE was

observed to vary from 6.7 to 10 depending on whether the microbial inoculant cells

are free or immobilized [48]. As revealed by Bhargava etal. [46] higher CEC of soil

permits greater sorption and immobilization of the metals. Depending on contami-

nant characteristics, different microbial-assisted phytoremediation mechanisms

require different nal electron acceptors. For example because of the highly reduced

state of petroleum hydrocarbons, the preferred and most thermodynamically rele-

vant terminal electron acceptor for microbial process is O2 while the degradation of

chlorinated solvents, depending on the degree of halogenation, is different from that

of petroleum hydrocarbons and other oxidized chemicals, and the preferred redox

condition is anaerobiosis [44].

1.8 Characteristics toConsider intheChoice ofaPlant

forMicrobial-Assisted Phytoremediation

A key aspect in biological remediation methods is the selection of appropriate

plant–bacteria partnerships for the remediation of polluted soils [3]. Some of plant

properties to be considered include: exceptional contaminant tolerance, ability to

quickly grow on degraded land, and rapid biomass production. For instance alfalfa

(Medicago sativa L.) that is often used in phytoremediation of contaminated soil is

a fast growing species. Another critical characteristic to be considered is the

E.T. Alori and O.B. Fawole

guarino@unisannio.it

composition of plant-recruited microbial communities. Plants that develop extensive

tap root system favor the establishment of rhizosphere microorganisms. Plants ideal

for phytoremediation should possess the ability to grow outside their area of collec-

tion, to produce high biomass, easy harvesting and accumulation of a range of heavy

metals in their harvestable parts [49]. Poplar and willow possess deep root systems,

produce great biomass, can be grown in a wide range of climatic conditions and

these explain why they are effective phytoremediator of polluted soil [46].

1.9 Conclusions

Soil pollutant could be organic or inorganic present in the hydrosoluble fraction

adsorbed onto particles or dissolved. Microbial-assisted phytoremediation remove,

destroy, sequester, or reduce the concentrations or toxic effects of contaminant in

polluted soils. Production of siderophores, biosurfactants, formation of biolms,

organic acids production, biomethylation, and redox processes and plant growth

hormones stimulation are mechanisms employed by microbial inoculants in phy-

toremediation. The number of available degrading microbes and the physical and

chemical properties of pollutants determine the success of microbial inoculants-

assisted phytoremediation. Exceptional contaminant tolerance, ability to quickly

grow on degraded land, ability to grow outside their area of collection, and rapid

biomass production are important plant characteristics to be considered in the

choice of plant for phytoremediation.

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1 Microbial Inoculants-Assisted Phytoremediation forSustainable Soil Management

guarino@unisannio.it

19© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_2

Chapter 2

Phytoremediation ofSalt-Impacted Soils

andUse ofPlant Growth-Promoting

Rhizobacteria (PGPR) toEnhance

Phytoremediation

KarenE.Gerhardt, GregoryJ.MacNeill, PerryD.Gerwing,

andBruceM.Greenberg

Abstract Soil salinization negatively impacts plant growth and soil structure,

which leads to environmental stress and agricultural/economic losses. Improved

plant growth during salt-induced ionic and osmotic plant stress is the key to success-

ful phytoremediation of salt-impacted sites. Using plant growth-promoting rhizo-

bacteria (PGPR) in PGPR-Enhanced Phytoremediation Systems (PEPS), positive

effects of PGPR on plant biomass and health have been observed in greenhouse and

eld experiments. Revegetation is arguably the most important aspect of salt phy-

toremediation and substantial biomass increases occur in PGPR-treated plants in

both sodic and saline soils. PGPR protect against inhibition of photosynthesis and

plant membrane damage, which suggests that they confer tolerance to plants under

salt stress. Using PEPS, decreases in soil salinity are observed due to uptake of

sodium and chloride from the soil into foliar plant tissue. Although rates of uptake

do not change due to PGPR inoculation, higher plant biomass due to PGPR enhance-

ment of plant performance leads to greater salt uptake on a per area basis relative to

that of untreated plants. Signicant improvements in plant growth and commensurate

K.E. Gerhardt, B.Sc., M.Sc., Ph.D. • B.M. Greenberg, B.Sc., Ph.D. (*)

Department of Biology, University of Waterloo, 200 University Ave. W,

Waterloo, ON, Canada, N2L 3G1

e-mail: kegerhar@uwaterloo.ca; greenber@uwaterloo.ca

G.J. MacNeill, M.Sc.

Department of Biology, University of Waterloo,

200 University Ave. W, Waterloo, ON, Canada, N2L 3G1

Department of Molecular and Cellular Biology, University of Guelph, 50 Stone Rd. E.,

Guelph, N1G 2W1 ON, Canada

e-mail: gmacneil@uwaterloo.ca

P.D. Gerwing, B.SA., M.Sc.

Earthmaster Environmental Strategies Inc.,

358, 58th Avenue SW, Suite 200, Calgary, AB, Canada, T2H 2M5

e-mail: perry.gerwing@earthmaster.ab.ca

guarino@unisannio.it

sodium chloride uptake, and the results of mass balance studies used to assess the

direct impact of ion uptake on actual observed changes in soil salinity, provide evi-

dence that phytoremediation of salt-impacted soil is feasible within acceptable time

frames using PEPS.

Keywords Field trials • NaCl • PGPR-Enhanced Phytoremediation System(s)

(PEPS) • Polyamines • Reactive oxygen species (ROS) • Revegetation • Salt

remediation

Abbreviations

ABA abscisic acid.

ACC 1-aminocyclopropane-1-carboxylic acid.

ACCD ACC deaminase.

Ca2+ calcium ion.

Chl a chlorophyll a.

Cl− chloride ion.

CT composite tailings.

ECe electrical conductivity of a soil-saturated paste extract.

F0 minimal uorescence.

Fm maximal uorescence.

Fm′ maximal uorescence in light-adapted tissue.

Fs steady-state uorescence.

Fv/Fm maximum quantum yield.

IAA indole acetic acid.

K+ potassium ion.

Mg2+ magnesium ion.

NaCl sodium chloride.

PAH polycyclic aromatic hydrocarbon.

PAM pulse amplitude modulated.

PEPS PGPR-Enhanced Phytoremediation Systems.

PGPR plant growth-promoting rhizobacteria.

PHC petroleum hydrocarbon(s).

PSII photosystem II.

qN non-photochemical quenching of uorescence.

qP photochemical quenching of uorescence.

ROS reactive oxygen species.

SAR sodium adsorption ratio.

SOS salt overly sensitive.

K.E. Gerhardt et al.

guarino@unisannio.it

2.1 Introduction

2.1.1 Overview ofPhytoremediation

Phytoremediation is a strategy whereby plants are used to extract, immobilize, con-

tain and/or degrade soil contaminants. Although the term “phytoremediation” was not

coined until the 1980s, the strategy has been employed for removing soil contami-

nants for at least 300 years [1, 2]. Rapid expansion occurred in this eld in the 1990s,

and phytoremediation has now become a useful strategy for on site and/or in situ

removal of many contaminants, including petroleum hydrocarbons (PHC), metals,

radionucleotides, munitions waste (e.g., trinitrotoluene) and salt [1, 3–5]. Microbe-

assisted phytoremediation, especially when used in conjunction with contaminant-

tolerant plant species and high-level agronomic practices, can be a particularly

effective green strategy for remediation and revegetation of impacted soils [6–13].

Plants have extensive rooting systems that can explore large volumes of soil to

allow for effective remediation of various contaminants within different soil types.

Typically, four types of phytoremediation processes for impacted soils are discussed

in the literature [1, 4, 8]. During phytoremediation, contaminants can be broken

down in the soil (e.g., rhizodegradation of PHC, also referred to as rhizoremedia-

tion) or taken up by the roots and stored in plant tissue (typically in the foliage, as

in phytoextraction of metals and salt). Some small molecules can be taken up by the

roots, and the unmodied or modied forms are then transported via the transpira-

tion stream to leaves, where they are released to the atmosphere via transpiration

(e.g., phytovolatilization of trichloroethylene). Various contaminants can be bound

within the rhizosphere (area immediately surrounding plant roots), making them

less bioavailable (phytostabilization), and therefore less harmful to biota. In addi-

tion to these four main processes, plant roots can also alter soil chemistry via pH

changes, which can further aid phytoremediation (e.g., breakdown of calcium car-

bonate in sodic soils provides calcium ions that can replace sodium ions at binding

sites in the soil and allows for leaching or uptake of sodium) [14].

2.1.2 Prevalence andSources ofSalt-Impacted Soils

Soil salts can occur naturally (e.g., weathering of geologic formations, encroach-

ment of seawater) or they can be released into the environment as a result of anthro-

pogenic activities (e.g., irrigation, upstream oil and gas exploration/production,

application of road salts) [15, 16]. Various environmental impacts associated with

excess salt in soil include degradation of chemical and physical properties of the

soil, diminished groundwater quality, and impaired plant growth. This results in

substantial global agricultural and economic losses, sustenance issues for subsis-

tence farmers, and ecosystem imbalances [17].

2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant Growth-Promoting…

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A signicant buildup of salt often occurs in soils due to crop irrigation, and

this has been suggested as “the rst man-made environmental problem” [18, 19].

Irrigation waters tend to have high concentrations of calcium, magnesium, and

sodium ions [16]. Use of this brackish water, particularly without adequate drainage

management, results in the accumulation of salts in the rooting zone of plants due

to evapotranspiration [17]. Calcium and magnesium tend to precipitate into carbon-

ates, leaving sodium as the most prevalent ion in the soil, and this negatively impacts

both plant growth and soil structure. Soil salinization affects 20% of irrigated land

worldwide, which equates to an area approximately the size of France (62 mil-

lionha) [20, 21]. The resulting annual crop value losses have been estimated to be

$27 billion (US) [17, 21]. Salinization, which occurs in virtually all geographic

regions, has been a problem for millenia and continues to be a global concern of

paramount importance: Soil salinity due to irrigation is thought to be a contributing

factor to the downfall of the Sumerian civilization more than 4000years ago, and

irrigated land continues to be degraded by salt at a rate of 2000ha/day [16, 22, 23].

Elevated salt levels in soils are as much of a problem for the upstream oil and gas

industry as petroleum-impacted soils [24–26]. Most petroleum was formed from the

remains of marine life that existed in ancient shallow seas. Consequently, oil deposits

often occur in reservoirs that contain water with dissolved salts (brine), and the brine

(which usually contains sulfates, bicarbonates, and chlorides of sodium, calcium, and

magnesium) is frequently co-extracted with the oil [24, 27]. Any leakage into, or on,

soils around the oil well will result in not only petroleum impacts, but also salt

impacts. Furthermore, salt may be used during oil extraction. For example, sodium is

often introduced during the extraction of bitumen from oil sands ore, and then winds

up in the tailings. This is a major concern in the Athabasca oil sands region of Alberta,

Canada where large volumes of uid ne tailings are produced and stored in tailings

ponds [25]. It was estimated that by the end of current upstream heavy oil operations,

more than one billion cubic meters of ne tailings will be stored in these ponds. To

reduce the stored volume, the composite tailings (CT) process is used, which involves

the addition of gypsum or alum as a coagulant [28]. During this process, water con-

taining high levels of salt is released from the CT, and this saline CT water makes

reclamation of the CT deposit areas difcult. Efforts to revegetate the CT are hin-

dered if salt from the CT water accumulates in the rooting zone.

Application of road salts (sodium chloride [NaCl], calcium chloride, potassium

chloride and magnesium chloride), particularly in large urban areas, also leads to

elevated soil salt levels in ecosystems adjacent to roads, snow removal dump sites,

and some salt storage facilities [29, 30]. An average of 5 × 106 tonnes of road salts

(primarily NaCl) are applied annually to Canadian roadways [29, 30]. This nega-

tively impacts physical and chemical properties of surrounding soils, which conse-

quently impacts associated biota. This problem was deemed critically important in

the Canadian Environmental Protection Act, 1999, which categorized road salts as

toxicants [30], and implemented new guidelines for their use (Code of Practice for

Environmental Management of Road Salts) [29].

K.E. Gerhardt et al.

guarino@unisannio.it

2.1.3 Soil Salt Chemistry

Based on a system developed by the US Salinity Laboratory [31], salt-impacted

soils can be broadly classied as either saline (high concentration of soluble salts),

sodic (high concentration of sodium), or saline-sodic (high concentrations of both

soluble salts and sodium). More recently, the USDA Natural Resources Conservation

Service classied salt-impacted soils into seven types that incorporate soil charac-

teristics that are observable in the eld as well as chemical analyses [32]. In this

chapter, the US Salinity Laboratory classications will be used when discussing soil

salt impacts. Because NaCl is the most prevalent salt contaminant in the environ-

ment [33], the term “salt” refers to NaCl in subsequent sections of this chapter,

unless specied otherwise.

One of the most common ways to measure total soluble soil salt concentration is

electrical conductivity of a saturated soil-water paste extract (ECe, measured as

dS/m) [31]. Soil sodicity can be calculated using the sodium adsorption ratio (SAR).

It is based on the ratio of sodium ions (Na+) to calcium ions (Ca2+) and magnesium

ions (Mg2+) in the soil, and takes into account the difference in adsorption strengths

of the ions to clay particles:

SAR Na

Ca Mg

=









+





++22

where the ionic concentrations are expressed in milliequivalents per liter in soil

extract solution in equilibrium [31, 32]. Saline soils have an ECe> 4 dS/m and

SAR<13in their saturation extract. Sodic soils have an ECe<4dS/m and SAR>13.

Saline-sodic soils have an ECe>4dS/m and SAR>13.

Saline soils tend to have white crusts formed from crystallized salts that have

precipitated at the soil surface. Sodic soils tend to have poor physical structure, low

permeability (i.e., restricted movement of water and air through the soil), and high

pH (7.8–8.5), all of which are detrimental to plant growth [32]. Poor structure is, in

part, because Na+ displaces Ca2+ and Mg2+, which are important for holding clay

lattices/particles together [34, 35]. In weakly aggregated soils, dispersion of soil

particles can ll soil pores and impermeable surface crusts can form after repeated

wet/dry cycles, which inhibit root penetration and growth. High pH limits the avail-

ability of some key plant nutrients and micronutrients such as phosphates, cobalt,

copper, iron, manganese, and zinc, which are all more bioavailable at pH<7. For

more detailed descriptions of characteristics and chemistry of sodic and saline-sodic

soils, please refer to Qadir etal. [36].

2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant Growth-Promoting…

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2.1.4 Responses ofPlants toSalt Impacts

Plants are perhaps the most vulnerable sector of the biosphere to salt. Rapidly mani-

festing drought stress symptoms can occur in salt-impacted soils despite the pres-

ence of adequate water, because the resulting increase in osmotic pressure diminishes

water uptake by plants [37, 38]. Over time, uptake of salt ions can lead to toxicity in

plant tissues (particularly accumulation of Na+; as well, chloride ions [Cl−] can

reach toxic levels in some sensitive species), and the presence of excess ions in the

soil can interfere with nutrient availability (e.g., high concentrations of Na+ in the

rhizosphere interfere with K+ uptake, due to the similar chemical nature of the ions,

and this leads to K+ deciency and growth inhibition in plants) [33, 37–40]. Salt

stress negatively impacts germination, plant growth, and reproduction by affecting

physiological processes such as photosynthesis, respiration, transpiration, and

enzyme function; membrane properties are affected, upregulation of the stress-

responsive hormones abscisic acid (ABA) and ethylene occurs, and reactive oxygen

species (ROS) are generated [41–45].

2.1.4.1 Uptake andTransport ofNa+, K+, andCl−

During phytoextraction of salt, ions are taken up from the soil into plant tissues. Na+

and Cl− are taken up by plants primarily through passive symplastic pathways

driven by concentration gradients and transpiration uxes [46, 47]. Ions are trans-

ported from the root cells to the leaves via the transpirational stream of the xylem

(Fig. 2.1) [33]. These ions are typically stored in the leaves, and little ion ow

occurs via the phloem down to the roots (Fig. 2.1) [38, 47, 48]. Ion homeostasis,

involving primarily Na+, K+, Ca2+ and Cl− is extremely complex, both at the cellular

and whole plant level [38, 40, 48, 49]. Various ion channels and pumps in plant

cells, many of them tissue-specic, regulate the ow of ions from the soil into roots,

translocation from roots to foliar tissue, and storage within the cells or excretion

from them [40, 50–54] (Fig. 2.2a). Phytoextracted salt can be removed from a site

by harvesting the foliar tissues with accumulated salt ions.

2.1.4.2 Salt Stress andROS Damage

Salt stress (both osmotic and ionic) frequently results in an increase in ROS, includ-

ing hydrogen peroxide, superoxide anion radicals, hydroxyl radicals, and singlet

oxygen [55–57]. Formation of ROS occurs primarily in chloroplasts; however, it also

occurs in mitochondria and peroxisomes [55, 56, 58]. During salt stress, cytosolic

polyamines are exported to the apoplast, where they are oxidized to ROS [59].

Excessive formation of ROS leads to oxidative damage of many cellular components,

including proteins, DNA, and lipids (e.g., membrane lipid peroxidation), ultimately

leading to growth inhibition or capitulation of plants [15, 60].

K.E. Gerhardt et al.

guarino@unisannio.it

2.1.4.3 Salt Stress andAcclimation Signaling Pathways

Although excessive salt-induced generation of ROS can impair metabolic processes,

leading to oxidative stress and cellular damage, ROS can also signal responses to

mitigate salt stress damage [56, 61–63]. Increases in antioxidant enzyme activities

have been correlated with salt tolerance [57]. For example, the antioxidant enzymes

catalase, superoxide dismutase, glutathione reductase, and glutathione peroxidase

are activated in response to salinity stress in the European olive [64]. The ROS sig-

naling pathways that result in acclimation to salt stress are integrated with numerous

other signaling pathways related to salt tolerance. These include calcium, hormone

and protein phosphorylation pathways, as well as complex interactions with poly-

amine pathways [56, 63].

Accumulation of polyamines is a key factor in achieving plant tolerance to salt

stress [63]. Polyamines such as spermine, spermidine, and putrescine play a signi-

cant dual role in ROS homeostasis by acting both as ROS scavengers, and as sub-

strates for amine oxidases in the apoplast that catalyze formation of ROS involved

in stress response signaling [59, 63, 65]. Some of the ways by which polyamines

inuence ion transport during salt stress via complex signaling pathways are shown

in Fig. 2.2b. For example, polyamines exported from the cytosol to the apoplast can

block non-specic cation channels in the plasma membrane to limit Na+ inux and

Fig. 2.1 Phytoextraction of salt. Uptake of ions from the soil to root epidermal cells occurs rst.

Ions are translocated via the root symplast to the xylem. Na+, Cl−, and other ions extracted from the

soil are transported through the xylem to leaf tissue, and are stored in vacuoles. There is minimal

ow of Na+ and Cl− back down the phloem

2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant Growth-Promoting…

guarino@unisannio.it

Fig. 2.2 Ion pumps, channels, and signaling in plant salt stress and adaptation. (a) Numerous ion

pumps and channels involved in salt stress and tolerance are shown. Not all of them are found in

all species, or in all cells, and the specics of ion conductance depend on a variety of conditions.

ABA abscisic acid, ACC 1-aminocyclopropane-1-carboxylic acid, ANN annexin-formed channel,

CAX cation/H+ exchanger, CCC cation-chloride-cotransporter, DA-NSCC depolarization-activated

non-selective cation channel (NSCC), DAO diamine oxidase, FV fast vacuolar channel, H2O2

hydrogen peroxide, HACC hyperpolarization-activated Ca2+ inux channel, HKT1 low-afnity Na+

histidine kinase transporter, HKT2 Na+/K+ histidine kinase symporter, KIRC K+ inward-rectifying,

guarino@unisannio.it

K+ efux, and cytoplasmic polyamines can inhibit cation channels in the tonoplast

to limit Na+ efux to the cytoplasm, thereby helping to maintain proper cellular K+/

Na+ ratios: this may be crucial for achieving salt stress tolerance [49, 63].

One well-researched signaling pathway for Na+ exclusion from cells was discov-

ered using the salt overly sensitive (SOS) line in Arabidopsis [16, 48, 66, 67] (see

Fig. 2.2). Following exposure to salt, an unidentied salt sensor in the root plasma

membrane perceives the stress and a Ca2+ spike is generated in the cytoplasm. This

activates a signal transduction cascade involving the SOS proteins: SOS3, a calcium

binding protein, activates SOS2, a kinase that phosphorylates the plasma membrane

antiporter, SOS1. Cytoplasmic Na+ is then transported out of root cells, either from

the cytosol to the apoplast (epidermal cells), or from the cytosol to the xylem (paren-

chyma cells) [68]. In leaves, a similar pathway exists, but SOS3 is replaced by

SCaBP8) [67]. SOS3 and SOS2 have been shown to play regulatory roles in salt

tolerance [69].

Salt stress can result in increased levels of the ethylene precursor 1- aminocyclop

ropane- 1-carboxylic acid (ACC), resulting in stress ethylene production [39, 70].

This response is mediated by ABA, and ultimately leads to leaf abscission, ridding

the plant of tissue that contains toxic levels of Na+ [56]. Upregulation of ABA also

promotes stomatal closure to avoid water loss during osmotic stress, but may cause

a shortage of CO2 for carbon xation, which leads to a decline in photosynthesis

[44, 55–57].

Signaling pathways involved in salt stress and subsequent acclimation are very

complex (some of the signaling pathways are illustrated in Fig. 2.2b). The linear SOS

pathway is the best understood, but it is not the only signaling pathway for adaptation

to salt stress [67, 69]. There is good evidence that crosstalk between SOS and ABA

signaling pathways occurs, and a complex signaling network with crosstalk between

polyamine, ROS, and ABA pathways has also been shown [63, 67]. Plant growth-

promoting rhizobacteria (PGPR) have been shown to positively inuence many

of these pathways and processes to mitigate salinity stress (see Sect. 2.1.5.3).

Fig. 2.2 (continued) KORC K+ outward rectifying channel, ROSIC non-selective voltage-independent

conductance, NHX Na+/H+ antiporter, NORC, •OH hydroxyl radical, PA polyamine, PAO poly-

amine oxidase, PEROX peroxiporin, Rboh respiratory burst oxidase homolog (an NADPH oxi-

dase), ROS-NSCC ROS activated non-selective cation channel, SOS1 Na+/H+ antiporter, SOS2

protein kinase, SOS3 Ca2+ sensor, SV slow vacuolar channel, V-ATPase vacuolar H+-ATPase,

VI-NSCC voltage- independent NSCC, VK K+-selective channel, VP1 vacuolar H+ pyrophospha-

tase. Hyperpolarization of the plasma membrane activates KIRC (more inux of K+ than N+). At

the onset of salinity stress, KORC is activated by membrane depolarization, allowing the inux of

Na+ and efux of K+. Details for other ion channels and pumps can be found in Sects. 2.1.4.1 and

2.1.4.3, and references therein. (b) Some of the signaling pathways involved in salt stress and

adaptation are outlined, with emphasis on PA/ROS-related pathways. Dotted lines with arrows

indicate some relevant sources of the ionic and molecular pools, solid lines with arrows indicate

positive regulatory actions, and lines with bars indicate negative regulatory actions. Salt stress

leads to a Ca2+ burst that activates the SOS pathway. Increased Ca2+ is perceived by SOS3, which

interacts with the kinase SOS2. This complex phosphorylates the SOS1 antiporter at the plasma

membrane, which leads to diminished accumulation of Na+ in the cytosol. It also leads to increased

activity of NHX at the tonoplast, which results in sequestration of excess Na+ in vacuoles. Further

details for signaling pathways can be found in Sects. 2.1.4.2 and 2.1.4.3, and references therein

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Detailed descriptions of the numerous interconnected signaling pathways, and the

salt ion channels, pumps and molecules involved in toxicity and tolerance are beyond

the scope of this review. For further details, please see Blumwald [71], Gao etal. [72],

Kronzucker and Britto [49], Kumar etal. [73], Kurusu etal. [74], Miller etal. [56],

Pottosin and Shabala [65], Saha et al. [63], Uozumi and Schroeder [75], and

Zhu [33].

2.1.4.4 Physiology ofSalt Tolerance inHalophytes andGlycophytes

There are many different parameters that have been used to dene halophytes in the

literature [15, 40]. Generally, they can be dened as plants that grow well in salt-

impacted soils. Plants that are not halophytes are frequently classied as glyco-

phytes. A more realistic view is not a division into two broad categories of plants,

but rather a continuum of salt tolerance ranging from extremely tolerant to extremely

sensitive plants [76]. Many major agricultural crops are sensitive to salt stress [57].

Salt tolerance can be assessed in terms of survival (more meaningful for perennials

than for annuals) and/or biomass production [77].

Halophytes can be obligate (absolute requirement for elevated salt habitats), fac-

ultative (can grow in salt-impacted soils, but optimum growth and health occurs in

soils with low or no salt) or habitat-indifferent (can ourish in soil with or without

salt) [15]. Glycophytes have varying sensitivities to salt, ranging from tolerant to

completely intolerant. Depending on the circumstances, a given plant might be

described as a facultative halophyte or a salt-tolerant glycophyte.

Halophytes have evolved different mechanisms that allow them to survive and

thrive in salt-impacted soils: salt exclusion (minimizing uptake), salt accumulation,

and salt excretion [78]. Some of these strategies are also employed by salt-tolerant

glycophytes. Salt exclusion mechanisms are varied and complex; however, the main

contributing factors are low permeability of root epidermal cell membranes, low net

uptake of Na+ by root cortex cells, and tight control of xylem loading via the peri-

cycle [15, 40, 51]. In salt accumulators, Na+ is taken up, transferred to leaf tissue

and sequestered in vacuoles to minimize damage to cytoplasmic components

(Figs. 2.1 and 2.2). Small organic osmolytes (compatible solutes), such as proline,

betaine, and mannitol, accumulate in the cytoplasm to maintain osmotic balance

within the cells, and some act as osmoprotectants to scavenge/quench ROS and

prevent damage to membrane structure, enzymes, and proteins [33, 38, 55]. Salt

excretion, prevalent in halophytes, is accomplished using leaf epidermal salt glands and

hairs that remove salt from mesophyll cells via secretion at the leaf surface [14, 79].

2.1.5 Remediation andPhytoremediation ofSalt

Remediation of salt-impacted soils has proven difcult and costly due to the absence

of a versatile in situ technology [24]. Often the impacted soil must be removed to

landll and replaced with clean soil. In addition to the physical challenges

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encountered with ex situ remediation (soil excavation and soil replacement), these

methods are also costly and unsightly. Various in situ treatments have also been

employed to remediate salt-impacted soils. Three widely used methods are leach-

ing, chemical amendments followed by leaching, and organic amendments [14, 24,

36]. For leaching, excess water is applied to the soil to move soluble salts from the

surface soil to lower horizons. This can lower ECe values in surface and rooting

zone soils, but not SAR, and is therefore effective primarily for saline soils [14].

Leaching results in valuable water resources being wasted, and diminishes soil sta-

bility and quality [14]. For sodic soils, numerous chemicals can be applied to the

soil to promote ion exchange, often replacing Na+ with Ca2+ at the cation binding

sites on clay particles. The amendments can be very costly, however, and leaching

is required afterward to remove the Na+ to lower soil horizons [14]. Organic amend-

ments can be used to increase dissolution of soil calcite and improve soil structure,

however these amendments are also costly and dissolution is a slow process [14].

Several mechanisms are involved in salt remediation when using plants (phytore-

mediation). Uptake of salt ions into plant tissue results in a decrease in soil ECe and,

possibly, SAR.Lowering soil pH via root exudates can increase the dissolution of

soil calcium carbonate (calcite), thereby providing Ca2+ to displace the adsorbed

Na+ in the soil. Displaced Na+ leads to improved soil structure with the resultant

uptake and removal of Na+ from soil by plants. Root growth and the associated

organic matter additions to the rhizosphere within impacted areas will increase

hydraulic conductivity of the soil, which increases the potential for natural leaching

of salt from upper to lower soil horizons [14, 36, 80, 81].

2.1.5.1 Advantages ofPhytoremediation ofSalt

Clearly, technologies are needed that can remediate salt-impacted soils in an envi-

ronmentally responsible and cost-effective way. Phytoremediation has numerous

advantages over conventional techniques for salt remediation. Some of the advan-

tages are greater environmental stewardship (e.g., soil is treated and reused, not

hauled to landll for disposal), ease of application, and lower cost. Using plants,

co-contaminants such as salt, PHC, and metals can be remediated simultaneously [82].

As an added benet, some crops that are grown for phytoremediation can be sold for

bioenergy sources, cellulose production, or livestock feed [14, 81].

2.1.5.2 Choosing Plants forPhytoremediation ofSalt

Numerous plant species have been shown to effectively decrease ECe and SAR in

salt-impacted soils [14, 81, 83, 84]. Plants chosen for phytoremediation must be

sufciently salt-tolerant to survive and grow in impacted soils. Some of the most

salt-tolerant halophytes are very slow growing, and consequently these plants do not

attain sufcient biomass to achieve phytoremediation in an acceptable time frame.

Many halophytes excrete salt ions through specialized leaf glands, and others drop

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older organs that have accumulated toxic levels of Na+ and other ions: neither of

these tolerance mechanisms leads to phytoremediation because the salt essentially

returns to the soil near the plant. Similarly, plants that exclude Na+ and Cl− from

roots are not efcient remediators of NaCl because the salt remains in the soil,

although it might be more easily leached due to plant-related improvements in soil

hydraulic conductivity.

A salt tolerance mechanism that is desirable for phytoremediation is uptake and

storage of Na+ and Cl− into above-ground tissues. Maintaining low concentrations

of cytoplasmic Na+ is a key factor in salt tolerance. As noted in Sect. 2.1.4.4, many

halophytes and salt-tolerant glycophytes sequester Na+ to leaf vacuoles to achieve

this [85, 86]. This prevents damage to cytosolic enzymes, and also counteracts the

low extracellular osmotic potential resulting from salt stress [33]. Ideally, if soils are

highly sodic, the chosen plant will have high Na+ uptake, but lower uptake of Ca2+

and Mg2+, which will lower SAR values [14]. Also, for effective phytoremediation,

the plants chosen should be suitable for repeated harvesting of the foliar tissues

containing phytoextracted salt.

2.1.5.3 PGPR-Enhanced Phytoremediation

One criterion that is essential for successful phytoremediation is substantial plant

biomass production. Unfortunately, as discussed in Sect. 1.4, plants growing in salt-

impacted soils are prone to the combined detrimental effects of water stress, ion

toxicity, and nutritional deciencies, which result in substandard plant growth.

Traditional plant breeding programs and genetic engineering have been employed

in attempts to improve salt tolerance in plants; however, the suite of genes and mul-

tiple pathways involved in salt tolerance, as well as the time involved to successfully

breed or engineer salt tolerant plants, make this a daunting task [55, 87, 88].

One strategy that has been utilized to overcome the challenges of abiotic stress-

ors is to employ plant growth-promoting rhizobacteria (PGPR); these soil microbes

can promote growth and health in plants during stress conditions [8, 10, 89–92].

PGPR accelerate plant growth under stress conditions by increasing plant tolerance

to elevated salt, PHC and/or trace metal levels, as well as other environmental

stressors such as saturated soil or drought conditions. This leads to rapid growth of

plants, including their roots. The vigorous plant growth that ensues leads to greater

proliferation of naturally existing microbes in the soil, resulting in a very active

rhizosphere that is typical of soils with normal plant growth. The substantial root

biomass that accumulates in the soil provides a sink which allows for rapid parti-

tioning of salt ions out of the soil, and their subsequent accumulation in the foliar

tissues of some plants.

PGPR have been shown to confer salt tolerance in a variety of plants, by amelio-

rating both the osmotic and ion toxicity effects of salt stress [6, 58, 89, 93–95].

Some PGPR confer salt tolerance via tissue-specic regulation of HKT1, a plasma

membrane Na+ uniporter [95] (see Fig. 2.2a). When plant growth inhibition is the

result of stress ethylene production, PGPR with ACC deaminase (ACCD) can be

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employed [70]. ACCD metabolizes ACC, a precursor of ethylene in the biosynthetic

pathway, thereby limiting the amount of stress ethylene that can be produced [6,

90]. Polyamines produced by PGPR have also been shown to lower stress ethylene

levels and mitigate osmotic stress [96, 97]. PGPR have been shown to promote

synthesis of antioxidants (including polyamines), and indole acetic acid (an auxin)

which can promote root growth [6, 58, 89]. Recently, PGPR were shown to regulate

a ROS-triggered caspase-like activity in rice; there was a concomitant decrease in

programmed cell death, a phenomenon previously linked to caspase-like activity

and salt-induced oxidative stress [58, 98]. Other mechanisms linking PGPR to salt

tolerance in plants include altered mineral uptake, which results in a benecial

increase in the cellular K+/Na+ ratio; and elevated production of quorum sensing

molecules, which can lead to alterations in the rhizosphere [92, 99–101].

2.1.5.4 Successful Remediation ofSalt-Impacted Soils

Numerous studies have been conducted to assess phytoremediation of salt-impacted

soils. For example, beet and millet were grown for 70days in the greenhouse, in saline

calcareous soil from Southern Ghor in Jordan [81]. Substantial amounts of Na+, K+,

and Cl− were taken up into above-ground tissues, decreases in EC1:1 of 54–69%

occurred, and better soil hydraulic conductivity was observed. Purslane has been shown

to remove considerable amounts of NaCl from saline soils, and was recommended as

an intercrop for salt removal in salt-sensitive fruit orchards based on pot experiments

[102]. Hue etal. [82] used material dredged from Pearl Harbor, Hawaii. This material

was amended with a soil from Oahu, Hawaii that was high in calcium and magnesium,

to achieve a nal ECe of ~18 dS/m. After growing a combination of two salt-tolerant

grasses and a legume for 3 months in a greenhouse, soil ECe decreased by ~50%. This

was attributed primarily to Na+ uptake by the legume and one of the grasses. Atriplex

halimus plants were grown for 90days in pot experiments using saline and saline-sodic

soils from the Ninavah province of Iraq [103]. Decreases in EC were observed in both

saline and saline-sodic soils (21 and 32%, respectively). Decreases in SAR were also

observed for both saline and saline-sodic soils (29 and 50%, respectively).

Field experiments were performed in the Khorezm Region of Uzbekistan [104],

where Chenopodium album and Apocynum lancifolium were grown in soil with ECe

values of ~10.5 and 13, respectively (top 15cm). Uptake of Cl− and Na+ in C. album

(105 and 34mg/g dry weight, respectively) was substantially higher than that of

A. lancifolium (49 and 12mg/g dry weight, respectively). Despite a root depth that

was less than half that of A. lancifolium, C. album had much greater salt uptake from

the soil (570kg/ha, compared to 130kg/ha for A. lancifolium). Thus, C. album was

deemed a good choice for remediation of salt-impacted soils, and was recommended

for integration into crop rotation programs, whereas A. lancifolium was deemed a

salt-tolerant species not suitable for salt remediation because, relative to C. album,

the rate of salt removal was deemed too low. For more results of phytoremediation

experiments in the greenhouse and eld, please see the following reviews of the

literature [8, 14, 81, 85].

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2.1.5.5 Obstacles Affecting Phytoremediation ofSalt-Impacted Soils

intheField

Efforts to translate phytoremediation research from successful laboratory and

greenhouse experiments to the eld have proven challenging [9]. Although there

have been many successful trials, there have also been numerous inconclusive and

unsuccessful attempts at phytoremediation of salt in the eld. A few general prob-

lems have emerged: different experimental conditions between the laboratory and

the eld, difculty in accurately assessing salt remediation, and length of time

required for salt remediation.

Numerous biotic and abiotic plant stress factors not present in laboratory and

greenhouse studies can result in signicant problems in eld applications of phytore-

mediation. These include, but are not limited to, variations in temperature, nutrients,

and precipitation; herbivory (insects and/or animals); plant pathogens; and competi-

tion by weed species that are native to the area [105]. Further, in the greenhouse, soils

are generally homogeneous; in the eld, contaminant concentrations vary across any

given site, resulting in “hot spots”. Factors such as root structure, soil structure,

organic composition of the soil, soil pH, moisture content, and microbial activity also

exhibit spatial variability at a given site, and can change over time [105, 106].

For salt, conventional means of assessing phytoremediation (e.g., decrease in

ECe in soil over time) may not be adequate to show that salt impacts are actually

decreasing, although in many cases active remediation may be occurring. Salt read-

ily migrates from lower soil horizons into the rooting zone of plants (i.e., the area

where phytoremediation takes place) due to evaporation and transpiration [107].

Thus, it can be difcult to assess remediation exclusively by measuring soil salt

levels in upper horizons. Assessing ion uptake into plant tissues and calculating

estimates of total salt uptake at a given site can provide an estimate of actual salt

removal and remediation over time.

Another challenge to phytoremediation of salt in the eld is the length of time

required to fully remediate the impacted soils [36]. Although this cannot be consid-

ered a failure of the technology, it is a disadvantage compared with traditional

methods such as excavation and soil removal. It has also been suggested that salt

remediation rates decrease over time, because in terms of mass balance, fewer salt

ions are removed from the soil when salt gets diluted in leaching water (natural or

applied) [14]. If salt uptake is the predominant removal mechanism, this should not

be a factor. In fact, the reverse should be true: as soil quality improves with each

successive growing season, plant root and shoot biomass should increase, providing

a greater sink for salt ions.

2.1.5.6 Revegetation asaMeasure ofSuccessful Phytoremediation ofSalt

As noted in Sect. 2.1.5.5, salt readily migrates from lower horizons to upper hori-

zons in the soil, and moves with water ow in general. When plants grow in soil,

this upward migration of water and salt is enhanced. Thus, it can be problematic to

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accurately assess remediation based on soil salt levels. Unlike heavy metals, which

can be highly toxic to humans and other animals at levels found in soils, NaCl is

generally not considered hazardous. Therefore, in the case of salt, the essential goal

of phytoremediation is to overcome plant salt stress. Ideally, plants that grow rap-

idly with high rates of salt uptake and accumulation (e.g., kallar grass and oats) can

be used to achieve both revegetation and salt removal from the soil [80]; however,

generic regulatory criteria that depend solely on diminishing soil ECe levels and

SAR in impacted soils may be too stringent and unnecessary in some cases. Because

soil salts (including NaCl) are generally not hazardous to humans and other ani-

mals, and plants are the most sensitive part of the biosphere, we propose that, rather

than achieving mandated levels of ECe and SAR, sustained revegetation of an

impacted site should be the goal of salt phytoremediation. In this case, achieving

75% sustainable plant productivity compared to reference sites should qualify as

successful phytoremediation of salt. This is in accordance with the reclamation

objectives of some Canadian and American regulatory bodies that seek to ensure a

self-sustaining ecosystem devoid of long-term toxicity, and to establish equivalent

land capability that existed prior to industrial activities [26, 108].

If revegetation is the goal, selection of plant species is important. Many halo-

phytes that accumulate large quantities of salt on a per mass basis grow too slowly

to provide sufcient biomass for revegetation within an acceptable timeframe. Also,

they may not be native to the site being remediated. Plants that exclude salt by limit-

ing uptake into the root, or plants that excrete the salt from aerial tissues, cannot

effectively remove salt from the soil. However, if the goal is only revegetation, these

species could be considered for use. In general, for revegetation, the goal should be

to achieve aggressive plant growth with species native to the impacted site.

2.2 PGPR-Enhanced Phytoremediation Systems (PEPS)

To ll the need for a versatile, green, in situ technology for remediation of cont-

aminated soils, PGPR-Enhanced Phytoremediation Systems (PEPS) have been

developed [9, 10, 83, 84, 91, 109–114]. To achieve successful PGPR-enhanced

phytoremediation, a skill set beyond being able to plant seeds is required. A funda-

mental understanding of soil science, contaminant chemistry, plant biology, soil

microbiology, agriculture, forestry, and regulatory guidelines is crucial for applica-

tion of this green technology. The key to successful remediation within an accept-

able time frame is to achieve vigorous plant growth because large amounts of

biomass are necessary for phytoremediation; however, this is generally difcult due

to suboptimal soil conditions (e.g., low organic content and poor soil structure) at

impacted sites. Phytoremediation is therefore facilitated by preparing high-quality

seed beds and utilizing other agronomic practices. After plant growth is established,

contaminant chemistry, including degradation and/or uptake of the contaminants

from soil, must be monitored. The standard PEPS protocol includes inoculation of

seeds with PGPR to accelerate plant growth under stress conditions (see Sect. 1.5.3),

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soil pre-treatment (tilling soils to achieve homogeneity, as well as fertilizing and

adding other required amendments to the rooting zone of plants), and adequate

monitoring of the site (including contaminant assessments).

Mixtures of grass species, including cereals, are most commonly used in

PEPS.Most Poaceae species (grass family) are facultative halophytes (also described

in the literature as salt-tolerant glycophytes): they can grow on salt- impacted soils,

but plant growth and health will be negatively impacted relative to growth in soils

with low or normal salt levels [15]. Specic grass species have been used because

they have been shown to be salt-tolerant, they accumulate salt in foliar tissue which

can be removed easily from impacted sites, and they produce substantial amounts of

root biomass. Using more than one plant species (co-cropping) can enhance overall

microbe-assisted phytoremediation because the unique characteristics and properties

of each plant species may support different microbial communities in the rhizo-

sphere, differentially penetrate the soil matrix, and have different temperature and

moisture optima, which increase the overall odds of success in the eld; co-cropping

has also been shown to limit weed proliferation and herbivory [115].

The PGPR used in PEPS are non-pathogenic, non-genetically modied soil bac-

teria (usually pseudomonads) that are present in the soils under remediation [9, 112].

These strains are naturally occurring, and express ACCD. They also synthesize

indoleacetic acid (IAA), which promotes root cell growth of host plants [116]. They

are sensitive to common antibiotics, do not grow at 37°C (i.e., they cannot prolifer-

ate in the human body), and are all classied as Biosafety Level 1 (the safest pos-

sible designation). They are ubiquitous in nature, common to soils around the world,

and pose no threat to humans, wildlife, or the environment. With PEPS, the PGPR

are used only via a seed treatment, whereby the seeds are treated in a controlled

environment. The plant roots of the treated seeds are thus inoculated with PGPR as

they pass through the seed coat during germination. Notably, PGPR, including those

used in PEPS, increase the number of root hairs in grass seedlings under stress, rela-

tive to plants without PGPR ([100], Greenberg etal. unpublished data). Root hairs

contribute substantially to the surface area of roots, and most of the ion uptake

(including Na+) occurs across the plasma membrane of the root hair epidermal cells

[51]. Thus, PGPR-treated PEPS plants have a greater capacity for Na+ uptake from

the soil than untreated plants, which generally corresponds to decreases in both soil

salinity and sodicity.

A mixture of PGPR can be used if the right combination of mixed microbial

strains can be found (e.g., [10, 58]). The rationale is that taxonomically different

PGPR have different optimum pH, temperature, and moisture requirements for col-

onizing rhizospheres/roots; and different PGPR may have different modes of action

for promoting plant growth that could be additive or synergistic in a microbial mix,

further increasing the odds of successful phytoremediation in the eld. Sometimes,

however, it is preferable to use a single strain of PGPR to avoid antagonistic effects.

Greenhouse experiments have been performed to ascertain whether or not different

PGPR strains should be used independently (e.g., CMH3) or in combination

(e.g., UW3+UW4) [83, 117].

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When possible, PEPS plant species are chosen that are native to the area in which

phytoremediation is being undertaken. This eliminates the ecological risk associ-

ated with introducing a non-native species to an ecosystem and facilitates native

habitat reconstruction/reclamation following remediation. Native PGPR are also

used, whenever possible. For instance, PGPR that have been isolated from the site

being remediated can be used. This provides multiple benets: PGPR isolated from

salt-contaminated soils are salt-tolerant, acclimated to the soil conditions in that

area, and may be more competitive in situ than non-native bacteria [118].

2.2.1 Development, Proof, andFull-Scale Application ofPEPS

In the initial stages of PEPS development, remediation of PHC (including large

recalcitrant polycyclic aromatic hydrocarbons [PAHs]), heavy metals (lead, copper,

and cadmium) and a pesticide (DDT) were the focus of the research [91, 113, 114,

119, 120]. The original process involved proven agronomic techniques, and plant

growth with PGPR [91, 113, 114], with both laboratory and small-scale eld trials

[9, 112]. A variety of monocot and dicot species were used in the initial plant growth

and phytoremediation experiments. Seeds were treated with various naturally occur-

ring, non-pathogenic Pseudomonas strains, both individually and in microbial

mixes. Although phytoremediation was observed in the absence of PGPR treat-

ments, enhanced remediation rates were observed with PGPR seed treatments.

A summary of the development, proof, and full-scale application of PEPS for PHC

remediation was published recently [10].

2.2.2 Adapting PEPS forSalt Remediation

As discussed in Sect. 2.1.2, soil salinization is as much of an environmental issue as

soil contaminated with compounds such as PHC (including PAHs) and metals [24–26].

For this reason, PEPS research was expanded to include phytoremediation of salt.

Laboratory, greenhouse, and eld experiments were conducted, resulting in the

adaptation of PEPS for salt remediation [83, 84, 109–111, 117, 121, 122].

2.2.2.1 Lab/Greenhouse Experiments

The effects of salt stress on plant growth, photosynthesis, and membrane integrity

were assessed in a series of greenhouse and laboratory experiments [109, 117, 121,

122]. Soils with a range of salinity (ECe) and sodicity (SAR) values were obtained

from sites in Saskatchewan, Canada. Three strains of PGPR, Pseudomonas sp.

UW3 (GenBank Accession Number KF145175), Pseudomonas sp. UW4 (GenBank

Accession Number CP003880), and Pseudomonas corrugata CMH3 (GenBank

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Accession Number KF041156), were used for seed treatment prior to phytoreme-

diation to promote plant growth and increase tolerance to salt [83, 123, 124]. All

these strains are naturally occurring, produce IAA and express ACCD, the enzyme

that consumes the precursor to ethylene, a plant stress hormone. UW3 and UW4

were isolated from unimpacted Ontario soils during a previous research project.

CMH3 was isolated from the rhizosphere of grasses grown on a highly saline soil

(ECe value of 20–50dS/m) at an upstream petroleum site in Saskatchewan. Details

of PGPR isolation, analysis, and identication; ACCD and IAA assays; bacterial

inoculation of seeds; greenhouse trials; uorescence assays; and electrolyte leakage

assays can be found in Chang etal. [83] and Greenberg etal. [109].

Effects ofSalinity andPGPR onPlant Growth

Previously, a decrease in biomass for wheatgrass grown without PGPR for 90 days

in saline soils (ECe= 30dS/m) was reported [109]. Biomass decreases for barley

and oats grown for 45 days in saline soils (ECe=9dS/m) were also reported previ-

ously [83]. Data for oats grown on saline (ECe=14dS/m) and sodic (SAR=24)

soils are provided here as other examples. A decrease in oat biomass due to salt

stress was observed in the absence of PGPR seed treatments. Oats without PGPR

had 40% lower shoot biomass (Fig. 2.3) and 50% lower root biomass [117] than

control plants grown on unimpacted soils (ProMix™). PGPR (UW3+UW4) com-

pletely alleviated the root and shoot growth inhibition caused by salinity.

In fact, the shoot biomass of plants treated with PGPR exceeded that of the con-

trols. UW3+UW4 improved the fresh weight of oat shoots (Fig. 2.3) and roots

[117] by ~100%, relative to untreated (−PGPR) plants after 20 days in sodic soil

(ECe, 3.2dS/m; SAR, 24) and 45 days in saline soil (ECe, 14 dS/m; SAR, 11). Under

the more saline conditions in wheatgrass experiments, PGPR (UW3 + UW4,

CMH3) ameliorated salt stress, but did not bring biomass levels back to those of

control plants grown under non-saline conditions [109]. Notably, the growth promo-

tion effect was much greater using a mix of UW3 and UW4 than using either UW3

or UW4 independently (Fig. 2.3a). When the kinetics of oat growth is examined,

with and without PGPR, it can be seen that PGPR protected the seedlings, espe-

cially during emergence and early growth phases (Fig. 2.4). This allows the plants

to become established in impacted soils. It has been suggested that protection of

young leaves is crucial for salt tolerance, due to the dearth of vacuoles available for

Na+ sequestration in these leaves, and the detrimental effects of Na+ on protein syn-

thesis and other processes crucial to plant growth [32, 79].

The studies described in this section, and those described in Sect. 1.5.3, show

that PGPR can improve plant growth on salt-impacted soils. This indicates that

crops and other plants not considered salt-tolerant can grow on saline soils with

PGPR inoculation. This also supports the concept that equivalent land use can be

achieved with PEPS and that phytoremediation of salt-impacted land via revegeta-

tion is feasible.

K.E. Gerhardt et al.

guarino@unisannio.it

Fig. 2.3 Effects of PGPR treatment on plant growth in salt-impacted soil. (a) A representative

photograph of oat growth after 20 days in sodic eld soil from a site in Saskatchewan, Canada with

low ECe (3.2dS/m) and high SAR [24]. “Control” shows the baseline normal plant growth in

ProMix™ (ECe< 2dS/m) growth medium, “-PGPR” shows plants grown in saline soil without

PGPR treatment, “UW3” shows plants that were grown from oat seeds treated with UW3 PGPR,

“UW4” were treated with UW4 PGPR and “UW3+4” were treated with a mix of UW3 and UW4

PGPR. (b) A representative photograph of oat growth after 45 days in saline eld soil with moder-

ate ECe (14dS/m) and SAR [11]. (c) Fresh weight (g) of oat shoot biomass after 45 days growth in

moderately saline eld soil (ECe=14dS/m, SAR=11). The results are expressed as means ± SEM

of four independent replicates (n = 4). Data were analyzed by one-way analysis of variance

(ANOVA) and Dunnett’s test. * indicates statistical differences (P< 0.01) in biomass relative to

untreated oats grown in saline soil

2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant Growth-Promoting…

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Alleviation ofSalt Inhibition ofPhotosynthesis inPGPR-Treated Plants

Numerous abiotic environmental stresses, including salinization, result in deleteri-

ous effects on photosynthesis in plants [16, 57, 91, 109, 125–127]. Inhibition of

photosynthesis is a good measure of the physiological state of the plant; therefore,

measurement of various photosynthetic parameters can be used as an indication of

the extent to which plants are salt-stressed. Indeed, negative impacts on plant growth

due to salt stress are often associated with a decrease in photosynthetic rate, possi-

bly the result of a decrease in stomatal conductance and the ensuing decrease in CO2

uptake [44, 128–131]. Osmotic stress, which occurs rapidly following plant expo-

sure to salt, results in a decrease in chloroplast volume and an increase in Na+

concentration in the cytosol and chloroplasts. This can lead to inhibition of the

photosynthetic electron transport chain [57, 125].

Chlorophyll a (Chl a) uorescence is a useful technique for assessing photosyn-

thetic activity [125, 132]. Chl a uorescence parameters obtained using pulse

amplitude-modulated (PAM) uorometry (e.g., Fv/Fm, yield, qP, and qN) can be

used to assess the efciency of photochemistry in plants and to study the effect of

salinity on photosynthetic electron transport [133]. Fm (maximal uorescence of

dark-adapted tissue) and F0 (minimal uorescence [background uorescence]) can

be used to calculate Fv/Fm ([Fm−F0]/Fm) which indicates the maximum quantum

yield of photosystem II (PSII) [133]. Optimal Fv/Fm values range from 0.79 to 0.83

for most plant species [134, 135]. Lower values indicate damage to the photosyn-

thetic apparatus, and resultant plant stress. Yield of steady-state photosynthesis

[(Fm′−Fs)/Fm′] can be calculated from the maximal uorescence in light-adapted

tissue (Fm′) and steady-state uorescence (Fs). Yield is a measurement of continu-

ous photosynthesis (i.e., the amount of light absorbed by PSII chlorophyll that gets

used in photochemical reactions) [136]. The parameter qP ([Fm′−Fs]/[Fm′−F0′]) is a

measure of photochemical quenching, which indicates the proportion of open (or

functional) PSII reaction centers [137–140]. Non-photochemical quenching of uo-

rescence, qN (1−[Fm′−F0]/[Fm−F0]), is related to the dissipation of energy as heat

and indicates the extent of photoinhibition [133, 139, 141].

10 20

Time (days)

UW4+UW3

Shoot fresh weight (g)

No PGPR

Fig. 2.4 Effects of PGPR

on kinetics of plant growth

in salt-impacted soil. Shoot

biomass (FW) of oats ±

PGPR in sodic eld soil

(ECe, 3.2dS/m; SAR, 24)

at 10, 20, and 30days

K.E. Gerhardt et al.

guarino@unisannio.it

An example of salt effects on photosynthesis is given in Table 2.1. Oats, with and

without PGPR treatment, planted on salt-impacted soil (ECe=30dS/m) and control

soil (ProMix, ECe<2dS/m), were grown for 20 days [122]. Various photosynthetic

parameters were measured using a PAM uorometer using methods published pre-

viously [109, 142]. For plants without PGPR, all Chl a uorescence parameters

(Fv/Fm, yield, qP and qN) showed signicant negative impacts due to growth on

saline soil (note: qN rises under stress conditions, while the other parameters fall

during stress) (Table 2.1). This suggests that the photosynthetic apparatus was dam-

aged and photosynthesis was impaired [136, 143, 144]. These data are in agreement

with our previously published results, and with those of numerous other researchers

using oats and other plant species [61, 91, 109, 145, 146].

The negative impacts of salinity on overall photosynthesis were largely allevi-

ated by PGPR treatment of plants (both UW3 + UW4 and CMH3): most Chl a

uorescence parameters (Fv/Fm, yield and qP) of the PGPR-treated plants had val-

ues that were similar to plants grown in control soil (Table 2.1). These results are

consistent with the improvements in plant growth on salt-impacted soils that were

observed in greenhouse experiments (Figs. 2.3 and 2.4). Similar relationships

between growth promotion and photosynthetic capacity were reported in Brassica

[130] and lettuce [44].

Oats are considered to have low tolerance to salt [147]. Despite the sensitivity of this

species, treatment with PGPR alleviated photosynthetic stress. Thus, PGPR seed treat-

ment can result in salt-sensitive species becoming more tolerant, thereby making them

candidates for phytoremediation of salt and/or revegetation of salt- impacted soil.

Effects ofSalinity andPGPR onCell Membrane Integrity

Plant cell membranes play an important role in the maintenance of the micro-

environment and metabolism of plant cells, and are often the rst targets of abiotic

plant stressors [41]. ROS-mediated membrane damage is a major cause of the

Table 2.1 Effects of salinity and PGPR on chlorophyll a uorescence of oats

Chlorophyll a

uorescence

parameters Control (ProMix™)

No PGPR

(salt soil)

UW3+UW4

(salt soil) CMH3 (salt soil)

Fv/Fm0.806±0.002*** 0.752±0.013 0.801±0.002*** 0.803±0.002***

Yield 0.690±0.005*** 0.488±0.030 0.711±0.020*** 0.706±0.022***

qP 0.910±0.004*** 0.729±0.030 0.865±0.010*** 0.877±0.006***

qN 0.301±0.016** 0.429±0.042 0.358±0.032 0.317±0.019*

PAM measurements were obtained after 20 days growth on ProMix™ (ECe< 2 dS/m) or salt-

impacted soil (ECe=30dS/m)±PGPR (UW3+UW4 or CMH3)

Fv/Fm (maximal PSII activity), Yield (steady-state PSII activity), qP (photochemical quenching;

indicates net energy storage), qN (non-photochemical quenching; indicates energy loss)

Results are expressed as means ± SEM of 12 independent replicates (n=12). Data were analyzed

by one-way analysis of variance (ANOVA) and post-hoc Dunnett’s tests. * (P<0.05), ** (P<0.01)

and *** (P<0.001) indicate signicant differences between values for “No PGPR (salt soil)”

relative to the other treatments

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cellular toxicity induced by salt stress in a variety of plants (see Sect. 2.1.4.2).

Salt- induced ROS lead to damage to plant cell membranes and increase their perme-

ability, allowing electrolytes that are contained within the membrane to leak into

surrounding tissues [148]. Therefore, maintaining cell membrane stability and

integrity is important for salt tolerance. The degree of damage to cell membranes

can be estimated by measuring electrolyte leakage from cells, by comparing the

electrical conductivity of the leaked contents (into water) from salt-stressed plant

tissues to that of control plant tissues [109, 148, 149].

As an example, data on electrolyte leakage in oat leaves is shown in Fig. 2.5. The

objective of the experiment was to assess cell membrane integrity following salt

stress, and to determine whether PGPR could ameliorate the damage. Oat seeds,

with and without PGPR treatment, were planted on moderately and highly impacted

saline soils (ECe = 12 and 18 dS/m, respectively) and control soil (ProMix™,

ECe < 2 dS/m) [122]. Shoots were removed from plants for electrolyte leakage

analysis after 12 days of growth. Electrolyte leakage was measured as electrical

conductivity (dS/m) of solutions containing ions that escaped from oat cells, pre-

sumably via damaged plasma membranes [109] (Fig. 2.5). The higher the EC

(dS/m) value of the receiving water, the greater the damage to plant membranes.

Fig. 2.5 Effects of salinity and PGPR on membrane damage in oats. Electrolyte leakage assays

were performed on excised oat leaves from plants grown for 12 days on ProMix™ (ECe<2dS/m),

moderately impacted saline soil (medium salt, ECe=12dS/m), or highly impacted saline soil (high

salt, ECe=18dS/m)±PGPR (UW3+UW4). Results are expressed as means ± SEM of six inde-

pendent replicates (n=6). Data were analyzed by one-way analysis of variance (ANOVA) and

post-hoc Bonferroni tests. * (P<0.05) indicates a signicant difference between values for “No

PGPR” relative to PGPR-treated plants

K.E. Gerhardt et al.

guarino@unisannio.it

Electrolyte leakage from plant tissues increased as soil salinity increased

(Fig. 2.5), indicating that plant membrane damage increased with salinity level. The

amount of electrolyte leakage was greatly diminished in PGPR-treated oats, indicat-

ing less damage to plasma membranes (Fig. 2.5). Similarly, Kang etal. [42] observed

a protective effect of PGPR following induction of high salt stress: leaves of PGPR-

treated cucumber had 21% less electrolyte leakage than control plants. The electro-

lyte leakage results indicate membrane damage due to salt stress, and are consistent

with the photosynthesis results in the previous section (Alleviation of Salt Inhibition

of Photosynthesis in PGPR-treated Plants). For instance, the lower yield and higher

qN values in the absence of PGPR inoculation indicate loss of thylakoid membrane

integrity relative to that in PGPR-treated plants.

2.2.2.2 Field Trials

The effects of salinity and PGPR on plant growth and salt uptake in eld experiments

were reported previously [83, 84, 110, 111, 117, 122]. Field trials were performed at

upstream oil and gas sites with poor quality soils of varying soil salinities (ECe,

2–40dS/m) and sodicities (SAR, 1–45) in Manitoba, Saskatchewan, Alberta, and the

Northwest Territories, Canada. Three strains of PGPR (UW3+UW4, CMH3) were

used to treat various grass species (tall fescue, tall wheatgrass, ryegrass, barley, oats)

prior to phytoremediation to promote plant growth and increase tolerance to salt.

Field trials were conducted over a period of two or three consecutive growing sea-

sons. Details of bacterial inoculation of seeds, eld trials, and analyses for Na+ and

Cl− in plant tissues can be found in previously published work [83, 84].

Effects ofSalinity andPGPR onPlant Growth

Results for the effects of salinity and PGPR on plant growth in the eld have been pub-

lished previously [83, 84]. Effects on plant growth were similar to those observed in the

greenhouse: increases in soil salinity led to decreased plant biomass production in the

absence of PGPR treatment. Treatment of seeds with PGPR alleviated the plant stress

such that root and shoot biomass and ground cover were comparable to control plants

grown in non-saline/sodic soils. In general, on saline soils, shoot biomass increases of

100–200% were observed in PGPR-treated plants, relative to untreated plants.

Uptake ofNaCl fromSoil

The NaCl concentrations in above-ground tissue of barley and oats from nine

upstream oil and gas sites in Saskatchewan were measured following a single grow-

ing season [83, 84]. On a per mass basis, above-ground plant NaCl concentrations

ranged from 22 to 97g/kg (DW). Generally, on a per mass basis, about 2–3 times

more Cl− than Na+ was stored in above-ground plant tissues. Notably, NaCl accu-

2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant Growth-Promoting…

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mulation in plant foliage was accompanied by decreases in soil salinity

(10–20%) at the eld sites. Results from the eld were in agreement with results

previously obtained in greenhouse experiments [83, 109]. Salt removal (kg/ha) by

barley and oats was comparable to literature values for various glycophytes, and

Na+ uptake was comparable to that of millet, another grass species [81].

Data from the Saskatchewan sites where PEPS was applied were compiled to

obtain average values for typical salt remediation using this remedial strategy [84].

Standard PEPS experimental protocols were similar to those detailed in Chang

etal. [83]. NaCl uptake into foliage averaged 29g/kg (DW), with three times more

Cl− than Na+ accumulation by weight in the foliage. An average of 150kg/ha of

NaCl was removed from the sites per harvest of above-ground biomass. A concomi-

tant average annual decrease in ECe of 15% was observed when the salt was only in

the top 30cm of soil (i.e., the rooting zone).

PGPR treatment did not result in increased NaCl uptake on a plant biomass

basis [83]. That is, the concentrations of salt in the foliage were similar with and with-

out PGPR treatment. However, the increases in plant biomass due to PGPR treatment

were substantial (generally 100–200%). This is in agreement with our results from

greenhouse experiments [109] and the ndings of other researchers. For instance,

Jesus etal. [14] indicated that a search of the literature showed biomass increases as a

result of PGPR inoculation, but there was not a reference that showed an increase in

salt phytoextraction on a biomass basis in any plant as a result of PGPR treatment.

Similarly, treatment of a perennial rhizome grass with a plant growth- promoting

mycorrhizal fungus did not increase uptake of Na+ and Cl− from salt-impacted soils

[150]. Because of the increases in biomass due to PGPR or fungal treatments, the net

effect was a much higher rate of NaCl removal from the soil for inoculated plants than

that of untreated plants [14, 150]. These results are in contrast with those of Ozawa

et al. [86]. They found that inoculation of a glasswort (a halophyte from the

Chenopodiaceae family that sequesters Na+ in valcuoles) with Pseudomonas pseudo-

alcaligenes did not increase fresh or dry weight of the glasswort shoots, but did

increase Na+ accumulation relative to uninoculated plants. This difference may be due

to dissimilar plant growth conditions in general, the plant species used (a succulent

marine halophyte), or the PGPR (an endophytic nitrogen- xing bacteria).

Little research has been done to determine the connection of ion uptake by plants

to actual observed changes in soil salinity in full-scale phytoremediation trials of

salt-impacted soils. This was investigated when PEPS was employed on a salt-

impacted (ECe = 5.97 dS/m) upstream petroleum site in Saskatchewan, Canada

[121]. Data from this eld trial were used to conduct mass balance studies, to deter-

mine the efcacy of PEPS on saline soils. Plant tissue collected over two successive

growing seasons was assayed for ionic content and these data were compared to

measured changes in soil salinity (ECe) for each eld season. Based on the amount

of ve predominant ions (Ca2+, Mg2+, K+, Na+, and Cl−) in the plant tissue samples,

removal of these ions from soil was measured, and the expected change in soil EC

was calculated. These values were used to determine how much of the observed

change in soil salinity could be attributed to ion uptake by PEPS plants during a

given eld season.

K.E. Gerhardt et al.

guarino@unisannio.it

Soil ECe decreased by 0.96 and 0.45dS/m in the rst and second year, respectively.

The mass of salt ion uptake into plant biomass and total annual biomass were com-

pared to the measured changes in soil salinity over the two eld seasons. Taking into

account the effect of each salt ion on the ionic strength of the soil solution, uptake

of soil salt ions into foliar plant tissue accounted for 60.5 and 76.8% of the change

in salinity in the rst and second year, respectively. Notably, only ve salt ions were

included in the mass balance calculations; therefore, the change in soil salinity that

was attributed directly to phytoremediation using PEPS was likely underestimated.

This research provided evidence that, for PEPS eld trials, the uptake of ions from

the soil into plant biomass plays a predominant role in soil salinity decreases, and is

not the result of water ux through the soil and movement of ions into deeper soil

horizons.

In general, phytoremediation research in the greenhouse and the eld has shown

that salt concentrations in the foliage tend to be fairly similar on a per mass basis,

independent of PGPR or fungal treatment, plant type, soil ECe, and SAR.However,

the increases in plant growth due to PGPR or fungal treatment tend to be large

(average shoot biomass increase of 150%), particularly in poor soils and those with

moderate to severe salt impacts. The extra biomass due to PGPR treatment will trans-

late to greater salt removal from the soil. We suggest that the key to salt phytoreme-

diation is to maximize growth with PGPR treatment or other means. Greater plant

biomass should result in higher rates of salt remediation each growing season when

PEPS are employed.

2.2.3 Feasibility ofSalt Phytoremediation Using PEPS

For salt remediation, PEPS are effective for several reasons: (1) The PGPR alleviate

plant stress and promote growth by conferring salt tolerance to the plants, as well as

conferring tolerance to potential co-contaminants such as PHC and metals. (2) The

PGPR protect plants against other potential abiotic stressors (e.g., cold) that result

in the production of stress ethylene and decreased rates of plant growth. (3) The

large amount of root biomass produced in the soil allows for effective partitioning

of NaCl out of the soil into the biosphere. (4) Foliar tissues of PEPS plants can be

harvested, thereby removing accumulated salt from the site. Harvested vegetation

will not have sufciently high levels of salt ions to be considered high-salt waste.

(5) PEPS are adapted to site-specic conditions (i.e., from the site in question),

which increases the chance of successful remediation.

Since 2009, PEPS has been deployed for full-scale remediation of several salt-

impacted sites in Manitoba, Saskatchewan, Alberta, and the Northwest Territories,

Canada [111]. Remediation goals were met at eight of these sites, either by lowering

soil salinity (ECe) levels to generic regulatory criteria, or by restoring plant growth

and productivity to equivalent land use (i.e., equivalent growth and productivity to

areas surrounding the site). We have observed that the ECe drops at a rate of appro-

ximately 15% per year when the salt is present only in the rooting zone [84].

2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant Growth-Promoting…

guarino@unisannio.it

The average amount of NaCl taken up into the leaves of PEPS grass plants is

29 g/kg (DW). An average of 150kg/ha NaCl is removed from a eld per crop

harvest. Thus, soils with an ECe of 10–15dS/m, spread to a depth of 0.5m (appro-

ximate rooting zone of grasses used in PEPS) can be remediated in about 5years

[111]. We note that as remediation proceeds, and soil salt levels drop and the soils

improve, the plants will grow better, which should lead to accelerated rates of reveg-

etation and remediation. More biomass will be produced per growing season, and

the levels of NaCl taken up by plants does not drop as the ECe decreases (see the

section entitled “Uptake of NaCl from Soil”). Given that research has shown that we

can successfully establish plant growth using PEPS on salt-impacted sites before

soil salt levels drop to generic regulatory criteria, phytoremediation based on reveg-

etation and equivalent land use will occur sooner than the 5 year estimate based on

salt uptake and biomass calculations.

2.3 Conclusions

Research described in this chapter indicates that salt phytoremediation is feasible

using PEPS and other systems. Rapid plant growth leads to revegetation of salt-

impacted sites, typically in less than 5 years. The calculations for the estimated time

required to remove NaCl from salt-impacted soils suggest that salt ions can be phy-

toextracted from soil at an acceptable rate, which will lead to unimpacted soil in the

long term. Revegetation and removal of salt from impacted soils should accelerate

in successive years of PEPS treatment as the soil improves, because this will facili-

tate increased plant growth over time, which in turn will provide a larger sink for

soil salts. Finally, revegetation may be the most important aspect of salt phytoreme-

diation, and in many cases can be considered a key measure of successful salt

phytoremediation.

Acknowledgements The work performed in the lab of B.M. G. was supported by grants from the

Natural Sciences and Engineering Research Council of Canada.

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2 Phytoremediation ofSalt-Impacted Soils andUse ofPlant Growth-Promoting…

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53© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_3

Chapter 3

Successful Integrated Bioremediation

System ofHydrocarbon-Contaminated Soil

ataFormer Oil Renery Using Autochthonous

Bacteria andRhizo-Microbiota

ValentinaSpada, PietroIavazzo, RosariaSciarrillo, andCarmineGuarino

Abstract The development of industrialized global economy have produced a

strong contamination by the petroleum-based products resulting from the activities

related to the petrochemical industry; in the last years, the hydrocarbons become

one of the major environmental problems. Bioremediation is a new approach based

on the use of microorganisms (bacteria and fungi) and plants, and it has been

researched extensively for possible applications related to hydrocarbon degradation

in the petroleum industry.

The scope of the application of this technology on soil of a former oil renery is

the production in situ of strong and diverse enzymatic activity such as to attack the

hydrocarbon molecules through various routes of enzymatic degradation. The appli-

cation of a remediation based on the biological degradation process by means of a

strategy of action based on in situ degradation principles of aerobic bacteria, fungi,

and plants either through biostimulation actions of the indigenous microbial popula-

tion, both by increasing the content of the same ora through further introduction of

native bacteria, fungi, and plants has the advantage of reducing the risks of residual

contaminants and/or inverse transformation.

Keywords Bioremediation • Total petroleum hydrocarbons • Autochthonous bacteria

• Land farming • Biostimulation • Phytoremediation • Bioaugmentation • In situ

treatments

V. Spada • R. Sciarrillo • C. Guarino (*)

Department of Science and Technologies, University of Sannio,

Via Port’Arsa 11, Benevento 82100, Italy

e-mail: guarino@unisannio.it

P. Iavazzo

Lande S.p.A.Environment\Heritage\Archaelogy, Via G.Sanfelice 8, Naples, Italy

guarino@unisannio.it

3.1 Introduction

In the last years, there is a widespread knowledge that the soil is an important com-

ponent of the environment, and it is not an inexhaustible resource. An improper use

of extracted organic substances from the ground can lead to a depletion of resources

and a possible loss into the environment. One of the major global problems is con-

tamination by the petroleum-based products, resulting from the activities related to

the petrochemical industry; especially in the past, when awareness of the health and

environmental effects connected with the production, use, and disposal of hazard-

ous substances were less well recognized than today [1]. Petroleum products are

principal components of our society, and increasing number of sites contaminated

by hazardous organic contaminants are detected. The industrialized countries have

regulated the emission of toxic substances into the environment and stated the need

to reclaim the now-contaminated environments because the pollution with petro-

leum and petrochemical products has been recognized as a signicant and serious

problem. Most components of petroleum oil are toxic and hazardous to the health of

plants, animals, and human, and it is easy to incorporate into the food chain; these

dangerous aspects have increased scientic interest in examining the distribution,

fate, and behavior of oil and long-term damage to aquatic and soil ecosystems and

natural resources. At rst, conventional technologies were used for the soil remedia-

tion, for instance, chemical oxidation, thermal desorption, and excavation with off-

site disposal in landll [2] but these technologies have shown many disadvantages

as more expensive with high energy consumption and can also lead to incomplete

decomposition of contaminants. Later, biological methods were applied in contrast

to traditional soil remediation technologies; they are environmentally friendly

approaches and cost-effective having a positive impact on public opinion and can

often be carried out in situ. With bioremediation, we identify a set of eco-friendly

techniques that use biological agents, such as bacteria, fungi, and green plants to

remove or neutralize hazardous substances in polluted site, known green technology

like land farming, biostimulation, phytoremediation, and bioaugmentation [3, 4].

Bioremediation offers many advantages over traditional remediation technologies

also because it can be applied in situ without the need to remove and transport the

contaminated soil and is usually less expensive and less labor intensive [5]. Land

farming is both ex situ technique in which contaminated soil is excavated and peri-

odically tilled until pollutants are degraded and on-site method because it is spread

the contaminated soil in a thin layer on the surface to be decontaminated in order to

stimulate aerobic autochthonous microbial activity and to facilitate degradation of

pollutants. Phyto- and bioaugmentation are a set of processes: fungi and bacteria

can detoxify and remove by breaking down pollutants such as hydrocarbons into

less harmful substances through their difference metabolic capabilities (enzymes

with biodegradative activity). Green plants (hyperaccumulators or metallophytes),

instead, can aerate polluted soil or stimulate enzymatic microbial activity, with

petroleum contaminants, and they can absorb heavy metals into their tops, which

are then harvested. The plants are able to tolerate phytotoxic level of heavy metals,

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and they can survive and reproduce in polluted soil developing systems to survive

and to adapt themselves at extreme environmental condition (increased the tolerance

to heavy metal ions or restricted the entry or root-to-shoot translocation) [6–11].

Through biostimulation we can add nutrients, oxygen, and electron acceptors or

donors to increase the population and the activity of naturally bacteria that are able

to degrade pollutants [12, 13] or to stimulate plants ability to adsorb inorganic com-

pound by environment respecting environment characteristics. Finally, into some

bioremediation strategy with a mix of pollutants, plant growth-promoting bacteria

(PGPB) are used because they can interact with roots of many different plants and

some of them colonize the interior of the plant as well, showing advantages for each

other [5, 14–17]. Each technique shows the ability to remediate a specic pollutant

but the removal of toxic compounds from the sites is further complicated when the

pollution is multiple and involves numerous classes of compounds, as heavy metals

or total petroleum hydrocarbons (TPHs), divided into aliphatic (C<18 and C>18)

and aromatic hydrocarbons. Bioremediation techniques are rapidly increasing

because it is a good alternative to conventional cleanup methods and has been used

in sites worldwide with success. It requires highly qualied staff, engineers, and

chemists who cooperate to improve remediation of polluted sites with knowledge,

great potential, and experiences into innovative technologies. Some researchers

demonstrated reducing of the contaminant concentration and ecotoxicity in the soil

via bioremediation processes made by autochthonous microorganisms [3, 18, 19].

Under the selective pressure of environmental pollution, only some microorganisms

are able to resist and degrade pollutants as TPHs. Contaminant are transformed and

breaking down by living organisms through their enzymatic metabolic processes,

but the biologic degradation is often a result of the action of consortium of microor-

ganisms (autochthonous bacteria of soil and rhizobacteria). Recent studies have

isolated and identied a large number of species of microorganisms that are able to

degrade a wide range of natural and xenobiotic compounds like hydrocarbonic or

aromatic molecules (Bacillus, Pseudomonas, Comamonas, Acinetobacter, Alcali-

genes, Streptomyces, Sphingomonas) and different fungi (Aspergillus, nonlignino-

lytic or ligninolytic fungi) [7, 20–22]. Each of them is able to use contaminants as

sole carbon source and to develop common biochemical pathways for degradation

(protein pattern or specic catabolic genes).

Until now, many works of soil bioremediation have been carried out in labora-

tory, instead the eld experiments are scarce [23, 24]. This biotechnological approach

has received a great attention in the recent years.

Below, we have shown our case study of an integrated bioremediation system of

hydrocarbon-contaminated soil from a decommissioned renery in Italy using

autochthonous bacteria and rhizo-microbiota. A total biological remediation pro-

cess allowed to overcome many of the restrictions linked to the application of

individual techniques achieving successful results with an in situ combined strategy

of different technologies, according to the site-specic features previously detected

in laboratory-scale assay. The main principle of selected strategy is represented by

stimulation of the aerobic degradation of autochthonous bacteria activity and their

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emphasizing, previously isolated in laboratory. Our Multi-Process System has

allowed to reduce the concentration of contaminants in short time respect to other

single method.

3.2 Integrated Bioremediation System ofHydrocarbon-

Contaminated Soil: Case Study

In order to close this section for readers and to explain the possibility of in situ

application of this multi-process system, we describe the aim of our work: it is the

remediation of a TPHs-contaminated soil in a former oil renery in northern Italy

by using an in situ application of an integrated bioremediation system with autoch-

thonous bacteria, rhizo-microbiota, and plants. The basic principle of the aerobic

biodegradation performed by autochthonous bacteria associated to a bioaugmenta-

tion step with the indigenous bacteria consortium previously isolated and character-

ized in laboratory.

According to the site-specic features previously detected in laboratory scale

assay, a total biological remediation allowed to overcome many of the restrictions

linked to the application of individual techniques achieving successful results with

an in situ combined strategy of different green technologies: land farming, biostimu-

lation, phytoremediation, and bioaugmentation as previously described. Experimental

design consists of different and specic steps that follow one another. After identify-

ing the polluted area, we carried out:

• detection of pollutant area and sampling

• laboratory activities split into isolation and identication of autochthonous bio-

degrading bacteria, quantization and characterization of TPHs, and mesocosm

trials in order to show the best combined technologies to remediate polluted

soils;

• application of bioremediation integrated method in situ (trial area). The third

step was made from: land farming, biostimulation, bioaugmentation, phytoreme-

diation, and biosparging and all performed based on the laboratory results and

monitoring.

Below we will analyze individually steps of our pilot work and they are summa-

rized in Table 3.1.

3.2.1 Detection ofPollutant Area andSampling

Our study was carried out with the soil from a decommissioned renery (about

400.000m2) located in Italy, contaminated by TPHs. On the basis of previous analy-

sis, we selected six points with different levels of contaminants at two different depths.

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Table 3.1 Activities performed into bioremediation integrated method (in laboratory and in situ techniques)

Activities Actions Targets

Laboratory activities Quantization of TPHs In order to determinate hazardous organic

contaminants detected into soil

Isolation and characterization

autochthonous microorganisms from

polluted soil

To identify indigenous biodegrading bacteria

to inoculate in mesocosm and in situ eld

In situ application

Preliminary operations To the surface Removing spontaneous vegetation,

stones, and inert materials

Prepare experimental in situ plane

Levelling of soil

Water drainage system

Land farming+

biostimulation

0–1.5m Soil plowed deeply and added of

fertilized

To promote aeration and stimulate and

enhance metabolism and the oxidation of

contaminants

Phytoremediation Planting of selected species Vigorous root system that are able to create

luxurious rhizosphere ideal for microbial

growth

Bioaugmentation

(PGPB)

Introduction of previously selected

indigenous bacteria and PGPB

To allows the increase of microbial biomass

Biosparging 1.5–2.5m Introduce air and nutrients

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These samples were used for laboratory analysis to detect the amount of total

petroleum hydrocarbons (especially the more recalcitrant fraction, C>12) and culti-

vable aerobic microbial population was evaluated (Figs. 3.1 and 3.2). Environmental

investigations carried out at different depths of pollutant soil highlighted values higher

than Italian regulatory limits provided by D.Lgs. 152/06.

Fig. 3.1 Core drill used to collected soil

Fig. 3.2 Example of collected soil for laboratory analysis

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3.2.2 Laboratory Activities

3.2.2.1 Analytical Analyses

Soil samples were collected using a core drill (Fig. 3.1) in order to arrive at two dif-

ferent depths (max 3 m) and arriving in laboratory the samples were stored at

−20°C until specic analysis. To determine the amounts of TPHs and their molecu-

lar characterization, soil samples were evaporated and analyzed with analytical

method that combines gas chromatography and mass spectrometry (GC-MS). We

have identied following groups, considered even for in situ experiments as: low

molecular weight hydrocarbons (C≤12), aliphatics (C13–C18 and C19–C36), aromat-

ics (C11–C22), and high molecular weight hydrocarbons (C>12). In all pollutant-

collected samples, the values of hydrocarbons are higher than Italian regulatory

limits provided by D.Lgs. 152/06.

3.2.2.2 Characterization ofAutochthonous Biodegrading Bacteria

Before starting with the isolation of hydrocarbon-degrading bacteria, we have made

a total viable count of cultivable aerobic native bacteria (for all points, we have

about 106 colony forming units, CFU); after opportunity dilutions and planting on

LB agar (incubated at 26°C for a week), the colonies were counted and values were

expressed as colony forming units. Enrichment cultures performed with contami-

nated soil and diesel as sole carbonic source (5%) allowed the selection and genomic

DNA was extracted from every isolates using specic DNA purication Kit

(Wizard® Genomic DNA purication Kit, Promega). 16S rRNA gene was used as

template with universal primers, F27 and R1492. Polymerase chain reaction (PCR)

was performed according to supplier’s instructions and the PCR-amplied DNA

was sequenced using an automated DNA sequencer (ABI 3500 Genetic Analyzer).

The partial 16S rRNA gene sequences from the isolates were deposited in the

GeneBank database in order to obtain the characterization of isolated strains.

A total of about 30 different bacterial strains were isolated and identied from

TPHs-contaminated soils (Table 3.2). The isolated strains were evaluated also for

the tolerance to the heavy metals and the capacity to produce auxine indoleacetic

acid (IAA), one of the most important plant growth-promoting molecules. Phylo-

genetic analysis (with BLAST analysis) showed high identity to strains belonging

to the phylum of Proteobacteria (10% α-proteobacteria, 30% β-proteobacteria, and

60% γ-proteobacteria).

Having characterized indigenous bacteria, the consortium was lyophilized and

stored until the next application in situ step.

Often, as in our study, soil bacteria owning hydrocarbons-degradation pathways,

also show plant growth-promoting features and actually the use of plants in con-

junction with hydrocarbons-degrading and plant growth-promoting bacteria (PGPB)

offers much more potential for the remediation of hydrocarbons-contaminated

soils [17]. Additionally, PGPB mitigate plant stress responses thus enhancing plant

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Table 3.2 Microorganisms isolated from the hydrocarbon-contaminated site; tolerances to heavy metals, capacity to production auxine as idoleacetic acid

(IAA) and capacity to grow on Diesel oil a sole C source; microorganism with same genera and specie names show genetic sequence difference, and it is

highlighted even by different responses at pollutants

Bacterial isolates/closest

described relative

CuCl2

(150mg/L)

NiCl2

(25mg/L)

HgCl2

(10mg/L)

AsIII

(10mM)

AsV

(100mM) Growth on

DIESEL oil

Production of the

IAA (mg/mL)Tolerance to metals

Acetobacter pasterianus + + −+ + − −

Achromobacter marplatensis + + − − −/+ − −

Achromobacter spanius + + − − −/+ − −

Achromobacter marplatensis + + −+−/+ − −

Comamonas koreensis + + − − +− −

Comamonas testosteroni + + − − +− −

Comamonas testosteroni + ++−+− −

Comamonas aquatica + + −+ + −53.54

Delftia acidovorans + + −+ + − −

Delftia acidovorans + + − − +− −

Ochrobactrum anthropi + + − − +− −

Pseudomonas stuzeri + + − − +− −

Pseudomonas brassicacearum + + −+ + + −

Pseudomonas brassicacearum + + −+ + + 13.44

Pseudomonas migulae + ++−/+ + + −

Pseudomonas mandelii + + + + + + 10.95

Pseudomonas frederiksbergensis + ++− − + 7.21

Pseudomonas mandelii + + − − −/+ + 17.90

Pseudomonas chloraphis + + − − − − −

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Pseudomonas chloritidismutans + + − − −/+ − −

Pseudomonas putida + + −+ + + 6.41

Pseudomonas putida + + + − − + 6.63

Pseudomonas resinovorans + ++−+− −

Pseudomonas resinovorans −/+ + − − −/+ − −

Pseudomonas alcaliphila −/+ + − − −/+ −19.73

Pseudomonas corrugata + + − − −/+ − −

Pseudoxanthomonas mexicana + + − − −/+ − −

Sphingobium abikonense −/+ + + −/+ −/+ + 4.19

Stenotrophomonas rhizophila + + − − +− −

Pseudomonas frederiksbergensis + + + − − + 63.95

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growth and development [25]. Some of these bacteria are able to degrade and

survive alone into TPHs-contaminated soil (as P. putida or P. mandelii). Many

scientists have studied the expression of specic pathway enzymes of several

Pseudomonas or Sphingomonas strains that are able to degrade a wide range of

natural or xenobiotic compounds or pesticides in no-toxic-substances [20].

3.2.2.3 Mesocosm Trials

After identication of indigenous biodegrading bacteria, we have designed a meso-

cosm trial in order to test different remediation methods using polypropylene trays

containing sieved soil and monitoring all environmental parameters (as temperature

and aeration).

Only eight of 12 samples were used for mesocosms and represented different

contamination levels. We have applied three different experimental conditions for

90days to verify the best bioremediation methods in order to use for in situ experi-

ments: rst, we have simulated site conditions and the soil is dampened by the rain,

called Natural Attenuation (NA); in the second set of trays, we have added N-P-K

mixture as nutrients only one time and irrigated twice a week with demineralized

water. So we have plowed soil to promote aeration and biodegrading bacteria activi-

ties, and this treatment is called Land farming (L). In the third set, we have treated

the soil with the same manner as the second one and after 60days we have added

indigenous bacteria strains that we have isolated and identied (LB). At the end of

90days, we have collected one sample for each tray, and we have quantied the

amounts of TPHs after treatments in order to dene the optimum method. Analytical

analyses were made at T0, T60, and T90 days for all trays and the positive effects of

major variation and decrease of hydrocarbons were highlighted after LB treatment.

3.2.3 In Situ Activities

3.2.3.1 Pilot Field

According to site-specic features and data obtained from mesocosm trials, a total

biological remediation process with an innovative and sustainable strategy based on

the combination of mechanical, microbial, and plant growth processes was applied;

this is an integrated bioremediation system.

For in situ experiments, we have used only a trial area of about 700m2 at two

different soil depths (Fig. 3.3), and we have applied integrated bioremediation

system for a period of 5 months, started after laboratory experiments, as mesocosm

trials with positive results which have shown different enzymatic degradation path-

ways by aerobic bacteria. Specically, we have worked in two depths (0–1.5m and

1.5–2.5m), in which we have applied simultaneously different biological remediation

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techniques. The main applied methods are: cleanup of soil, land farming, biostimu-

lation, phytoremediation, and bioaugmentation with the injection of specic com-

mercial mix of PGPB and nally biosparging (Fig. 3.4).

Before starting with all activities of bioremediation set, we have applied some

operations that are necessary to clean up and prepare the eld: mowing of spontane-

ous vegetation, removal of stones or inertial materials; these operations are inserted

into total bio-approach.

Fig. 3.3 Pilot area

Polluted area (400000

m2)

Field area 700 m2

Landfarming

Application of integrated

system of bioremediation

in-situ

Biostimulation

Bioaugmentation

(PGPB)

Phytoremediation

0-1,5 m depth

1,5-2,5 m depth Biosparging

Fig. 3.4 Experimental design of in situ application of integrated bioremediation system

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3.2.3.2 Land Farming

At rst, land farming method was applied in order to promote aeration and to bios-

timulate the oxidation of contaminants by indigenous bacteria. Soil was plowed

deeply, weekly and for 40days, up to about 1.5m, using a mechanical excavator

(Fig. 3.5). Then, levelled eld for the following activities and to facilitate the pro-

cessing of the hydraulic-agricultural settlement of soil, we have created system for

the outow of excess water and draining surface one; all these operations were

developed into a period of 2 months.

3.2.3.3 Biostimulation

After the preparation of soil steps, nutrients were added to the soil at different time

and the selection of type and amount were made according to biological and agro-

nomic conditions. Biostimulation method was performed in order to increase the

degrading activities of indigenous community and to ensure rapid root develop-

ment. The mycorrhization was developed in two steps: rst, before planting, in

order to create the best condition for future plant growth (about 100g/m2 and at

20cm of depth); second, after planting, directly applying mix to the roots manually

and wetting the surface to promote the germination of fungal spores (Fig. 3.6).

Finally, we have fertilized with slow-release ternary fertilizer containing Nitrogen-

Phosphorus- Potassium (N-P-K) about 100g/m2 of fertilizer were uniformly distrib-

uted by hand in order to promote degrading and energy metabolisms of autochthonous

bacteria and create optimal condition for following plants.

Fig. 3.5 Moving top-soil with excavator

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3.2.3.4 Phytoremediation

In phytoremediation step, we have selected in reference to literature and preli minary

studies of the soil (chemical characteristics and amount of pollutants) and planted:

Festuca arundinacea Schreb., Phragmites australis (Cav.) Trin. ex Steud. (autoch-

thonous), and Populus nigra L. (autochthonous) (Table 3.3). All selected plants have

a vigorous root system that is able to create an optimal structural function for micro-

bial growth. Phytoremediation has been recognized an effective and eco-sustainable

method to remove inorganic and some organic molecules by employing a variety of

mechanisms often to support microbial degradation [26].

After doing pump system, we have planted the vegetable species into pilot area

split in two parts: parcel A P. australis+F. arundinacea and parcel B P. nigra+F.

arundinacea (Fig. 3.7). With a frequency of two times a week, we have irrigated the

plants during the dry period with a frequency of two times a week.

3.2.3.5 Bioaugmentation

Bioaugmentation method was divided into two steps: preparation of inoculum to

inject starting by isolated and characterized biodegrading consortium bacteria plus

selection of commercial mixture of PGPB and then the injection of each mixture.

We have used similar growth condition and enrichment steps of laboratory step

for preparation of bacteria inoculum for in situ application, starting from consor-

tium prior lyophilized. The bacterial consortium was grown using bioreactors

Fig. 3.6 In situ eld application of fertilizer and mycorrhizae manually

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Table 3.3 Plants used for phytoremediation and their main feature

Application in Phytoremediation Structural functions Planting technique Distribution in situ

Festuca

arundinacea

Change the composition of the

rhizospheric microbial population

promoting the degradation

capacity of organic compounds

High coverage density and

deep root system (up to 80cm)

Hand seeding Uniform distribution

Phragmites

australis

Mostly used for constructed

wetland

Deep root system (60–80cm):

the long rhizomes allow the

creation of oxidized micro-

zones easily colonized by

aerobic bacteria

Planting of

autochthonous

rhizomes

70×70cm

Populus

nigra

Widely used in rhizodegradation

for the intensive phytoremediation

performed by roots

Preservation of aquifers made

by large root system

Transplant of

autochthonous trees

1×1m

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(BIOSTAT® Cplus, Sartorius) that are able to control growth condition and to keep

constant for all time (28°C, pH6.8, ow air 10 (middle) and we added 5% diesel as

sole carbonic source). We have achieved a nal concentration of 109 CFU/mL

(Fig. 3.8). The injection was realized using a ow-controlled injector pole which

allowed precise distribution at low pressure and the pole was pointed in proximity

to the root system and the pump was necessary to assure homogenous air distribu-

tion. The injector pole was pointed in proximity to the root system (20–30cm of

depth) and about 220mL/m2 of consortium were inoculated in the whole pilot eld

(Fig. 3.9). By adding bacterial consortium, we increase the microbial concentration

and stimulate autochthonous bacteria distributing air ow in order to enhance the

pollutants biodegradation. It is advantageous to increase both the tolerance and the

resistance to variations in natural environment.

In addition to autochthonous consortium, we have selected even a commercial

microgranular consortium in according to site-specic features and added into rhi-

zosphere with action like plant growth promoting. Commercial inoculum was made

from mycelium and vital spores of arbuscolar micorrhizal fungi of Glomus sp.

enriched by natural microorganisms (Trichoderma sp., Bacillus spp., Streptomyces

sp., Beauveria sp., Metharhizium sp) that are able to stimulate and emphasize

microbial degradation of TPHs.

Fig. 3.7 Pilot eld after planting and pump system

Fig. 3.8 Bacterial

consortium

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3.2.3.6 Biosparging

The biosparging system is the same used to biostimulate (Fig. 3.10) and add the

microbial consortium; we have stimulated the previously detected activity of

autochthonous bacteria in saturated zone distributing an air ow in order to enhance

the biodegradation. All over the pilot eld, we have realized nine dug-out (every

25m). Along each dugout, a horizontal piping formed by two coaxial tubes was laid

at about 1.80–2 m b.g.s. The external tube (ø = 160 mm) was provided with

Draining pumpAir injection system

Ditch

Biosparging pipe

Fig. 3.10 Layout (name of

samples analyzed) and area

divided with biosparging

system

Fig. 3.9 Direct inject into eld with injector pole

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microslots along the buried length. The inner microslotted tube (ø=63mm) was

connected to a low pressure air compressor at one side and to a draining pump at the

opposite side in order to purge out the excess water that was then stocked in a tank

container for further dump. The air compressor worked in continuum at about

120mbar of pressure for 80days. Pump pressure was estimated to overcome hydro-

statical pressure and to assure homogenous air bubble distribution along the whole

piping length in the overheading soil layer (Fig. 3.11). At the same time, contami-

nants volatilization due to air sparging was taken into account and evaluated as

negligible. Finally, to increase oxygen distribution process into saturated zone, we

have injected a commercial product, with a high quality calcium peroxide powered,

and widely used to enhance aerobic bioremediation processes due to the slow

release of oxygen and heat when in contact with water.

3.3 Integrated Process Monitoring andInSitu Results

Periodically, the efciency of applied remediation methods were evaluated monitor-

ing the concentration of pollutants into soil samples (at 50cm of depth) at different

times: T0, before remediation process, T1 (after land farming and biostimulation,

Fig. 3.11 Structure of biosparging system into pilot eld

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before phytoremediation and bioaugmentation, day 50), T2 (after phytoremediation

and bioaugmentation, day 100), Tf (at the end of the trial, day 150). Besides, the

plant growth was conducted with visual monitoring at the end of the process (Tf )

and determinated the total microbial biomass (CFU) (Fig. 3.12).

The goal of our study was to show mainly the efciency of different integrated

technologies applied in situ by decreasing TPHs amount followed by an increase of

biodegradation activity of autochthonous bacteria, in a short time (150 days).

Moreover, we have used plants with potential role into phytoremediation approach

[27–29].

At rst, we have evaluated the plants growth and their activity, even if the short

range time of experiments did not allow to develop all parts of them. At Tf (the end

of trial), the suitable rooting and the rst plants growth signs were detected: Festuca

arundinacea covered homogenously the whole pilot eld; little trees of Populus

nigra had all responded well to the transplant and began to form the rst dormant

buds, and rst shoots of Phragmites australis began to rise up. These parameters did

not yet show effect of plants into remediation process because probably the root

system of plants was still at the early stage of growth but they are able to grow in

contaminated area.

Autochthonous bacteria proliferation and the increased of microbial degradation

activity in soil were enhanced by the use of integrated technology, in fact, land

farming and biostimulation were used to inject oxygen in order to improve environ-

mental conditions for microbial growth and bioaugmentation to increase the total

biomass (Fig. 3.13).

In Table 3.4 and Fig. 3.14, we have showed microorganisms content at different

steps of integrated strategy, with an increase at the end of trial, after bioaugmenta-

tion of microbial consortium that are able to remediate TPHs.

The total compounds present in soil before, during and after the application of

Multi-Process System were analyzed using GC-MS and matched with the limits

established by Italian environmental legislation (D.Lgs. 152/06). The preliminary

characterization has showed the amounts of PAHs; BTEX and heavy metals (HMs)

(Table 3.5) were below the established limits by law.

Fig. 3.12 Soil sampling points (in red the points corresponding to T0, T1, T2, Tf; in blue only Tf)

V. Spada et al.

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Fig. 3.13 Sampling of soil

with little core drill in

order to monitor the effects

of multi-process steps

Table 3.4 Autochthonous bacteria proliferation (values of CFU)

Sample

(depth) T0T1T2TfΔ (CFU/g dry)

A1 (0–1) 1.27E+05 1.63E+07 1.57E+07 6.79E+06 6.66E+06

A1 (1–2) 1.27E+07 2.94E+07 1.1E+07 – −1.61E+06

A2 (0–1) 2.63E+06 4.46E+07 2.13E+07 – 1.86E+07

A2 (1–2) 2.65E+07 5.91E+07 2.35E+07 – −2.98E+06

A3 (0–1) 5.63E+06 5.95E+07 1.25E+07 – 6.88E+06

A3 (1–2) 5.17E+07 6.11E+07 1.81E+07 – −3.36E+07

B1 (0–1) 1.119E+07 5.49E+06 5.12E+07 – 3.94E+07

B1 (1–2) 9.3E+06 1.06E+08 1.95E+07 – 1.02E+07

B2 (0–1) 2.28E+07 2.11E+07 5.77E+06 3.19E+07 9.09E+06

B2 (1–2) 4.61E+07 7.17E+07 3.07E+07 5.63E+06 −4.04E+07

B3 (0–1) 1.04E+07 6.33E+06 3.19E+07 3.33E+07 2.3E+07

B3 (1–2) 2.70E+07 3.1E+07 1.59E+07 – −1.11E+07

TPHs represented the only contaminant class present into area and at signicant

concentration level; a specic analytical method was allowed to identify four differ-

ent groups: low molecular weight hydrocarbons (C≤12), aliphatics (C13–C18 and

C19–C36), aromatics (C11–C22), and high molecular weight hydrocarbons (C>12)

(Table 3.6). Long chain hydrocarbons are less biodegradable and more recalcitrant

to biological actions as well as they are less volatile and soluble in the water

(decreased of migration into environment). Changes in the residual of THPs of soil

samples treated are showed in Fig. 3.15, quantied in top and deep layer. We have

reported the contamination values as percentage ratio compared to starting contami-

nation (T0). In panel (a) we have showed a signicant TPHs concentration decrease

(about 50%) after land farming and biostimulation step (T1). The pollutants removal

3 Successful Integrated Bioremediation System ofHydrocarbon-Contaminated Soil…

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speed after 50days is decreased due to the lower amounts in soil. Each stage of the

applied technological process led to a reduction value of TPHs content always

higher than 40%, highlighting a signicant removal speed at each phase. At the end

of the trial test (Tf), the remediation speed was found higher than 85% with respect

to the initial contamination. Instead, the removal rate showed in panel (b) is too slow

compared to the rst step of remediation into top layer (T1); this is due to a lower

soil aeration which leads to a slower activation of biological processes. The follow-

ing phases of biosparging and bioaugmentation showed a faster removal rate,

comparable to that observed for top layer, conrming the high efciency of the two

combined processes. Finally, despite the rst slower phase, we observed a remedia-

tion of about 80% (Tf) with respect to the initial contamination (T0).

TPHs concentrations were always below the Italian legal limits at the end of the

trial (Tf), both top or deep layer. All the results did not show any substantial differ-

ence in removal rate between C13–C18 and C19–36 fractions, conrming the effective-

ness of the selection of autochthonous microorganisms in the laboratory. The further

step of consortium cultivation, carried out in laboratory, likely allowed a high spe-

cic adaptation to the particular substrate made available at that stage: the high

selective pressure generated a strong adaptation of microorganisms, with a potential

considerable removal rate at optimal conditions of growth in view of full-scale

remediation.

All results of bioremediation method are monitoring after 150days and this time

is too short to analyze the activity of plants; however, we have shown the rooting

and the germination of plant species (Fig. 3.16). So, we can estimate that the root

system is in an early growth stage and therefore, it is not able to inuence the reme-

diation process.

Fig. 3.14 Autochthonous bacteria proliferation

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Table 3.5 Total amount of inorganic compound into soil at T0 and limit values (D.Lgs 152/06)

Sample

(depth) As Cd

(III)

(IV) Fe Mn Hg Ni Pb Cu Sn Zn

T0 (mg/kg)

A1 (0–1) 11.2 <0.20 15.0 <0.20 11,309 765 0.74 25.4 18 13 0.86 98

A1 (1–2) 5.1 <0.20 13.4 <0.20 5988 272 0.61 21.7 3 4 <0.50 61

A2 (0–1) 9 <0.20 18 <0.20 11,585 981 0.88 30.8 23 16 0.92 115

A2 (1–2) 6.6 <0.20 20.6 <0.20 8415 503 0.79 34.4 4 5 <0.50 80

A3 (0–1) 8.9 <0.20 17.9 <0.20 11,407 768 0.89 28.4 86 20 2.31 283

A3 (1–2) 4.5 <0.20 16.2 <0.20 5942 306 0.93 27.2 5 5 <0.50 92

B1 (0–1) 10.4 <0.20 14.5 <0.20 10,556 996 0.80 25.1 10 10 0.71 89

B1 (1–2) 5.4 <0.20 18.4 <0.20 7334 531 0.88 29.8 4 6 <0.50 85

B2 (0–1) 7.6 <0.20 15.8 <0.20 10,579 741 0.85 27.8 53 27 1.33 137

B2 (1–2) 5.2 <0.20 17.3 <0.20 7544 329 1.70 32 6 9 1.02 160

B3 (0–1) 8.2 <0.20 23.6 <0.20 13,748 729 0.87 30.8 74 26 1.71 137

B3 (1–2) 5 <0.20 31 <0.20 10,026 399 3.46 44.9 6 12 2.21 359

Limit

D.Lgs.

152/06

50 15 800 15 – – 5 500 1000 600 350 1500

Table 3.6 Total amount of TPHs T0 and Tf; we do not show low molecular weight hydrocarbons

(C≤12) and aromatics C11–C22 because we have parameters too low (data for C≤12: sample A2

(0–1)=89.9mg/kg and B1 (1–2)=32.7mg/kg; for other samples, we have values lower than legal

limit)

Aliphatics C13–C18

(mg/kg)

Aliphatics C19–C36

(mg/kg) TPHs (C>12) (mg/kg)

Sample

(depth)

T0T1T2TfT0T1T2TfT0T1T2Tf

A1 (0–1) 204 171 103 80 363 326 242 116 567 497 345 196

A1 (1–2) 157 88 13 1 314 233 142 8 471 321 155 9

A2 (0–1) 2357 1415 172 72 4173 732 287 187 6530 2147 459 259

A2 (1–2) 353 287 94 44 376 297 142 96 729 584 236 140

A3 (0–1) 1291 524 506 160 2239 1232 723 328 3530 1756 1229 488

A3 (1–2) 306 214 120 20 332 262 153 52 638 476 273 72

B1 (0–1) 1405 651 223 133 1852 943 536 206 3257 1595 759 339

B1 (1–2) 439 391 113 1 477 427 122 10 916 818 235 11

B2 (0–1) 902 463 353 203 1520 781 616 216 2422 1244 969 419

B2 (1–2) 281 222 75 75 403 354 274 74 684 576 349 149

B3 (0–1) 1429 876 723 624 1958 1269 1078 678 3387 2145 1801 1314

B3 (1–2) 224 226 418 238 307 317 359 179 531 543 777 417

3 Successful Integrated Bioremediation System ofHydrocarbon-Contaminated Soil…

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F. arundinacea has covered homogenously all pilot eld; in parcel A, young

shoots of P. australis are visible and in parcel B, the little tree of P. nigra have

responded well to the transplant operation.

3.4 Conclusion

In our study, we applied a total biologically integrated bioremediation system in situ

based on aerobic degradation by microorganisms (autochthonous and inoculated)

and plants, on petroleum hydrocarbons-contaminated soil at a former renery. Our

100

50 100

Aliphatics C13-C18

Aliphatics C19-C36

Hydrocarbons C>12

Aliphatics C13-C18

Aliphatics C19-C36

Hydrocarbons C>12

Operational time (d)

Residual contamination(%)

Top Layer

DeepLayer

Operational time (d)

150 200 050 100 150

200

Fig. 3.15 Removal trend expressed as percentage over trial time (a) top layer; (b) deep layer

Fig. 3.16 Images of plants at Tf; (a) P. australis, (b) F. arundinacea, and (c) P. nigra

V. Spada et al.

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results have highlighted that the integration of different bioremediation methods

increased the effects of aerobic hydrocarbon-degrading bacteria activities, with the

developed of optimal environment conditions for the microorganisms growth and

low energy consumption, before laboratory trials and after in situ.

Our integrated technology was designed and developed in order to produce a

diversied enzymatic degradation activity to reduce the amount of TPHs molecules

without risk of pollutant residues and inverse toxic transformation. The primary

laboratory steps are important to establish the best environment condition for plants

and bacteria activities as well as the optimum condition for technology transfer in

large scale.

So, not all remediation steps are involved into TPHs-degradation, in particular,

the presence of the plants could play a potential role as support for biological activ-

ity, how reported in literature [15, 27], and contributing to a more development of

suitable environmental conditions, reducing time of remediation. A well-developed

root system and the oxygen and fertilizer injection could support microbial activity

in less time just as the trial is monitored step by step in order to highlight the poten-

tial of bioremediation system with the contribution of phytoremediation. The micro-

bial degradation is linked to enzymatic activities of microorganisms into soil, and

the bioremediation of polluted area is based on catabolic metabolism that is capable

of using organic contaminants as carbon source and energy. The organic compounds

can be completely degraded to carbon dioxide and water, or mineralized or biotrans-

formed into less toxic compounds. The integrated bioremediation system is the pos-

sible future in situ application in order to clean up contaminated areas (in large eld

scale) and decontaminated areas can return to society.

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77© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_4

Chapter 4

Phytoremediation ofPetroleum-Contaminated

Soil inAssociation withSoil Bacteria

PrayadPokethitiyook

Abstract Unprecedented progress in industrial activities over the last century has

directly contributed to the discharge of huge amounts of petroleum hydrocarbons

into the environment. It has been estimated that about 1.7–8.8 million metric tons of

oil is released into the environment every year. More than 90% of this oil pollution

is caused by accidents due to human errors and also deliberate disposal of the waste

containing hydrocarbons. Generally, petroleum and its products get into the envi-

ronment through natural seepages, transportation, accidental spills, deliberate

disposal, offshore production, and breakage of pipelines. Presence of petroleum

hydrocarbon compounds in the environment can affect both on human health and

the environment. Therefore, their presence in nature is of great concern today, and

they need to be cleaned from the environment in the best possible way. Many

research works have been carried out to determine the eco-toxicity of these pollut-

ants but biological method has been reported to be more suitable to determine the

possible hazards of pollutants in soil on the ecological and environmental bases.

Keywords Rhizosphere bacteria • Plant–microbe interactions • Bioremediation •

Phytoremediation • Petroleum • PAH

4.1 Introduction

Unprecedented progress in industrial activities over the last century has directly

contributed to the discharge of huge amounts of petroleum hydrocarbons into the

environment. It has been estimated that about 1.7–8.8million metric tons of oil is

released into the environment every year [1]. More than 90% of this oil pollution is

caused by accidents due to human errors and also deliberate disposal of the waste

containing hydrocarbons. Generally, petroleum and its products get into the envi-

ronment through natural seepages, transportation, accidental spills, deliberate dis-

posal, offshore production, and breakage of pipelines [2–4]. Presence of petroleum

P. Pokethitiyook, Ph.D. (*)

Department of Biology, Faculty of Science, Mahidol University,

272 Rama VI Road, Phayathai, Ratcha Thewi, Bangkok 10400, Thailand

e-mail: prayad.pok@mahidiol.ac.th

guarino@unisannio.it

hydrocarbon compounds in the environment can affect both on human health

and the environment [5]. Therefore, their presence in nature is of great concern

today, and they need to be cleaned from the environment in the best possible way.

Many research works have been carried out to determine the eco-toxicity of these

pollutants but biological method has been reported to be more suitable to determine

the possible hazards of pollutants in soil on the ecological and environmental

bases [6].

The effort to clean up these contaminants in the environment through various

green technologies has become a prioritized search for both the scientic commu-

nity and the industries. One of the promising technologies is phytoremediation [7].

Phytoremediation can be performed in the soil, air, groundwater, or surface water

environment depending on the plants or the process settings. However, the toxicity

of low molecular weight petroleum hydrocarbons is considered to inhibit plant

growth and development. Furthermore, total carbon concentration of roots signi-

cantly decreased with increasing petroleum concentration [8]. Moreover, petroleum

hydrocarbons had been found to be positively correlated to the abundance of bacte-

rial genes responsible for biodegradation in the soil system [8–10]. Therefore, bio-

logical remediation of hydrocarbon-contaminated soils by environmental scientists

and engineers need to be explored further.

Bioremediation and phytoremediation of petroleum-contaminated soils have

been studied extensively in the past few decades and have been proved to be effec-

tive techniques [11, 12]. Rhizoremediation, a subset of phytoremediation, is the use

of synergy between plant and its associated rhizosphere microbes to degrade pollut-

ants in soil, has been recently found to be an effective technology. Several studies

on remediation of petroleum contamination conducted both in situ and ex situ using

plant–microbe interactions have proven to be effective [13–15].

Several studies have concentrated on the plant–microbe interaction using indig-

enous microorganisms [16–18]. The use of plants in cooperation with hydrocarbon

(HC)-degrading bacteria or plant growth-promoting bacteria (PGPB) offers an

enhanced potential for the bioremediation of TPH-contaminated soil [19–22].

Hence, plant-associated bacteria, such as rhizosphere bacteria (RB) and endo-

phytic bacteria (EB), have been shown to contribute to biodegradation of toxic

organic compounds in polluted soil and could have potential for improving the tol-

erance of plants in phytoremediation due to their possession of alkane or benzene

biodegradation pathways and their metabolites [19, 23]. Additions of these oil

degrading bacteria to the root zone might be able to enhance the remediation ef-

ciency of plants as well [24, 25]. Since this is a new innovation in environmental

biotechnology, a lot remains to be explored to make the technology even more

effective. Examples of the role of rhizosphere bacteria in the improvement of plant

tness in petroleum-contaminated soils as well as the site-specic selection of

plants for soil remediation by promoting rhizosphere bacteria for eld use will be

described.

P. Pokethitiyook

guarino@unisannio.it

4.2 Petroleum Spills into theEnvironment

Petroleum hydrocarbons are generally divided into two groups: aliphatic and aro-

matics. Aliphatic hydrocarbons are the compounds whose carbon atoms are joined

together in straight or branched open chains but not in rings. Aliphatic hydrocarbons

(alkanes, alkenes, and alkynes) in gasoline, crude, diesel, and lubricating oils consti-

tute a substantial part of organic contamination in the environment [26]. This set of

contaminant comprises saturated and unsaturated hydrocarbons having linear

or branched open-chain structures. When total petroleum hydrocarbon (TPH) is

directly released to water through spills or leaks, certain TPHs fractions will oat in

water and form thin surface lms. Other heavier fractions will accumulate in the

sediment at the bottom of the water, which may affect bottom-feeding sh and

organisms. If the TPH spills occur in soil, physico-chemical processes will inuence

the fate and behavior of it [27]. The properties of hydrocarbons especially in crude,

diesel, and lubricating oils will inuence the degradation by microorganisms.

Diesel oil is composed of middle end distillates of crude oil with boiling points

between 200 and 300°C.Hydrocarbons in diesel oil are generally found to be in the

C8 to C26 range, which comprises an estimated 60–90% alkanes and cycloalkanes,

less than 5% alkenes and 10–30% aromatics [9]. Lubricating oil is an important

product obtained from the residue of crude oil, which accounts for 60% of crude oil

derivatives. It is a petroleum product typically characterized by a very high boiling

point of more than 350 °C. Its typical carbon ranges are C20 to C45+ comprising

around 90% alkanes and 10–30% aromatics [28]. Chemical structures of various

categories of hydrocarbons are shown in Table 4.1 [28].

Apart from the much visible and attention gaining large-scale accidental dis-

charge of petroleum into both terrestrial and marine environment, the seemingly

insignicant regular discharge from efuents, urban runoff, cleaning operations,

and other oil treatments make up an estimated 90% of the total petroleum pollution

brought about by anthropogenic activities. On the other hand, localized large-scale

discharge like tanker accident and pipeline breaks make up 5–10% of the total

anthropogenic petroleum spill. Of the petroleum spills, taking into account the

amount of petroleum handled being more on land, discharge on land can be more or

even greater than into the marine environment [11].

4.3 Fate ofHydrocarbons inSoil

Petroleum spill on land is followed by rapid vertical inltration downward until it

meets the water table. Once it reaches the water it spreads out laterally over it. Two

important features of that inuences the percolation of total petroleum hydrocar-

bons (TPHs) are its viscosity and porosity of the soil. Light petroleum like gasoline

inltrate rapidly into porous soil while the case is not the same with heavy

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

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lubricating oil spills [11]. Most soils are described to be a multiphasic system

characterized by the presence of ionic solid mineral matrix and associated organic

matter that is enveloped by a lm of water (Fig. 4.1). In unsaturated soils, techni-

cally termed as vadose zone, a gas phase occupies the pore spaces while in the

spaces in saturated soils are occupied by aqueous phase. Fresh spills of TPHs on

land are initially subject to volatilization, especially from the less porous surfaces

(Fig. 4.1) while the heavier hydrocarbons may be partially oxidized by auto-, ther-

mal-, and photo-oxidation in addition to biodegradation [12].

Petroleum hydrocarbons released into the environment are subject to degradation

process with time. The processes that degrade TPHs include evaporation, leaching

(transfer to the aqueous phase), chemical oxidation, and microbial degradation [13].

Petroleum hydrocarbons by virtue of their nature are generally biodegradable. Even

the ones in natural reservoirs, the site of petroleum formation, are subject to biodeg-

radation [14]. The alteration in the composition of petroleum hydrocarbons in the

soil brought about by various physical, chemical, and biological factors are collec-

tively called weathering [15].

Table 4.1 Chemical structure of various categories of hydrocarbons

Hydrocarbon classication Description

Chemical structure and

example

Aliphatic

Alkanes Carbon chain with single bond H

n-Butane

Alkene Carbon chains with at least one

carbon–carbon double bond

C=C

Butadiene

Alkynes Carbon chains with at least one

carbon–carbon triple bond

HC ≡ CCH

CH3

1-Butyne

Cycloalkanes Single-bonded carbon ring

structure

Cyclohexane

Aromatics

Monoaromatics The benzene ring made up of six

carbon atoms with alternating

single and double bonds

Benzene

Polycyclic aromatic

hydrocarbon (PAH)

Aromatic compounds having

two or more benzene rings fused

together

Naphthalene

P. Pokethitiyook

guarino@unisannio.it

4.4 Treatment ofPetroleum-Contaminated Soil

Over the years, many treatment methods have been developed and practiced to treat

petroleum-contaminated soil, which can be broadly classied into physical, chemi-

cal, and biological processes [17].

4.4.1 Phytoremediation

Phytoremediation is a term used to describe a set of technology employed to clean

up contaminants utilizing plants. The term “phytoremediation” was rst used in

1991, which was the product of research efforts in constructed wetlands, oil spills,

and agricultural plant accumulation of heavy metals [25]. It is also dened to be a

technology that uses plants with its associated rhizosphere microorganisms to

remove, transform, or stabilize contaminants found in soils, sediments, and water

bodies. At present, the technology is used for decontaminating many categories of

contaminants including petroleum hydrocarbons [16].

Fig. 4.1 Principles of aerobic degradation of hydrocarbons by microorganisms (Fritsche and

Hofrichter [18])

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

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There are various mechanisms proposed by researchers over the years that have

been theorized and experimented in phytoremediation (see [25]).

4.4.2 Microbial Degradation ofPetroleum Hydrocarbons

Biological degradation of pollutants is founded on the principles that sustain all

ecosystems. The processes involve circulation, transformation, and accumulation of

energy and matter in nature. The soil microbes that predominantly degrade use TPH

as a carbon source and electron donor for generating energy. Through laboratory

tests and microbial characterization of bioremediation works, many bacteria have

been identied to be active in biodegradation among which a dominating population

is found in the genus Pseudomonas (Table 4.2) [18]. Although many bacteria are

capable of degrading contaminants, a single bacterium may not possess the enzy-

matic capability to degrade all pollutants in soil. For the complete and successful

cleanup of contaminated sites characterized by complex environment and contami-

nant composition, mixed microbial consortiums are needed [18, 24].

4.4.3 Bioremediation

Bioremediation can be dened as the use of microorganisms such as bacteria to

remove environmental pollutants from soil, water, or gases [26]. It can also be

dened as the utilization of the natural ability of microbes to use waste materials in

their metabolism and change them into harmless end products. Bioremediation

requires special kind of bacteria and also special operation conditions to accelerate

the natural biodegradation rates by overcoming the limiting factors. It is the con-

trolled manipulation of environment to produce proper enzymes for catalyzing the

desired reactions to break down contaminants. Basically, it is the application of

chemistry in a more intricate manner as it involves the crucial role of specic

enzymes to run the reactions, which are introduced into the system by specic

microorganisms [24]. The primary aim of bioremediation is to degrade the TPHs

fully by microbes to carbon dioxide and water. This technology comes with added

advantages over other methods of treating pollutants. Some notable advantages are

low-cost operation, reduced health, and ecological ramications [17]. The terms

and technologies involved are summarized in Table 4.3.

Table 4.2 Predominant

bacteria in soil samples

polluted hydrocarbons

Gram-negative bacteria Gram-positive bacteria

Pseudomonas spp. Nacardia spp.

Acinetobacter spp. Mycobacterium spp.

Alcaligenes spp. Corynebacterium spp.

Flovobacterium/Cytophage group Arthrobacter spp.

Xanthomonas spp. Bacillus spp.

P. Pokethitiyook

guarino@unisannio.it

Bioremediation technology is labeled to be efcient and cost-effective method to

clean up TPH-contaminated soil in both in situ and ex situ after excavation of the

soil. If in situ treatment is not feasible accounting to any reason, especially of envi-

ronmental concern, the TPH-contaminated soils are removed from the site of con-

tamination and are treated biologically using land treatment units, composting,

biopiles, or slurry bioreactors [15].

Advantages of bioremediation

• In situ treatment is possible

• Permanent removal of contaminants

• Economically cheap and feasible

• Positive public acceptance

• Long-term liability risk eliminated

• Minimum disturbance to the site of contamination

• Can be clubbed with other treatment methods

Disadvantages of bioremediation

• Some pollutants cannot be broken down by biological processes

• Extensive monitoring should be put in place

• Site-specic requirements

• Toxicity of contaminants hamper the method

• Potential production of unknown by-products in the process [24].

4.4.3.1 Mechanism ofMicrobial Degradation ofTPHs

In bioremediation, hydrocarbon substrate serves as the food (carbon) source for

energy and growth of microorganisms, which is made available following two major

ways: oxidation and/or reduction. However, hydrocarbons being already reduced

chemically and stable, further reduction is not the main mode for bioremediation,

even under anaerobic conditions [14].

Table 4.3 The terms and technologies involved in bioremediation technology (Juwarkar etal. [87])

Terms Technology involved

Bioaugmentation Addition of bacterial cultures to a contaminated medium; frequently

used in bioreactors and ex situ systems

Biolters Use of microbial stripping columns to treat air emission

Biostimulation Stimulation of indigenous microbial populations in soils and/or

groundwater; may be done in situ or ex situ

Bioreactors Biodegradation in a container or reactor; may be used to treat liquids

or slurries

Bioventing Method of treating contaminated soils by drawing oxygen through the

soil to stimulate microbial growth and activity

Composting Aerobic, thermophilic treatment process in which contaminated

material is mixed with a bulking agent; can use static piles, aerated

piles, or continuously fed reactors

Land farming Solid-phase treatment system for contaminated soils may be done in

situ or in a constructed soil treatment cell

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

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Basically, aerobic respiration of hydrocarbons entails the need for enzyme

oxygenase. The role of oxygenase is to incorporate molecular oxygen into the

reduced hydrocarbon substrate. The initial products are alcohols that are subject to

sequential oxidation producing carboxylic acids, which undergo β-oxidation [14,

15, 18]. Microorganisms engaged in biodegradation of TPHs require oxygen at two

points in the metabolic pathway; rst at the initial oxidation of the substrate and

then at the end of the respiratory chain [18].

To have the fast and better degradation of TPHs under aerobic conditions, some

characteristics of aerobic microbes must be met.

1. They must have metabolic process to increase the contact between the microbes

and the TPHs. For biodegradation, TPHs must be bioavailable to the microbes.

For example, TPHs must be accessible and also in the form that microbes can

start working on. For example, TPHs must be water soluble by the oxygenase

enzymes or by biosurfactants produced by microbes themselves.

2. The rst step of degradation requires microbes to work on TPHs by enzyme

activation and incorporating oxygen into the chains. It is an oxidative process by

oxygenases and peroxidases enzymes.

3. TPHs are converted into intermediates compounds via the metabolic pathways

of tricarboxylic acid (TCA) cycle and β-oxidation.

4. Production of cell biomass from central intermediary metabolites such as acetyl-

CoA, succinate, and pyruvate. Sugar is produced from gluconeogenesis, i.e.,

used for biosyntheses and growth.

The aerobic degradation of hydrocarbons is illustrated in Fig. 4.1 [18].

4.4.4 Rhizoremediation

The treatment of hydrocarbon-contaminated soils by the combined ability of plants

and their associated microorganisms is referred to as rhizoremediation and has been

demonstrated to be the primary mechanism responsible for plant-mediated hydro-

carbon degradation. Basically, it is the breakdown of soil contaminants by microbial

activity, which is enhanced in the plant root zone [27]. With the plant exerting

changes in the physical, chemical, and biological properties of the soil effecting

degradation of contaminants, rhizoremediation is a parallel and inseparable term

from phytoremediation [29]. It is also known by other terminologies like plant-

assisted degradation, plant-assisted bioremediation, plant-aided in situ biodegrada-

tion, and enhanced rhizosphere biodegradation [25].

It is a treatment technology that combines phytoremediation with bioremedia-

tion. Besides the plant itself undertaking phytodegradation of the TPHs, it increases

microbial numbers in the rhizosphere that undertakes biodegradation. This phenom-

enon is termed rhizosphere effect. A previous study done to see the effect of rye-

grass and alfalfa microbial population and diversity in petroleum-contaminated soil

found a signicant increase in heterotrophic bacteria in planted soils as compared to

bulk soil over a 7-week period [9].

P. Pokethitiyook

guarino@unisannio.it

A symbiotic relationship, based on evolutionary signicance, between plants and

the microorganisms play a key role in the degradation process. Root exudates from

plants, besides directly degrading contaminants, act as substrate for soil microor-

ganisms thereby enhancing microbial activity that results in increased rate of bio-

degradation. On the other hand biotransformation of contaminants by microorganisms

helps detoxify chemicals that may be deleterious to the plant itself [25, 30].

4.5 Phytoremediation ofPetroleum-Contaminated Soil

It has long been recognized that plants can remove metal contaminants from the soil

and water during the last three decades. However, the abilities of plants to tolerate

and degrade petroleum hydrocarbons are rather new. It has opened up possibilities

for researchers to explore deeply into more details on how to use plants more effec-

tively for remediation of TPH-contaminated soil [31, 32]. Plants vegetated in con-

taminated soil can uptake small quantity of TPHs and accumulate them in the root

and shoot parts [33, 34]. Once inside the plant, these TPHs may have multiple fates;

some TPH compounds can be sequestered in root tissue, some can be transported

into shoots and leaves, where they can be stored in the vacuole or volatilized into the

surroundings [33, 35]. The metabolic processing to clean up any xenobiotic com-

pounds by plants was compared to mammalian liver and was termed the “Green

Liver” by [36]. The green liver treats xenobiotics into three phases: transformation

by enzyme activities (Phase I), conjugation to form moieties of conjugates (Phase

II), and storage of the nal products in the vacuoles (Phase III). However, Phase III

in mammals the conjugates are excreted in urine or feces instead of being stored [36].

Considering this ability of plants to store toxic chemicals, one can therefore use

plants as the sink for chemical hazard materials.

Soil is normally heterogeneous and form from weathered bedrock. The way in

which TPH compounds partition differently among the different soil horizons depends

on their individual constituents and the disposal sites. Some TPHs cannot move con-

siderably into plants from soils due to partitioning coefcient of that substance. How

much TPHs can dissolve into water before being transported into plants depends on

the n-octanol/water partition coefcient, Kow. The higher the Kow, the more nonpolar

the compound is. Log Kow is generally used as a relative indicator of the tendency of

an organic compound to adsorb to soil. Log Kow values are generally inversely related

to aqueous solubility and directly proportional to molecular weight. Many TPHs can-

not move considerably into plants from the soil when log Kow>4 [37].

Vegetation growing on a soil can signicantly affect many of these characteris-

tics and responses. Depending on the nutrient sources, for example, plant roots can

make the soil near them either more acidic or more alkaline than the soil at a dis-

tance from the root. This is because the root exchanges anions or cations with the

soil as part of the root’s uptake of essential plant nutrients [38]. Smiley [38] mea-

sured the rhizosphere pH (pHr) of eld and container-grown wheat plants and com-

pared it with the non-rhizosphere pH (pHb). The pHr, was generally lower than pHb

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

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when ammonium was supplied as a fertilizer, higher when nitrate was supplied, and

remained relatively unchanged when both forms were added together. In the rhizo-

sphere, plants support TPH-degrading microbes involving in the biodegradation of

TPHs [39–42]. A number of plant and grass species were studied for the abilities to

tolerate and remove TPHs from the soil. Among them, alfalfa (Medicago sativa var.

Harpe), ryegrass (Lolium multiorum L.), birdsfoot trefoil (Lotus corniculatus var.

Leo), sorghum (Sorghum bicolor L.), maize (Zea mays L.), Bermuda grass (Cynodon

dactylon L.), legumes, and beggar ticks (Bidens cernua L.) have been shown to be

suitable for the TPH removal [3, 43–45]. All of these plant species have one thing in

common the brous root system. Plants used for the cleanup of soil contaminated

with TPHs should enclose the ability to tolerate their high concentrations and pos-

sess the extensive root system.

The previous above studies have been focused on the use of grasses for the remedia-

tion of TPHs due to their ability to tolerate the high concentration of the TPHs, exten-

sive brous root system, large root surface area, and deep penetration of the root system

into the soil [21, 46]. Processes and mechanisms taking place in the areas surrounding

the roots provide an ideal environment for TPHs degradation. These processes include

the exchange of gases, provision of water, and the increase in the bioavailability of

TPHs by decreasing the surface and volume of soil micropores [47, 48]. There was also

the enhancement of bacterial population, diversity, and some activities. Overall activi-

ties consequently in favor to TPH biodegradation [7].

Apparently, plants enhance soil microbial population and activity through the

release of organic compounds, e.g., amino acids, sugars, enzymes, organic acids,

and carbohydrates, or the so-called Root Exudates, from the root system [49, 50].

Several compounds released by roots act as inducer for microbial genes or co-

metabolite involving in TPH biodegradation [51, 52]. It was reported that root

exudates supported the development of high diversity of bacteria containing known

hydrocarbon-degrading genes [53]. Since considerably higher numbers and diver-

sity of HC-degrading bacteria were observed in rhizosphere soil as compared to

bulk soil (the fungal abundance is 10–20 times higher and the bacterial abundance

2–20 times higher) [54–56]. Therefore, enhanced phytoremediation of TPH-

contaminated area might be due to an increase in the population and activities of

TPH-degrading bacteria in the rhizosphere [34, 57, 58].

4.6 Phytoremediation ofPetroleum-Contaminated Soil

inAssociation withSoil Bacteria

The remediation of soils containing diverse organic pollutants, including organic

solvents, pesticides, and petroleum, is possible with the use of plants and their rhi-

zosphere processes or the so-called phytodegradation. Phytodegradation of petro-

leum hydrocarbons may be enhanced by bacterial activities. In this process, plants

interact with soil microorganisms by providing nutrients in the rhizosphere which

leads to an increased microbial activity and degradation of organic pollutants.

P. Pokethitiyook

guarino@unisannio.it

A narrow zone of soil affected by the presence of plant roots is dened as rhizo-

sphere [50]. The rhizosphere is known to be a center of microbial activities. This is due

to an increase in nutrient supply for microbes by the release of some extracellular

organic compounds from the root system, namely, exudates and mucilage [59].

Therefore, rhizosphere is a soil matrix with a high microbial diversity resulting in a high

microbial diversity. This microbial activity in turn affects the root development and

plant growth in general. In general, the microbes serve as mediator between the plant

and the soil. Since, in general, plant requires soluble mineral nutrients but often soil

contains the necessary nutrients in low concentrations and in complex and inaccessible

forms. Thus rhizosphere microorganisms, as a mediator, can provide a critical link

between plants and soil plus organic compounds attached to soil (Fig. 4.2, [50, 60]).

4.6.1 Plant Growth-Promoting Rhizobacteria (PGPR)

Rhizosphere bacteria continuously metabolize various organic compounds from root

exudates. As a result, there are quantitative and qualitative alterations of the released

root exudates. Bacteria in the rhizosphere can signicantly inuence the nutrient

supply of plants by competing for mineral nutrients and by mediating the turnover

and mineralization of organic compounds. Therefore, bacteria in the rhizosphere can

be a leading control of the turnover of nutrients in the soil [61]. Rhizosphere bacteria

can inuence plant growth also directly by releasing a variety of compounds, e.g.,

phytohormones or antimicrobial compounds [62] or biofertilizers [63].

ROOT SOIL

Bacteria

Mycorrhizal fungi

increased nutrient

supply for

microorganisms

RHIZOSPHERE

Root exudates

Organics

organic

compounds

(e.g., exudates

and mucilage)

• layer of soil (~1-5 mm) surrounding the root

• r/s ratios: fungi 10-20, bacteria 2-20

• high diversity and activity of

microorganisms

Fig. 4.2 Rhizosphere microorganisms as a critical link between plants and soil (adapted from

Hrynkiewicz and Baum [50])

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

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Plant growth-promoting rhizobacteria (PGPR) represent a wide variety of soil

bacteria and can be plant specic but varies over time [64, 65]. Diversity of bacteria

is affected by the plant age, the season, and the soil conditions [66]. For a long

period, PGPR were largely applied in agriculture for facilitating plants to uptake

nutrients from the environment or preventing plant diseases [67]. The combined use

of plants and pollutant degrading and/or PGPR is relatively a new concept in

the eld of bioremediation of contaminated soil and water [46, 57, 67, 68].

PGPR can increase the availability of nutrients to plants by enzymatically nutri-

ent mobilization from organic matters or soil and the production of siderophores

[69, 70]. Some rhizosphere bacteria also produce siderophores which can be

absorbed as the bacterial Fe3+-siderophore complex by a number of plant species in

the deciency of iron [63]. Microbial siderophores in the rhizosphere can signi-

cantly contribute to the biocontrol of soil-borne pathogens due to their competitive

effects [71] and the mobilization of metals to plants [70].

Some PGPR are acting as biofertilizers. Biofertilizer is dened as a substance

which contains living microorganisms which, when applied to seed, plant surfaces,

or soil, colonizes the rhizosphere or the interior of the plant and promotes growth by

increasing the supply or availability of primary nutrients to the host plant. Whether

the existence of a microorganism increases the growth of plants by replacing soil

nutrients or making nutrients more available (by solubilization of phosphates) or

increasing plant access to nutrients (by increasing root surface area), as long as the

nutrients available to plants have been enhanced by the microorganism, the sub-

stance that was applied to the plant or soil containing the microorganisms, is referred

to here as a biofertilizer [63].

PGPR producing extracellular degrading enzymes are major decomposers of

organic matter. They contribute essentially to the soil aggregation and nutrient avail-

ability [72]. In soils with low phosphate, bacteria facilitate the release of phosphate

ions from low-soluble mineral P crystals and from organic phosphate sources. These

bacteria slowly release organic acids that dissolve the P crystals and exude enzymes

that split organophosphate [63, 73].

PGPR are usually in contact with the root surface and improve growth of plants

by several mechanisms, e.g., enhanced mineral nutrition and disease suppression [73].

PGPR can also promote the root growth. Rhizobacteria produces phytohormones

such as indole-3-acetic acid (IAA), cytokinins, gibberellins, ethylene which pro-

mote cell division and cell enlargement, extension of other morphological changes

of roots [50].

Petroleum hydrocarbon pollutants can be biodegraded by plants through

biochemical reactions taking place within the plants and in the rhizosphere. The

remediation of soils containing diverse TPHs, including crude oil, fuel oil, and lube

oil, is possible with the use of plants and their rhizosphere processes (phytodegrada-

tion) [50]. Phytodegradation of organic pollutants may be enhanced by bacterial

activities. In this process, plants interact with PGPR by providing nutrients in the

rhizosphere which leads to an increased microbial activity and degradation of TPHs

as describe earlier.

P. Pokethitiyook

guarino@unisannio.it

Endophytic bacteria, bacteria colonizing healthy plant tissue intercellularly and/

or intracellularly without causing any apparent symptoms of disease, can produce

extracellular enzymes, including pectinase, cellulose, lipoidase, proteinase, pheno-

loxidase, and lignin catabolic enzymes. All these enzymes are necessary to pene-

trate and colonize the host plants. Degradation of organic pollutants also occurred

by rhizosphere and/or endophytic bacteria [74].

The investigation of site-adapted cultivable microorganisms in unfavorable or

contaminated soils will contribute to identify the site-specic microbial populations

and to provide fundamental knowledge and strain collections for subsequent selec-

tions and applications of plant growth and site remediation promoting microbial

strains.

4.6.2 Microorganisms andIts Selection forthePromotion

ofPlant Growth andSoil Bioremediation

Applications of bacterial inoculation provide a great challenge in the future to

increase plant growth and remediate contaminated soils. PGPR bacteria are the

most important group capable of improving phytoremediation of petroleum hydro-

carbons contaminated soil [21, 47, 75]. They are ubiquitous in the environment and

play an important role in biodegradation of TPH contaminants from the soil, water,

and air [5, 44, 76]. However, several obstacles must be overcome to achieve the suc-

cessful applications of such treatments. Theoretically, microbial inoculum should

be relatively universal for various plants and soils and its effectiveness should be

relatively easy to evaluate on a standard scale. Practically, many experiments were

plant-specicity and soil-specicity instead of being universal [63].

Information on the diversity of microorganisms at polluted sites is supposed to

be valuable for a future selection of microbial inoculum for those sites. Information

on microbial diversity and activity may not only provide evidence of ecosystem

degradation but it might also be a valuable source of information for future applica-

tion as inoculums for PGPR bacteria. Some molecular techniques, e.g., denaturating

gradient gel electrophoresis (DGGE) and terminal restriction fragment length

polymorphism (T-RFLP) can provide detailed information on the taxonomic and

phylogenetic relationships of bacteria found on the contaminated sites. This infor-

mation can describe the co-evolution between plants and bacteria in the eld [77].

For PGPR to have a benecial effect on plant growth through an enhancement of

the nutrient status of their host, there obviously needs to be an intimate relationship

between the PGPR and the host plant. However, the degree of intimacy between the

PGPR and the host plant can vary depending on where and how the PGPR colonizes

the host plant. As rhizobacteria themselves can be categorized into two groups:

(1) rhizospheric and (2) endophytic bacteria [63].

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

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Soil microorganisms (95–99%) is known to be at least so far nonculturable [78].

However, the basic criterion for the selection and application of this microbial inoc-

ulum useful for plant-growth promotion is cultivable and easily multiplication of

bacteria. Information of critical factors inuencing plant–microbe interactions with

TPHs in soils could lead to an improved selection of bacterial inoculum for a

bacterial- assisted phytoremediation of TPHs. A fundamental basis for the subse-

quent on-site applications of selected microorganisms is their safety for the environ-

ment and humans. Therefore, before eld applications, all selected microorganisms

have to be precisely identied and toxicologically assessed. Few microbial taxa

have been reported so far for their capability to promote plant growth at contami-

nated soils. Also little is known on the microbial diversity which might be relevant

to promote plant growth in those soils. In general, numerous species of soil bacteria

which inhabit the rhizosphere can promote plant growth [63], e.g., by enzymatic

nutrient mobilization from soil or organic matter (mostly P and N) and production

of siderophores [70]. The study on barley (Hordeum vulgare L.) have uncovered

that PGPR can contribute essentially to soil aggregation and nutrient availability

which is often important for contaminated soils [72]. Therefore, enzyme activities

can be suitable selection criteria for microbial inoculum for plant growth promotion

in disturbed soils.

Microbial enzyme activities in the soil were predominantly measured as total

potential activities rather than at the level of isolates within a community. As a mat-

ter of facts, investigations of single strains are also necessary for the selection of

potential inoculum [66]. Acid phosphatases contribute to the P mobilization from

organic matter. These enzymes cause the release of phosphate from a variety of

substrates as inositol phosphate, polyphosphates, and phosphorylated sugars into

the soil solution [79]. The production of these enzymes is species- and strain-

dependent and often stimulated by deciency of mineral phosphate. Beside the

phosphatase activity, cellulolytic and pectolytic activities have been used for selec-

tion of microorganisms for promotion of plant growth and mycorrhiza formation.

High cellulolytic and pectolytic activities of mycorrhizal fungi and rhizosphere

bacteria allow the disintegration of living and dead plant tissue and, consequently,

can enable microorganisms to enter roots. High cellulolytic and pectolytic activities

can be used as a distinguish factor in selecting the rhizobacteria.

In many rhizospheric relationships, the PGPR actually attached to the surface of

the plant. Scanning electron micrograph of bacteria on the surface of plants roots is

a good scientic tool for proof checking of the microbial existence [63, 80]. In

endophytic relationships, microorganisms actually reside within apoplastic spaces

inside the host plant. Although there is rare evidence of endophytes occupying intra-

cellular spaces [81].

Quantitative PCR (qPCR) has been used to monitor the presence of specic

HC-degrading bacteria in any environment (e.g., by looking at the abundance of

alkB gene in the rhizosphere and endosphere of plants growing in TPH-contaminated

soil) and to monitor dened functional activity (e.g., alkane-degrading alkB gene

P. Pokethitiyook

guarino@unisannio.it

expression during phytoremediation of HC-contaminated soil). Studies regarding

the abundance and expression of alkB and CYP153 genes in rhizosphere, and shoot

and root interior of plants vegetated in HC-contaminated soil indicated that bacteria

carrying these genes are not only able to colonize the rhizosphere and plant interior

but are also metabolically active in HC degradation [21, 82–84]. Moreover, these

studies concluded that survival and metabolic activities of HC-degrading bacteria

varied distinctly between different strains, plants species, plant development stages,

and plant compartments. Greater numbers of HC-degrading bacteria possessing

alkB and tol genes were also found at the vegetative growth stages of ryegrass

(Lolium multiorum L.) [8].

It has been concluded that the use of bacteria with both pollutant degrading as

well as plant growth-promoting properties worked better than using the bacteria

having either pollutant degrading or plant growth-promoting properties only. PGPR

showing 1-aminocyclopropane-1-carboxylate (ACC) deaminase activity can dec-

rease ethylene amounts produced by plants under stress and consequently reduce

stress symptoms leading to improved plant growth and development [84, 85].

Inoculation of plants with bacteria possessing both HC-degrading and plant

growth-promoting activities, with both seed inoculation and soil method, has been

successfully applied in the laboratory, greenhouse, and eld for their mediation of

HC-contaminated soil and water [46, 75, 82, 83]. Most PGPR promote plant growth

through their ability to x N2 in situ [63]. A list of these PGPR is shown in Table 4.4.

Some recent successful examples of rhizobacteria application for the phytoreme-

diation of petroleum hydrocarbon-contaminated soil are shown in Table 4.5. The

potential benets of using genetic engineered bacteria to improve recalcitrant

organic pollutants biodegradation are summarized by Newman and Reynolds [86].

Field trials will need to be done to determine if this advantage remains stable in the

eld-grown plants. Moreover, the concept of releasing engineered bacteria into the

environment must be addressed and monitor with a rigorous surveillance program.

Although these organisms have been transformed using naturally occurring bacte-

rial genes, their function in the host system might be different or distorted.

Table 4.4 Plant growth-promoting rhizobacteria (PGPR) for which evidence exists that their

stimulation of plant growth is related to their ability to x N2 (Vessey [63])

PGPR Relationship to host Host crops

Azospirillum sp. Rhizospheric Maize, rice, wheat

Azoarcus sp. Endophytic Kallar grass, sorghum, rice

Azotobacter sp. Rhizospheric Maize, wheat

Bacillus polymyxa Rhizospheric Wheat

Burkholderia sp. Endophytic Rice

CyanobacteriaaRhizospheric Rice, wheat

Gluconacetobacter diazotrophicus Endophytic Sorghum, sugarcane

Herbaspirillum sp. Endophytic Rice, sorghum, sugarcane

aNumerous species; predominantly of the genera Anabaena and Nostoc

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

guarino@unisannio.it

Table 4.5 Examples of successful remediation of hydrocarbons from soil by combined use of plants and rhizobacteria

Plant used Rhizobacteria Bacterial characteristics Reference

Maize (Z. mays L.) Pseudomonas sp. UG14Lr Hydrocarbon degradation Chouychai etal. [88]

Italian ryegrass (L.

multiorum var. Taurus)

Pantoea sp. strain BTRH79 Hydrocarbon degradation

and ACC deaminase

activity

Afzal etal. [46]

Alfalfa (M. sativa L.) R. meliloti (strain ACCC17519) Hydrocarbon degradation Teng etal. [20]

Maize (Z. mays L.) Rhizobacterium, Gordonia sp.

S2RP-17

Hydrocarbon degradation,

ACC deaminase, and

siderophore synthesizing

activities

Hong etal. [89]

Sorghum (S. bicolor)Sinorhizobium meliloti Hydrocarbon degradation,

auxin production

Golubev etal. [90]

Ryegrass (L. multiorum)Acinetobacter sp. strain Hydrocarbon degradation Yu etal. [91]

Italian rye grass (L.

multiorum var. Taurus)

and birdsfoot trefoil (L.

corniculatus var. Leo)

Pantoea sp. strain BTRH79,

Pseudomonas sp. strain

ITRH76

Hydrocarbon degradation Yousaf etal. [44]

Winter rye (Secale cereale

L.), alfalfa (M. sativa L.)

Azospirillum brasilense SR80 Hydrocarbon degradation,

indole-3-acetic acid (IAA)

production

Muratova etal. [92]

Italian ryegrass (L.

multiorum var. Taurus)

Rhodococcus sp.strainITRH43 Hydrocarbon degradation Andria etal. [93]

Sorghum (S. bicolor

L.Moench)

S. meliloti P221 Phenanthrene degradation,

indole-3

Muratova etal. [94]

Acetic acid (IAA)

production

Maize (Z. mays L.) P. putida MUB1 Hydrocarbon degradation Chouychai etal. [19]

P. Pokethitiyook

guarino@unisannio.it

Annual ryegrass (L.

perenne), tall fescue (F.

arundinacea var. Inferno),

barley (Hordeumvulgare)

Pseudomonas strains, UW3

and UW4

ACC deaminase production Gurska etal. [75]

Rice (Oryza sativa L.) Acinetobacteria sp. Hydrocarbon degradation Li etal. [95]

Barley (H. sativum L.) P. putida KT2440 Hydrocarbon degradation Child etal. [96]

Barley (H. sativum L.) Mycobacterium sp. strain KMS Hydrocarbon degradation Child etal. [97]

Wheat (Triticum spp.) Pseudomonas sp.GF3 Phenanthrene degration Sheng and Gong [98]

Wheat (Triticum spp.) A. lipoferum sp. Hydrocarbon degradation

and indole-3

Muratova etal. [99]

Acetic acid (IAA)

production

Common reed (P. australis)P. asplenii AC ACC deaminase production Reed and Glick [100]

White Clover (T. repens L.) R.. leguminosarum Hydrocarbon degradation Johnson etal. [101]

Tall fescue grass

(F. arundinacea)

A. brasilense Cd, Enterobacter

cloacae CAL2, P. putida UW3,

P. putida, Flavobacterium sp.,

P. aeruginosa

Hydrocarbon degradation

and ACC deaminase

activities

Huang etal. [82]

Barley (H. sativum L.) P. uorescens, P. aureofaciens Hydrocarbon degradation Anokhina etal. [102]

Barmultra grass

(L. multiorum)

P. putida PCL1444 Naphthalene- degrading

bacteria

Kuiper etal. [103]

4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

guarino@unisannio.it

4.7 Conclusion

The use of rhizosphere bacteria contributes signicantly to the improvement and

sustainability of agriculture and agroforestry as well as the phytoremediation of

organic contaminated soils. Selection and promotion of desirable rhizospheric pro-

cesses requires a fundamental understanding of the complex microbial interactions

in the rhizosphere. Rhizobacteria belong to the microorganisms in the rhizosphere,

which contribute essentially to increase the soil fertility and remediate chemically

contaminated soils.

Inoculation of soils with selected plant growth and soil remediation PGPR has the

capacity to improve the plant tness in polluted soils with unfavorable conditions and

increase the biodegradation of organic pollutants. The successful use of such inocu-

lum in the contaminated eld with natural environmental conditions and competition

will be a great challenge. In this regard, for efcient petroleum hydrocarbon remedia-

tion, it is of primary importance that the inoculated hydrocarbon- degrading bacteria

colonize the rhizosphere and/or plant cells so as to initiate their effects on plant growth

and hydrocarbon biodegradation. Rhizobacteria and endophytic bacteria showing

hydrocarbon degradation capacity and/or plant growth- promoting ACC deaminase

activity are more effective in petroleum hydrocarbon phytoremediation since they are

enhancing plant growth and simultaneously encouraging hydrocarbon degradation.

The bacterial ACC deaminase activity accelerates root growth, as a result a better

access to nutrients and water and consequently faster initial growth, which enable

plants to better counteract stress responses caused by hydrocarbon contamination.

At present, it seems necessary to use always site-specic selections of inoculum

since a general suitability of inoculum for diverse site conditions seems rather

unlikely. Combined use of plant and effective and specic rhizobacteria seems to be

a more promising technique for the remediation of petroleum hydrocarbon-

contaminated soil as compared to only bioaugmentation (only use of microorgan-

isms) and phytoremediation (only use of plants).

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4 Phytoremediation ofPetroleum-Contaminated Soil inAssociation withSoil Bacteria

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Part II

Higher Plants in Biomonitoring

and Environmental Bioremediation

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103© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_5

Chapter 5

The Use of Higher Plants in Biomonitoring

and Environmental Bioremediation

Svetlana Vladimirovna Gorelova and Marina Vladimirovna Frontasyeva

Abstract This chapter provides basic information on the use of higher plants for

biomonitoring and bioremediation in the world. It contains a large amount of mate-

rial of the authors’ own research on the possibility of using woody plants for bio-

monitoring and phytoremediation of environment anthropogenic pollution with

heavy metals. The species of woody plants are revealed, which are recommended

for use in biomonitoring of anthropogenic pollution of the environment in temperate

latitudes (the study of biogeochemical parameters of leaves): Acer platanoides,

Aesculus hippocastanum, Betula pendula, Cotoneaster lucidus, Populus nigra, and

Salix fragilis. The following species are recommended for phytoremediation of

soils from heavy metals: Betula pendula, Cotoneaster lucidus, Syringa vulgaris,

Sorbus aucuparia, Philadelphus coronarius, and Larix sibirica. The species of

woody plants—bioindicators of air and soil pollution by heavy metals—are

revealed. The chapter also shows the signicance of the statistical analysis for the

detection of the main element pollutants of the environment.

Keywords Biomonitoring • Bioremediation • Heavy metals • Woody plants •

Bioaccumulation • Soil and air pollution • Statistical analysis

S.V. Gorelova, Ph.D. (*)

Institute of Advanced Training and Professional Retraining of Education Employees

of Tula region, Lenina Street, Tula 300041, Russia

e-mail: salix35@gmail.com

M.V. Frontasyeva, Ph.D.

Joint Institute for Nuclear Research, Frank Laboratory of Neutron Physics,

Sector of Neutron Activation Analysis and Applied Research,

Str. Joliot-Curie, 6, Dubna, Moscow Region 149080, Moscow, Russia

e-mail: marina@nf.jinr.ru

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104

5.1 The Use of Higher Plants for Biomonitoring

(Basic Information)

Biomonitoring is a system of control and obtaining quantitative characteristics of

biological objects (biomonitors) in time for the assessment of the environmental

changes. The main objective of biomonitoring is to prevent the adverse effects of

sequences of the environmental changes and forecasting of developments of events

at the level of individual populations as well as of biogeocenosis and the biosphere

as a whole.

Monitoring can be carried out at different levels: molecular, tissue, organ, organ-

ism, population, species, ecosystem, and biosphere. Depending on the venue, it can

be local, regional, national, and international. It may also vary according to the

objects of research and may dene the parameters, being passive (carried out

directly in the wild nature) or active (requires setting of the experiment by a

researcher) [1–5].

To conduct biomonitoring, the major role belongs by the choice of the environ-

mental monitors (markers) of the environmental state, the development of uniform

methods of sampling, sample preparation, and the proper selection of analytical

methods for different types of contaminants. Due to the fact that the spectrum of the

environmental pollutants comprises more than 400,000 items, use of chemical

methods of analysis only is too costly, and it does not allow to get the whole picture

of their cumulative impacts on biota; so it is more cheaper to apply methods of bio-

indication, biological testing, and biomonitoring. However, none of biological

object may be a universal indicator or monitor sensitive to various substances to the

same degree.

Basic requirements for plant biomonitors are summarized as follows [2, 3]:

1. Widespread and long vegetation period and high degree of bioaccumulation of

elements of the environment (passive biomonitoring), the ability to good growth

in standardized conditions (active biomonitoring)

2. A clearly marked and reproducible response to certain changes in the environ-

ment and bioaccumulation of toxic elements in an amount reecting the situation

in the environment

3. High sensitivity to pollutants (diagnosis effect at low levels of contamination)

There are no universal biomonitors that meet the requirements with respect to all

possible contaminants; therefore, an important task for biomonitoring is the selec-

tion of species that can be used for biomonitoring of various parameters of the

environment.

Markers for biomonitoring may be molecular mechanisms: the study of the struc-

ture of DNA changes, the genetic response, and synthesis of substances [6].

Biomonitoring can be carried out at a biochemical and physiological level: determina-

tion of the content of low molecular antioxidants, involved in detoxication mecha-

nisms when the radicals produced under stress. Such antioxidants are ascorbic acid,

glutathione and proline [7–13]. The stress level can be determined by the change of

activity of antioxidant enzymes superoxide dismutase, peroxidases, catalase, and

S.V. Gorelova and M.V. Frontasyeva

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glutathione reductase [14–18]. However, these plant reactions are not always specic

to certain toxicants and depend on the species of the plant [19, 20].

The most commonly used is an identication sign of quantication of chloro-

phylls and carotenoids in plants [21–23], as well as their ratio and the response of

the light phase of photosynthesis [24–26]. At the level of organelles, membrane

structure, chloroplasts, and mitochondria (transmission electron microscopy, TEM)

are known [27]. When studying the plants at the tissue level, histochemical meth-

ods are applied using dyes that are specic to a particular metal, which helps to

determine the localization in the tissue and way of their movement and accumula-

tion in the plant [28, 29]. As biomarkers in model experiments оn determination of

the effects of various concentrations of toxic substances in the environment on the

plant, individual organs of plants can be used, where the biomass growth (shoots),

parameters such as germination and vigor (seeds), the root test, and denition of

tolerance index are studied [30–33]. In passive monitoring at the organ level, the

development and percentage of leaf necrosis and chlorosis, development of defor-

mation of shoots and leaves, and modications of the leaf blade (the appearance of

the blades in simple leaves, the absence of leaf share, threadlike leaves, etc.) are

determined. Besides, one can determine the leaf square, the degree of xeromor-

phism, and determine the percentage of dead shoots and dry crown of trees [34, 35].

At the organism level, vitality of species in the altered environmental conditions is

determined [21, 36, 37]. However, when it comes to polymetallic pollution of the

environment with heavy metals and metalloids, most signicant is determination of

elemental (biogeochemical) composition of plants and plant organs, which may

reect the situation in the environment [36], if the plant is an indicator [37–40]: it

adsorbs and bioaccumulates metals in the process of growth, develops mechanisms

of resistance to toxic elements, and does not belong to excluders (which exclude)

or bioaccumulators in accordance with the classication proposed by AJM Baker

[41, 42].

From the plant physiology point of view, biomarker of the pollution stress effect

at the level of phytocoenosis, to some extent could be chlorophyll uorescence [24,

25]. For biomonitoring of ecosystems, geobotanical methods are also applied: anal-

ysis of the number and types of species and their vitality, crown density, and density

of herbaceous (or moss-lichen) cover and the analysis of the presence of anthropo-

genic weeds in phytocenosis, which makes it possible to conclude about the degree

of digression of the community.

According to their response to the content of toxic components in the environment,

bioindicators and biomonitors may be sensitive (respond to the impact of a signicant

deviation from the norm) or bioaccumulative (feedback manifests itself gradually, and

pollutant accumulates in the body or individual organs and tissues) [1–3].

Most often to biomonitor atmospheric deposition the higher spore plants – mosses

– are used. The idea of using terrestrial mosses for the analysis of atmospheric deposi-

tion of heavy metals has been proposed in the late 1960s of the twentieth century by

Rühling and Tyler [43–45]. It is based on features of moss anatomic structure and

physiology. The leaves of moss are composed of 1–3 layers of cells, they lack cuticles

on the leaves preventing the penetration of pollutants, they have no roots, and they

readily absorb water and nutrients from wet and dry deposition by rhizoids.

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Mosses effectively accumulate heavy metals and other compounds due to the

large specic surface area and slow growth. As a passive biomonitor in most cases,

they help to identify the impact of pollutants at the ecosystem level. Ideas of moss

monitoring in Europe have been developed by Rühling etal. [46, 47], Steinnes [48,

49], Steinnes and Andersson [50], Steinnes etal. [51], Steinnes and Frontasyeva

[52], Rühling and Steinnes [53], Berg and Steines [54], Schröder etal. [55], and

Harmens etal. [56–60].

Since the 1970s, in the Scandinavian countries, and in the last 20 years in the

Eastern, Central, and Western Europe, passive briomonitoring receives support of

targeted state grants and programs, and it is held regularly every 5 years in the

framework of the UN Convention on Long-Range Transboundary Air Pollution

(LRTAP) [53, 58, 60–64]. Coordination of moss biomonitoring in Europe, Russia,

and Asia is carried out through the United Nations program (UNECE ICP Vegetation).

Based on the monitoring results, the atlases of atmospheric deposition of pollut-

ants are edited and published, which allow estimating the cross-border transfer of

elements, reveal sources of pollution and their impact on the environment, as well

as trace the retrospective distribution of elements in the atmosphere [58, 60, 61, 65].

In Russia, conducting biomonitoring rst started in the northwestern regions:

Leningrad region [66, 67], Kola Peninsula, and Karelia [54, 68]. Since the late

1990s of the twentieth century, biomonitoring was carried out on the basis of the

analytical complex of the Joint Institute for Nuclear Research for a number of cen-

tral regions of Russia: one-time study conducted in Tula region [69, 70], Tver,

Kostroma, part of Moscow and Ivanovo regions [71–74], Ural [75, 76], Udmurtia

[77], as well as Kaliningrad region [78–82].

In 2014 the coordination of the moss surveys in the UNECE ICP Vegetation has

been transferred from the UK to Russia, Joint Institute for Nuclear Research (Dubna,

Moscow Region) to M. V. Frontasyeva, so far JINR has direct access to the member-

states in which the UNECE ICP Vegetation is interested in the Caucasus region and

Asia: Azerbaijan, Georgia, Kazakhstan, Mongolia, Vietnam, and Moldova in the

southern east. Currently, the study area of atmospheric deposition by passive bio-

monitoring greatly increased, and GIS mapping and transport models build on the

data submitted by teams from different countries; it will be possible to make more

global conclusions on the transboundary transport of substances (2015–2016 moss

survey) and to create a database on the content of elements in mosses on a global

scale, which can be replenished in the future [83]. In addition to atmospheric depo-

sition of heavy metals, this method also allows evaluating the contamination of

nitrogen, persistent organic pollutants (POPs), and radionuclides [58, 59, 83–87].

Besides higher spore plants for biomonitoring, woody and herbaceous plants of

genera Gimnospermae and Angiospermae can be used. They reect the state of soil,

air and water may due to change of their biochemical, physiological, and morpho-

logical parameters and their ability to bioaccumulate the toxic elements from the

environment [21–23, 34, 35, 88–105].

Using the higher seed plants for biomonitoring purposes has advantages over the

use of spore plants: they are easily identied (e.g., than mosses) and grow in urban

ecosystems, and some of them have extensive habitat areals and can be used for the

diagnostics of transboundary transport of elements between countries and continents.

S.V. Gorelova and M.V. Frontasyeva

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Table 5.1 Higher plants for bioindication and biomonitoring of the environment (Applied Ecobiotechnology [106]; Gorelova [107]; Gorelova etal. [97, 104, 105])

Species Used in bioindication, symptoms Used in biomonitoring

Substance to which

the given type reacts

Taráxacum ofcinále +, necrosis and chlorosis of leaves, the

change of physiological parameters

(photosynthetic pigments content of low

molecular weight antioxidants)

+, accumulation of heavy metals in the body

depending on the degree of pollution

Heavy metals, ozone

Gladiolus gandavensis, Tulipa

gesneriana, Iris germanica,

Petroselinum crispum

+, regional and apical necrosis,

accumulation of uorine in the dry matter

−HF

Urtica urens +, necrosis strips on the underside of leaves −Peroxyacetyl

Nicotiana tabacum, Spinacia oleracea,

Glycine max, Trifolium pratense,

Trifolium angustifolium and subsp.

+, necrotic changes in leaves (spots),

pink spots on the leaves (reaction on

pollution by ozone), interveinal necrosis

(reaction on nitrogen oxides)

+ O3, NO2

Poa annua +, necrosis strips on the leaves + NO3−

Medicago sativa, Fagopyrum

esculentum, Plantago major, Pisum

sativum, Trifolium incarnatum, Pinus

sylvestris, Quercus subsp., Platanus

subsp., Populus subsp., Acer subsp.,

Fagus sylvatica, Fraxinus excelsior

+, interveinal point and extensive necrosis

and chlorosis. Marginal necrosis and

chlorosis, early death of needles and dry

crown dieback (Pinus)

+ SO2

Spinacia oleracea, Phaseolus vulgaris,

Latuca sativa

+, chlorosis, change the structure of

chloroplasts, photosynthesis violation

Cl2

Lepidum sativum +, reduction of biomass, germination and

growth energy, dying roots

NaCl, heavy metals

Pleurozium schreberi, Hylocomium

splendens, Polytrichum sp.,

Brachythecium sp., Hypnum

cupressiforme, Sphagnum subsp.

+, accumulation of atmospheric deposition Radionuclides Sr90,

Cs137, K40

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Species Used in bioindication, symptoms Used in biomonitoring

Substance to which

the given type reacts

Pleurozium schreberi, Hylocomium

splendens, Polytrichum subsp.,

Brachythecium subsp., Hypnum

cupressiforme, Sphagnum subsp.

+, necrosis at high concentrations of heavy

metals in the environment

+, accumulation in organs from atmospheric

deposition

Heavy metals and

metalloids

Fagus sylvatica, Ailanthus altissima,

Robinia pseudoacacia, Picea abies

needle chlorosis, the appearance

of dead branches in the crown

+, accumulation of toxic elements in leaves Heavy metal

Aesculus hippocastanum +, regional and interveinal necrosis +, accumulation of Cu, Ni, Pb. Reduction of amount

of photosynthetic pigments, decrease of ascorbic acid

content, increased synthesis of phenolic compounds

Heavy metals

Betula pendula +, asymmetry of leaves, necrosis +, Mn, Ni, Zn, Cd, Pb, reduction of the amount of

photosynthetic pigments, decreased ascorbic acid

content, synthesis of phenolic compounds

Heavy metals

Populus nigra +, decrease in vitality, necrosis of leaves,

early defoliation, dry crowns

+, Fe, Ni, Zn, Cd, Pb, reduction of the amount of

photosynthetic pigments, decrease of ascorbic acid

content, synthesis of phenolic compounds

Heavy metal

Tilia cordata +, regional and focal necrosis +, reduction of photosynthetic pigments, reduction of

ascorbic acid, increased synthesis of phenolic compounds

SO2, NO2, Na Cl,

heavy metals

Acer platanoides +, regional and focal necrosis, damage by

fungi

+, accumulation of Fe, Mn reduction of

photosynthetic pigments, reduction of ascorbic acid,

increased synthesis of phenolic compounds

Heavy metals

Quercus sp. +, necrosis of leaves, reduced vitality,

insect damage

+, accumulation of heavy metals Complex (SO2, NO2,

heavy metals)

Thuja occidentalis +, death of needles, dying of shoots,

reduced vitality, loss of landings

+, accumulation of heavy metals V, Cr, Fe, Ni, As,

and Mo

SO2, NO2, Na Cl,

heavy metals

Juniperus scopulorum Skyrocket +, death of needles and shoots, reduced

vitality

+, accumulation V, Cr, Fe, Ni, and Sb SO2, NO2, Na Cl,

heavy metals

Taxus baccata +, dying of shoots, reduced vitality, loss of

landings

+, accumulation of Cr, Fe, Ni, Zn, As, and Cd in

shoots

SO2, NO2, Na Cl,

heavy metals

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The list of species used for biomonitoring and bioindication purposes on a global

scale is relatively large. Some of these species are shown in Table 5.1.

5.2 The Use of Higher Plants for Bioremediation

(Basic Information)

5.2.1 Phytoremediation

Phytoremediation is the restoration of ecosystems, or their individual components,

contaminated with heavy metals, radionuclides, NaCl, petroleum ore, and other toxic

organic products by the use of herbaceous and woody plants. The advantage of phy-

toremediation over other methods of purifying ambient environment is its relatively

low cost (50–100 times less), the ability of the implementation in situ, environmental

safety, and the ability of further use of obtained biomass to extract valuable elements

[108]. One of the drawbacks is the time required for full soil recovery.

Selection of plants for phytoremediation is determined by their ability to absorb

toxic compounds from the soil or water systems or transfer them from the surface to

volatile forms, the growth rate, and the volume of produced biomass during vegeta-

tion and depth of root system.

For phytoremediation, plant hyperaccumulators could be used which accumulate

high concentrations of toxic compounds (heavy metals, nonmetals, radionuclides)

in biomass [109, 110]. They have developed mechanisms to adapt to high concen-

trations of metals in organs: representatives of galmain ora (Viola lutea var. cala-

minaria, Thlaspi Zn), “tin ora” Trietaris europea, Gnaphalium suaveolens,

accumulating Ni, Alyssum bertolonii, Sebetaria, copper acuminators Cyanotis

cupricola, Sopubia metallorum, Gypsophila patrinii, and others.

Such plants, as a rule, are usually characterized by low biomass. Lately more and

more attention of scientists is directed to the use of plants with the medium potential

for bioaccumulation of toxic elements, but creating more biomass in the process of

vegetation (e.g., C4 plants and woody plants) [111–116]. So far the feasibility of

using plants-accumulators of heavy metal is determined by the metal accumulation

rate (mg/kg of biomass), multiplied by their biological productivity (kg/ha per

year). Economically justiable plants for phytoremediation are those in which the

yield of biomass reaches at least 250kg/ha per year and metal content in biomass of

at least 1% (dry weight) [108].

There are several ways to plant uptake of toxic elements from the environment:

5.2.2 Phytoextraction

Phytoextraction is a process of conversion of heavy metals or metalloids by plants into

the form of complex compounds (chelates) and their accumulation in tissues and

organs (overground or root system) [117].

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To increase the ability of plants to absorb heavy metals from soils, chelators (e.g.,

ethylenediaminetetraacetic acid, citric acid and oxalic acid, malic acid, salicylate,

succinate, tartrate, and other compounds) or inoculation of plants by symbiotrophic

microora (fungi and bacteria), are used [118–120]. Furthermore, the studies are

known where the genes of bacterial cells are introduced to the organism of higher

plants to increase their ability to absorb heavy metals from the substrate.

The effectiveness of phytoextraction is inuenced by several factors: by the content

of humus in the soil (the binding of toxic components into complexes increases; their

availability for plants decreases), activity of soil microorganisms, introduction of sor-

bents (iron oxides, manganese, organics, clay, y ash, coal, vermiculite sawdust, and

others) into soil, the pH value of the soil solution (pH reduction leads to an increase in

mobility of many heavy metals, Cd, Zn, Ni, etc., as well as binding them with organic

soil components), liming (this leads to reduction of the solubility of Fe, Cu, Ni, Co,

Zn, and Cd and their availability for plants), introduction of organic acids and com-

plexing agents (enhances uptake of heavy metals by plants), application of plant

growth stimulators (heteroauxsin, succinic acid, etc.), and interaction between the

ions in the soil solution (formation of insoluble compounds) [106].

5.2.3 Phytotransformation and Phytodegradation

Phytotransformation is an ability of plants to convert toxic compounds (organic

pollutants—xenobiotics) in the process of the plant fermentative enzymatic

reactions to nontoxic form and their subsequent transfer to the vacuole or bind-

ing to lignin and other components of the cell.

5.2.4 Rhizodegradation (Rhizosphere Biodegradation Ore

Phytostimulation)

Rhizodegradation (rhizosphere biodegradation ore phytostimulation) is decomposi-

tion of toxic organic compounds in the soil in the process of enzymatic degradation in

the interaction of the rhizosphere of plants and microorganisms. Thus, the roots of

plants affect xenobiotic root exudates, stimulate the increase of the number of micro-

organisms in the rhizosphere, and accelerate the transfer of toxic compounds in the

root zone due to the difference in osmotic pressure between root cells and the soil

solution [121, 122]. It is used for soil purication from oil products (not more than

2%): the PAH, PCBs, other hydrophobic aromatic compounds, and pesticides [106].

5.2.5 Phytovolatilization

Phytovolatilization is conversion of toxic components into nontoxic volatile

compounds using enzymatic reactions in biochemical cycles of plants and their

subsequent release (selenium, mercury) (Liriodendron tulipifera) [106].

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5.2.6 Phytostabilization

Phytostabilization is transfer of metals into the insoluble stable compounds due

to synthesis and release by plant compounds that reduce the spread of pollutants

(binding to lignin or organic soil components) (conversion into insoluble forms)

[123] and precipitation of heavy metals and metalloids (Cd, Cr, Cu, Hg, Pb, Zn,

andAs) in the root zone in the form of carbonates, phosphates, and hydroxides. It

is used as a step in soil remediation together with the introduction of lime,

organic fertilizers, and structurants (phosphates, synthetic resin, clay, bentonite,

y ash, zeolites, aluminosilicates, hydroxides of Fe, Al, and Mn) [106].

5.2.7 Rhizofiltration

Rhizoltration is absorption, concentration, and precipitation of heavy metals and

hazardous chemicals by plant roots. This is a most often used technology for water

purication from toxic substances and radionuclides. For rhizoltration rafts in

ponds with terrestrial and aquatic plants (in situ) or special tanks for water treatment

with platforms (gratings) for plants (ex situ) are used [124–126].

An example of the integrated use of living organisms for bioremediation of water

is so called “living machines”: a system of tanks for anaerobic treatment, aerobic

treatment using microorganisms and planktonic animals, containers with higher

plant hydrophytes and hygrophytes (Lemna minor, Eichornia crassipes, Phragmites

australis, Typha latifolia, Gliceria uitans, Calla palustris, Alisma plantago-aquat-

ica, Sagittaria spp., and others), through which the contaminated water ows. As a

result, the water is puried from organic and inorganic pollutants.

After rhizoextraction of rhizoltration, the biomass of plants containing metals

can be used for extraction of metals by chemical means (Ni, Cu, Au) or for energy

generation [127–129].

At present, scientists intend to create greenbelts of the plant in the industrial zones,

which serve as a barrier to heavy metal and serve as phytoremediation of the environ-

ment by absorbing heavy metals and radionuclides from the air and soil. The most

promising for this are woody plants which possess a combination of features: a deep

(or surface) root system of a large volume, a large volume tree crown (height 1.5–

30m), ability to accumulate a large biomass of leaves during the growing season, the

possibility of absorption and bioaccumulation of heavy metals (mainly Pb and V) by

leaves from atmospheric deposition, durability, and possibility to use wood.

See Table 5.2 for higher plants v.

Many woody plants meet all the requirements of biomonitors listed by Markert:

• High abundance

• Widespread

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Table 5.2 Higher plants used for phytoremediation (Baker and Brooks [109]; Wenzel etal. [130];

Glass [131]; Palmer et al. [132]; Prasad [108, 133, 134]; Trace elements [135]; Applied

Ecobiotechnology [106]; Favas and Pratas [136])

Species (genera) Accumulated elements Substrate

Acacia dealbata Cu, Pb Soil

Agrostis tenuis Cu, Pb, Zn Soil phytostabilization

Agrostis capillaris Cu, Pb, Zn Soil phytostabilization

Agrostis lanatus Fe, Cu, Zn, As, Pb Soil

Alyssum sp. Ni Soil

Alnus glutinosa Cu, Pb Soil

Amaranthus retroexus,

Amaranthus tricolor

137Cs, Zn Soil

Armeria maritima Pb Soil

Atemisia absinthium Zn, Cu, Cr Soil

Artemisia vulgaris Zn, Ni, Cu Soil

Atriplex prostrata NaCl Soil

Alisma plantago-aquatica,

Calla palustris, Gliceria

uitans, Sagittaria spp

HM, organic

compounds

Storm water ditch, sewage wetlands,

water

Berberis Xenobiotics Soil rhizodegradation

Beta vulgaris Ni, Cu, Zn, Cr Soil

Brassica canola 137Cs Soil

Brassica juncea Pb, Zn, Cr, Cd, Ni,

Cu,90Sr, Se, U

Soil phytoextraction (U—only with

organic acids), phytotransformation

(Cr+6-Cr+3); phytostabilization

Brassica nigra Zn, Pb Soil

Buxaceae Ni Soil

Calamagrostis epigejos Pb, Zn, Cu Soil

Cardamonopsis hallerii Heavy metals (HM)

(Zn, Cd)

Soil, hydroponics

Сynodon dactylon Xenobiotics Soil rhizodegradation

Eucalyptus sp., Eucalyptus

globulus

Na, As, Cu, Pb Soil

Eichhornia crassipes Pb, Cu, Cd, Fe Storm water ditch, sewage wetlands,

water

Festuca arundinacea Xenobiotics Soil rhizodegradation

Festuca rubra Pb, Zn Soil phytostabilisation (with CaCO3)

Fagopyrum esculentum Ni Soil

Juncus compressus Zn, Cd, Pb In the roots

Haumaniastrum katangense Co Soil

Chenopodium album Zn, Cu Soil

Helianthus annuus Cr, Mn, Cd, Ni, Zn, Cu Soil, storm water ditch, sewage

wetlands, water (rhizoltration)

Helianthus annuus Pb, U, 137Cs, 90Sr, Cu

(mutant forms)

Soil

Hydrocotyle umbellata Pb, Cu, Cd, Fe Soil

Kochia scoparia Radionuclides (RN) Soil

(continued)

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• Easy to identify

• Easily available

• Analytically accessible and low detection and determination thresholds with

current analytic technology

• Accumulation of pollutants

Table 5.2 (continued)

Species (genera) Accumulated elements Substrate

Lemna minor Pb, Cu, Cd, Zn Storm water ditch, water

Linum usitatissimum Cu, Ni, Cd, Cr, Pb Soil (phytoremediation+raw

material for plant ber)

Lycopersicon lycopersicum Pb, Zn, Cu Soil

Lolium perenne Radionuclides,

xenobiotics

Soil rhizodegradation

Medicago sativa Ni (HM), Pu,

xenobiotics

Soil (symbiosis with bacteria)

Melilotus ofcinalis Zn, Ni, Cu Soil

Miscanthus giganteus Cu, Ni, Cd, Cr, Pb Soil (phytoremediation+raw

material for plant ber)

Morus sp. Xenobiotics, HM Soil rhizodegradation

Phalaris arundinacea Cd, Cu, Zn, Pb Storm water ditch, sewage wetlands,

water

Polygonum sp., P. sachalinense Cd, Pb, Zn, 137Cs, 90Sr, Cu Soil

Populus sp. Hg, Fe, Ni, Zn, Cd, Pb,

herbicides

Soil phytoextraction, soil

rhizodegradation

Pinus pinaster Fe, Zn, As, Pb, W Soil

Phalaris arundinacea Cd, Cu, Zn, Pb Storm water ditch, sewage wetlands,

water

Quercus ilex, Quercus

rotundifolia, Quercus suber

Ni, As, W, Zn, Pb Soil

Rhus typhina Polycyclic aromatic

hydrocarbons

Soil rhizodegradation

Salix sp. Ni, Zn, Cd, Pb,

perchlorate

Waste water, ltrates

Scirpus sylvaticus Cd, Cu, Zn, Pb Storm water ditch, sewage wetlands,

water

Secale cereale Zn, Pb Soil (only with the introduction

of bacteria Rhodococcus equi)

Silene latifolia Zn, Cu Soil

Sorgo bicolor Zn, Cu, Pb Soil

Trifolium sp. Xenobiotics Soil rhizodegradation

Typha latifolia Cu, Zn, Cd, Pb Storm water ditch, sewage wetlands,

water

Thlaspi caerulescens Zn, Cd Soil

Thuja occidentalis V, Cr, Fe, Ni, As, Mo Soil

Urtica dioica Cu, Ni, Cd, Cr, Pb Soil

Vetiveria zizanioides Cr, Cu, Ni, Zn, As, Cd,

Water, soil

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And features mentioned by Bargagli [3]:

• Long vegetation period

• Clearly marked and reproducible response to certain changes in the environment

In addition, they have a number of advantages for phytoremediation [136, 137]:

• A high yield of biomass (at a density of 10,000–20,000 per hectare) to 15tonnes

of dry matter/ha per year, which enables efcient phytoextraction with moderate

amounts of accumulation of toxic elements [138]

• Famous cultivating agricultural technology (Salix and Populus trees grown on a

short rotation system: the harvest in 3–5 years with a total duration of 30 years

of cultivation), which can be adapted for use at contaminated lands

• Ability after bioremediation to be used as biofuels (direct combustion, anaerobic

digestion processing, fermentation in liquid fuels), for the production of wood,

ethanol, biogas, biofortied, biochar, chipboard, paper, constructions, and tech-

nical production [139–141]

• Ability to use in urban landscapes

• Creation of greenbelts and phytocenoses for remediation [142–144]

• The stabilization of the substrate under cultivation: soil protection from water

and air erosion; prevent metal leaching to protect surface and groundwater

5.3 Possibilities of Woody Plant Use for Biomonitoring

of Antropogenic Pollution of Environment

Selection of species for environmental assessment is dictated by a number of neces-

sary conditions: they must be sufciently widespread in the study area and well

reect the state of the environment by changes of qualitative or quantitative charac-

teristics (e.g., the development of necrosis and chlorosis, the change of physiologi-

cal parameters, morphological or anatomical changes, changes at the molecular

level, etc.). Based on these characteristics the assessment of the environment can be

carried out which includes physiological and biogeochemical characteristics of spe-

cies [19–23, 34, 35, 88–105].

An important issue is the expansion of the list of species of woody plants, which

can be used for phytoremediation of environment in conditions of complex pollution

of air and soil by heavy metals in industrial centers.

We carried out integrated monitoring of ecosystems with varying degree of

anthropogenic load at the territory of a model region of the central zone of Russia—

Tula region. The parameters of the woody plants growing in natural habitats (forests

and forest-steppe ecosystems) and in polluted urban environment of the regional

center of industry (protection zones of motorways, the territory of the metallurgical

enterprises) were determined [103].

The parameters of bioaccumulation of toxic elements of trees growing in

contaminated and clean areas of Tula region (Russia) were studied.

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5.3.1 Study Area: Objects of Investigation

The model region of the study was Tula region and the city of Tula. In the study

area, three natural territories are located: coniferous-deciduous forests in the north,

deciduous forests in the center of the region, and forest steppe and steppe in the

south. The region is characterized by a well-developed industry: mechanical engi-

neering and metal working, chemical industry, defense industry, ferrous metallurgy,

construction materials, light industry, and food industry. Districts of Tula region are

characterized by a varying degree of anthropogenic impact. Tula industry (ferrous

metallurgy and mechanical engineering) and Novomoskovsk (chemistry) account

for more than 2/3 of the regional production. A great contribution to the pollution of

the region is from the chemical industry centers Schekino, Efremov, and Aleksin

and the center of the electricity Suvorov. The remaining 20 districts of the region

account for 10% of industrial production. According to the concentration of indus-

trial enterprises, the Tula region is the second after Moscow, and it is among the ve

most ecologically unfavorable regions of Russia, ten times exceeding the amount of

emissions to the atmosphere of the surrounding Kaluga and Oryol regions. 94% of

all emissions are due to the city of Tula and Aleksinskiy, Suvorovskiy, Efremovskiy,

Novomoskovskiy, Uzlovsky, and Schekinsky districts where the largest number of

industrial enterprises is clustered. 52% of pollutants in the atmosphere fall to the

share of industrial enterprises.

The regional center—the city of Tula—is located 180km south of Moscow. This

ecosystem includes the city area of 154km2 and a population more than 500,000; it

represents an area with developed metallurgical, chemical, engineering, and defense

industries with the city’s infrastructure and network of roads with heavy trafc. The

rst stage of investigations was focused on revealing the geochemical anomalies of

soil and the analysis of atmospheric air. The results of these investigations showed

that more than 40% of the territory of the selected ecosystem was characterized by

excess of maximum permissible levels (MPL) of the set of heavy metals in the envi-

ronment (Fig. 5.1).

The main element pollutants of urban (Tula city) soils were:

Mn (in sampling point 1 up to 50% of soil exceeded MPL)

Fe (high gross concentration all over)

Cu (24% soil exceeded MPL up to 3–6 times)

Zn (28% of soil showed excess of MPL by 15–62%)

As (38% of soil excess of MPL by 36–62%)

Pb (12% soil excess of MPL by 10–50%)

The total index for grading soil contamination identied 20 areas of moderately

hazardous category (28% of soils) and 4 of extremely dangerous category (6% of

soil) [145]. The map (Fig. 5.1) presents geochemical anomalies in soils of the city

on the total pollution index. In the most polluted areas, the analysis of atmospheric

air was carried out. The high content of Fe in the form of oxides and sulfates at all

sampling points was revealed, which exceeded the MPL average concentrations of

Fe by several hundreds of times. The copper concentration was higher than the

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Fig. 5.1 Сharacterization of Tula city soil pollution

maximum single MPL by 56% of the surveyed zones and exceeded the daily aver-

age by 1.5–3.3 times and maximum single—in 3–9 times. Pb content exceeds the

average daily rate of MPL sampling points close to Kosogorsky Metallurgical

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Works (ferromanganese production), the Mogilev square (Pedagogical University),

and at the intersection of two main roads of the city (Krasnoarmeyskiy Avenue)

[104, 105, 145].

To investigate the possibility of using woody plants for biomonitoring the envi-

ronmental situation, seven areas with different levels of anthropogenic pollution

of the region and the regional center—the city of Tula—were chosen (Fig. 5.2,

Table 5.3). The objects of biogeochemical parameters in the study (content of ele-

ments in the leaves) were native tree species: Tilia cordata, Acer platanoides,

Salix fragilis, and Picea abies.

Fig. 5.2 Characterization of Tula region soil pollution and sampling points

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The sampling sites to determine the suitability of species for phytoremediation of

soils insanitary-protective planting of metallurgical enterprises of Tula are: point I, JSC

“Kosogorsky Metallurgical Works” (KME) (ferromanganese production), and point II,

complex of enterprises of JSC JV “Tulachermet” and “Vanadium” (Tulachermet) (pro-

duction of pig iron, vanadium, and chromium). For relatively pristine (background or

control) zone, the area of the Central Park of Culture and Leisure was chosen. The

distance between point I and control zone is 2–3km and between point II and control

zone is 5–6km. Distance of sanitary-protective planting from aerosol emission sources

is 30–400m.

Table 5.3 Characterization of different areas of Tula region

Sampling point Description of environmental conditions

City of Tula, sanitary-

protective plantings along

roads

Industrially developed urban ecosystem with a high level of

technogenic pollution, exceedance of MPL of HM metals in soils

by 40% of the territory, a high level of dust and exceedance of

the MPL of HM in the air sampling points

Tula, sanitary-protective

zone of metallurgical

enterprises, KME and

Tulachermet

In soil samples of the sanitary-protective plantings observed, an

excess of MPL of HM on a number of elements, Mn (twofold),

Pb (1.5 times), and Zn (2 times), was observed

Kulikovo Pole The area is located in forest-steppe vegetation zone of Tula

region and characterized by low level of human impact (a

historical place reserve museum area). Agricultural using with

fertilizers is the main form of anthropogenic activity for the soils

Yasnaya Polyana The area of museum reserve “Yasnaya Polyana” is located in

deciduous forests and inuenced of metallurgical and chemical

enterprises (Kosogorsky metallurgical plant, Shchekinoazot)

Plavsk town The city is located in the forest-steppe part of the region. The

state of ecology is affected by a distillery “Plavsky” emissions of

which contaminate the river

Belev town The city is located in coniferous-deciduous forest area and

experiences recreational and vehicle load, and among the

industrial enterprises is the plant Transmash

Belevskiy area (forest) The region is located in a strip of coniferous-deciduous forests,

and there are no large industrial enterprises

Novomoskovsk town The city with high level of industrial pollution (Nitrogen, Procter

& Gamble—Novomoskovsk, Knauf Gypsum Novomoskovsk,

Orgsintez, Polyplast, Novomoskovskaya GRES)

Suvorov town The industrial city is located in coniferous-deciduous forest area

and inuenced of Cherepetskaya hydropower station,

Cherepetskaya precast concrete plant, Mitinskaya Iron Works

and recreation

Suvorovskiy area

(Varushizi)

The city is located in coniferous-deciduous forest area and

inuenced of Cherepetskaya hydropower station

Efremovskiy area (Shilovo) The region is located in the forest-steppe part of the region. The

enterprises have a negative impact on the environment and are

the production of synthetic rubber and household chemicals

(Novomoskovskbytkhim and Procter & Gamble) and Efremov

thermal power station

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The objects of investigation of woody plant feasibility for bioremediation are

seven tree species and eight shrubs dominating in the sanitary-protective zone of the

metallurgical enterprises: Sorbus aucuparia, Acer platanoides, Populus nigra,

Aesculus hippocastanum, Tilia cordata, Larix sibirica, Betula pendula, Crataegus

sanguinea, Crataegus monogyna, Cornus alba, Cotoneaster lucidus, Symphoricarpos

albus, Syringa vulgaris, Philadelphus coronarius, and Physocarpus opulifolius.

5.3.2 Sampling, Sample Preparation, and Methods

of Research

Sampling to determine the ability of wood to bioaccumulation of toxic elements was

carried out in the third decade of July during the vegetation peak over the perimeter

of the tree crowns at a height of 1.5–2m in the plant communities of different districts

of the region and urban ecosystems of the city of Tula. The minimum number of trees

(shrubs) in each type of sampling points was ten. The minimum number of leaves

from each tree (shrubs) was ten.

Leaves of woody plants were washed in running water, and then they were

washed twice in distilled water. This way of sample preparation, as opposed to the

use of unwashed samples, in our opinion, allows to avoid large errors in sample

preparation that may occur due to loss of the dust particles in the course of opera-

tions, packaging, grinding, weighing, and pressing samples and eliminates depen-

dence on climatic factors (washings by rains, the wind emission) during the sampling

and before it. It allows to perform a comparative description of the research results.

Washed samples were dried at room temperature and brought to constant weight

in an oven at a temperature of 60°C.The samples were averaged and were packed

in paper bags with a label. Sample preparation for instrumental neutron activation

analysis (INAA) (grinding, weighing, pressing, and packing containers of samples)

took place in a chemical laboratory sector neutron activation analysis LNP JINR.

Part of the elements in plant samples (Mn, Fe, Ni, Cu, Zn, Cd, Pb) was deter-

mined in the laboratory of chemical analysis of the Geological Institute (GIN RAS)

by atomic absorption spectrometry using “QUANT-2A” (KORTEK, Moscow)

equipped with deuterium corrector of nonselective absorption and relevant hollow-

cathode lamps; determination of heavy metals in the samples was carried out in

accordance with standardized methods [146]. Determination of Zn, Pb, Cu, and Cd

was performed in “propane-air” ame and Fe, Mn, and Ni in “acetylene-air” ame.

Quality control was provided by using certied reference materials IAEA-SOIL-7,

IAEA-336 (lichen), SRM 1572 (citrus leaves), and SRM 1575 (pine needles).

INAA of plant samples was carried at the IBR-2 reactor at JINR LNP using acti-

vation with epithermal neutrons along with the full energy spectrum. To determine

the long-lived isotopes, samples of leaves of about 0.3g were packed in aluminum

foil. The containers with samples were irradiated for 4–5 days in a cadmium- screened

channel (epithermal neutron activation analysis). After exposure, the samples were

repacked in clean plastic containers for measurement of induced activity.

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Induced gamma activity of the samples was measured twice: after 4–5 days after

irradiation (for determination of As, Br, K, La, Na, Mo, Sm, U, and W) and after 20

days (to determine Ba, Ce, Co, Cr, Cs, Fe, Hf, Ni, Rb, Sb, Sc, Sr, Ta, Tb, Th, Yb, and

Zn). Measuring time was 40–50min and 2.5–3h, respectively.

To determine the short-lived isotopes of elements (Al, Ca, Cl, I, Mg, Mn, and V),

samples of 0.3 g weight were packed in polyethylene packs and irradiated for

3–5min. Induced gamma activity of the samples was measured after 5–7min of

cooling twice: for 3–5 and 10–5min, successively.

Measurement of the induced gamma activity was carried out by gamma spec-

trometers with Ge (Li)—detectors with a resolution of 2.5–3keV of the gamma line

1332keV of 60Co and HPGe detector with a resolution of 1.9keV of the gamma line

1332keV of 60Co.

A software package developed in the Frank Laboratory of Neutron Physics of

JINR was used for processing gamma spectra of induced activity and calculating the

elemental concentrations. The concentrations of elements were determined by

relative methods (by comparison with the standards) [147]. Certied reference

materials (pine needles, NIST) were irradiated and measured together with

samples.

The uncertainties in elemental determinations of Na, K, Cl, As, Sr, Fe, and Pb are

in the range of 5–10% and for V, Ni, Cu, Se, Mo, Cd, and Sb are 30%.

So far for vegetation there are no identied MPLs and the data on the elemen-

tal content in the different studies are very different in dependence on the used

methods and sample preparation [21, 34, 90–95] (some studies are made using

unwashed plant material), to assess the biochemical characteristics of the investi-

gated samples, they were compared with the average data of the Reference plant

(RP) [148].

5.3.3 Results and Discussion

The results carried out in seven districts of the Tula region showed that the leaves of

woody native species can be used as bioindicators and biomonitors of biogeochemi-

cal parameters in determination of the degree of anthropogenic load on

ecosystems.

The two studied species in the cities accumulate more chlorine in the leaves than

in the steppe and forest communities. Thus, the chlorine content in the leaves of

Tilia cordata and Acer platanoides in the towns of Plavsk, Novomoskovsk, and Tula

varied in interval of 3270–6400mg/kg, that is, 1.5–3 times higher than the critical

concentrations and mean values for vegetation [148–150] and 2–17 times higher

than the values for forest and steppe of the region (370–1520mg/kg) (Fig. 5.3). The

accumulation of high concentrations of chlorine in leaves of trees in urban areas

may be due to the use of NaCl on the sidewalks followed by washing the salt melt

water into the soil in winter and early spring as well as by deposition of the aerosol

particles due to the impact of the chemical industry.

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Acer platanoides, Salix fragilis, and Picea abies well reect an increase in the

concentration of vanadium in the environment. So, in Belev and in deciduous forest

near the city, as well as in the city of Tula (sanitarian-protective zone of motor-

ways), V content in the leaves and needles of the enlisted species ranged from 0.7 to

1.5mg/kg of dry matter, that is, 1.3–3 times higher than in the reference plant (RP)

(0.5 mg/kg). It could be connected with aerosol emissions from enterprises of

“Vanadium,” Instrument Design Bureau, NGO “Fusion” (Tula), and JSC

“Transmash” (Belev) (Figs. 5.4, 5.5, and 5.6).

The concentration of chromium in leaves of all investigated deciduous woody

plants was 1.5–2.7 times higher than in RP (2.3–5.0mg/kg dry weight) at all sam-

pling points examined except Belevskiy area (Figs. 5.4, 5.5, and 5.7). This fact may

be an evidence of air emission of this element, and it also conrms our previous

assumption that woody plants concentrate more chromium in organs rather than

herbaceous plants [38–40, 148, 149].

The given fact evidences that the woody plants are sensitive indicators to the

content of the given element in the environment in time and they can reect the spa-

tial distribution of an air emission and absorb elements from deeper soil horizons.

Two deciduous species Acer platanoides and Salix fragilis react to high iron

content in the soil and air (Figs. 5.4 and 5.5). The greatest sensitivity characterizes

Acer platanoides (element content of the leaves increases up to 1250mg/kg (Tula),

that is, twofold higher than the concentrations of toxic element for vegetation [149,

150] and eight times higher than the RP). Such intense absorption of iron along with

other heavy metals can lead to the development of necrotic changes in the leaf and

to reduction of vitality of the species in terms of polymetallic soil pollution of

industrial cities.

Chlorine content in the leaves of woody plants

(Tula region), mg/kg

6000

5000

4000

3000

2000

1000

Acer platanoides Tilia cordata Salix fragilis

Reference Plant

Kulikovo Pole

Yasnaya Polyana

Belevskiy (forest)

Belev

Plavsk

Novomoskovsk

Suvorov

Efremovskiy (Shilovo)

Tula

Fig. 5.3 The chlorine concentration in the leaves of woody plants from Tula region ecosystems

with varying degrees of anthropogenic load

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Bioaccumulation of copper in leaves of woody plants in Belevskiy area (city and

forest), Yasnaya Polyana, and the regional center—the city of Tula—reached 24–53mg/

kg of dry weight, that is, 2.5–5 times higher than the values of RP.The high concentra-

tion of the element in the wood plants is conditioned by its high content in the air and

soil due to the impact of metallurgical industry and metalworking [104, 105, 145].

Accumulation of arsenic was noticed in the leaves of deciduous trees at the sam-

pling points in Novomoskovsk and Efremovskiy area (Shilovo village) (Figs. 5.4

and 5.7). Its concentrations of 0.21–0.37mg/kg of dry weight are 2–3.5 times higher

than in the reference plant. Picea abies needles accumulate in two times less arsenic

than RP regardless of the point of sampling (Fig. 5.6).

Fig. 5.4 The content of heavy metals and metalloids in Acer platanoides leaves grown in different

areas of Tula region

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The trend in accumulation of cadmium is similar to the arsenic one: Picea abies

does not accumulate element in the needles, while all deciduous woods are investigated

at all points, except with sampling growing in sanitary-protective zone along Tula

roads and in the forest area of Belevskiy, accumulated from 0.16 to 0.92mg/kg of the

element in the needles that exceeded by 3–18 times the average data for plants

(Fig.5.6). The highest concentration of the element was observed in Acer platanoides

leaves growing in the Kulikovo Pole, as well as in the leaves of Salix fragiles (Figs.5.4

and 5.5). The best biomarker for this element is Salix fragilis. Thefact of high

Fig. 5.5 The content of heavy metals and metalloids in the leaves of Salix fragilis (Tula region)

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accumulation of cadmium by leaves of woody plants can be explained by air emission

of the element by motorway, as well as, apparently, by better degree of absorption of

the elements at alkaline soils (Efremovskiy area (Shilovo), Kulikovo Pole). The trend

of low bioaccumulation of cadmium in leaves of trees growing in Tula, in the soils of

multi-element anomaly, can be explained by the antagonism of the ions when accu-

mulated from the environment, as well as the low level of cadmium in the soil.

Taking into consideration the difference in bioaccumulation of toxic elements by

leaves of plants growing in different districts of the region, one may conclude that

the chosen plant for bioindication and biomonitoring reects the environmental

Fig. 5.6 The content of heavy metals and metalloids in the needles of Picea abies in different

ecosystems of Tula region

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situation in ecosystems with different levels of anthropogenic load and can be used

as biomonitors.

At the same time the largest range of changes in biogeochemical activity is typi-

cal for species Acer platanoides and Salix fragilis, which are preferably used for

biomonitoring based on biogeochemical parameters of variability options when

changing of the main sources of environmental pollution occurs. Aboriginal species

of Gymnospermae—Picea abies—is less favorable for the chosen purposes. Below

are details of the biogeochemical activity of the studied species (Table 5.4).

Fig. 5.7 Heavy metal bioaccumulation in leaves of Tilia cordata (Tula region)

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Analysis of correlation bindings at bioaccumulation of elements by leaves of

plants revealed strong correlation between accumulated elements. Al correlates

with V, Zn, Sm, Hf, Th, and U (accumulation of elements of technogenic soil

contamination); Mn with Cu; Cd with W; V with Th and U; and Zn with Th and U

(impact factor of metallurgical defense and metal processing industries).

5.4 Accumulation of Heavy Metal in Conditions

of Polymetallic Contamination of Industrial Areas

(Metallurgical Plants) by Woody Plants of Moderate

Climate: Possibilities of Their Application for Soil

Phytoremediation

At present a lot of investigations on the use of woody plants for phytoremediation of

soil from heavy metals were undertaken. The genera Saliсaceae: Populus and Salix

[113, 137, 151–158], Pinus [136, 159–161], Acer [162, 163], Betula [164–168],

Quercus [136, 169, 170], Morus alba [115], Acacia retinoides and Eucaliptus tor-

quata [171, 172] for soil of subtropical climate were investigated.

For phytoremediation of the environment from radionuclides the resistant to

them Juglans mandshurica and characterized by high ecological plasticity

Phellodendron amurense were used.

An important issue is the expansion of the list of species of woody plants, which

can be used for phytoremediation of environment in conditions of complex air and

soil heavy metal pollution in industrial centers (in a temperate continental climate).

Prior to the beginning of our biogeochemical studies, we have evaluated the

vitality of species of sanitary-protective plantations of metallurgical enterprises and

highways, the presence of necrotic and chlorotic leaf damage, as well as their ability

to accumulate dust emissions on the leaf surface.

In assessing the vitality, the scale was proposed by T.V. Chernenkova [94]: 1

point, “healthy”; 2 points, “weakened”; 3, “severely weakened”; 4 points, “mori-

bund”; and 5 points, “deadwood.” Assessment of the vitality showed that the most

morphologically adapted species to the conditions of polymetallic contamination

are Larix sibirica, Syringa vulgaris, Caragana aerorescens, Ligustrum vulgare

(vitality 1), Philadelphus coronarius (vitality of 1–2), Sorbus aucuparia, and Acer

platanoides (vitality 2). All studied species in the sanitary-protective plantations

showed necrotic and chlorotic leaf change (from 7 to 98%—the most in the genus

Populus), expressed in the point and edge necrosis and interveinal chlorosis.

Table 5.4 Biogeochemical activity of leaves of woody plants reecting the anthropogenic load on

ecosystems

Species Elements which accumulated at elevated technogenic loads

Acer platanoides Cl, V, Mn, Fe, Ni, Cu, As, Cd

Tilia cordata Cl, Mn, Ni, Cd

Salix fragilis Cl, V, Mn, Fe, Ni, Zn, As, Cd

Picea abies V, Mn, Fe

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Manifestation of damage in the leaves of various species was different.

For example, Sorbus aucuparia, Betula alba, Acer platanoides, Tilia cordata, and

Cotoneaster lucidus showed the appearance of the point of necrosis on the leaf blade

(Figs. 5.8, 5.9, and 5.10). In Aesculus hippocastanum, Cotoneaster lucidus, Crataegus,

Cornus alba—edge and interveinal necrosis of leaves appear (Figs. 5.8 and 5.10).

All these symptoms may be indicative of the direct damage of leaf tissue by toxic

concentrations of iron, manganese, nickel, and chromium [34]. The development of

necrosis of leaves can be used for bioindication of the environment at affected areas

by the polymetallic pollution with the help of the leaves of wood. Leaf chlorosis

may be caused by an imbalance in the leaves with an excess of magnesium ions of

other substituents. At the level of the leaf anatomy noted manifestation of was

observed kseromorphism symptoms (reduction of the leaf square, Fig. 5.9), An

increase in the number of stomata, sheet thickness and diameter of the stomata to

compensate for exchange in dusty conditions and high concentrations of heavy met-

als in the environment of the sampling sites took place [35].

The ability to accumulate dust by leaves of woody plants is different for trees and

shrubs and is ranging from 8 to 206mg/dm2 for trees and from 17 to 423mg/dm2

for shrubs.

The maximum ability to accumulate dust in sanitary-protective plantations belongs to:

• Tilia cordata (till 115mg/dm2)

• Populus nigra (till 115mg/dm2)

• Salix caprea (till 141mg/dm2)

Fig. 5.8 Necrosis and chlorosis of tree leaves growing in the area of inuence of metallurgical

enterprises (Sorbus aucuparia, Aesculus hippocastanum, Tilia cordata)

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Fig. 5.9 Decrease of leaf area of the Symphoricarpos albus, which grows in the area of inuence

of metallurgical enterprises (right) in comparison with the control zone (left)

Fig. 5.10 Necrosis and chlorosis of shrub leaves growing in the area of inuence of metallurgical

enterprises (Cornus alba, Cotoneaster lucidus, Crataegus monogyna)

S.V. Gorelova and M.V. Frontasyeva

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129

• Sorbus aucuparia (till 206mg/dm2)

• Crataegus sanguninea (till 107mg/dm2)

• Philadelphus coronarius (till 149mg/dm2)

• Symphoricarpos albus (till 153mg/dm2)

• Syringa vulgaris (till 189mg/dm2)

• Cornus alba (till 296mg/dm2)

• Ligustrum vulgare (till 423mg/dm2)

We have determined the dust particle content on the surface of leaves using the

method of electron scanning microscopy in the laboratory of geochemistry and min-

eralogy of soil of Federal State Institute of physical, chemical, and biological prob-

lems of pedology of RAS (operator E.I. Elmov).

The distribution of elements on the leaves surface is random, but the greatest

number of dust particles concentrates at the bottom, along the edge of the leaf sur-

face and along the main veins of leaves (Fig. 5.11). Analysis of the spectra showed

Fig. 5.11 Distribution of dust emissions on a fragment of leaf surface of Crataegus sanguinea

which grows in the area of metallurgical production (scanning electron microscope)

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130

that most of the adsorbed particles are those compounds of organic nature, and

about 1–3% accounted for by soil components, the proportion of heavy metals Fe,

Cu, Mn, Cd, Zn, and Pb on the leaf surface may vary from 0.02 to 5% for each ele-

ment. The method might be the qualitative for analysis of components of aerosol

emissions and allows one to establish the presence and ratio of the components on

the surface of the leaf that reects the air pollution of the studied zone.

5.4.1 Bioaccumulation of Heavy Metals by the Trees

and Shrub Leaves

5.4.1.1 Manganese

Analysis of the Mn content in leaves of woody plants showed that its concentration

ranges from 23 to 385mg/kg (Table 5.5, Fig. 5.12). That is a rather strong variation

of the element probably depending on the species features, growing conditions, and

the element concentrations in the environment . Low concentrations of the element

(23–150 mg/kg) are characteristic for species such as Aesculus hippocastanum,

Tilia cordata, Betula pendula, Cornus alba, Physocarpus opulifolius, and

Philadelphus coronarius growing in urban conditions in relatively pristine areas.

However, the impact of emissions from the steel industry resulted in increase of the

content of elements by two times in Philadelphus coronarius and Tilia cordata

(sample point KME), four times in Aesculus hippocastanum, and eight times in

Betula pendula and reached a value of 340mg/kg, that is, on average, 1.5 times

higher than in the reference plant [148]. However, for Larix sibirica in excess of the

concentration of the element in the environment overrelatively permissible concen-

trations, increase of its content in the leaves was not noticed. This may serve as

evidence for inclusion of protective mechanisms in the absorption and transport of

this element by plant root system, and it also can be caused by antagonism with Fe

ions in the process of element uptake by root system.

The role of Betula pendula, Pinus sylvestris, and Larix sukaczewii in the absorp-

tion of Mn from heaps of mining enterprises has been described [34]. The maximum

element concentration in the leaves of the studied species of woody plants was

characteristic for Acer platanoides, Betula pendula, Tilia cordata (180–340mg/kg),

and Cotoneaster lucidus (336 mg/kg dry weight). When the concentration of ele-

ments in soil exceeds maximum permissible level (MPL) by a factor of 4.7 (site

affected by metallurgical enterprises), the contents of the element in the leaves of

Acer platanoides, Tilia cordata, and Betula pendula increased by 2–7 times, but not

as signicantly as in Aesculus hippocastanum.

Thus, among the studied species, the maximum accumulation of Mn in the leaves

is characteristic of Acer platanoides, Betula pendula, and Cotoneaster lucidus,

which is a peculiarity of species.

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Table 5.5 Content of heavy metals in the leaves of woody plants, mg/kg of dry weight

Element Mn Fe Ni Cu Zn Cd Pb

RP (Markert [148])/terrestrial vegetation

(Kabata-Pendias [149])

200 150 1.5 10 50 0.05 1

15–500 18–1700 0.1–3.7 5–30 1.2–73 0.08–028 0.1–10

Woody plants (zone of metallurgical

enterprises and ore dumps) Acer rubrum,

Betula pendula, Pinus sylvestris, Picea,

Lukina, Nikonov [90] Chernenkova [94],

Kulagin, Shagieva [34]

197–1055;

157–1050;

195–1130

88–698;

2140–4790;

1630–3680

14–98;

4.5–16;

0.9–11

16–37;

5.1–57;

9.3–26

19–54;

5.6–56; 93

0.14–0.64;

0.42–0.97

1.2–16.3

Acer platanoides I (KME) 180±22 810±7 0.7±0.3 5.7±0.4 31±6 0.09±0.01 2.6±0.2

II (Tulachermet) 260±28 1250±41 3.2±0.1 10.1±0.9 44±8 0.16±0.02 0.6±0.1

Control zone 120±17 310±16 2.2±0.2 8.2±0.7 38±7 0.11±0.02 0.6±0.1

Aesculus

hippocastanum

I 53±16 1065±54 4.5±0.7 15.4±1.3 41±9 <0.04 5.6±0.2

II 91±11 1388±78 1.5±0.2 12.8±0.8 33±8 <0.04 0.6±0.1

Control zone 83±14 230±12 1.1±0.2 10.7±0.8 32±5 <0.04 0.6±0.1

Betula pendula I 240±19 800±28 4.3±0.2 6.3±0.5 156±24 0.15±0.02 5.6±0.2

II 340±7 940±85 6.1±0.4 6.8±0.4 99±12 0.14±0.03 0.5±0.1

Control zone 45±8 180±24 3.1±0.2 4.3±0.2 53±8 <0.04 0.5±0.1

Populus I 190±21 1580±62 5.0±0.3 6.8±0.5 171±19 0.58±0.03 8.1±0.2

II 92±17 500±27 7.2±0.4 10.9±0.9 127±16 0.66±0.04 <0.5

Control zone 27±5 195±14 1.5±0.1 4.6±0.2 124±18 0.27±0.02 <0.5

Sorbus aucuparia I 110±16 1780±54 1.2±0.1 6.2±0.4 29±4 0.08±0.01 6.9±0.2

II 120±14 2570±78 1.8±0.2 6.0±0.3 29±4 0.04±0.01 <0.5

Control zone 51±7 550±24 1.5±0.1 8.4±0.7 29±3 0.10±0.01 0.9±0.1

Tilia cordata I 306±54 5082±72 3.5±0.2 7.6±0.6 35±3 0.09±0.01 1.1±0.1

II 130±14 970±35 1.8±0.1 8.0±0.5 41±3 0.08±0.01 <0.5

Control zone 130±15 220±25 2.3±0.3 9.0±0.6 41±4 0.10±0.01 0.7±0.1

Larix sibirica I 110±17 1210±64 0.9±0.1 6.0±0.4 33±2 0.09±0.01 15.2±0.3

II 120±32 3640±126 4.2±0.4 5.1±0.4 42±4 <0.04 1.6±0.1

Control zone 410±38 840±32 3.1±0.4 6.5±0.5 41±3 0.06±0.01 1.9±0.2

Excess concentration (toxic concentration)

(Cabata-Pendias, 1989)

≥500 ≥500; >1000 10–100 20–40 ≥300 ≥60

Normal concentration (range of normal

regulation) (Cabata-Pendias, 1989)

20–60 19–250 0.1–2.1 3–12 20–60

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Mn content in the leaves of shrubs

mg/kg 320

280

240

200

160

120

KME

RP, Markert, 1992

Crataégus sanguinea

Crataegus monogyna

Cornus alba

Cotoneaster lucidus

Syringa vulgaris

Symphoricarpos albus

Philadelphus coronarius

Physocarpus opulifolius

Tulachermet city road control zone

Fig. 5.12 Comparison of Mn content in the leaves of shrubs growing near metallurgical enter-

prises and city road with RP

5.4.1.2 Iron

The results of investigations showed that the Fe content in leaves of woody plants in

urban conditions, in the absence of additional sources of pollution (emissions of

enterprises), varies from 48 to 2062mg/kg of dry weight. In the absence of addi-

tional pollution from industries, the minimum iron content is characteristic for

leaves of Populus nigra (48–195mg/kg of dry weight) (Table 5.5). The maximum

from trees was observed in Larix sibirica needles (840mg/kg) and Cotoneaster

lucidus leaves (687mg/kg of dry weight). The content of Fe in the leaves of Aesculus

hippocastanum and Tilia cordata is characterized as close to the average of 142–

290mg/kg, which exceeds the average values for the reference plant [148]. The

leaves of Betula pendula are characterized by a minimum range of Fe concentra-

tions (154–180mg/kg).

Moreover, the maximum content of Fe was observed in leaves of Larix sibirica,

Sorbus aucuparia, Tilia cordata, Cornus alba, and Cotoneaster lucidus (Table 5.5,

Fig. 5.13).

In general, Larix sibirica, Sorbus aucuparia, Tilia cordata, Aesculus hippocasta-

num, Acer platanoides, Populus nigra, Cotoneaster lucidus, Cornus alba, and

Crataegus sanguina can be considered the best bioaccumulators of Fe when its

content in soil is high. The content of Fe in the leaves of these woody species

growing in industrial areas is 1250–8930mg/kg, that is, 10–60 times higher than the

values for the reference plant (Table 5.5).

According to some authors, organ concentrators of iron are the roots and bark of

trees growing in dumps of polymetallic deposits [94]. The content of Fe in the organs

S.V. Gorelova and M.V. Frontasyeva

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133

(Betula, Pinus) amounts to some thousand ppm [34, 90, 94]. Thus, one should

expect that the total accumulative ability of trees in the industrial site in relation to

iron would be even higher due to accumulation of iron in the perennial organs.

The ratio of Fe/Mn, apparently, is a decisive factor of plant resistance to toxic

elements [149]. In this connection, the establishment of relationships in the intake

of a given pair of elements in different plant species is of interest. The results

obtained show that the ratio of pairs of elements in the accumulation of leaves of

woody plants has species peculiarities. Thus, the lowest ratio of Fe/Mn is charac-

teristic for leaves of A. platanoides, 0.5–2.5; Populus nigra in the Balkan countries,

1–1.4 [98, 99]; Betula pendula, 1–4; and Syringa vulgaris and Crataegus san-

guina, 1–1.1 (in buffer zone conditions). A high content of iron relative to manga-

nese is specic for Aesculus hippocastanum, 2.5–7; Tilia cordata, 2–12; and all

other species of investigated shrubs, 3–14. Under the impact of polymetallic con-

tamination, the ratio increases by up to 5in Acer platanoides, 7–16in Tilia cor-

data, 8in Populus nigra, 16–21in Sorbus aucuparia, 11–30in Larix sibirica, 65in

Crataegus sanguine, 9–165in Cornus alba, and 52in Cotoneaster lucidus leaves.

There species should be preferably used for phytoremediation of soils from an

excess of Fe.

5.4.1.3 Zinc

The concentration of Zn in leaves of woody plants in urban ecosystems is ranging

from 16 to 175mg/kg (Table 5.5, Fig. 5.14). The minimum content of the element

in the leaves is characteristic for Aesculus hippocastanum, 16–32 mg/kg; Tilia

Fe content in the leaves of shrubs

mg/kg

9000

8000

7000

6000

5000

4000

3000

2000

1000

KME Tulachermet city road control zone

RP, Markert, 1992

Crataégus sanguinea

Crataegus monogyna

Cornus alba

Cotoneaster lucidus

Syringa vulgaris

Symphoricarpos albus

Philadelphus coronarius

Physocarpus opulifolius

Fig. 5.13 Comparison of Fe content in the leaves of shrubs growing near metallurgical enterprises

and city road with RP

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cordata, 19–41mg/kg; and Crataegus sanguina, Cornus alba, and Cotoneaster luci-

dus, 18–31mg/kg. Close values of element contents in the leaves are also character-

istic for A. platanoides. All seven species exhibit a relative resistance with increasing

content of Zn in the environment (site of emissions of metallurgical enterprises) that

may be due to the known antagonism between the Fe and Zn ions. Data on the

antagonism intake of Fe and Zn in the plant organisms [149] are not conrmed for

all types of woody plants.

For example, in A. platanoides, Aesculus hippocastanum, and Tilia cordata,

increase in accumulation of iron in the leaves takes place along with even a slight

accumulation of zinc, i.e., iron absorption dominates over the absorption of zinc.

However, for the species accumulators of Zn, Betula pendula and Populus nigra,

an increased concentration of all three element antagonists Fe, Mn, and Zn—in the

site of polymetallic pollution—was revealed. The concentration of Zn increased up

to 153–176mg/kg that exceeds the average for terrestrial plants by a factor of 3–3.5.

Such undiscriminating absorption of three elements at their high concentrations in

the soil by the given species may be due to the absence of barrier function of the root

system to absorption of Fe, Mn, and Zn that is a characteristic peculiarity of hyper-

accumulators [29].

5.4.1.4 Nickel

Accumulation of Ni by leaves of woody plants was low and ranged from 0.1 to

4.5mg/kg of dry weight within urban ecosystems (Table 5.5, Fig. 5.15). The val-

ues of Ni content in the leaves of trees exeeds average values for the reference

plants by a factor of 1.5–2 of the species Acer platanoides, Betula pendula, Tilia

cordata, and Larix sibirica. However, in the site affected by the metallurgical

Zn content in the leaves of shrubs

mg/kg 60

KME Tulachermet city road control zone

RP, Markert, 1992

Crataégus sanguinea

Crataegus monogyna

Cornus alba

Cotoneaster lucidus

Syringa vulgaris

Symphoricarpos albus

Philadelphus coronarius

Physocarpus opulifolius

Fig. 5.14 Comparison of Zn content in the leaves of shrubs growing near metallurgical enterprises

and city road with RP

S.V. Gorelova and M.V. Frontasyeva

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135

industry, the content of the element in the leaves of trees increased by a factor of

1.5–2 for the species Acer platanoides, Tilia cordata, and Betula pendula and

more than a factor of 4–4.5 for Aesculus hippocastanum and Populus nigra (up to

4.3–7.2mg/kg), exceeding the values for the reference plants by 3–5 times. High

concentrations of nickel were found for shrubs Crataegus, Cotoneaster lucidus,

and Syringa vulgaris in control zone with 4.9–16.6mg/kg of dry weight. In the

sampling sites affected by the metallurgical industry, the content of the element in

the leaves of shrubs increased by a factor of 2–5 for such shrubs as Crataegus

monogina (KME) and Cotoneaster lucidus. However, the total concentration of

nickel in the tree leaves did not exceed the threshold of phytotoxicity of the ele-

ment and was located in the middle of toxic concentrations for species of shrubs

such as Crataegus monogyna and Cotoneaster lucidus. High concentration of

heavy metals (nickel in particular) can cause necrosis of Crataegus leaves in case

of low activity of antioxidant system [149].

5.4.1.5 Lead

Accumulation of Pb by leaves of studied trees and shrubs was low and ranged within

0.5–2.7mg/kg (Table 5.5, Fig. 5.16). In the areas affected by emissions of metal-

lurgical enterprises, when the concentration of the element in soil exceeds the maxi-

mum permissible level, Tilia cordata showed the greatest resistance, the concentration

of elements in the leaves of which increased slightly compared with the relatively

clean area. Possible low bioaccumulation of lead by Tilia leaves is due to the pecu-

liarities of leaf surface. However, in the leaves of Aesculus hippocastanum, Betula

pendula, Populus nigra, Sorbus aucuparia, Larix sibirica (sample point KME),

Philadelphus coronarius, and Cotoneaster lucidus (sample point Tulachermet), the

Ni content in the leaves of shrubs

mg/kg 35

KME Tulachermet city road control zone

RP, Markert, 1992

Crataégus sanguinea

Crataegus monogyna

Cornus alba

Cotoneaster lucidus

Syringa vulgaris

Symphoricarpos albus

Philadelphus coronarius

Physocarpus opulifolius

Fig. 5.15 Comparison of Ni content in the leaves of shrubs growing near metallurgical enterprises

and city road with RP

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136

concentration of element in the leaves in the impact site of metallurgical enterprises

increased by a factor of 9–16 and reached the values of 5.6–17mg/kg, which is

higher than in the “reference plant” by a factor of 5–17. However, in our case, the

general element concentrations in leaves of most woody plants do not exceed the

values for herbaceous terrestrial plants growing in nonmetallic areas (1–10mg/kg)

[149]. Only for Larix sibirica and Philadelphus coronarius grown under conditions

of metallurgical enterprises, the value of lead concentrations in the leaves is above

the average values (15–17mg/kg). It is known that the transport of lead in the roots

is passive, and the major part of the element is kept at a certain level in roots and in

the tissue accumulators (inner and outer cortex) [29], but in its sufcient quantity, it

can be absorbed by leaves from the air [149]. This is probably the main way of

absorption and accumulation of lead by leaves of woody plants [173]. This conrms

the fact of the higher lead content in the tree leaves of the cities of the Balkan coun-

tries, where sampling is performed in urban heavy trafc areas in comparison with

the parks in Russia. For example, high concentration of many metals and toxic ele-

ments in dust in Soa was observed, e.g., up to 192mg/kg of Pb, 8mg/kg of As,

123mg/kg of Cu, 710mg/kg of Zn, etc. [98].

It was established that leaves of woody plants are suitable for biomonitoring of

elements in urban environments, and in the Mediterranean for that purpose such

species as Quercus ilex was used [2, 3, 88, 174]. However, as it was demonstrated

in our study, most of the examined species of trees and shrubs react to the polyme-

tallic pollution accumulating some of these elements.

Of all studied species, only in Aesculus hippocastanum, Betula pendula, Tilia

cordata, Populus nigra, Crataegus sanguine and monogina, and Cornus alba, com-

bined effect of polymetallic contamination with Mn, Fe, Ni, Pb, and Zn causes sta-

ble morphological changes, namely, decrease of vitality, appearance of necrosis,

and chlorosis of leaves (Figs. 5.8 and 5.10; Table 5.6).

Pb content in the leaves of shrubs

mg/kg 16

KME Tulachermet city road control zone

RP, Markert, 1992

Crataégus sanguinea

Crataegus monogyna

Cornus alba

Cotoneaster lucidus

Syringa vulgaris

Symphoricarpos albus

Philadelphus coronarius

Physocarpus opulifolius

Fig. 5.16 Comparison of Pb content in the leaves of shrubs growing near metallurgical enterprises

and city road with RP

S.V. Gorelova and M.V. Frontasyeva

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137

Table 5.6 The possibility of using the studied species of woody plants for bioremediation of soils

contaminated with heavy metals (Gorelova etal. [98, 99, 101, 102, 145, 175])

Species

Vitality/

steadiness to

polymetallic

pollution

Morphological

changes

Ability to bioaccumulate heavy

metals under the inuence of

polymetallic contamination,

mg/kg/concentration factor in

relation to the reference plants

Aesculus

hippocastanum

2–3/–, the death

of up to 30% of

trees at planting

Regional and

interveinal necrosis

of leaves

Fe—970–1388/6.5–9

Ni—4.5/3

Cu—12.8–15.4/1.3–1.5

Pb—5.6 /5

Betula pendula 1–2/+ Point necrosis of

leaves

Mn—240–340/1.2–1.7

Fe—800–940/5–6

Ni—4.3–6.1/3–4

Zn—99–156/2–3

Pb—5.6/5

Populus nigra 3/– Necrosis of leaves,

dry branches in the

crown, dieback

Fe—500–1580/3–11

Ni—5–7.2/3–5

Zn—127–171/2.5–3.4

Cd—0.58–0.66/11–13

Pb—8.1/8

Sorbus

aucuparia

2/+ Necrosis of the the

leaf margin, chloroses

Fe—1780–2570/11–17

Pb—6.9/7

Larix sibirica 1/+ −Fe—1210–3640/8–24

Crataegus

sanguinea

2/+ in the

absence of

pruning

25–28% necrotic

spots on the leaves

Cl—1720–2560/2

V—2.9–5.5/6–11

Fe—1780–3400/12–23

Ni—4.5–5.5/3–3.7

Cu—26/2.5

Crataegus

monogyna

2/+ in the

absence of

pruning

10–37% regional and

interveinal necrosis

and chloroses of

leaves

V—4.7/9

Fe—457–2470/3–16

Ni—16.6–34.7/11–23

Cu—14/1.4

Cornus alba 2/+ depends on

the emissions

components

24–28% necrosis of

the the leaf margin,

chloroses, pest insect

damage (aphid)

Cr—6.5/4

Fe—2050–8930/14–59

Cotoneaster

lucidus

1–2/+ 9% regional and

interveinal necrosis

of leaves, chloroses

Mn—336/1.5

Fe—2860–7260/19–48

Ni—5–23/3–15

Cd—0.123/2.5

Pb—1.4–8.1/1.4–8

Symphoricarpos

albus

1–2/+ 4–5% point necrosis

of leaves

V—4.4/9

Cr—4.2–5.6/3–3.5

Syringa vulgaris 1/+ 4–5% necrosis of the

the leaf margin

Cl—4460/22

V—3/6

Cr—5.2/3.4

Fe—810–2310/5–15

Philadelphus

coronarius

1/+ 7%—necrosis of the

the leaf margin,

chloroses

Cl—4830/24

V—5.3/10

Fe—430–735/3–5

Pb—1.4–17/1.4–17

5 The Use ofHigher Plants inBiomonitoring andEnvironmental Bioremediation

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138

This fact demonstrates preference of using the given species for bioindication. It

is known that the meaning of bioindicator and biomonitor is not identical [148]. Due

to this fact, the choice of biomonitors should be based on the whole set of features.

The recommended biomonitors in heavy metal pollution areas are Populus nigra,

Betula pendula, and Cotoneaster lucidus as concentrations of all studied elements

in polymetallic pollution increase sharply in these species (Table 5.6). The other

species can be used for biomonitoring selectively, given the species specicity of

the absorption of elements.

On the contrary, for phytoremediation it is more reasonable to use species which

have directed adaptive changes and which preserve normal vitality in the conditions

of polymetallic contamination.

The number of accumulating elements in this case diminishes, but duration of

detoxication of the environment increases. According to the results obtained, Betula

pendula, Sorbus aucuparia, Tilia cordata and Larix sibirica, and Cotoneaster luci-

dus belong to such species.

5.4.2 Transfer Factor of Elements in Trees and Shrub Leaves

from the Soil

Heavy metals enter plants in two ways: by absorbing root system and by uptake

through the aboveground organs. An objective criterion, which characterizes the

efciency of accumulation of chemical elements, is the factor of biological accumu-

lation or transfer factor (TF) [176]. To identify possible sources of contamination, it

is also important to know the level of air pollution [177]. Part of pollutants are rather

hygroscopic and can penetrate the epidermis and stomata in the form of tiny parti-

cles as well as a concentrated solution, causing water shortages and promoting early

defoliation [178].

The results of the calculation of transfer factor (TF) for woody plants in the buf-

fer site of metallurgical enterprises are shown in Tables 5.7 and 5.8. For most heavy

metals, the values of TF are within 0.01–0.2. However, for some species, specicity

of accumulation of elements in leaves relatively their concentration in the soil was

observed. For example, for Populus leaves known as accumulator of heavy metals,

the TF for Cd > 1 and it is 5–15 times higher than for other species. These values of

TF may be associated with foliar absorption. Poplar leaves secrete sticky substances

that promote the transition of insoluble compounds of aerosol particles in the solu-

ble forms of active transport in the leaves cells [94]. However, under the increase of

Cd content in the soil, we observed decrease of TF in ve times (sampling point I),

which is clearly associated with increased barrier function of the root system.

Because of bioaccumulation from the soil, the content of Zn in the Populus species

was higher than its total content in the soil. However, when Zn concentration

exceeded the maximal permissible level in the soil by two times, as well as in the

case of cadmium, the TF decreased due to barrier function of the roots.

S.V. Gorelova and M.V. Frontasyeva

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139

Table 5.7 TF of woody plant leaves in buffer site of metallurgical enterprises (Tula, Russia)

Species Sampling point Mn Fe Ni Zn Cd Pb

Acer platanoides I 0.03 0.06 0.01 0.13 0.08 0.03

II 0.29 0.03 0.08 0.35 0.62 0.02

Control point 0.16 0.03 0.08 0.58 0.65 0.02

Aesculus

hippocastanum

I 0.01 0.06 0.10 0.17 0.03 0.06

II 0.10 0.02 0.04 0.26 0.15 0.02

Control point 0.11 0.02 0.04 0.49 0.24 0.02

Betula pendula I 0.03 0.05 0.09 0.66 0.13 0.06

II 0.37 0.03 0.16 0.79 0.54 0.02

Control point 0.06 0.02 0.11 0.82 0.24 0.02

Populus nigra I 0.03 0.11 0.11 0.73 0.50 0.09

II 0.10 0.01 0.19 1.01 2.54 0.02

Control point 0.03 0.02 0.05 1.91 1.59 0.02

Tilia cordata I 0.01 0.07 0.07 0.15 0.08 0.01

II 0.14 0.03 0.05 0.33 0.31 0.02

Control point 0.17 0.02 0.08 0.63 0.59 0.02

Table 5.8 TF of shrub leaves in buffer site of metallurgical enterprises (Tula, Russia)

Species Samling point Mn Fe Ni Zn Cu Cd Pb

Symphoricarpos

albus

Control point 0.14 0.04 0.09 0.94 0.22 0.0013 0.019

I 0.02 0.01 0.09 0.11 0.10 0.0002 0.009

II 0.09 0.08 0.07 0.38 0.11 0.0002 0.024

Syringa vulgaris Control point 0.22 0.02 0.18 0.98 0.37 0.0018 0.015

I 0.03 0.06 0.08 0.12 0.15 0.0007 0.006

II 0.09 0.05 0.06 0.25 0.19 0.0004 0.020

Cotoneaster lucidus Control point 0.24 0.06 0.18 0.41 0.31 0.0013 0.026

I 0.01 0.32 0.65 0.06 0.10 0.0003 0.016

Philadelphus

coronarius

Control point 0.06 0.05 0.06 0.39 0.27 0.0003 0.028

I 0.01 0.05 0.04 0.09 0.10 0.0002 0.015

II 0.12 0.07 0.05 0.11 0.28 0.0003 0.652

Сrataegus

monogina

Control point 0.03 0.04 0.59 0.47 0.21 0.0012 0.037

I 0.01 0.23 0.97 0.09 0.30 0.0005 0.008

II 0.04 0.07 0.09 0.17 0.12 0.0005 0.049

Сrataegus

sanguineа

Control point 0.18 0.02 0.17 0.48 0.25 0.0014 0.027

I 0.01 0.01 0.15 0.21 0.17 0.0004 0.013

II 0.03 0.05 0.12 0.15 0.68 0.0001 0.026

Cornus alba Control point 0.06 0.02 0.06 0.39 0.20 0.0003 0.025

I 0.01 0.61 0.04 0.09 0.11 0.0002 0.016

II 0.01 0.08 0.04 0.15 0.09 0.0002 0.016

Physocarpus

opulifolius

Control point 0.07 0.02 0.05 0.55 0.30 0.0014 0.022

I 0.01 0.02 0.05 0.10 0.15 0.0015 0.008

5 The Use ofHigher Plants inBiomonitoring andEnvironmental Bioremediation

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140

A reduction of the TF, while excess of the maximal permissible levels (MPL) in

the soil of the total content of Ni, Zn, and Cd for A. platanoides; of Mn, Zn, and Cd

for Aesculus hippocastanum; and of Mn, Zn, Cd, and Pb for Tilia cordata, is also

observed (Table 5.7). For phytoextraction of heavy metals from soils, most advis-

able species are those in which the barrier mechanisms are not working at high

concentrations of elements in the environment; however, the mechanism of physi-

ological adaptation and TF increases compared to the background site. When select-

ing plants for phytoremediation, it is important to consider also species specicity

in the accumulation of individual elements. The results obtained showed that with

increasing Fe concentration in the medium in all species studied increased the value

of TF for the transport from soil to leaves (Tables 5.5 and 5.6). For the sampling site

(I), an increase in the value of TF for Fe may be due to foliar absorption element

from aerosol particles (the concentration of elements in the air increases). However,

due to the fact that iron is an element, concentrated mainly in the roots [108], for all

kinds of woody plants, TF values for Fe would be considered low compared with

the values for Cd and Zn, due to the barrier function of endoderm and low transport

capacity of elements in the acropetal direction.

The average TF for investigated woody plants is as follows:

Mn—0.01…0.37 and maximal for Betula pendula

Fe—0.01…0.61, maximal for Cornus alba

Ni—0.04…0.18, but for Crataegus monogyna 0.97

Zn—0.06…0.48, maximally for Cornus alba

Cu—0.09…0.68, maximally for Crataegus sanguinea

Cd—0.0002…0.0018

Pb—0.009…0.049 but for Phyladelphus coronarius 0.65

For the species which TF of elements is greater than 0.2 (Tab. 5.7, 5.8) the biore-

mediation of soils from heavy metals will last for 2–10 years (depending on leaves

biomass formed during the growing season, water content in leaves, nutrition areas,

and the species values of TF) in the absence of additional receipt of heavy metals

from the environment.

5.5 Statistical Analysis of the Results

One of the methods to identify the main trends in the intake of elements from the

environment is multivariate statistical analysis. We carried out correlation, cluster,

and factor analyses of the results of the study, which allowed to identify the main

elements—environmental pollutants and their groups. Figure 5.17 shows the den-

drogram of groups of elements, which woody plants bioaccumulated from the

environment.

Results of cluster analysis clearly distinguished group of elements, which can be

divided into several categories:

Group 1: Ca-K are elements-antagonists which play an important role in creating an

osmotic pressure in plant cells and the regulation of processes in the plant.

S.V. Gorelova and M.V. Frontasyeva

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141

Group 2: K-Si are biologically essential elements. Si can be included in the compo-

sition of cell walls of plants.

Group 3: Combined elements are organogenic (essential elements) Ca, K, Si, Fe, Cl,

and Mg.

Group 4: Combined elements—environmental contaminants of the region Fe, Cl,

Mn, Sr, and Ti, associated with soil and ferrous metallurgy.

Group 5: Soil elements and pollutants associated with the processing of ores.

Group 6: The components of the metallurgical and defense industry that can be

combined into a conglomerate entering the plant, including root uptake from soil

resuspention and atmospheric deposition.

Correlation analysis (Table 5.9) of bioaccumulation of elements by woody plants

growing along highways of industrially contaminated city reects the bioaccumula-

tion characteristics of plants and components of the environmental pollution from

the activity of enterprises, highlighting element group and links between them.

Ca-Mg are divalent essential elements with the same way of transport play an

important role in the life of plants (Mg is a component of chlorophyll, the regulator

of photosynthesis processes, Ca is a regulator of cellular immunity, enzymatic pro-

cesses in the cell, component of microtubules and is a cell wall component).

Elements that can substitute each other in the biochemical processes in plants are

K-Rb and Ca-Sr.

Elements associated with the soil particles are Na-Mg, Na-Al, and Ca-Al.

20000

40000

Linkage Distance

60000

80000

1ES

Si Mn Ti Ba Zr Ni VLaNbSmCsU Sc Na

AIThSbHfAsSeMoErCrRbZnSrMgFeK

Fig. 5.17 Cluster analysis of bioaccumulation of the elements from the polluted environment by

woody plants

5 The Use ofHigher Plants inBiomonitoring andEnvironmental Bioremediation

guarino@unisannio.it

Table 5.9 Correlation analysis of the element bioaccumulation in woody plants growing near highways

Na Mg Al K Ca Sc V Cr Mn Fe Ni Co Cu Zn As Se Rb Sr Zr Mo Ag Cd Nd Sm Eu Tb Yb Hf Ta W Hg Th U

Na 1.00

Mg 0.56 1.00

Al 0.74 0.63 1.00

K 0.47 0.46 0.37 1.00

Ca 0.48 0.74 0.57 0.35 1.00

Sc 0.16 0.05 0.30 0.31 0.13 1.00

V 0.42 0.34 0.72 0.05 0.34 0.14 1.00

Cr −0.15 −0.16 0.08 −0.04 −0.24 0.26 0.17 1.00

Mn 0.09 0.34 0.32 0.02 0.32 0.12 0.52 −0.07 1.00

Fe 0.45 0.52 0.82 0.29 0.44 0.31 0.78 0.23 0.53 1.00

Ni −0.21 −0.23 −0.02 −0.07 −0.28 0.04 0.17 0.86 −0.06 0.17 1.00

Co 0.17 0.11 0.31 0.05 0.06 0.49 0.27 0.62 0.08 0.30 0.28 1.00

Cu 0.19 0.13 0.12 0.21 0.24 −0.01 0.01 −0.23 0.16 0.00 −0.27 0.09 1.00

Zn 0.36 0.14 0.14 0.09 0.11 −0.14 0.00 −0.42 0.14 0.01 −0.40 −0.22 0.25 1.00

As 0.36 0.37 0.26 0.27 0.38 −0.08 0.07 −0.68 0.12 0.13 −0.58 −0.44 0.11 0.35 1.00

Se 0.09 0.12 0.27 0.19 0.07 0.62 0.17 0.54 0.03 0.31 0.28 0.59 −0.08 −0.14 −0.23 1.00

Rb 0.26 0.29 0.17 0.65 0.19 0.17 −0.08 −0.12 −0.12 0.12 −0.11 −0.08 0.19 −0.01 0.30 0.13 1.00

Sr 0.49 0.45 0.51 0.29 0.68 0.14 0.31 −0.32 0.13 0.36 −0.33 −0.12 0.07 0.16 0.54 0.17 0.44 1.00

Zr 0.01 0.01 0.19 0.17 −0.06 0.54 0.12 0.74 −0.05 0.25 0.35 0.84 0.06 −0.34 −0.50 0.68 0.04 −0.17 1.00

Mo 0.06 0.01 0.03 −0.06 0.07 −0.11 −0.06 −0.16 0.15 −0.03 −0.14 −0.13 0.02 0.00 0.03 −0.11 −0.06 0.05 −0.15 1.00

Ag −0.06 −0.03 0.12 0.07 −0.14 0.47 0.09 0.76 −0.04 0.20 0.40 0.78 −0.01 −0.20 −0.56 0.72 −0.06 −0.24 0.92 −0.18 1.00

Cd 0.11 0.24 0.28 0.34 0.18 0.37 0.22 0.54 0.10 0.29 0.28 0.61 0.07 −0.16 −0.43 0.55 0.09 −0.05 0.72 −0.11 0.76 1.00

Nd −0.04 0.19 0.16 0.13 0.06 0.39 0.06 0.36 0.14 0.21 0.08 0.46 0.03 0.20 −0.19 0.66 0.01 −0.03 0.57 −0.14 0.74 0.65 1.00

Sm 0.36 0.44 0.68 0.24 0.47 0.37 0.64 −0.08 0.57 0.77 −0.13 0.13 0.07 −0.08 0.27 0.17 0.14 0.46 0.07 −0.02 −0.02 0.19 0.10 1.00

Eu −0.23 −0.10 −0.10 0.18 0.01 0.58 −0.06 0.17 −0.01 −0.02 0.10 0.20 −0.12 −0.32 −0.01 0.57 0.25 0.20 0.32 −0.17 0.34 0.33 0.40 0.22 1.00

Tb −0.06 0.10 0.15 0.02 0.11 0.30 0.11 0.04 0.29 0.23 −0.09 0.24 0.02 0.33 −0.04 0.42 −0.08 0.03 0.20 −0.12 0.43 0.39 0.76 0.19 0.30 1.00

Yb −0.23 −0.13 −0.03 0.02 −0.22 0.31 0.04 0.75 −0.07 0.09 0.64 0.46 −0.21 −0.14 −0.48 0.64 −0.08 −0.27 0.60 −0.22 0.78 0.62 0.66 −0.13 0.49 0.46 1.00

Hf −0.06 0.00 0.14 0.04 −0.11 0.44 0.09 0.76 −0.05 0.21 0.39 0.79 −0.02 −0.21 −0.57 0.67 −0.12 −0.28 0.94 −0.15 0.98 0.75 0.72 −0.04 0.25 0.38 0.73 1.00

Ta −0.18 0.00 −0.10 0.06 −0.07 0.19 −0.06 0.15 0.03 −0.04 0.03 0.23 −0.02 0.21 −0.17 0.43 −0.06 −0.15 0.33 −0.17 0.55 0.51 0.77 −0.09 0.45 0.71 0.68 0.50 1.00

W 0.24 0.25 0.42 0.22 0.32 0.24 0.34 0.18 0.20 0.39 0.07 0.33 0.05 0.07 0.07 0.48 0.19 0.40 0.31 0.00 0.35 0.46 0.49 0.41 0.34 0.40 0.33 0.29 0.35 1.00

Hg −0.02 0.07 0.11 0.23 0.10 0.53 0.01 0.22 −0.08 0.14 0.03 0.32 −0.05 −0.29 0.11 0.54 0.27 0.29 0.51 −0.31 0.42 0.30 0.35 0.26 0.69 0.13 0.38 0.39 0.31 0.34 1.00

Th 0.55 0.50 0.80 0.34 0.39 0.47 0.57 0.34 0.31 0.78 0.14 0.46 −0.09 0.08 0.06 0.54 0.14 0.35 0.44 −0.04 0.44 0.44 0.45 0.60 0.08 0.32 0.25 0.43 0.09 0.51 0.20 1.00

U 0.65 0.50 0.81 0.30 0.38 0.36 0.65 0.11 0.24 0.63 0.00 0.36 0.08 0.05 0.24 0.33 0.15 0.41 0.23 −0.02 0.18 0.31 0.19 0.53 0.03 0.11 0.03 0.18 −0.04 0.39 0.18 0.72 1.00

guarino@unisannio.it

143

Contaminants of city soil and particulate emissions into the atmosphere from

components of metallurgical production and processing of ore are Fe-V, Fe-Mn,

Cr-Ni, Cr-Co, and Sr-As.

Groups with a high correlation between heavy metals and rare earth element-

pollutants of the atmosphere and soil of the city, originate from the sources of plant

pollution—enterprises of the defense industry, instrumentation, and metallurgy.

Multivariate statistical analysis revealed three factors (Table 5.10):

Factor 1 is associated with ores used for production of steel and alloys.

Factor 2 can be attributed to teсhnogenic pollution (metallurgical production)

and soil particles.

Factor 3 is associated with physiological activity of plants.

Table 5.10 Factor analysis of

elements bioaccumulation by

woody plants

(continued)

Factor 1 Factor 2 Factor 3

Na −0.14 0.73 0.11

Mg −0.08 0.72 0.30

Al 0.11 0.92 0.05

S−0.13 0.17 0.04

Cl 0.15 0.29 0.32

K 0.05 0.35 0.54

Ca −0.16 0.67 0.34

Sc 0.53 0.22 0.34

Ti −0.12 0.44 0.00

V 0.13 0.77 −0.21

Cr 0.83 0.00 −0.37

Mn −0.03 0.56 −0.01

Fe 0.23 0.85 −0.05

Ni 0.53 −0.07 −0.44

Co 0.78 0.27 −0.11

Cu −0.13 0.14 0.25

Zn −0.31 0.20 0.13

As −0.60 0.29 0.49

Se 0.77 0.17 0.27

Br −0.03 0.30 0.62

Rb −0.07 0.12 0.60

Sr −0.24 0.51 0.47

Zr 0.90 0.10 0.03

Mo −0.21 0.05 −0.14

Ag 0.96 0.03 0.02

Cd 0.76 0.23 0.21

In −0.28 0.16 0.16

Sb −0.14 0.56 0.21

I−0.09 0.64 0.04

5 The Use ofHigher Plants inBiomonitoring andEnvironmental Bioremediation

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144

5.6 Conclusion

• Woody plants can be good bioindicators of the environmental pollution with

heavy metals (air and soil) affecting the morphological parameters: the develop-

ment of necrosis and leaf chlorosis between 25 and 98% for unresisting species

to pollution. For the purposes of bioindication, species of woody plants such as

Populus nigra, Tilia cordata, Aesculus hippocastanum, Cornus alba, Сrataegus

monogina, and Crataegus sanguinea can be used.

• Analysis of the dust particles on the surface of woody plant leaves using method

of electron scanning microscopy can be used as the qualitative method for analy-

sis of components of aerosol emissions and allows one to establish the presence

and ratio of the components on the leaf surface that reects the air pollution.

• Woody plant species can be used as biomonitor of technogenic emissions due to

their ability to bioaccumulate Сl, V, Mn, Fe, Ni, Zn, Cu, As, Cd, and Pb which

characterize the anthropogenic pollution of soil and air (compared to background

values or control zone):

• Acer platanoides—Cl, V, Mn, Fe, Ni, Cu, As, and Cd

• Aesculus hippocastanum—Ni, Cu, As, and Pb

• Betula pendula—Mn, Fe, Ni, Zn, Cd, and Pb

• Cotoneaster lucidus—Mn, Fe, Ni, Cu, Cd, and Pb

• Crataegus monogyna—Fe and Ni

• Larix sibirica—Fe and Pb

• Philadelphus coronarius—Pb and Sb

• Populus nigra—Mn, Fe, Ni, Zn, Cd, and Pb

• Salix fragilis—Cl, V, Mn, Ni, Zn, As, and Cd

• Tilia cordata—Cl, Mn, Ni, and Cd

Table 5.10 (continued) Factor 1 Factor 2 Factor 3

Ba 0.23 0.39 −0.04

Cs 0.54 0.38 0.16

Nd 0.72 0.11 0.34

Sm 0.04 0.72 0.23

Eu 0.46 −0.20 0.62

Tb 0.43 0.13 0.28

Dy 0.08 0.25 0.43

Yb 0.84 −0.15 0.06

Hf 0.94 0.04 −0.04

Ta 0.54 −0.17 0.38

W 0.38 0.38 0.38

Au −0.04 0.20 0.14

Hg 0.46 −0.02 0.59

Th 0.43 0.77 0.06

U 0.18 0.77 0.07

Expl. var 9.35 8.25 4.12

Prp. totl 0.21 0.19 0.09

S.V. Gorelova and M.V. Frontasyeva

guarino@unisannio.it

145

Most of the studied species are good biomonitors to study contamination of soil

and air with iron.

• Woody plants which form a large biomass of leaves per season are able to absorb

heavy metals to the extent exceeding several times the values characteristic for

the reference plants. This allows to recommend them for phytoremediation of the

environment from heavy metals:

• Aesculus hippocastanum—Ni, Cu, As, and Pb

• Betula pendula—Mn, Fe, Ni, Zn, Cd, and Pb

• Crataegus sanguinea and C. monogina—Cl, V, Fe, Ni, and Cu

• Cornus alba—Cr and Fe

• Cotoneaster lucidus—Mn, Fe, Ni, Cd, and Pb

• Syringa vulgaris—Cl, V, Cr, Fe, and Cu

• Sorbus aucuparia—Fe and Pb

• Philadelphus coronarius—Pb and Sb

• Populus nigra—Fe, Ni, Zn, Cd, and Pb

• Larix sibirica—Fe and Pb

• Analysis of element transfer from the soil into the leaf biomass of woody plants

(TF) has shown that for a number of elements (Ni, Cd, Zn), increase of their con-

tent in the soil leads to decreased transfer of these elements in the leaves of woody

plants. That fact could be a sign of the barrier function of the root system.

Relatively high values of TF (0.2–1) for the elements Fe (species Crataegus san-

guinea and C. monogina, Cotoneaster lucidus, Cornus alba), Cu (species Crataegus

sanguinea and C. monogina, Philadelphus coronarius), Zn (species Acer platanoi-

des, Aesculus hippocastanum, Betula pendula, Crataegus sanguinea, Betula pen-

dula, Symphoricarpos albus, Syringa vulgaris, Populus nigra), Cd (species Acer

platanoides, Betula pendula, Populus nigra, Tilia cordata), and Pb (Philadelphus

coronarius) in woody plants conrm the possibility of their use for phytoremedia-

tion from the enlisted elements.

• When selecting plants for phytoremediation, it is important to consider that spe-

cies of Aesculus hippocastanum, Crataegus sp., Cornus alba, and Populus nigra

are not resistant to the integrated pollution with heavy metals. Species Betula

pendula, Cotoneaster lucidus, Syringa vulgaris, Philadelphus coronarius,

Sorbus aucuparia, and Larix sibirica are resistant to high level of pollution (have

normal vitality) and can be used in the creation of sanitary-protective zones of

metallurgical enterprises and greenbelts.

• Methods of statistical analysis (correlation and factor analysis) in the processing

of the biogeochemical composition of leaves of woody plants allow to clearly

reveal group of elements polluting the study area.

Acknowledgments The nancial support by the

• RFBR grant mob_st. № 09-05-90722 scientic work of the Russian young scientist Svetlana

Gorelova in the Geological Institute of Russian Academy of Sciences on the project: “Study of

biogeochemical variability of plants under conditions of intense anthropogenic soil pollution by

using modern physical and physical-chemical methods of analysis”

5 The Use ofHigher Plants inBiomonitoring andEnvironmental Bioremediation

guarino@unisannio.it

146

• The Black Sea Economic Cooperation—Project Development Fund (Research Contract No.

BSEC/PDF/0018/11.2008—01.2010) «Revitalization of urban ecosystems through vascular

plants: assessment of technogenic pollution impact»). BSEC (Союз Балканских государств)

(Болгария, Греция, Россия, Румыния, Сербия, Турция) The Serbian scientic group also

acknowledges support by MSTD RS (Contract No. 141012).

• RFBR grant r_center_a 13-05-97508 “The study of adaptive characteristics and buffer role of

woody exotic species in the migration of toxic elements in the urban ecosystems”

• RFBR 13-05-97513 r_center_a “Evaluation of the sustainability of the state and prognosis of

natural and anthropogenic ecosystems in Central Russia (Tula Region)”

• RFBR grant r_center_a 15-45-03252 “Biomonitoring of air pollution by industrial emissions in

forest and forest-steppe ecosystems of the Central regions of Russia (example of Tula region)”

The authors are grateful

• for participation in our research, support and fruitful collaboration to the head of the labora-

tory of chemical and analytical investigations of the Geological Institute of the Russian

Academy of Sciences Sergey Lyapunov and Senior Research Scientists Anatoly Gorbunov and

Olga Okina;

• for collection of plant material in ecosystems Tula region, for participation in our reseach, fruit-

ful discussions and scientic inspiration to the head of the RFBR grant 13-05-97513 r_center_a

“Evaluation of the sustainability of the state and prognosis of natural and anthropogenic eco-

systems in Central Russia (Tula Region)”, Ph.D. of Biological Sciences, Associate Professor of

Tula State University and my friend Elena Volkova;

• for help in the statistical data treatment to Lecturer of Radiation Ecology, Radiation Protection

Expert of Radiation Protection and Civil Defense Department, Nuclear Research Center,

Egyptian Atomic Energy Authority, Cairo, Egypt, Wael Badawy;

• for participation in the study of soil, air pollution and woody plants to my diploma students

(L.N. Tolstoy Tula State Pedagogical University)

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5 The Use ofHigher Plants inBiomonitoring andEnvironmental Bioremediation

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A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_6

Chapter 6

Phytoremediation Applications

forMetal- Contaminated Soils

Using Terrestrial Plants inVietnam

BuiThiKimAnh, NgyuenThiHoangHa, LuuThaiDanh, VoVanMinh,

andDangDinhKim

Abstract In the past few decades, the association of economic growth and mining

activities has led to an increase in areas of heavy metal-contaminated soils in

Vietnam. As a developing country, Vietnam has the limited nancial source for envi-

ronmental restoration, so phytoremediation, a low cost and ecologically sustainable

remedial technology, is considered to be a relevant option. To promote the applica-

tion of phytoremediaton for heavy metal-contaminated soils in Vietnam, there have

been several research programs conducted during the last decade. The studies iden-

tied two arsenic (As) hyperaccumulators, Pteris vittata and Pityrogramma

calomelanos, and four grasses suitable for treatment of lead (Pb)- and zinc (Zn)-

contaminated soils, Eleusine indica, Cyperus rotundus, Cynodon dactylon, and

Equisetum ramosissimum, of which E. indica was found as Pb hyperaccumulator.

All of these species are indigenous and naturally adapted to heavy metal-

contaminated habitats. Three plant species, P. vittata, P. calomelanos, and E indica

and one introduced plant species, Vetiveria zizanioides, were subjected to further

evaluation of their heavy metal removal potential under greenhouse and eld condi-

tions. The results of greenhouse experiments showed that two fern species, P. vittata

and P. calomelanos, are effective in the accumulation of soil As in roots and fronds;

E. indica can absorb high concentration of both Pb and Zn in roots. Under eld

conditions, the combination of P. vittata, P. calomelanos, and V. zizanioides or

P. vittata, E. indica and V. zizanioides is very effective in treatment of soils con-

taminated with low or moderate concentration of As and Pb in short time (3years).

B.T.K. Anh • D.D. Kim (*)

Institute of Environmental Technology, Vietnam Academy of Science and Technology,

18 Hoang Quoc Viet, Hanoi, Vietnam

e-mail: dangkim.iet@gmail.com

N.T.H. Ha

VNU University of Science, Vietnam National University, 334 Nguyen Trai, Hanoi, Vietnam

L.T. Danh

College of Agriculture and Applied Biology, University of Can Tho, Can Tho, Vietnam

V. VanMinh

University of Education, University of Da Nang, Danang, Vietnam

guarino@unisannio.it

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Through these studies, phytoremediation has been demonstrated to be feasible for

the remediation of heavy metal-contaminated soils in Vietnam.

Keywords Mining activity • Heavy metal-contaminated soil • Indigenous hyperac-

cumulator • Potential species for phytoremediation • Bac Kan and Thai Nguyen

provinces

6.1 Introduction

An increasingly industrialized global economy and rapid rise in world population

over the last century have led to dramatically elevated releases of anthropogenic

chemicals, particularly heavy metals, into the environment [1]. The annual world-

wide release of heavy metals reached 22,000 metric ton (t) for cadmium, 939,000t

for copper, 783,000t for lead, and 1,350,000t for zinc over recent decades [2].

Sources of heavy metal released into soil environments include mining, smelting of

metalliferous, electroplating, gas exhaust, energy and fuel production, fertilizer and

pesticide application, sewage sludge treatment, warfare and military training [3].

Hard-rock mining that is the largest producer of heavy metal waste takes place in all

of the continents of the world with the exception of Antarctica [4].

Heavy metal-contaminated soils have caused serious problems threatening eco-

logical systems and human health, recently attracted considerable public attention.

Several metals, such as Cu and Zn, are essential for biological systems and must be

present within a certain concentration range [5], at high concentrations they will

become toxic. Other metals, such as Cd, As, Hg, and Pb, have not been found to have

any function in plants and animals, and very toxic for biological life even occurred at

low concentrations. Metals can act in a deleterious manner by blocking essential

functional groups, displacing other metal ions, or modifying the active conformation

of biological molecules [6]. Exposure to high levels of these metals can cause adverse

effect on human and wildlife [7]. Toxic heavy metals can mutate DNA resulting in

carcinogenic effects in animals and human [8, 9]. Lead causes neurological damage

in children leading to reduced intelligence, loss of short-term memory, learning dis-

abilities, and coordination problems [7]. The effects of arsenic include cardiovascu-

lar problems, skin cancer and other skin effects, peripheral neuropathy [10]. Cadmium

accumulates in the kidneys and is responsible for a wide range of kidney diseases

[10]. The principal health risks associated with mercury are damage to the nervous

system, with such symptoms as uncontrollable shaking, muscle wasting, partial

blindness, and deformities in children exposed in the womb [10].

Concentrations of heavy metals that have exceeded safety levels in soil should be

treated [11]. There are several methods used for soil remediation, including

chemical, physical, and biological techniques. Physical treatments involve removal

from contaminated sites (soil excavation), deep burial (landlling), and capping,

while chemical methods use strong acids and chelators to wash polluted soils.

B.T.K. Anh et al.

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159

These approaches are expensive, impractical, and at times impossible to carry out,

as the volume of contaminated materials is very large. Furthermore, they irrevers-

ibly affect soil properties, destroy biodiversity, and may render the soil useless as a

medium for plant growth [7]. Recently, phytoremediation that refers to a diverse

collection of plant-based technologies using either naturally occurring or geneti-

cally engineered plants to clean contaminated environments [12] represents a novel,

environmentally friendly, and cost-effective technology and attracts the attention of

publics and scientists worldwide.

The idea of using plants to extract metals from contaminated soil was reintro-

duced and developed by Utsunamyia [13] and Chaney [14] and the rst eld trial on

Zn and Cd phytoextraction was conducted by Baker etal. [15]. Some plants which

grow on metalliferous soils have developed the ability to accumulate massive

amounts of indigenous metals in their tissues without symptoms of toxicity [15].

Depending on storage sites of heavy metals in plants, phytoremediation technology

can be used for containment (phytostabilization) and removal (phytoextraction)

purposes [16]. Phytostabilization involves plants to stabilize contaminants by heavy

metal retention in roots. Phytoextraction uses plants to absorb metals from soils and

translocate them to harvestable shoots where they are collected.

In Vietnam, the increase in mining activities associated with the economic growth

has resulted in the increased areas contaminated with heavy metals in recent years.

Mining, ore processing, and disposal of tailings provide obvious sources of heavy

metal contamination in the mine area and surroundings. The contaminated soils

require prompt remediation, and phytoremediation is considered to be one of the

best demonstrated available technologies for such purpose [17]. Field applications

of phytoremediation have only been reported in developed countries in spite of its

cost-effectiveness and environment-friendliness. In most developing countries, it is

yet to become commercially available technology possibly due to the inadequate

awareness of government and public about its inherent advantages and principles of

operation [18]. Since the last decade, therefore, there have been several groups of

Vietnamese scientists studying on the use of plant for removal of heavy metals from

soils in order to promote the application of phytoremediation. This chapter sum-

marizes the recent research and application related to phytoremediation of heavy

metal-contaminated soils in Vietnam, including investigation and selection of native

hyperaccumulators of arsenic (As), cadmium (Cd), lead (Pb), and zinc (Zn) natu-

rally grown on heavy metal-contaminated habitats; and evaluation of heavy metal

removal potential of selected plants under greenhouse and eld conditions.

6.2 Selection ofPotential Plants forHeavy Metal Removal

fromSoils

The selection of plants species for phytoremediation is possibly the most important

factor determining the heavy metal removal efciency. Other aspects such as the

ecological and environmental protection should be taken into account as selecting

the phytoremediating plants.

6 Phytoremediation Applications forMetal-Contaminated Soils Using Terrestrial…

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The success of phytoextraction is dependent on two important characteristics of

plants: the ability to produce large quantities of biomass rapidly and the capacity to

accumulate large quantities of environmentally important metals in the shoot tissue

[19–22]. Hyperaccumulators have been characterized by high heavy metal accumu-

lating potential, small size, and slow growth, while the common non- hyperaccumulators

have low potential for metal bioconcentration that is often traded off by the produc-

tion of signicant biomass [23]. In environmental aspect, most of hyperaccumulators

are classied as weedy species that can be invasive and endanger the harmony of

ecosystem in the new environments, while some crops are palatable and pose a risk

to grazing animals. Therefore, the choice of metal hyperaccumulators or common

non-accumulator species for phytoremediation is one of the most debated controver-

sies in the eld.

Many researchers have supported for the use of non-accumulator species, while

others have promoted the application of natural hyperaccumulators. In the study of

Ebbs etal. [23], Brassica juncea (also known as Indian mustard) was more effective

in removing Zn from soil than Thlaspi caerulescens (a well-known Zn hyperaccu-

mulator) although the Zn concentration in its biomass was about one-third the con-

centration of Zn in Thlaspi caerulescens. The advantage is due primarily to the fact

that B. juncea produces ten times more biomass than T. caerulescens. Nevertheless,

Chaney etal. [24] analyzed the rate of Zn and Cd removal by non-accumulators

crops and came to the remark that these crops could not remove enough metal to

support phytoextraction. In addition, the high concentrations of heavy metals at

many contaminated sites may cause toxicity to crop species and signicant biomass

reduction. In support of this, several maize (one of the most productive crops)

inbred lines have been identied which can accumulate high levels of Cd [25].

However, these lines were susceptible to Zn toxicity and, therefore, could not be

used to cleanup soils at the normal Zn:Cd ratio of 100:1 [24]. In addition, when

appropriate disposal is an important regulatory concern, the use of lower biomass

producing hyperaccumulator species would be an advantage because less conta-

minated biomass will have to be handled. Moreover, the use of native plants for

phytoremediation is more effective because such plants respond better to the stress

conditions at the site than would plants introduced from other environments [26].

Consequently, the selection of native and hyperaccumulators for phytoremediation

purposes is one of the most important steps to ensure the success of phytoremedia-

tion programs.

In order to select the indigenous hyperaccumulators of heavy metals, there were

two investigations conducted at mining sites in Bac Kan and Thai Nguyen province,

northern Vietnam, where the most mining activities are done in Vietnam. Soil

analyses of mining sites in this region showed that soils have been heavily contami-

nated with a range of heavy metals, including Mn, Zn, As, Cd, and Pb (Table 6.1).

Therefore, the objective of these surveys was to search for hyperaccumulators of

Mn, Zn, As, Cd, and Pb.

B.T.K. Anh et al.

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6.2.1 Bac Kan Province, Northern Vietnam

6.2.1.1 Site Description

The mine site is situated in Cho Don district, Bac Kan province, northern Vietnam

(Fig. 6.1). This is one of the biggest Pb–Zn mines in Vietnam. Mining activities

started from eighteenth century and still have been active currently [27]. Rainy sea-

son starts from April to September, and dry season from October to March. The

average rainfall is around 100–600 and 8–22mm month−1 in rainy and dry season,

respectively. Humidity is 76–88% and 35–45% in rainy and dry season, respectively

[27]. The highest and lowest average temperature is 31–36°C and 10–11°C, respec-

tively [27]. The main ore minerals are galena (PbS), sphalerite (ZnS), pyrotin (FeS),

pyrite (FeS2), chalcopyrite (CuFeS2), and arsenopyrite (FeAsS) [27]. In addition,

high concentration of Mn was obtained in Pb–Zn ore and in sphalerite which was

9892–20,500mg kg−1 and 0.09–0.23% [28, 29]. High concentrations of Pb, Zn, As,

and Mn may result in the leaching of these heavy metals into the surrounding envi-

ronments via mining activities.

6.2.2 Plant Accumulation andTranslocation ofHeavy Metals

High concentrations of heavy metals in the soil and water may result in high levels

of these elements in the collected plant samples. The concentrations of all heavy

metals varied greatly among sites and plant species [30]. The highest concentrations

Table 6.1 Family, species, and number of plant samples around and outside of the mine site

STT Code name Family Species n

1 Age Asteraceae Ageratum houstonianum Mill. 12

2 Bid Asteraceae Bidens pilosa L. 6

3 Dip Athyriaceae Diplazium esculentum (Retz.) Sw. 9

4 Ele Poaceae Eleusine indica (L.) Gaertn 9

5 Hou Saururaceae Houttuynia cordata Thunb. 9

6Kyl Cyperaceae Kyllingia nemoralis 12

7 Lee Poaceae Leersia hexandra Sw. 9

8Lyg Lygodiaceae Lygodium exuosum (L.) Sw. 6

9 Nep Lomariopsidaceae Nephrolepis cordifolia (L.) Presl. 9

10 Pte Pteridaceae Pteris vittata L. 24

11 Sac Poaceae Saccharum spontaneum L. 9

12 Sci Cyperaceae Scirpus juncoides Roxb. 9

13 Sel Selaginelaceae Sellaginella delicatula (Desv.) Alst 15

14 The Thelypteridaceae Thelypteris noveboracensis 9

15 Thy Poaceae Thysanolaena latifolia 12

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Fig. 6.1 Map showing the location of the Bac Kan sampling sites

B.T.K. Anh et al.

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of heavy metals (mg kg−1-DW) in the plant root were found in K. nemoralis for Mn

(2130), N. cordifolia for Zn (3780), P. vittata for As (861), L. hexandra for Cd

(13.3), and A. houstonianum for Pb (2080); those in the shoot were found in

K. nemoralis for Mn (1990), N. cordifolia for Zn (1710), P. vittata for As (2300),

A. houstonianum for Cd (19.0), and E. indica for Pb (2010) (Tables 6.2 and 6.3).

Among all plant species in the present study, highest concentrations of Mn, Zn, and

As in the root and shoot were observed in K. nemoralis, N. cordifolia, and P. vittata,

respectively.

Almost all collected plant species accumulated higher concentrations of Zn, As,

and Pb than their toxicity threshold levels in plants. The concentrations of Mn, Zn,

As, Cd, and Pb in the shoot of A. houstonianum, E. indica, and H. cordata were

within and above the toxic levels for plant species (Tables 6.3 and 6.4). In addition,

all plant species can adapt very well with the soil that highly contaminated with

multiple heavy metals (Table 6.4). These results may indicate that plant species

growing on the site contaminated with heavy metals were tolerant of these metals.

P. vittata showed great potential for accumulating As in the shoot (Table 6.2). The

concentrations of As in the shoot of P. vittata L. were signicantly higher than those

in other plant species. Concentrations of As in the root were signicantly higher than

those in B. pilosa, D. esculentum, K. nemoralis, L. exuosum, S. spontaneum, and

T. noveboracensis. Mn concentrations in root of P. vittata were signicantly higher

than those in B. pilosa, L. exuosum, and S. spontaneum. However, those in the shoot

were signicantly lower than those in H. cordata and K. nemoralis. Cd concentrations

in the root of this species were lower than those in H. cordata and S. juncoides.

E. indica accumulated the highest concentration of Pb in the shoot among 15

collected plant species (Table 6.2). Pb concentration in E.Indica roots was higher

than in roots of B. pilosa L., K. nemoralis, L. exuosum, N. cordifolia, S. sponta-

neum, and S. delicatula. Concentrations of the heavy metals in the root and shoot of

E. indica were signicantly higher than those in L. exuosum, those of As were

signicantly higher than those in D. esculentum, K. nemoralis, L. exuosum, and

T. noveboracensis. Cd concentrations in the root and shoot were signicantly higher

than those in B. pilosa, L. exuosum, S. spontaneum, and T. latifolia.

6.2.3 Potential Plant Species forPhytoremediation

ofContaminated Soils

The typical characteristics of an ideal plant species for phytoremediation are as fol-

lows: (1) a hyperaccumulator of metals which in aboveground tissues; (2) a high

and fast-growing biomass and be repulsive to herbivores to avoid the escape of

accumulated metals to the food chain; (3) BCF and TF values higher than 1; (4) a

widely distributed, highly branched root system; (5) easy to cultivate and a wide

geographic distribution; and (6) relatively easy to harvest [31, 32].

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Table 6.2 Concentrations (mgkg−1-DW) of Mn, Zn, and As in plant growing in and outside of the mine (N=6–24)

No Code name

Mn Zn As

Root Shoot Root Shoot Root Shoot

1 Age 565 (85.2; 371–704) 555 (94.3; 238–1030) 514 (28.9; 197–1130) 563 (37.8; 152–1210) 272 (4.72; 162–350) 204 (15.9; 203–209)

2 Bid 130 (96; 115–155) 213 (142; 196–238) 191 (123; 149–265) 216 (112; 178–286) 183 (32.2; 156–213) 440 (34.5; 285–528)

3 Dip 1600 (221; 1190–2010) 215 (80.4; 81.7–349) 574 (56.2;544–604 ) 176 (26.3; 98–255) 82.5 (3.69; 73.3–91.7) 17.2 (19.0; 10.2–24.2)

4 Ele 385 (102; 142–628) 139 (90.4; 130–147) 1159 (57.6; 178–2140) 372 (41.9; 119–625) 339 (4.77; 230–448) 364 (1.55; 204–524)

5 Hou 987 (193; 941–1030) 710 (307; 434–985) 537 (53.8; 465–609) 319 (50.5; 234–404) 416 (2.99; 284–548) 240 (0.78; 197–283)

6Kyl 1150 (210; 418–2130) 1440 (303; 1120–1990) 453 (76.0; 267–795) 239 (63.0; 173–284) 103 (0.99; 84.8–119) 57.2 (0.78; 29.6–97.5)

7 Lee 326 (277; 300–352) 147 (113; 118–176) 638 (138; 433–844) 170 (158; 142–197) 368 (43.5; 278–458) 191 (41.8; 182–199)

8Lyg 119 (82.6; 97.2–149) 204 (84.7; 187–225) 38.1 (26.3; 31.3–43.8) 71.5 (21.7; 53.6–96.1) 20.1 (0.83; 16.3–23.1) 18.6 (13.4; 14.7–21.5)

9 Nep 614 (187; 566–662) 299 (156; 240–357) 2690 (183; 1590–3780) 1322 (158; 934–1710) 299 (52.0; 259–340) 176 (39.0; 158–193)

10 Pte 634 (254; 177–1200) 165 (98.1;65.0–358 ) 1150 (110; 247–3710) 388 (80.6; 155–1430) 464 (135; 291–861) 1180 (149; 251–2300)

11 Sac 138 (134; 129–146) 112 (120; 75.0–150) 455 (123; 179–731) 167 (74.3; 106–237) 185 (54.0; 173–197) 174 (32.6; 166–182)

12 Sci 574 (183; 215–933) 301 (135; 225–376) 204 (126; 165–242) 73.0 (41.7; 59–87) 469 (0.83; 237–700) 190 (13.4; 183–197)

13 Sel 476 (153; 148–717) 278 (207; 117–430) 341 (89.3; 222–438) 246 (131; 206–278) 250 (87.7; 91–362) 89.3 (69.6; 78–164)

14 The 446 (131; 217–676) 212 (141; 159–365) 874 (93.4; 300–1450) 325 (112; 168–581) 48.2 (39.5; 42.1–54.3) 33.1 (25.5; 29.8–56.5)

15 Thy 304 (289; 110–532) 155 (118; 112–212) 233 (126; 196–288) 161 (94.6; 117–190) 333 (18.3; 172–480) 128 (16.9; 63.1–184)

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Table 6.3 Concentrations (mgkg−1-DW) of Cd and Pb in plant growing in and outside of the mine

(N=6–24)

No Code name

Cd Pb

Root Shoot Root Shoot

1 Age 6.72 (0.50;

3.35–12.6)a

7.52 (0.56;

1.35–19.0)

1299 (11.8; 770–2080) 844 (4.87;

631–1070)

2 Bid 1.41 (0.75;

1.23–2.14)

0.60 (0.36;

0.52–0.75)

411 (74.1; 297–485) 505 (58.5;

378–563)

3 Dip 10.9 (0.71;

8.98–12.8)

0.71 (0.17;

0.70–0.72)

1307 (37.5; 949–1670) 200 (8.90;

154–246)

4 Ele 6.55 (0.75;

3.75–9.35)

5.14 (0.34;

2.93–7.34)

1840 (65.1;1810–1870 ) 1300 (15.3;

595–2010)

5 Hou 12.3 (0.15;

11.1–135)

9.22 (0.08;

8.99–9.46)

843 (11.1; 606–1080) 1060 (5.11;

999–1130)

6Kyl 4.17 (0.15;

2.95–4.91)

2.18 (0.08;

1.70–2.53)

556 (21.0; 281–712) 386 (17.0;

116–702)

7 Lee 11.3 (0.54;

9.44–13.3)

0.77 (0.36;

0.56–0.97)

1910 (17.0; 1880–1940) 357 (8.79;

211–503)

8Lyg 0.31 (0.12;

0.25–0.42)

0.18 (0.04;

0.12–0.24)

43.4 (8.29; 28.7–52.4) 9.42 (1.03;

8.44–10.1)

9 Nep 7.72 (0.64;

7.24–8.20)

8.68 (0.35;

7.95–9.41)

366 (63.7; 276–456) 501 (74.3;

492–510)

10 Pte 5.25 (0.48;

0.94–10.6)

0.72 (0.24;

0.37–0.99)

1070 (234; 453–1840) 544 (83.5;

92.0–781)

11 Sac 0.47 (0.38;

0.37–0.58)

0.25 (0.20;

0.25–0.38)

563 (42.0; 438–688) 807 (65.8;

781–832)

12 Sci 14.8 (0.12;

11.8–17.7)

5.27 (0.06;

4.58–5.97)

1650 (8.29; 1520–1780) 793 (1.03;

721–864)

13 Sel 2.59 (0.57;

2.40–2.79)

2.20 (0.35;

0.57–3.50)

359 (13.7; 234–613) 408 (6.86;

142–865)

14 The 2.22 (0.35;

2.16–2.28)

0.74 (0.27;

0.73–0.75)

865 (78.2; 842–888) 411 (89.9;

191–631)

15 Thy 0.70 (0.34;

0.47–0.85)

0.51 (0.15;

0.25–0.70)

667 (118; 474–886) 367 (23.5;

242–511)

aAverage (reference, min–max)

Table 6.4 Concentrations (mgkg−1) of heavy metals in the soil in and outside of the mine area

Site Mn Zn As Cd Pb

19270±350 7150±1420 2290±440 70.2±15.1 8780±790

2 4940±290 5780±1790 5630±2910 93.7±18.0 9090±1940

3 3620±810 1720±370 2450±660 17.2±5.24 4360±1200

4 1730±220 1570±390 489±307 3.71±1.79 3360±910

5 2410±1010 1470±130 2130±810 4.72±2.22 2470±580

6 4010±850 2010±740 1550±220 25.1±9.7 3340±720

7 3210±770 4620±840 538±301 37.0±25.3 4180±960

8 4270±260 6210±300 858±66 75.2±5.7 8290±710

Reference 817±167 88.8±2.5 4.77±0.24 1.00±0.05 70.8±21.3

Values present means±standard deviations (N=3–9)

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Hyperaccumulators are dened as plants with leaves able to accumulate at least

100mg kg−1of Cd; 1000mg kg−1of As or Pb; or 10,000mg kg−1 of Mn or Zn (dry

weight) when grown in a metal-rich environment [33–35]. Among all plant species,

hyperaccumulation levels (mg kg−1-DW) were obtained in P. vittata (1180) for As

(Table 6.2), in E. indica (1300) and H. cordata (1060) for Pb (Tables 6.2 and 6.3).

Of which, P. vittata has been reported as a well-known hyperaccumulator of As [36].

A. houstonianum and E. indica have been reported to hyperaccumulate Pb [32, 37].

H. cordata is a hyperaccumulator of As (1140mg kg−1-DW); however, to the best of

our knowledge, no previous study has reported the hyperaccumulation of Pb in

H. cordata.

Bioconcentration factor (BCF) values of Mn, Zn, As, Cd, and Pb of 15 plant spe-

cies varied within 0.04–0.76, 0.04–0.78, 0.01–1.19, 0.02–2.00, and 0.02–0.46,

respectively (Table 6.5). BCF values of plants for Mn, Zn, and As at the uncontami-

nated site were signicantly higher than those at the mine site. This is possibly due

to low concentrations of heavy metals in associated soils outside of mining area

(Table 6.4). BCF values of Mn and Zn were correlated (p < 0.05). BCF values

higher than 1 were only observed in H. cordata (2.00) and P. vittata (1.19) for Cd

and As, respectively. This result reected high accumulation capacity of heavy met-

als by these species. Most BCF values were found to be lower than 1. This is pos-

sibly due to the existence of heavy metals in various geochemical forms in soils

(water-soluble, exchangeable, bound to carbonate, bound to Fe-Mn oxide, bound to

organic matter, and residual forms) [38–40]. In addition, the possible source of

heavy metals was derived from a sulde deposit, consequently, these heavy metals

are assumed to partially exist as suldes. The occurrence of heavy metals in suldes,

Table 6.5 Bioconcentration factors of plant growing around the mine

Code Mn Zn As Cd Pb

Age 0.11±0.08a0.17±0.08 0.09±0.05 0.28±0.12 0.12±0.03

Bid 0.11±0.08 0.11±0.08 0.35±0.15 0.12±0.11 0.39±0.22

Dip 0.12±0.10 0.13±0.07 0.05±0.03 0.29±0.19 0.06±0.02

Ele 0.07±0.04 0.08±0.01 0.43±0.39 0.63±0.45 0.46±0.28

Hou 0.76±0.34 0.73±0.59 0.12±0.04 2.00±1.87 0.38±0.05

Kyl 0.22±0.01 0.11±0.02 0.01±0.01 0.14±0.03 0.04±0.02

Lee 0.45±0.19 0.27±0.16 0.52±0.42 0.07±0.08 0.19±0.11

Lyg 0.21±0.13 0.78±0.30 0.36±0.11 0.19±0.07 0.02±0.02

Nep 0.07±0.03 0.26±0.11 0.40±0.26 0.61±0.67 0.24±0.02

Pte 0.10±0.15 0.20±0.15 1.19±0.50 0.10±0.11 0.11±0.06

Sac 0.04±0.01 0.07±0.07 0.12±0.03 0.02±0.01 0.26±0.02

Sci 0.10±0.03 0.04±0.01 0.15±0.01 0.58±0.47 0.28±0.13

Sel 0.18±0.22 0.26±0.27 0.23±0.23 0.31±0.30 0.09±0.08

The 0.05±0.03 0.08±0.04 0.12±0.13 0.13±0.05 0.09±0.05

Thy 0.06±0.03 0.05±0.03 0.11±0.06 0.05±0.08 0.08±0.03

Average 0.18 0.22 0.28 0.37 0.19

aMean±standard deviation

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combined with the fact that the poor structure of soil developed in mine tailings may

reduce metal availability to root over short periods of time [41].

Translocation factor (TF) values of Mn, Zn, As, Cd, and Pb in 15 plant species

varied within 0.12–1.72, 0.28–1.88, 0.22–2.98, 0.07–1.12, and 0.18–3.46, respec-

tively (Table 6.6). It is noted that most TF values of heavy metals in this study were

lower than 1. This is in line with the result reported by Stoltz and Greger [42] that

most of the plant species growing on mine tailings have a restricted translocation of

metals and As to the shoot. The restriction of upward movement from root to shoot

can be considered as one of the tolerance mechanisms [43]. The average TF values

of Mn, Zn, As, Cd, and Pb of plants growing at the uncontaminated site were 0.28,

0.50, 3.53, 0.30, and 0.41, respectively. The TF value of Cd was signicantly lower

than that of Mn, As, and Pb. TF values of plants for Mn, Zn, and As at the uncon-

taminated site were signicantly higher than those at the mine site. The translocation

of Mn from root to shoot in D. esculentum, P. vittata, and T. noveboracensis was

signicantly lower than that in other species. TF values of As in B. pilosa, E. indica,

and P. vittata were signicantly higher than those in other plants. Signicantly

higher TF values of Pb in B. pilosa, H. cordata, N. cordifolia, and S. spontaneum

than those in other species were also observed. B. pilosa showed the high capacity

to translocate multiple heavy metals from the root to the shoot (Table 6.6).

Among all plants collected in the present study, P. vittata is the most widely

distributed species. The results of the present study were in agreement with the

previous study that P. vittata L. is an efcient As hyperaccumulator [36]. The high-

est concentrations of As, Pb, Zn, Mn, and Cd in shoot of P. vittata L. were 358,

1430, 2300, 0.99, and 784mg kg−1-DW, respectively. TF values exceeded 1 were

Table 6.6 Translocation factors of plant growing around the mine

Code Mn Zn As Cd Pb

Age 0.97±0.08a1.15±0.08 0.84±0.05 0.81±0.12 0.75±0.23

Bid 1.64±0.08 1.14±0.08 2.41±0.15 0.43±0.11 1.23±0.22

Dip 0.12±0.10 0.31±0.07 0.22±0.03 0.07±0.09 0.18±0.02

Ele 0.57±0.04 0.48±0.01 1.03±0.39 0.78±0.75 0.70±0.28

Hou 0.73±0.34 0.58±0.59 0.61±0.04 0.76±0.87 1.35±0.35

Kyl 1.65±0.01 0.63±0.02 0.54±0.11 0.53±0.03 0.63±0.22

Lee 0.46±0.19 0.28±0.16 0.54±0.42 0.07±0.08 0.19±0.11

Lyg 1.72±0.13 1.88±0.30 0.78±0.11 0.58±0.07 0.22±0.09

Nep 0.50±0.03 0.52±0.11 0.59±0.26 1.12±0.67 3.46±0.62

Pte 0.27±0.15 0.45±0.25 2.98±0.50 0.29±0.11 0.43±0.16

Sac 0.80±0.01 0.43±0.07 0.94±0.23 0.51±0.11 1.50±0.42

Sci 0.72±0.03 0.39±0.12 0.54±0.11 0.36±0.27 0.48±0.13

Sel 0.63±0.22 0.75±0.37 0.43±0.23 0.85±0.40 0.99±0.28

The 0.41±0.03 0.31±0.04 0.64±0.13 0.33±0.15 0.48±0.15

Thy 0.65±0.03 0.70±0.03 0.39±0.16 0.70±0.28 0.54±0.23

Ave r 0.18 0.22 0.28 0.37 0.19

aMean±standard deviation

6 Phytoremediation Applications forMetal-Contaminated Soils Using Terrestrial…

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168

obtained for As. In addition, P. vittata L. has considerable biomass, grows fast, and

propagate easily [36, 44, 45]. Therefore, this plant has high potential for phytoreme-

diation of multi-metals, especially for As [44, 45], Zn and As [46], Cd and As [47],

and multiple heavy metal-contaminated soils [32].

Of the three Pb hyperaccumulators identied in this study, E. indica (L.)

accumulated highest concentrations of Pb in the shoot. H. cordata had the highest

translocating factor of Pb from root to shoot (TF=1.35). TF values of E. indica and

A. houstonianum and BCF values of all Pb hyperaccumulators were lower than 1.

This study was conducted to assess the phytoremediation potential of plants grow-

ing on a site contaminated with heavy metals. Results of this research indicated that

among 15 plant species being collected, P. vittata L. is a good candidate for phytore-

mediation of As; A. houstonianum, E. indica, and H. cordata are potential species

for phytoremediation of Pb. Further studies are required to conrm the phytoreme-

diation potential of those plant species through greenhouse and eld experiments as

well as to establish the agronomic requirements and management practices in order

to investigate their whole phytoremediation possibilities.

6.3 Selection ofIndigenous Plants Suitable

forPhytoremediation inThai Nguyen Province

The study was performed at four mining sites located at two districts of Thai Nguyen

province, northern Vietnam: Tan Long (Zn/Pb mine) and Trai Cau (Fe mine) site in

Dong Hy district, Ha Thuong (Ti/Sn mine) and Yen Lang (coal mine) in Dai Tu

district. Soil samples were collected at the same place with plant samples (Figs. 6.2

and 6.3)

This research was conducted to determine soil concentrations of As, Pb, Cd, and

Zn at four mining sites of Thai Nguyen province as well as to identify indigenous

potential plants for phytoremediation. Total 33 indigenous plants and 12 soil in situ

plant samples in these areas were collected for heavy metal analysis. The soils of

surveyed mining areas contained 181.2–6754.3mg kg−1 As, 235.5–4337.2mg kg−1

Pb, 0.8–419mg kg−1 Cd, and 361.8–17565.1mg kg−1 Zn depending on the charac-

teristics of each mining site. As compared to the upper limit of As (15 mg kg−1), Cd

(1.5 mg kg−1), Pb (70 mg kg−1), and Zn (200 mg kg−1) for industrial soil in Vietnam

[48], these soils are much higher than standard values.

The collected 33 plant species can grow at the mine tailings or in the soils

affected by mining waste. The heavy metal concentrations in their roots and shoots

of these plant species were evaluated. In the total of these selected plants, only six

potential indigenous plant species of Thai Nguyen province was presented in the

Table 6.7. The results showed that two ferns, Pteris vittata and Pityrogramma calo-

melanos were capable of accumulating high arsenic concentrations. As concentra-

tions in shoot and root of P. vittata were 5877 and 2643mg kg−1, respectively, while

these values of P. calomelanos were 2426 and 2256 mg kg−1. Remarkably, a large

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Fig. 6.2 Location of survey areas in Thai Nguyen province

Fig. 6.3 Some sampling sites in Bac Kan and Thai Nguyen

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Table 6.7 Heavy metal concentration in shoots and roots of six potential indigenous plant species of Thai Nguyen province

Plant species

As (mg kg−1) Pb (mg kg−1) Cd (mg kg−1) Zn (mg kg−1)

Shoot Root Shoot Root Shoot Root Shoot Root

Pteris vittata L. 5876.5±99.6 2642.5±72.3 9.4±1.4 10.2±1.7 0.4±0.1 1.3±0.3 152.3±12.7 220.5±23.5

Cynodondactylon L. 44.1±3.5 765.5±23.2 538.5±25.2 4579.6±88.5 4.4±0.7 34.6±4.8 912.8±42.5 15.6±3.8

Eleusine indica L. 25.2±2.8 236.0±15.1 664.5±45 4638.2±210.4 9.3±1.3 26.5±2.5 4346.8±157.9 3108.7±213.5

Equisetum ramosissimum

(Vauch)

28.2±2.6 34.3±4.1 455.2±32.6 1025.7±65.8 9.2±1.5 29.0±3.6 1346.2±130.2 3756.9±145.7

Cyperus rotundus L. 19.9±1.7 37.7±3.6 941.3±35.2 1560.2±113.5 7.2±2.1 9.5±2.4 1201.4±147.3 2194.4±155.7

Pityrogramma

calomelanos L.

2426.3±104.5 2256.0±123.4 49.9±5.6 85.4±7.4 1.0±0.3 1.1±0.5 368.6±15.7 230.8±24.6

Values are mean±standard deviation of three replicates

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amount of As from roots of these ferns transported to shoot, facilitating its removal

from the soil. Many research results have reported that they are two As hyperaccu-

mulating ferns [28, 29, 32, 36, 49–51]. None of the collected plant species had high

Cd accumulating ability.

Zn accumulating ability in some investigated plant species was quite high.

E. ramosissimum, C. rotundus, and E. indica can accumulate Zn in their shoots with

1346, 1201, and 4347mg kg−1, respectively, and in their roots with 3757, 2194, and

3109mg kg−1 Zn, respectively. As indigent plants, they can easily adapt to the local

conditions being also potential for phytoremediation.

Our ndings in Thai Nguyen province indicate that two ferns P. vittata and

P. calomelanos are suitable for As treatment in the mining soil of Ha Thuong, Dai

Tu, district (Table 6.7). Four grasses, E. indica, C. dactylon, C. rotundus, and

E. ramosissimum are potential for Pb, Zn removal from soils. Some research results

reported that E. indica is Pb hyperaccumulator [32, 37].

6.4 Some Research Results inGreenhouse Experiment

ofPotential Plant Species

Based on the screening results, three species, namely, P. vittata, P. calomelanos, and

E. indica were selected with an introduced plant Vetiveria zizanioides and a crop

plant Brassica juncea for evaluation under greenhouse conditions.

6.4.1 Pteris vittata andPityrogramma calomelanos

The obtained results from greenhouse experiments showed that Pteris vittata and

Pityrogramma calomelanos can grow in the mining soil containing 15,146ppm As.

Although they are As hyperaccumulators, the plants still also have the ability to

accumulate Cd, Pb, and Zn. Pteris vittata and Pityrogramma calomelanos can toler-

ate 5000 and 4000 mg kg−1 Pb (concentration of Pb was established by adding

Pb(NO3)2 in the garden soil); 1200 and 300mg kg−1 Cd (concentration of Cd was

established by adding Cd(NO3)2 in the garden soil), respectively. The highest level

of As accumulation in Pteris vittata and Pityrogramma calomelanos are 6042

and 4034 mg kg−1 (in the fronds); 3756 and 2256 (in the roots), respectively.

Concentration of As, Cd, Pb, and Zn in Pteris vittata were comparable to those

found by An etal. [46] and Ha etal. [32]. From 3 to 4months after growing there is

appropriate time for harvesting plant biomass if applied in practical processing

(Fig. 6.4).

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6.4.2 Eleusine indica

Eleusine indica can be used for remediating the soil contaminated with Pb and Zn.

The results of the survey showed that this plant can grow in the waste area of lead,

zinc processing factory. Analyzing Pb and Zn concentration in soil and plants

showed that if soil contained 4316.9 mg kg−1 Pb, there would be 664.5 and

4638.2mg kg−1 Pb in shoots and roots of the plant, respectively; if soil contained

1000mg kg−1 Zn, there would be 761.6 and 2011.3mg kg−1 in shoots and roots,

respectively. Eleusine indica could grow well at the concentration of Pb and Zn (in

the form of Pb(NO3)2 and Zn(NO3)2), respectively. Other studies have found Eleusine

indica (L.) higher accumulating Pb in the shoots [32, 37].

6.4.3 Vetiveria zizanioides

In mining soil contaminated with Pb from 1400.5 to 2530.1mg kg−1, Vetiveria ziza-

nioides still grew well after 90-days treatment. Some characteristics of plant grow-

ing on Pb-contaminated soil such as height, root length, biomass, and the chlorophyll

concentration increased more than those on control soil (soil without Pb). Pb con-

centration analysis in soil after this experiment showed that the Pb extraction effect

from the contaminated soil by Vetiveria zizanioides could reach from 87% to 92.6%.

However, the average Pb accumulation in its shoots and roots were not high being

only 24 and 349mg kg−1, respectively. This species also can accumulate As and Cd

taken from soil. Many of our further experimental results conrmed feasibility of

using Vetiveria zizanioides as phytostabilization agent for Pb, Cd, and As in con-

taminated soils. Some research results also reported that Vetiver grass has the ability

to accumulate wide range of heavy metals [52–55].

Fig. 6.4 Pot experiments of potential plant species

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6.4.4 Brassica juncea

As, Pb, and Cd accumulations of Brassica juncea were quite high. All three heavy

metals can be accumulated in roots more than in the shoots. In trace concentration,

heavy metals can stimulate plant growth, but at higher concentrations (Cd>25mg

kg−1, As> 200mg kg−1, and Pb >2000mg kg−1) they inhibited plant growth. Pb

accumulation in shoots and roots of Brassica juncea grown on 2000mg kg−1Pb soils

were 1325 and 2546.2mg kg−1, respectively. The concentration of Pb accumulated

in Brassica juncea shoots in this study was similar to that reported in study of

Lombi etal. [56] and Jae etal. [57]. When cultivated on soils containing 25mg kg−1

As and Cd, concentration in shoots and roots were 185.6 and 228.9mg kg−1 for As,

185.6 and 228.9mg kg−1 for Cd, respectively. Brassica juncea can be used to remove

As, Pb, and Cd concentration in contaminated soil but it should be noted that this

plant is also a popular green vegetable. Therefore, the use of this plant species for

phytoremediation is limited due to the risk of poisoning human through consump-

tion of its heavy metal-contaminated leaves.

6.5 Field Evaluation ofHeavy Metal Accumulating Potential

oftheSelected Terrestrial Plants

6.5.1 Study atHa Thuong andTan Long Mines, Thai Nguyen

Province

The eld study was performed at Ha Thuong and Tan Long mine site. Selection of

the experimental sites was based on three criteria: (1) areas affected by mining

activities, containing high concentration of heavy metals As, Pb, Cd, and Zn;

(2) potential of indigenous plants for phytoremediation; (3) local conditions suitable

for operation model.

6.5.2 Ha Thuong Field Experimental Site

The analysis of soils collected at Ha Thuong Ti/Sn mine site showed very high con-

centration of As (4521mg kg−1), moderate concentration of Pb and Zn (235 and

463 mg kg−1, respectively), low concentration of Cd (4.5mg kg−1), and low pH

(2.3). Concentration of As, Cd, Pb, and Zn in the polluted soils was301.4, 3, 3.4, and

2.3 times higher than the permitted standards for agricultural soils, respectively

(Table 6.8). At this site, there is no plant species survived, except Pityrogramma

calomelanos. The source of soil contamination is from tin mining wastewater dis-

charged daily into the drain near this site. Spreading of contaminants has often

occurred in rainy season, when the whole area is totally submerged in water for

several hours or longer with frequency of 3–5 times per year.

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Table 6.8 Soil characteristics of Ha Thuong area before and after growing P. vittata, P. calomelanos, and V. zizanioides

Times As (mg kg−1) Cd (mg kg−1) Pb (mg kg−1) Zn (mg kg−1) pH (KCl) OM (%) CEC (cmolkg−1)

S0 (0month) 4521±324 4.5±0.9 235±67 463±85 2.3±0.8 0.21±0.13 3.1±0.6

S1 (8months) 2317±389 2.3±0.7 115±51 216±37 6.7±0.8 1.8±0.5 11.1±1.3

S2 (12months) 2011±215 1.8±0.8 70±12 191±21 7.2±0.6 2.3±0.6 10.3±0.8

S3 (24months) 1360±180 0.8±0.5 25±8 112±34 7.3±0.5 3.6±0.7 12.1±2.1

S4 (36months) 956±87 0.3±0.1 9±2 60±12 7.2±0.4 4.1±0.5 11.8±3.2

Note: OM organic matter, CEC cation exchangeable capacity; n=5, values are presented in mean±standard deviation. Allowable limits of As, Cd, Pb, and Zn

in agricultural soil recommended by the Vietnam National Technical Regulation are 15, 1.5, 70, and 200mg kg−1dw, respectively

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The experimental area at Ha Thuong mine site is 700m2. Before experiment

started, several plant species (Sesbania sesban, Reynoutria japonica, Senna alata)

had been planted in this area for creating a favorable environment, and CaO had

been added to raise soil pH to 7. Three plant species (Pteris vittata, Pityrogramma

calomelanos, and Vetiveria zizanioides) were tested at this site. Concentration of

heavy metals and As over 3-year period is presented in Table 6.8 (Fig. 6.5).

6.5.3 Tan Long Field Experimental Site

Soils at Tan Long experimental site contained very high concentration of Pb and Zn

(3470 and 3191mg kg−1, respectively), moderate concentration of As (213mgkg−1),

low concentration of Cd (52mg kg−1), and high pH value of 8.2. At this site, Pteris

vittata was found the most popular, while other species, such as Pityrogramma calo-

melanos, was also detected but with less number as compared to Ha Thuong site.

Tan Long experimental site has an area of 740m2. Vetiver and elephant grass

were cultivated around the experimental site to control erosion and leaching. Three

plant species were used at this site, including Pteris vittata, Vetiveria zizanioides,

and Eleusine indica.

Concentrations of As, Cd, Pb, and Zn (mg kg−1 dw) in the soils at Ha Thuong and

Tan Long experimental site were determined at 0, 8, 12, 24, and 36months after

cultivation of selected plants (Tables 6.8 and 6.9). To increase the efciency of phy-

toremediation, mycorrhiza fungi, EDTA, phosphorous, and organic fertilizers were

applied at two experimental sites. In general, soil concentrations of heavy metals

and As were markedly reduced over 3years at both experimental sites. Particularly,

soils contaminated with low or moderate concentration of heavy metal and As

(Cd, Pb, and Zn at Ha Thuong site, As and Cd at Tan Long site) were effectively

remediated to contain the level of heavy metals that are below the limits of Vietnam

National Technical Regulation. It should be noted that the removal effectiveness of

the heavy metals from the soil depends on the plant species; plant biomass; the

added of mycorrhiza fungi, EDTA, P, organic fertilizers; plant–microorganisms

relationship and soil leaching (Fig. 6.6).

6.6 The Uptake Capacity forHeavy Metals ofVetiveria

zizanioides atField Conditions

Khanh Son landll site is located above hill area of Lien Chieu District, Da Nang

City. This area was the municipal solid wastes dumping site of Da Nang City since

1992, and closed in 2006. The studied site was selected at a dumping area inside

the landll where the solid wastes were kept for 2years, covered with 0.5m of

surface soil. The solid wastes were already decomposed and mixed. The second

experimental site was selected at waste disposal point with an area of 1500m2

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located in residential area of Hoa Minh ward, Lien Chieu district, Danang city. The

studied site was the place used for holding and recycling the second-hand cars.

Vetiver grass was cultivated at a density of 20 seedlings.m−2 for both experimental

sites. Heavy metal concentrations in shoot (stem and leaves) of vetiver were deter-

mined at 3, 6, and 12months after cultivation of vetiver.

Heavy metal concentrations accumulated in vetiver shoot were gradually dimin-

ished with time (Table 6.10). The highest concentration of Zn, Cu, and Pb in stems

and leaves of Vetiver grown at Khanh Son landll were 342.4, 30.3, and 5.6mgkg−1,

respectively. At Hoa Minh waste landll, vetiver accumulated the highest Zn and Pb

concentration of 36.4 and 6.4mgkg−1, respectively. The concentrations of Zn, Cu,

and Pb in shoot of vetiver grown at eld condition were higher than those of vetiver

grown under greenhouse condition. The concentration of heavy metals in stem and

leaf of Vetiver was highest after 3months of transplanting at both Khanh Son and

Hoa Minh areas. After 12months of growth, the amount of Zn accumulated by

vetiver was 0.9 gm−2year−1 and 1.5 gm−2year−1 at Khanh Son landll and Hoa Minh

waste disposal site, respectively (Table 6.11).

In terms of physical–chemical characteristics and the concentration of heavy

metals in soil, the obtained results showed that the contents of organic matter (OM)

and total nitrogen (Nts) increased at both experimental sites after the experiment was

completed. The amount of organic matter increased from 9% to 13% and the total

nitrogen increased from 23% to 68% at the end of experiment compared with those

Fig. 6.5 Ha Thuong experimental site before and after growing Pteris vittata, Pityrogramma calo-

melanos, and Vetiveria zizanioides

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Table 6.9 Soil characteristics of Tan Long area before after growing P. vittata, V. zizanioides and E. indica

Times As (mg kg−1) Cd (mg kg−1) Pb (mg kg−1) Zn (mg kg−1) pH (KCl) OM (%)aCEC (cmolckg−1)b

S0 (0month) 213±54 52±11 3470±123 3191±231 8.2±1.3 0.6±0.2 4.3±0.5

S1 (8months) 127±23 31.2±1.9 2135±121 2365±237 7.6±0.7 3.8±0.7 15.7±2.4

S2 (12months) 93±16 22.5±2.8 1924±202 2132±131 7.8±0.5 3.3±0.6 16.5±2.8

S3 (24months) 47±9 10.2±1.5 1367±185 1512±214 7.9±0.4 4.6±0.8 17.1±3.1

S4 (36months) 16±7 5.7±1.3 954±96 1034±123 7.7±0.6 5.1±0.6 16.8±5.3

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at the beginning of experiment. In contrast, the contents of bioavailable phosphorus

in both areas reduced from 12% to 23%. In addition, the amount of bioavailable

potassium in the soil at Khanh Son increased 31%, whereas this amount at Hoa

Minh reduced 5%. These results were also consistent with the results reported in the

study of Phien and Tam [58]. Signicantly, the amount of heavy metals in soils at

the end of experiment was lower than that at the beginning of experiment. The

reduction for Zn, Pb, and Cu were 13–16%, 7–12%, and 17%, respectively.

6.7 Conclusion

Mining activities in Vietnam have resulted in large areas of land contaminated with

high concentrations of heavy metals and As. The contaminated soils require immedi-

ate remediation to control adverse effect of contaminants on human and environment.

Fig. 6.6 Phytoremediation in Tan Long mining site

Table 6.10 Concentrations of heavy metals (ppm) in aerial parts (stems and leaves) of Vetiver

Places Heavy metal

Periods of experiment

3months 6months 12months

Khanh Son Landll Zn 342.4±3.4 305.4±6.5 287.5±7.1

Cu 30.2±0.9 27.37±1.8 23.2±2.8

Pb 5.6±0.5 5.76±0.3 4.1±0.1

Hoa Minh waste disposal site Zn 336.4±3.9 321.6±0.9 310.5±3.7

Pb 6.4±0.1 6.3±0.1 5.6±0.2

Table 6.11 The amount of studied-heavy metals per 1m2 after 12months at the eld conditions

Places

Amount of heavy metals (g/m2)

Zn Cu Pb

Khanh Son Landll 0.931 0.075 0.013

Waste disposal site at Hoa Minh 1.469 – 0.026

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Among several available remedial technologies, phytoremediation is the most

appropriate because the technology is simple, cost-effective, and environmentally

friendly. Several research programs have been conducted since the last decade in

order to search for indigenous hyperaccumulators of As, Cd, Pb, and Zn and evaluate

the selected plant species for phytoremediation purpose under greenhouse and eld

conditions.

Among plant species being collected in Bac Kan and Thai Nguyen, P. vittata and

P. calomelanos are good candidates for phytoremediation of As. A. houstonianum,

E. indica, H. cordata, C. dactylon, C. rotundus, and E. ramosissimum are potential

species for phytoremediation of Pb and Zn. The mixed cultivation of P. vittata,

V. zizanioides, and E. indica at Tan Long mine site, and P. vittata, P. calomelanos,

and V. zizanioides at Ha Thuong mine site together with application of mycorrhiza

fungi, EDTA, phosphorous, and organic fertilizers, showed very promising results.

Concentrations of As, Pb, and Zn were signicantly reduced over 3-year period. It

can be concluded that the mixed cultivation of the selected plants can be used to

remediate As-, Pb-, and Zn-contaminated soils.

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183© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_7

Chapter 7

Essential Elements andToxic Metals inSome

Crops, Medicinal Plants, andTrees

ElenaMasarovičová andKatarínaKráľová

Abstract Plants play an ever-increasing role not only for providing safe and

healthy food for a growing world population but also for new biotechnologies

including phytoremediation of areas contaminated by toxic metals, phytofortica-

tion used in functional foods preparation and nanoagrochemicals application in

agriculture. Since species of genus Brassica are not only important crops but they

have use for technical purposes, we evaluated these important crops from the aspect

of nutrition or toxic metal responses. Medicinal plants are presented as a source of

natural substances widely used in pharmaceutical, food and cosmetics industries

and potentially also in phytoremediation technology. Therefore, we analyzed the

effect of bioelements and toxic metals on growth and physiological processes of this

important group of the plants. Trees (both forest and fast growing trees) as one of

the world’s most abundant raw materials for industrial products and renewable

energy as well as their non-production functions (reducing erosion, moderating the

negative climatic changes, and phytoremediation procedures) are outlined. We have

emphasized that plant responses to different nutrient and toxic metal conditions are

expressed through structural composition and physiological processes. Results from

experiments with above-mentioned plants treated with bioelements and toxic metals

are shortly presented. Here, we used ion form of elements (Cd, Cr, Cu, Hg, Ni, Pb,

Zn) and elements as complexes (Cu, Cd, Fe, Se, Zn). Finally, we stressed that both

scientists and politicians will have to accept fundamental bioethical principles to

ensure the sustainable development of human society as well as essential protection

of the environment and nature.

Keywords Bioelements • Crops • Growth • Physiological processes •

Phytoremediation • Phytofortication • Medicinal plants • Nanoagrochemicals •

Toxic metals • Trees

E. Masarovičová (*)

Faculty of Natural Sciences, Department of Soil Science, Comenius University in Bratislava,

Ilkovičova 6, SK-84215 Bratislava, Slovakia

e-mail: masarovicova@fns.uniba.sk

K. Kráľová

Faculty of Natural Sciences, Institute of Chemistry, Comenius University in Bratislava,

Ilkovičova 6, SK-84215 Bratislava, Slovakia

e-mail: Kata.Kralova@gmail.com

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7.1 Introduction

Plants play an ever-increasing role for providing safe and healthy food for a growing

world population and for replacing limited, and expensive, fossil resources as feed-

stock for the production of energy and industrial materials. The strategic agenda for

the plant research outlines the approach that can contribute to full consumer

demand for safe, sustainable, and healthy food. Novel plants aim at delivering non-

allergic foods and foods with longer shelf lives, better nutritional composition, and

more varied tastes. These plants may need less input in terms of water, fertilizer or

pesticides and will be more stress-resistant, mainly against drought or seasonal

instabilities caused by climate change. Farmers should increase agricultural produc-

tivity while decreasing its environmental footprint. Plants (both crops and trees) or

plant waste will in the future be an important source for the production of energy,

biofuels, and biopolymers, replacing the use of fossil fuels as feedstock. However,

new technologies must be applied within systems that are both economically and

environmentally sustainable.

One reason for interest in plant–metal interaction has been the recent attention on

the use of plants either to remediate toxic metal-contaminated soils or increasing the

bioavailable concentrations of essential nutrients in edible portions of food crops

through agronomic intervention or genetic selection. In addition, since plants are

known to interact with different metals, they have been used for the “green biosyn-

thesis” of metal nanoparticles. Such bioinspired methods are dependable, environ-

mentally friendly, and benign. In general, phytoremediation, phytofortication, and

metal nanoparticle biosynthesis are thus natural green biotechnology with using

crops, medicinal plants as well as trees.

7.2 Crops

In general, there are top ve crops produced in the world (sugar cane, maize,

wheat, rice, and potatoes) and ten crops that feed the world (maize, potatoes, sweet

potatoes, yams, cassawa, soybeans, sorghum, plantain, wheat, and rice) [1]. Thus

it is very difcult to discuss all these most important crops from the aspect of nutri-

tion or toxic metal responses. Therefore, we decided to evaluate the species of

genus Brassica that is not only an important crop but it has use for technical pur-

poses. Moreover, some of rapeseed genotypes can be used in both phytoremedia-

tion and phytofortication technologies. Authors of this chapter have also some

experiences with rapeseed. Many experiments with this species have been done in

our laboratory or under eld conditions and many papers we published, too (see

“References”).

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7.2.1 General Characteristic ofGenus Brassica

Brassica is a genus from family Brassicaceae that species are informally known as

cruciferous vegetables, cabbages or mustard plant and commonly used for food

(e.g. cabbage, cauliower, broccoli or Brussels sprouts) and include a number of

weeds, both wild taxa and escapees from cultivation. It counts over 30 wild species

and hybrids as well as many cultivars and hybrids of cultivated origin. However, it

should be stressed that there is some disagreement among botanists on the classi-

cation and status of Brassica species and subspecies. More of them are seasonal

plants (annuals or biennials) and some are even small shrubs. Species of genus

Brassica have been the subject of much scientic interest for their agricultural and

food importance (e.g., [2, 3]). Six species, B. carinata, B. juncea, B. oleracea,

B. napus, B. nigra, and B. rapa, evolved by the combination of chromosomes from

three earlier species, as described by the “Triangle of U theory”—theory about the

evolution and relationships between members of the plant genus Brassica (e.g., [4]).

This genus is native in the wild in Western Europe, the Mediterranean, and temper-

ate regions of Asia and many wild species grow as weeds, especially in North

America, South America, and Australia. Almost all parts of plants are used for food,

including the roots, stems, leaves, owers, buds, and seeds. Some forms with white

or purple foliage or owerheads are also grown for ornamental intention. There is

also very important use of rapeseed oil for technical purposes, especially as a bio-

fuel. It seems very promising the use of some rapeseed varieties for phytoremedia-

tion of soils contaminated by toxic metals (see Sect. 7.5 of this chapter).

7.2.2 Effect ofCadmium andSome Bioelements

onBrassica Species

Brassica juncea L. (Indian mustard) together with Vigna radiata (L.) Wilczek

(mung bean) are important crops in the poorer countries—mainly of Asia, whereby

the leaves, the seeds, as well as the stem of Indian mustard are edible. For their high

nutritive values they are source of proteins, Ca, P, certain vitamins and some culti-

vars possess excellent aroma (cf. Betal etal. [5]). Šimonová etal. [6] determined the

effect of different Cd concentrations (6–120μmol dm−3) on Hill reaction activity

(HRA) of isolated chloroplasts, content of chlorophylls (Chls) and carotenoids

(Cars) as well as both Cd uptake and accumulation in plant organs. Seeds of studied

crops were grown in thermostat on wet cellulose wadding at 80% of relative air

humidity and air temperature of 25±1°C for 4 days. Then the seedlings were trans-

ferred into the plastic containers lled with 2dm3 of Hoagland solution (control

variant) and cultivated in growth chamber under relative air humidity 60–70%, day/

night temperature 25/20 ± 1 °C, and 12 h photoperiod with irradiance

100μmolm−2s−1 PAR.The solution was continuously aerated. In the growth stage

of primary true leaves in both species, the seedlings were transferred to the Hoagland

7 Essential Elements andToxic Metals inSome Crops, Medicinal Plants, andTrees

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solution with above- mentioned Cd concentration. The response of plants to the Cd

effect was evaluated 5 days after Cd application. It was found that Cd stress inhib-

ited HRA of both species; the mung bean showed a higher sensitivity to Cd treat-

ment than Indian mustard. The leaves of Cd-treated plants possessed lower content

of Chls and Cars, whereby negative effect increased with Cd concentration. A dif-

ference between studied crops was also found in Cd uptake and Cd accumulation.

In both species, Cd was accumulated more in the roots than in the shoots, with

higher accumulation in B. juncea than in V. radiata.

Rapeseed (Brassica napus L. var. napus; syn: Brassica napus L. ssp. oleifera) is

other species from the genus Brassica that belongs in structure of actual agriculture

to the perspective and economically interesting crops also under Slovakian climatic

condition (cf. Tóth and Hudec [7]). Following cereals and maize it is the third most

important crop in Slovakia. Rapeseed has large-scale utilization, mainly as a food,

further in pharmaceutical and chemical industry and oil extrusions are important

nourishing component in animal fodder. Moreover, this species is used as a techni-

cal plant for production of renewable biofuels (FAME—fatty acid methyl ester) or

PPO (pure plant oils) (in detail see Masarovičová etal. [8]). It should be stressed

that rapeseed is not only an essential crop, technical and melliferous benecial

plant, but this species will become a perspective functional crop in the near future.

From such plants it is possible (using phytofortication biotechnology) to prepare

functional food fortied by substances with high nutritional value [8–10]. Because

of higher accumulation of some toxic metals (mainly Cd) into the root and shoot,

rapeseed was assigned to the plant species potentially used also in phytoremediation

technology [11, 12]. In our earlier paper [13] we investigated production potential

of chosen rapeseed cultivars from the aspect of soil quality (applied agrotechnology

and plant protection), whereby were tested cultivars represented three production

regions—maize production region (MPR), rapeseed production region (RPR), and

potato production region (PPR) of Slovakia. Following six rapeseed cultivars were

tested: ES ASTRID (medium-early and low type cultivar, France, PPR), ATLANTIC

(medium-early and high type cultivar, France, PPR), CALIFORNIUM (medium-

early to early and medium-high type of cultivar, France, MPR), LABRADOR (late

and low to medium-high type of cultivar, France, RPR), MANITOBA (late and

medium-high type of cultivar, France, RPR), and OPONENT (late and high type of

cultivar, Czech Republic, MPR). Field experiment was realized at the experiment

area of Centre for Research of Crop Production, Research Institute of Crop

Production in Borovce near Piešťany, West Slovakia. Since agrotechnology as well

as plant protection of rapeseed is complicated, we recommend to read it in detail in

the above-mentioned paper. However, it is important to stress that the soil on experi-

mental area is Phaeozem formed from loess, pH of the soil was 5.5–7.2; humus

content was 1.8–2%; potassium content was good; phosphorus content was medium;

and magnesium content was high (Melich II). From production parameters were

estimated seed production (yield), from qualitative parameters were determined oil

content (in %) and oil production (in t/ha). It was found that the lowest seed yield,

oil content as well as oil production had genotype Californium. High seed yield was

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estimated for genotypes Oponent, Atlantic and ES Astrid, whereby Oponent had

both high oil content and oil production. Content of the most important fatty acid

from the aspect of the FAME quality—oleic acid ranged from 60.02 (Labrador) to

65.80% (Atlantic). Oponent had also higher content of oleic acid (62.12%). Seeds

of all studied genotypes had low content of linolenic acid (8.33–9.59%), which is

important for oxidation stability of fatty acid methyl ester (biocomponent for bio-

diesel). These analyses were realized according to actual EU standards in the

accredited institution, State Veterinary and Food Institute in Bratislava, Slovakia.

From this aspect the most suitable genotype seems to be Atlantic. In the case of all

genotypes content of erucic acid ranged from <0.01 to 0.05%, which is in agree-

ment with corresponding EU standards. Considering all studied quantitative and

qualitative parameters, it should be concluded that for analyzed soils and

under Slovakian climatic condition suitable genotype seems to be Czech genotype

Oponent.

Later [14, 15] we evaluated production potential of the same six rapeseed culti-

vars from the point of various climatic condition. From production parameters was

estimated seed production, from qualitative parameters were determined oil con-

tent, oil production, and content of the substantial fatty acids in the seed. Air tem-

perature and atmospheric precipitation were chosen as the most important climatic

factors for rapeseed production. Mean month values of these factors in the 2007/2008

years were compared with the values found for long-term period 1961–1990 years.

Based on these data, it could be concluded that in general weather course was favor-

able for all important stages of the growth (ontogenetical development): sowing,

owering as well as seed maturing. In August 2007, when rapeseed was sown, both

air temperature and precipitation were supernormal, September and December were

substandard cold with supernormal precipitation. Winter months (January and

February) were supernormal warm and with partially limited precipitation. Spring

months till harvest were 1.5–2.5°C over month average of air temperature, and

precipitation were 75–164% of normal values. High seed production was estimated

for cultivars Oponent, Atlantic and ES Astrid, whereby Oponent had both high oil

content and oil production. Content of the most important fatty acid, oleic acid,

ranged from 60.02 (Labrador) to 65.80% (Atlantic). Oponent had also higher con-

tent of oleic acid (62.12%). Seeds of all studied cultivars had low content of linole-

nic acid (8.33–9.59%) and content of erucic acid ranged from <0.01 to 0.05%,

which is in agreement with corresponding EU standards. Considering all studied

quantitative and qualitative parameters, it was concluded that for Slovakian climatic

condition suitable cultivar seems to be Czech cultivar Oponent. Tatarková etal. [16]

determined in homogenized soil samples coming from the experimental area

contents of phosphorus and inorganic-, nitrate- and ammonium nitrogen. The exper-

imental plot was evaluated in terms of basic nutrients content. The highest sensitive

response to the applied fertilization and N, P soil content in spring and consequently

the highest yield and seed production was found for cultivar Labrador: yield

4.68tha−1 and seed production 2.28dkgplant−1.

7 Essential Elements andToxic Metals inSome Crops, Medicinal Plants, andTrees

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7.2.3 Response ofSome Crops toToxic Metal Application

Roots (especially different types of roots) are important for plants in terrestrial eco-

systems because these have different functions. Most of plants form one or more

orders of lateral root branches that vary in their thickness, branching patters, growth

rates, capacity for secondary growth and structural features, as well. Higher orders

of lateral roots are generally thinner, shorter, and do not live as long as of lower

orders. Young roots with living epidermal cells and root hairs are often considered

to be responsible for the most direct nutrient uptake (e.g., Nyambane and Mwea

[17]). Therefore, roots of some species (e.g., cucumber and lettuce) are recom-

mended for toxicity testing and environmental assessment [18].

In connection with uptake of elements (ions) from the soil into the root and trans-

port of them from the root in the shoot, there were established two important fac-

tors: bioaccumulation factor (BAF) and translocation factor (TF). BAF is dened as

the ratio of metal content (concentrations) in plant dry mass (μgg−1d.m.) to those

in soils (μgg−1 soil) or in solution (μgmL−1). TF is the ratio of metal content in the

shoot (μgg−1d.m.) and in the root (μg g−1 d.m.). TF has been used to determine

the effectiveness of plants in translocating metal ions from the root to the shoot.

Both above-mentioned factors have to be considered for categorization (classica-

tion) of metallophyte (in detail see Masarovičová etal. [9]).

As mentioned above phytotoxicity testing of toxic metals based on evaluation of

reduction of root and shoot growth of plants is widely used (e.g., [19, 20]). We inves-

tigated the effects of seven metal ions (Cd(II), Cr(VI), Cu(II), Hg(II), Ni(II), Pb(II),

and Zn(II)) on length of roots of ve rapeseed (Brassica napus L. subsp. napus)

cultivars registered in Slovakia (Atlantic, Baldur, Californium, Oponent, and Verona)

and evaluated their phytotoxic effect using IC50 values [21]. From rapeseed cultivars

Atlantic is a medium-early and high type genotype suitable for potato production

region; Californium is a medium-early to early and medium-high type of genotype

suitable for maize production region; Oponent is a late and high type of genotype

suitable for potato production region and Verona is a late and medium-high genotype

suitable for maize and potato production regions and Baldur is a medium-early and

medium-high type of genotype suitable for all production regions. In general, the

toxicity of metal ions decreased in the following order Cu > Cr >Hg > Cd > Pb > Ni

> Zn. Atlantic, Baldur, and Californium were more sensitive to Cd than to Ni, for

Oponent and Verona higher toxicity exhibited Ni. The sensitivity of studied cultivars

treated with toxic metals decreased as follows: for Cd: Atlantic>Californium>Ver-

ona>Baldur>Oponent; for Cr: Atlantic=Californium=Verona>Baldur=Opon-

ent; for Pb: Atlantic>Verona>Californium>Baldur>Oponent, for Zn: Atlantic

>Californium>Oponent>Verona>Baldur, for Cu: Atlantic>Californium=Vero-

na>Baldur>Oponent, for Hg: Oponent>Californium>Atlantic=Verona>Bald-

ur and for Ni: Oponent=Atlantic>Verona>Baldur>Californium.

From the studied rapeseed cultivars, Atlantic and Californium were found to be

most sensitive to tested metals. On the other hand, high tolerance to metal treatment

was determined for Baldur. Czech cultivar Opponent showed high tolerance to Cd,

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Cr, Cu, and Pb, but it was sensitive to Hg and Ni. The obtained results showed that

root tolerance index can serve as good biomarker for evaluating the relative toxicity

of toxic metals to different rapeseed cultivars.

Our results are in agreement with ndings of [22] who estimated following tox-

icity rank for root growth inhibition of maize plants: Cu2+> Hg2+ > Cd2+ > Zn2+,

whereby the toxicity showed correlation with the afnity of metal ions to –SH

groups. The tested metal ions are known as mitotic poisons. In the concentration

range 0.1–10mmoldm−3 Cu2+ exhibited toxic effect on chromosomal morphology

in maize and induced c-mitosis, anaphase bridges, and chromosome stickiness [23].

According to Doncheva [24], treatment with copper results in the interruption of

progression of nuclei at the crucial G1/S transition point of the cell cycle, when Cu

prevents their entry into mitosis and the effect of copper on root meristem cell pro-

liferation is reected in the decreased root growth. Cu2+ decreased mitotic index

which reects the frequency of cell division in Vicia faba root meristematic cells

and numerous micronuclei, chromatid bridges and lagging/lost chromosomes found

in the meristematic cells of V. faba indicated the clastogenic effect of Cu2+ [25].

Copper-induced root growth inhibition of Allium cepa L. involved disturbances in

cell division and DNA damage, whereby microtubules were one of the target sites

of Cu toxicity in root tip meristematic cells, and Cu exposure substantially impaired

microtubule arrangements [26]. Genotoxic effect of copper was determined also in

Triticum aestivum L. ([27] and rye (Secale cereale) roots [28]. Ni [29] and Zn [30]

were also found to decrease mitotic index in Zea mays L. and cytotoxic effects of

Hg on root tip cells of Cicer arietinum L. were estimated by Cavusoglu etal. [31],

while short-term exposure to Cr (VI) caused cytogenetic damage in root tip meri-

stems of barley seedlings [32]. Adverse effects of Cr(VI) on mitotic index were

described also by Chidambaram etal. [33] and Eleftheriou etal. [34]. Inhibitory

effect of Cd2+ ions on mitotic index as well as on active mitotic index for Hordeum

vulgare and Setaria italica was estimated by Yadav and Srivastava [35] and

Amirthalingam et al. [36] who reported that mitotic divisions in root of Vigna

unguiculata have been withheld when the Cd stress increased and DNA damage in

cells manifested in strand breakage, removal of nucleotides and variety of modica-

tions in organic bases of nucleotides due to ROS formation was observed.

In plants copper is an essential element, usually bound to proteins and involved

in numerous processes where it participates on catalyzing redox reactions, whereby

only 2% of plant Cu occurs in its free form. Cu in plastocyanin is indispensable in

the electron transport chain and it is also a component of the active cytochrome c

oxidase complex in the mitochondrial electron transport chain. It is involved also in

the light reaction of photosynthesis as well as in cell detoxication by Cu-containing

enzymes glutathione-S-transferase or Cu-Zn superoxide dismutase. However, at

high concentrations it is toxic and inhibits plant growth, photosynthesis and respira-

tion, impairs protein synthesis, induces ROS formation, causes water loss and inac-

tivates key metabolic enzymes [37–39]. Thus, as the site of inhibitory action of Cu

the donor and acceptor side of photosystem (PS)II have been suggested (e.g., [40–44]).

Copper also damages chloroplasts either by inducing iron deciency or by replacing

Mg in the chlorophyll by Cu [45–47]. Application of 50 and 100μmoldm−3

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Cu resulted in signicantly decreased B. napus plant growth, biomass, photosyn-

thetic pigments, and gas exchange characteristics and reduction of activities of anti-

oxidant enzymes superoxide dismutase (SOD), peroxidase (POD), ascorbate

peroxidase (APX), and catalase (CAT) along with protein contents was observed as

well accompanied with the increased malondialdehyde (MDA) and H2O2 in both

leaves and roots [48]. On the other hand, it was found that selenium can protect

rapeseed seedlings not only from Cd-induced oxidative stress [49] but it can also

alleviate Cu toxicity in rape [50]. Yurela [51] overviewed main features concerning

copper function, acquisition and trafcking network as well as interactions between

copper and other elements. Review focused on the adverse effects of Cu excess on

growth and yield of essential food crops was presented by Adrees etal. [52]. On the

other hand, Brassica sp. have ability to bioaccumulate heavy metals and can be used

to reduce the level of contaminants in the soil (phytoremediation), and thus to clean

up and prepare soils for cultivation [53–55].

Cu translocation within B. napus plants from external solution was found to be

low and excess Cu signicantly decreased other microelement content, such as Fe

and Mn in plants, caused reduction of photosynthetic pigments and inhibition of

plant growth, whereby longer exposure to Cu resulted also in accumulation of

highly reactive oxygen species, whereby B. napus was found to be more sensitive to

Cu-induced stress than B. juncea [56, 57]. Copper was considerably more toxic

against B. napus plants than Zn, it was retained in the roots and was poorly trans-

ported to shoots, while Zn proved to be highly mobile, it was concentrated in the

upper leaves and actively transported. While high Cu concentrations slowed strongly

Zn uptake by the roots but practically did not change its movement over the plant,

Zn concentrations facilitated Cu uptake by the roots and its transfer to shoots [58].

External Zn stress resulted in increased Zn content in rapeseed plants, while signi-

cant lower concentrations of P, Cu, Fe, Mn, and Mg were estimated, especially in

roots, and also increased lipid peroxidation was determined [59]. In B. napus plants

excess Cu (200μmoldm−3) induced chlorosis on young leaves similar to Fe de-

ciency symptoms [60].

In an experiment, we evaluated the effect of copper on some production and

biochemical characteristics of 3 weeks old B. napus plants (cv. Verona) which were

for 7 days hydroponically cultivated in the presence of CuSO4 (0.5–60μmoldm−3).

At lower Cu concentrations (0.5–3μmoldm−3) a signicant increase of biomass

(both plant organs), with highest stimulation at application of 0.5μmoldm−3 Cu(II)

was observed [61]. Similarly, treatment with low Cu(II) concentrations resulted in

increase of fresh weight of Helianthus annuus L. seedlings and Zea mays L. root

growth [23, 62] and increased rape growth grown on Cu-polluted paddy soil [63].

However, in the concentration range 6–60μmoldm−3 CuSO4 we estimated notable

reduction of biomass. This is in agreement with ndings of Zaheer etal. [64] who

observed that exposure to 50 and 100 μmol dm−3 Cu signicantly reduced the

growth, biomass production, chlorophyll content and soluble proteins of B. napus

seedlings, causing also enhanced production of H2O2, MDA and electrolyte leakage

in leaf and root tissues of rapeseed plants.

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In our experiment Cu-induced stress manifested by chlorosis was observable

already at treatment with 12 and 24μmoldm−3 and leaves of plants treated with

60μmoldm−3 were wilted and desiccated. Ali etal. [65] estimated leaf chlorosis and

lesser biomass yield in hydroponically cultivated Brassica oleracea var. capitata

plants already at concentrations ≥2 μmol dm−3 Cu. Due to Cu(II) treatment we

observed also signicant decrease in the content of Chla, Chb, and Cars already at

concentration 6μmol dm−3. Cu is known to interfere with the biosynthesis of the

photosynthetic machinery by modifying the pigment and protein composition of

photosynthetic membranes [66] and to modify nutrient uptake. Feigl et al. [57]

observed that excess Cu signicantly decreased other microelement content, such as

Fe and Mn in the shoots of B. napus what resulted in decreased concentrations of

photosynthetic pigments in B. napus leaves and was accompanied with more intense

growth inhibition. The reduction in Phaseolus vulgaris leaf Chl concentration due

to the Cu-mediated Fe deciency was explained also by Pätsikkaä et al. [45].

According to Purakayastha etal. [67], root length emerged as the powerful param-

eter to dictate the uptake of metals by Brassica spp.

With increasing Cu concentration from 0.5 to 60μmoldm−3, Cu concentration in

rapeseed plant organs, cv. Verona linearly increased, in roots from 222.7 to

9249mgkg−1 and in shoots from 6.2 to 47.9mgkg−1 [61]. Hence, the amount of Cu

accumulated in roots was 36 to 193 times higher than in shoots and TF values

showed a decrease from 0.028 (0.5μmoldm−3) to 0.005 (60μmoldm−3). While the

portion of Cu allocated in shoots from the total Cu amount accumulated by plant

was 27.6% at application of 0.5μmoldm−3, at treatment with 60μmoldm−3 it rep-

resented only 8.4%. These results are in agreement with several researchers. For

example, compared to control plants treatment with 50, 100, and 150μmol dm−3

resulted in 5.1, 6.3, and 7.6 times higher Cu concentration accumulated in the leaves

of Sinapis alba after 10 days exposure, whereby for B. napus plants it was 3, 5, and

7 times higher [68]. Signicantly higher accumulated Cu amount in roots of

B. napus compared to shoots was observed previously also by Rossi etal. [69].

Greater Cu content in root than in shoot indicates adoption of exclusion mechanism

to tolerate the toxicity in which the roots accumulate the metals preventing its sub-

sequent transport to the shoots [70, 71].

The oxidation state of chromium signicantly affects its toxicity. The oxidation

state of Cr is important because the common triplet oxidation state (CrIII) is not toxic

as compared to the hexavalent form [72]. However, Cr(III) which is thought essen-

tial for animals in trace amounts, is toxic to plants even at low concentration: it is

reported to cause severe oxidative damage and exhibit adverse effects on plant

growth, water balance, and pigment content [73–75]. Cr can inhibit δ-aminolevulinic

acid dehydratase, an important enzyme involved in chlorophyll biosynthesis, and Cr

mostly in the form of Cr(VI) can replace Mg ions from the active sites of many

enzymes and deplete Chl content [76]. It is believed that Cr(III) enters the Fenton

reaction, whereby the catalytic activity of Cr(III) is much higher in a Fenton reac-

tion system compared to other metals like Co(II), Cd(II), Zn(II), Mn(II) and Fe(III)

but lower than Cu(II) [77].

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Cr(VI) stress inhibited growth of B. juncea L. plants and was directly interrelated

with its accumulation, whereby treatment with Cr(VI) resulted in the modulation in

activities of various stress markers (SOD, POD, APX, glutathione reductase (GR),

dehydroascorbate reductase (DHAR), and lipid peroxidation) [78]. Pandey etal. [79]

observed signicant increases in lipid peroxidation and tissue concentration of H2O2

in B. juncea plants exposed to 2 and 20μmoldm−3 Cr(VI). Signicant reduction of

B. napus plant growth, biomass, Chl contents, and Cars as well as soluble protein

concentrations, while considerable increase of SOD, guaiacol peroxidase, CAT, and

APX activity and MDA levels was estimated as a result of Cr(VI)-induced stress

[80, 81]. H2O2 may act as a signal that triggers defense mechanisms which in turn

protects canola seedlings from Cr(VI)-induced oxidative damage [82]. Zaimoglu

etal. [83] estimated that due to Cr(VI) application after a signicant increase, a

sharp decrease in the activity of APX and GR in Brassica juncea L. and Brassica

oleracea L. plants occurred and the researchers suggested that a coordinated

increase in APX and glutathione reductase activities in both Brassica species under

Cr stress play a role as signals to protect the plants from Cr-induced stress. Induction

of phytochelatins along with antioxidant defense system in response to Cr stress

suggested the cumulative role of phytochelatins and antioxidants in conferring tol-

erance against accumulated Cr in B. juncea [84]. Increased concentrations of some

of antioxidant enzyme activities in leaves and roots of four B. napus cultivars

exposed to Cr was observed by Gill etal. [85] who also estimated that application

below 400μmol dm−3 Cr caused changes in leaf ultrastructures like broken cell

wall, immature nucleus, a number of mitochondria, ruptured thylakoid membranes

and large size of vacuole and starch grains, while at concentrations exceeding

400μmol dm−3 Cr damage of roots in the form of disruption of Golgi bodies and

diffused cell wall was estimated. At treatment of eight canola cultivars with

100 μmol dm−3 Cr(VI) the estimated Cr concentrations in aerial parts of plants

ranged from 255.0 to 705.8μg Cr g−1 d.w., whereby the cultivar with the highest

accumulated Cr amount showed the lowest levels of chlorophyll content and highest

levels of lipid peroxidation [86]. The Cr contents in stem, leaf, and root of B. juncea

plants usually were heightened with increased concentrations of Cr(VI) in soil and

average Cr concentration in the leaves of Laifengjiecai and Sichuanhuangzi culti-

vars growing 70 days in the soil spiked with 300mgkg−1 Cr(VI) reached 167.30 and

197.60 mg kg−1, while the maximum Cr contents in plant shoots were 1.71 and

2.81mg/plant, indicating that portions of Cr removed by plant shoots were 0.23 and

0.38% of Cr content in treated soil [87]. Signicant translocation of Cr from the

roots into the above-ground parts was estimated in soybean treated with Cr(VI) [88].

High activities of antioxidant enzymes supported by high Cr concentrations in roots

and aerial parts established the Indian mustard as a potential hyperaccumulator and

a hypertolerant species to Cr stress [89]. Karuppanapandian and Manoharan [90]

found that uptake and translocation of Cr in Vigna mungo L. plants was relatively

higher during rst 12h of treatment with 100μmoldm−3 Cr(III) as well as Cr(VI)

and Cr-treated roots of Vigna mungo L. retained 15 times more Cr than the shoots.

Bluskov etal. [91] found that B. juncea plants grown on soils supplemented with

100mgkg−1 of Cr (III) or Cr(VI) concentrated Cr mainly in the roots and removed

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about 48 and 58 μg Cr/plant. Cr was localized, and probably sequestered, in

epidermal and cortical cells in the roots and epidermal and spongy mesophyll cells

in the leaves and Cr (III) was detected, primarily as acetate in the roots and oxalate

in the leaves. Thus, B. juncea was found to be able to detoxify more toxic Cr (VI).

Similarly, in mesquite which could be classied as a hyperaccumulator of Cr, sup-

plied Cr(VI) was uptaken by the mesquite roots, however, the data analyses of the

plant tissues demonstrated that it was fully reduced to Cr(III) in the leaf tissues [92].

Eichhornia crassipes (water hyacinth), supplied with Cr(VI) in nutrient culture,

also accumulated non-toxic Cr(III) in root and shoot tissues, whereby the reduction

of Cr(VI) to Cr(III) appeared to occur in the ne lateral roots and Cr(III) was subse-

quently translocated to leaf tissues where it was bound to oxalate ligands [93].

Conversion of CrO42− in the root to Cr(III) by several plants was reported by Zayed

et al. [94], whereby translocation of both Cr forms from roots to shoots was

extremely limited and accumulation of Cr by roots was 100-fold higher than that by

shoots and did not depend on the applied Cr species. Chelates and organic acids

were found to enhance Cr(III) accumulation, but toxic effects were not avoided and

Cr (III) complexes were as toxic to plants as Cr(VI) [95]. Han etal. [74] found that

chromium from Cr(VI)-contaminated soils was more phytotoxic to B. juncea plants

than that from Cr(III)-contaminated soils and caused growth retardation, reduced

the number of palisade and spongy parenchyma cells in leaves, clotted depositions

in the vascular bundles of stems and roots, and increased number of vacuoles and

electron dense materials along the walls of xylem and phloem vessels. Studies on

Cr(III) and Cr(VI) speciation in the xylem sap of maize plants showed that Cr(III)

and Cr(VI) were present as mobile and soluble anionic organic complexes, probably

Cr (III)-citrate in the xylem sap [96].

Gong etal. [97] found that Cr(VI) inhibited PS II in Chlamydomonas reinhardtii

mainly through damaging the oxygen evolving complex (OEC) and blocking the

electron transfer from QA to QB. On the other hand, Gupta etal. [98] observed in

7-day-old seedlings of Brassica juncea that Cr (VI) promoted PS II-mediated pho-

toreactions and found that Cr enhanced tolerance of PS II to alkaline pH.Deterioration

of oil quality by Cr(VI) application was manifested by reduction of oleic, linoleic,

and linolenic acid contents of lipids and increased erucic acid content in B. napus

due to increasing Cr(VI) concentration [99].

Physiological changes induced by Cr stress in plants were overviewed by Hayat etal.

[100] and chromium interactions in plants were summarized by Shanker etal. [101].

In our experiment we investigated the response of hydroponically cultivated 3

weeks old rapeseed (Brassica napus L., cv. Verona) plants to Cr(III)-nitrate applied

in the concentration range from 6 to 480μmol dm−3 [102]. Application of Cr(III)

resulted in reduced length, fresh mass as well as dry mass of rapeseed plant organs,

whereby the inhibitory effect increased with increasing metal concentration in the

external solution, the roots being more sensitive to Cr(III) treatment than the shoots.

Signicant reduction of the concentration of assimilation pigments (Chl a, Chl b,

Cars), proteins and reduced thiol groups in rapeseed leaves was also estimated after

treatment with Cr(III), while the concentration of thiobarbituric acid reactive sub-

stances (TBARS) increased with increasing Cr(III) concentration. Even though the

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reduction in root growth could be due to inhibition of root cell division or elongation,

or due to the extension of the cell cycle [103], damage to the root apparatus due to

Cr(III) treatment might cause an unbalanced supply of nutrients and/or an alteration

of their role in anabolic pathways, which ultimately arrest normal physiological and

developmental processes [104] and it can be at high Cr concentrations connected

also with the collapse and subsequent inability of the roots to absorb water from the

medium [105]. Changes in plant water relations resulting in decrease in physiologi-

cal availability of water due to Cr(III) treatment were reported by Pandey and

Sharma [106]. The reduction in shoot height might be mainly due to the reduced

root growth and consequently lesser nutrients and water transport to the above parts

of the plant. Decrease in chlorophyll concentration in leaves of Cr(III)-treated

B. napus plants could be explained by reduced leaf tissue concentration of Fe

because Cr(III) impairs the Fe requiring steps of chlorophyll and heme biosynthesis

[107]. The decrease in protein contents in Cr(III)-treated plant at higher concentra-

tions of Cr was probably due to adverse effects of ROS, which may cause degrada-

tion of a number of proteins [80, 108]. Cr(III) ions at increased concentrations can

interfere with several metabolic processes and decrease of concentration of reduced

thiol groups in the leaves of rapeseed plants due to increasing supply of Cr(III)

indicated that thiol-rich peptides were consumed and detoxication mechanisms in

plants failed to eliminate toxic effects of chromium [102]. Cr(III) can be endoge-

nously reduced to Cr(II) by biological reductants such as cysteine and NADPH and

in turn, the newly formed Cr(II) reacts with H2O2 producing hydroxyl radicals and

causes tissue damage [109]. Cr(III)-induced oxidative stress was reected in the rise

of TBARS levels [102] what is in agreement with ndings of Karuppanapandian

and Manoharan [90]. The BAF values determined for rapeseed (cv. Verona) roots

varied in close range from 1241 to 913 and they were by one till three orders higher

than those determined for shoots which increased with increasing Cr(III) concentra-

tion in hydroponics from 1.92 (6μmoldm−3) to 30.18 (480μmoldm−3). Low mobil-

ity of Cr(III) within the plants was reected by low values of translocation factors

ranging from 0.0015 to 0.0330 [102].

Cd2+ and Hg2+ ions are known to inhibit photosynthetic electron transport (PET)

and they were found to interact with the intermediates Z+/D+, i.e., with the tyrosine

radicals on the donor side of PS II situated in the 161st position in D1 and D2 pro-

teins, with the primary donor of PS I (P700), whereby the oxidation of Chla dimer

in the reaction center of PS I occurred yet in the dark and with the manganese clus-

ter which is situated in the oxygen evolving complex. These ions damaged also all

mechanisms, i.e., direct, cyclic and non-cyclic reductions of P700+ and formed

complexes with amino acid residues constituting photosynthetic peptides what was

suggested as possible mechanism of their inhibitory action [14, 110]. Singh and

Singh [111] and Fodor etal. [112] situated the site of Cd2+ inhibitory action in the

site of QA or QB on the acceptor side of PS II.Inactivation of PS II activity by Ni2+

compounds at donor side of PS II due to interaction with Z•/D• intermediates and

manganese cluster in OEC was estimated as well [113–115] and inhibitory effects

of Ni2+ ions on photosynthetic apparatus could be also connected with their ability

to form complexes with amino acid residues in photosynthetic proteins [116].

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In contrast to Cd2+ and Hg2+ which bind readily to SH-groups, Ni would rather bind

to aromatic nitrogen [117].

Application of 10mgCdkg−1d.m. of soil caused visible symptoms of chlorosis

on B. napus leaves and a statistically signicant decrease in aerial biomass, whereby

biomass decrease depending on the soil type decreased in the following order: min-

eral neutral > organic neutral > mineral acidic > organic acidic and addition of Cd

to soil increased the Cd2+:Ca2+ and Cd2+:Mg2+ ionic ratios [118]. Armas etal. [119]

observed that low Cd concentrations stimulated growth of B. juncea plants, while

application of large Cd concentrations resulted in growth inhibition and increase in

lipid peroxidation due to Cd treatment, which was always greater in shoots than in

roots, whereby an increase in guaiacol peroxidase, ascorbate peroxidase and cata-

lase activities was estimated. Sikka and Nayyar [120] denoted that signicant reduc-

tion in the dry matter yield of B. juncea occurred with application of ≥5mgCdkg−1

soil. Application of Cd to soil decreased the content of micronutrients in plants, but

signicant reduction occurred only for Fe at rates beyond 50mgCdkg−1 soil. Since

no visual toxic symptoms were observed on the leaves of B. juncea grown in a

sandy loam soil polluted with Cd (5–80mgkg−1 soil), it could be concluded that Cd

may accumulate in this vegetable without visual evidence of its presence. However,

accumulated Cd content by plant organs increased with increasing concentration of

applied Cd, being much higher in roots than in shoots. While the relationship of Cd

with Zn and Fe was synergistic in both roots and shoots at the lower rates but antag-

onistic at higher Cd application rates, in the case of Mn and Cu, the relationship was

negative and antagonistic.

Study of in vitro grown callus and seedlings of B. juncea treated with equimolar

Cd concentrations showed that the overall activity of antioxidative enzymes (SOD,

CAT, and APX) was found to be higher in callus in comparison to seedlings of

B. juncea indicating that calli were more tolerant toward Cd-induced oxidative

stress [121]. Verma etal. [122] found that treatment with Cd resulted in an increase

in ionically bound cell wall peroxidase activity in roots of seedlings which showed

direct correlation with increased level of H2O2 in roots.

B. juncea was found to accumulate more Cd in the shoots compared to B. napus,

whereby excess Cd increased the lipid content of B. juncea leaves grown in the pres-

ence of Cd, but did not affect fatty acids composition, while in B. napus leaves an

alteration in the lipid composition as well as a decrease in the lipid contents was

estimated [123, 124]. Higher level of Cd caused signicant accumulation of proline,

gradual increases in activities of antioxidant enzymes such as catalase and peroxi-

dase along with increased lipid peroxidation and decreased the concentrations of

soluble proteins and chlorophylls [125].

Cd induced alteration in lipid prole of developing B. juncea L. seed: total and

non-polar lipids decreased regularly with increasing Cd doses; a positive correlation

was found between saturated fatty acid (palmitic acid, stearic acid) and Cd concen-

tration, while unsaturated fatty acids (oleic acid, linoleic acid, linolenic acid) were

found to be decreasing with increasing Cd concentration. Based on increased ratio in

the saturated/unsaturated acids due to Cd treatment it can be assumed that the synthe-

sis or activity of olelyl-CoA desaturase enzyme was affected signicantly [126].

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Goswami and Das [127] investigated Cd phytoremediation ability of Indian

mustard and found that the highest shoot and root Cd accumulation (10,791 and

9602μg g−1d.w., respectively) was achieved with application of 200mgkg−1 Cd

and the maximum leaf Cd accumulation was 10,071.6gg−1d.w. at 100mgkg−1Cd,

after 21 days of treatment. Harmful Cd-induced effects on growth, photosynthesis,

and oxidative stress in Brassicacae sp. could be alleviated by application of Se [49],

salicylic acid, and Ca [128, 129].

Cd treatment of B. juncea plants was associated with a rapid accumulation of

phytochelatins in the root, where the majority of the Cd was coordinated with sulfur

ligands, probably as a Cd-S4 complex, while Cd moving in the xylem sap was coor-

dinated predominantly with O- or N-ligands and Cd translocation to the shoot

appeared to be driven by transpiration [130].

Nickel is essential for higher plants in low concentrations but becomes toxic to

plants when applied in excess causing plant growth inhibition, chlorosis, necrosis,

and wilting. Its entry in root system can occur via passive diffusion (cation transport

system) and active transport, using the magnesium ion transport system, or by high-

afnity Ni transport proteins [131] and toxic effects of Ni are reected in distur-

bance of mineral nutrition, photosynthesis, water relations, respiration and nitrogen

metabolism of plants [132–134]. Due to Ni-induced oxidative stress which is con-

nected with ROS formation, oxidation of macromolecules in plant tissues [133] and

impairment of membrane function resulting from lipid peroxidation occurs [135,

136]. Krupa etal. [137] indicated an indirect effect of Ni on photosystems related to

the disturbances caused by the metal in the Calvin cycle reactions and downregula-

tion or even feedback inhibition of electron transport by the excessive amounts of

ATP and NADPH accumulated due to non efcient dark reactions.

Amari etal. [138] showed that xylem transport rate of Ni in B. juncea increased

with increasing Ni supply and a positive correlation was established between Ni and

citrate concentrations in the xylem sap and shoots of plant accumulated also signi-

cant concentrations of malic acid and histidine. Khan etal. [139] reported that H2O2

alleviated Ni-inhibited photosynthetic responses through increase in use-efciency

of nitrogen and sulfur, and glutathione production in B. juncea plants. Due to appli-

cation of Ni proline and MDA in the leaves of B. juncea increased with increasing

metal concentration, while soluble protein content was decreased [140]. However,

genotypic variation in phytoremediation potential of B. juncea plants exposed to Ni

stress was estimated [141, 142]. Wang etal. [143] found that at the exposure of

Brassica juncea L. var. megarrhiza to the metal concentration of 300μmoldm−3 as

much as 98% of the Cu and 79% of Cd were retained in the roots, while Ni was rela-

tively uniformly distributed between leaves (32%), stems (29%), and roots (39%),

whereby the dominant storage compartments for Cd and Cu in the stems and leaves

were the cell wall and soluble fractions and the soluble fraction was the dominant

storage compartment for Ni in stems and leaves.

An overview focused on the Ni uptake, essentiality and toxicity in plants was

presented by Yusuf et al. [144]. Screening of ve Brassica species (B. juncea,

B. campestris, B. carinata, B. napus, and B. nigra) for hyperaccumulation of Zn,

Cu, Ni, Pb, and Cd showed that B. napus accumulated highest amount of Pb, Ni, and

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Cd. Succinic acid was characterized and quantied as one of the dominant organic

acids in root exudates of promising Brassica species indicating probable role of this

acid in metal acquisition through complexation [145].

Marchiol etal. [146] evaluated bioconcentration factors related to roots (BCFR) and

shoots (BCFS) of B. napus, B. juncea, Raphanus sativus, and B. carinata grown on a

substrate contaminated by several heavy metals caused by the use of contaminated

irrigation water and found that BCFR was >1 for all the species for Cd, Cu, Ni, and Zn

without signicant differences among species, while BCFS were lower than 0.5.

Mercury is toxic metal which may directly inhibit enzymes by interacting with

protein –SH groups causing protein conformation changes and subsequent enzyme

inactivation [117] and binding of Hg to small biothiols such as glutathione and cys-

teine was conrmed, as well [147, 148]. Excess Hg results in visible symptoms of

phytotoxicity, such as reduced growth, chlorosis, etc. [149, 150], and physiological

disorders in plants [151], PET inhibition [110, 152, 153], closure of leaf stomata

and physical obstruction of water ow in plants [154, 155], inhibition of nutrient

uptake [156], and induction of oxidative stress in plants [157, 158]. According to

Kopittke et al. [159], the toxicity of Hg(II) with median toxic concentration

0.47μmoldm−3Hg against plants grown in solution was higher than the toxicity of

Cu, Cd, As, Co, Ni, Zn, and Mn. The mobility of Hg within the plant is very low

[149, 160, 161], it remains predominantly immobilized in roots where it can be non-

specically absorbed to cell walls or sequestered in complex form with phytochela-

tins in the vacuoles of the root cells [162] and it was reported that even 95–99% of

the Hg taken up by the roots did not reach the shoots [160, 163]. Chen etal. [148]

reported that Hg in plant organs of 10-day-old seedlings of Brassica chinensis

which were exposed for 3 days to 200 μmol dm−3 HgCl2 were estimated

26,089gg−1d.m. in roots and 2839gg−1d.m. in shoots.

B. juncea plants which were grown hydroponically in an Hg-spiked solution

effectively generated an enzymatic antioxidant defense system, especially CAT to

scavenge H2O2, resulting in lower H2O2 levels in shoots with higher mercury con-

centrations and tested cultivars demonstrated an efcient metabolic defense and

adaptation system to Hg-induced oxidative stress. A majority of Hg was accumu-

lated in the roots and low translocation of Hg from roots to shoots was observed

[157]. Elevated Hg concentrations (≥2mgdm−3) resulted in signicant reduction in

both biomass and leaf relative water content and caused signicantly changed leaf

cellular structure represented by thickly stained areas surrounding the vascular bun-

dles, reductions in the number of palisade and spongy parenchyma cells and reduced

cell size and clotted depositions [164]. Up to hundreds of ppm Hg accumulated in

the roots of Indian mustard plants grown with soil contaminated by HgS were

observed by Su etal. [165]. In Brassica juncea L.Czern & Coss, cv. Pusa Jai Kisan

reduction in growth, activities of antioxidant enzymes such as SOD, CAT, APX, and

GR were enhanced with increase of applied Hg concentration. The Hg-induced

alterations in growth were connected with increase in lipid peroxidation (MDA

and H2O2), while the enhanced activities of antioxidant enzymes secured protection

of plants from Hg-induced oxidative stress [166]. In plants of B. juncea co-treated

with Hg and Se high molecular-weight Se/Hg-containing compounds were found

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primarily in the plant root extract which may be protein- associated [167]. Tolerance

of Brassica juncea to mercury can be enhanced also by carbon monoxide [168].

We compared the response of 3 weeks old hydroponically cultivated rapeseed

(Brassica napus L., cv. Verona) plants which were exposed for 7 days to Cd, Ni

(6−120μmoldm−3), and Hg (6−60μmoldm−3) [169]. While local symptoms of

chlorosis were observed at treatment with 6μmoldm−3 Ni, 12μmoldm−3 Hg, and

60μmoldm−3 Cd, plants exposed to 60μmoldm−3 Ni or Hg showed considerable

chlorosis, their leaves were wilted and roots were brownish. Reduction of root and

shoot length due to metal treatment was only moderate and application of

120μmoldm−3 Cd and Ni caused 19% reduction of root length and 13% (Cd) or 6%

(Ni) reduction of shoot length compared to the corresponding control plants. On the

other hand, the same reduction (22%) of both plant organs was obtained with appli-

cation of 60 μmol dm−3 Hg. In general, reduction of dry mass of plant organs

increased with increasing metal concentration in external solution; nevertheless,

more signicant reduction showed root dry mass. The toxicity of metals which was

reected in reduced root dry mass decreased as follows: Hg (IC50: 26.1μmoldm−3)>Cd

(IC50: 78.3μmoldm−3)>Ni (IC50: 84.1dm−3) and the same rank of metals toxicity

was obtained for shoot dry mass. Applied metals reduced shoot water content,

whereby water content in the above-ground part of plants after application of

60μmoldm−3 Cd, Ni, and Hg represented 76.9%, 94.9%, and 74.4% of the control

and due to treatment with 120μmoldm−3 Cd and Ni it declined to 63.0% and 87.3%,

respectively. The concentrations of Chla and Chlb as well as Cars in leaves of young

rapeseed plants decreased as the metal supply was increasing, whereby at treatment

with 60μmoldm−3 the mean reduction of pigment concentrations compared to the

control was about 50% for Cd and Ni and >60% for Hg.

Protein concentration in rapeseed leaves showed exponential decay with increas-

ing metal concentration in hydroponics and at the highest applied metal concentra-

tion the estimated reduction of protein concentrations compared to control plants

was 52.7% (Cd), 39.1% (Ni), and 43.6% (Hg). Concentration of TBARS also rose

with increasing metal concentration and at the treatment with 120μmoldm−3 it was

1.64 (Cd) and 2.28 (Ni) times higher than that of the control, at application of

60μmoldm−3 Hg it was 2.91 times higher.

In the studied concentration range 0−120μmoldm−3 Cd or Ni and 0−60μmoldm−3

Hg, metal concentration accumulated in the roots showed linear increase with

increasing external metal concentration. The most effective metal accumulation in

roots showed Hg, the lowest one Ni. While linear increase of metal concentration in

shoots in the whole investigated concentration range was estimated for Ni and Hg,

gradual saturation of shoot tissue with Cd was observed at concentrations higher

than 24μmoldm−3. The levels of Hg in rapeseed shoots were the lowest in the whole

studied concentration range.

BAFs determined for roots ranged from 1436 to 952 for Cd and from 410 to 225

for Ni, showing a decrease with increasing external Cd and Ni concentration, while

the corresponding BAFs determined for Hg varied only slightly (2227–2567). BAF

values related to shoots ranged from 68.6 to 29.1 for Cd, from 29.7 to 40.4 for Ni,

and from 7.4 to 12.7 for Hg. The effectiveness of metal translocation from root to

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shoot increased as follows: Hg < Cd <Ni. The TF values estimated for Hg

(0.0033−0.0057) were by one to two orders lower than TFs determined for Cd

(0.0478−0.0306) and Ni (0.0724−0.1795), respectively.

Reduction of dry mass of plant organs due to application of studied metals (Cd,

Ni, Hg) estimated in the present experiment which was observed previously also by

many researchers in a wide spectrum of model plants is connected mainly with

inhibition of root cell division (e.g., [170]), inhibition of chlorophyll synthesis [171,

172], lower absorption of macronutrients and microelements from cultivation

medium [173, 174], inhibition of photosynthesis [110, 175, 176] and replacement of

Mg in chlorophyll by the studied metals [46]. Kopittke etal. [159] reported that

across a range of plant species and experimental conditions, the phytotoxicity of the

trace metals followed the trend Hg > Cd > Ni and median toxic concentrations were

0.47μmoldm−3 for Hg, 5.0μmoldm−3 for Cd, and 19μmoldm−3 for Ni.

Gradual decline in shoot water content of rapeseed plants, cv. Verona with

increasing metal concentration, which was considerably higher at Hg or Cd treat-

ment compared to Ni application, could be connected with changes in plant roots

that inhibit water uptake what results in reduced physiological availability of water

[177, 178].

Changes in plant water relations causing physiological drought in B. juncea

L. plants due to exposure to Cd connected with changes in plant water relations

were observed by Singh and Tewari [125] and Gajewska etal. [179] who observed

decline in relative water content due to treatment with higher Ni concentrations. Hg

treatment was found to inhibit water uptake through aquaporins in plasma mem-

branes in higher plants and decrease in transpiration and water use efciency in

plants occurred due to inhibition of water channels in wheat root cells [154]. Hg

rapidly and signicantly decreased the pressure-induced root water ux in tomato

plants exposed to Hg [155]. Similarly as in our experiment, the strong decrease of

the concentration of assimilation pigments with increasing metal (Cd, Ni or Hg)

concentration was observed by several researchers (e.g., [76, 180–186]).

This decrease can be caused for example by toxic effect of Cd on photosynthetic

pigments causing degradation of Chl and Cars as well as inhibition of their biosyn-

thesis [187, 188] and by inhibition of biosynthesis of photosynthetic pigments

resulting in disturbances in electron transport rates of PS I and PS II and subsequent

generation of oxygen free radicals [189]. The reduction in pigment contents due to

Ni toxicity could be attributed to δ-aminolevulinic acid utilization due to inhibition

of Chl biosynthesis by creating nutrient imbalances [76] and due to replacement of

Mg2+ ions by toxic metals [186].

The decrease in protein contents in metal-treated plants is connected with effects

of ROS, which may be due to degradation of a number of proteins [108] as a result

of increased protease activity [190] as well as with the effect of these metals on

nitrate reductase activity [76] and it may also be a consequence of an inhibition of

cell division in young cells which are characterized with particularly intensive pro-

tein synthesis [191]. Reduction in soluble protein content in Cd-treated plants was

also reported by Costa and Spitz [192] and Mohan and Hosetti [193], while Maleva

etal. [194], Duman and Ozturk [195], and Ali etal. [196] published similar results

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concerning Ni-treated plants and protein degradation due to Hg treatment was

observed in wheat chloroplasts by Panda and Panda [197].

Presented results concerning increase of TBARS levels due to treatment with Cd,

Ni, and Hg are in agreement with ndings of several researchers, e.g., Ansari etal.

[166] who informed about Hg-induced increase in lipid peroxidation causing altera-

tions in growth of B. juncea, Maheswari and Dubey [198] who observed elevated

levels of H2O2 and TBARS in rice (Oryza sativa L.) seedlings treated with 200 and

400μmoldm−3 Ni or with results concerning increased TBARS levels due to treat-

ment with Cd in B. napus [180], Brassica juncea [199], Brassica campestris [200],

and Pisum sativum plants [201].

Brassica varieties grown in hydroponics which were tested for Cd accumulation

in shoots accumulated 200–600 mg Cd kg−1 dry mass after treatment with

10μmoldm−3. Rapeseed plants, cv. Verona used in our study, which were exposed

to 12 and 24μmoldm−3 Cd accumulated in the shoots 115.2 and 191.7mgCdkg−1d.m.,

respectively. Consequently, it can be concluded that for Verona cultivar lower levels

of toxic metals in the shoots are characteristic, indicating its tolerance against Cd.

In this experiment Hg accumulation in the roots of Verona cv. was more than by two

orders higher than that in the shoots what is in agreement with previous ndings that

the translocation of Hg from root to shoot is appreciably limited and predominant

amount of mercury uptaken by plants remains immobilized in roots [149]. TF val-

ues estimated by Marchiol etal. [146] for B. napus plants (cv. Kabel) grown 60 days

on multi-contaminated soil containing 38.6mgkg−1 Cd and 46.9mgkg−1 Ni were

0.0053 for Cd and 0.0009 for Ni, while TF values evaluated for considerably

younger rapeseed plants of Verona cultivar grown in hydroponics in the presence of

120μmoldm−3 Cd and Ni factors were considerably higher, 0.0306 and 0.1795,

respectively [169], what could be connected with better bioavailability of tested

metals in plants cultivated in hydroponics as well as with appropriate levels of all

essential nutrients required for growth of rapeseed plants.

7.2.4 Effect ofMetal Complexes onCrops

Chelators can signicantly affect biological activity of metal ions. The chemical

form of the metal is a very important determinant in understanding the quantitative

aspects of metal toxicity and it signicantly affects uptake of metal by plants.

Differential uptake and toxicity of ionic and chelated copper in Triticum aestivum

was reported by Taylor and Foy [202] and positive effect of chelating agent applica-

tion for more effective Cu, Zn, and Pb uptake by several plants was described in

[203]. Cysteine, histidine and aspartate or glutamate are major cellular ligands of

Zn that form tetrahedral coordinations [204] and these ligands bind to Zn with a

greater afnity and with more stability than to Fe, thereby protecting the sulfhydryl

groups against oxidation [205]. Vacuolar sequestration of Zn by high levels of vacu-

olar citrate may be a central mechanism in the accumulation of Zn in plants, exposed

to either low or high levels of this metal [206] and the high root zinc concentration

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and decreased zinc translocation from root into the shoot in pea plants treated with

zinc and succinate suggested that succinate facilitated the formation of metal-succi-

nate complexes in the roots [207]. The application of Zn complexes also signi-

cantly increased Zn uptake by maize plants [208]. In leaves mainly Zn-citrate

complexes exist even though malate is more abundant and in xylem sap citrate and

histidine are the prevalent ligands of Zn [209]. Organic acids, most notably nicoti-

anamine, and specialized proteins bind iron before it can be inserted into target

molecules for biological function and inside the cells and generation of highly toxic

hydroxyl radicals by iron redox reactions is avoided by intricate chelation mecha-

nisms [210]. In our research we focused our attention on the effects of different

metal complexes and metal chelates on PET, growth of alga Chlorella vulgaris and

growth of vascular plants (maize and mustard) as well as on metal accumulation in

plant organs of Zea mays L.

In pyruvideneglycinatocopper(II) complexes the chelate-forming ligand is the

dianion of pyruvideneglycine (condensation product of pyruvic acid and glycine).

Three donor atoms of Schiff base form with the Cu(II) ion two ve-membered che-

late rings affording a structure which is sufciently stable also in aqueous solutions.

The planar arrangement of donor atoms around the Cu(II) ion of the complexes with

Schiff bases results in strong stability increase of Cu(II) oxidation state. PET-

inhibiting activity of pyruvideneglycinatocopper(II) complexes with S-donor

ligands thiourea (tu), ethylenethiourea (ettu) and chlorophenylthiourea (cphtu)

expressed by IC50 value was lower than that of CuCl2⋅2H2O (11.8μmol dm−3):

18.4 μmol dm−3 for Cu(pyrgly)(tu)(H2O), 23.3 μmol dm−3 for Cu(pyrgly)(ettu)

(H2O), 35.1μmoldm−3 for Cu(pyrgly)(cphtu) [211]. Increasing lipophilicity of the

additional molecular ligand led to activity decrease. Toxic effect of these Cu(II)

compounds applied in the concentration 100 μmol dm−3 on maize growth was

reected in reduction of dry mass of roots, stems and leaves related to the corre-

sponding control plants, whereby the most inhibited parameter was dry mass of

leaves, while dry mass of roots and stems was only slightly affected. The smallest

toxic effect on reduction of leaf dry mass exhibited Cu(II) chelate with most

lipophilic ligand, i.e., L = N-(2-chlorophenyl) thiourea and the treatment with

100μmoldm−3 of this compound resulted in increased accumulation of Cu in the

above-ground part of the plant representing approximately 10% from the total

uptaken Cu amount. The toxic effects of these Cu(II) chelate are probably due to the

substitution of their additional ligands with N-, S- or O-donor ligands present in

proteins of plant cells.

Using EPR spectroscopy as the site of action of Cu chelates in the photosynthetic

apparatus the Z+/D+ intermediate and oxygen evolving complex, both situated on

the donor side of PS II was estimated and also interaction of Cu(II) complexes with

aromatic amino acids residues of proteins was conrmed by uorescence measure-

ments [40, 41]. The IC50 values related to PET inhibition in spinach chloroplasts

estimated for a set of aqua(aryloxyacetato)copper(II) compounds (aryl = substituted

phenyl) varied in the range from 4.58 to 22.59μmoldm−3, whereby the most effec-

tive compounds contained Cl substituents in their molecule [40]. On the other hand,

great difference between IC50 values of simple carboxylate copper compound

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[Cu(sal)2(H2O)2]⋅2H2O (IC50=12.6μmoldm−3) [212] and chelates potassium [aqua-

(N-salicylidene- -glutamato)cuprate, K+[Cu(sal-L-glu)(H2O)]⋅2H2O (IC50 = 1967

μmoldm−3), diaqua(N-pyruvidene-β-alaninato)copper(II) monohydrate, [Cu(pyr-β-

ala)⋅H2O]⋅H2O (IC50 = 261 μmol dm−3) and potassium [(isothiocyanato)-

(N-salicylidene-β-alaninato)cuprate(II)], K+[Cu(sal-β-ala)NCS] (IC50=1007 μmol

dm−3) [213] could be connected with higher stability of chelates in aqueous solu-

tion. Similar results were obtained in the study of antialgal activity of above-men-

tioned Cu(II) compounds [214]. In comparison with CuCl2⋅2H2O and Cu(sal)2⋅4H2O,

the lower inhibitory effect of K[Cu(sal-L-glu)]⋅2H2O on growth of maize was con-

rmed, too [212]. In the concentration range of higher metal concentrations the

most sensitive parameter to Cu(II) toxicity was the primary root growth, whereby

the formation of lateral roots and root hairs was also pronouncedly suppressed. The

IC50 values obtained for the inhibition of root and shoot dry mass were higher than

those obtained for the inhibition of root and shoot growth. While the effects of

CuCl2⋅2H2O and Cu(sal)2⋅4H2O were comparable, K+[Cu(sal-L-glu)]⋅2H2O exhib-

ited signicantly lower toxic effect on investigated production characteristics, with

the exception of primary root inhibition. Because the stability constants for com-

plexes of copper with salicylic and glutamic acid are relatively high (logK1=10.6

and 7.85 respectively), it could be assumed that the biological activity of the studied

Cu(II) complexes could correlate with the ability of these complexes to exchange

their additional ligands (H2O molecules) with N-, S- or O-donor ligands occurring

in proteins of Zea mays L.Organic ligands can also contribute to better transport of

metal ions through the lipophilic regions of cell membranes. The lower inhibitory

effect of K+[Cu(sal-L-glu)(H2O)]⋅2H2O probably results from higher stability of

this anionic chelate complex in aqueous solutions and it could be also assumed that

due to sterical conditions the access of copper bound in this complex compound

with polydentate ligands to its site of action is more intensively limited than that of

copper with monodentate ligands. These ndings are in accordance with our previ-

ous ndings [40, 41, 213] that the different coordinating modes of acidoligands

pronouncedly affect the biological activity of Cu(II) compounds.

A set of anti-inammatory Cu(II) complexes with biologically active ligands of

the type CuX2⋅H2O and CuX2Ly, where X = ufenamate (N-(α,α,α-triuoro-m-tolyl)

anthranilate), mefenamate (2-((2,3-dimethylphenyl)amino)benzoate)), niumate

(2-(α,α,α-triuoro-m-toluidino) nicotinate), naproxenate (6-methoxy-α-methyl-2 -

naphthaleneacetate); L = nicotinamide, N,N-diethylnicotinamide, ronicol

(3-hydroxymethylpyridine), caffeine, methyl-3-pyridylcarbamate; and y= 1 or 2

was investigated related to inhibition of oxygen evolution rate (OER) in the suspen-

sions of Chlorella vulgaris [214] and spinach chloroplasts [215]. The anionic X

ligands increased the inhibitory effect while the effect of the L ligands was not sig-

nicant. Taking into account the X ligands, the inhibitory activity decreased in the

order ufenamate ~ niumate > mefenamate > naproxenate, i.e., the most active

inhibitors were compounds containing uorine atoms in their molecules. The PET

inhibiting activities of these compounds in spinach chloroplasts were approximately

two to three orders higher (IC50: 6.6–14.2 μmol dm−3) than those determined

for OER inhibition in C. vulgaris (IC50: 0.976–2.291mmoldm−3) what could be

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connected with the fact that in C. vulgaris for reaching the site of action the inhibitor

must penetrate through the outer and inner algal membranes, while in partially bro-

ken spinach chloroplasts used in this study the corresponding inhibitor could

directly interact with the thylakoid membranes.

Six chelate cuprates of the composition M+[Cu(TSB)(X)]− containing tridentate

Schiff base dianion ligands (TSB2−) of N-salicylideneaminoacidato type (derived

from α-alanine or β-alanine, valine, phenylalanine), additional pseudohalogeno

ligands (NCS−or NCO−), and M (K, NH4 or Na) as well as six molecular (N-

salicylidene-β- alaninato)copper(II) complexes of the composition [Cu(sal-β-ala)

(L)] with additional organic molecular ligands (L = imidazole (im), pyrazole (pz),

pyridine (py), quinoline (quin), urea (ur) or thiourea (tu)) were investigated on their

effects on reduction of chlorophyll content in statically cultivated green alga

Chlorella vulgaris and PET inhibition in spinach chloroplasts [216]. The differ-

ences in immediate toxic effects of all studied Cu(II) complexes on the PET inhibi-

tion in spinach chloroplasts were relatively small and more signicant effect of

individual ligands on the biological activity was not observed. The inhibitory effec-

tiveness of the majority of the tested compounds (with the exception of compounds

Cu(sal-β-ala)(im) and Cu(sal-β-ala)(tu)) was approximately by two orders lower

than that of CuCl2⋅2H2O and Cu(CH3COO)2⋅H2O.As mentioned above, the lower

inhibitory effect of both types of Cu(II)-chelate complexes probably resulted from

their higher stability in aqueous solutions.

The IC50 values of six molecular Cu(II) complexes with additional organic

ligands and N-salicylidene-β-alaninato(2-) ligand related to reduction of chloro-

phyll content in C. vulgaris varied in the range from 21.0μmoldm−3 (Cu(sal-β-ala)

(quin)⋅H2O) to 58.7 μmol dm−3 (Cu(sal-β-ala)(tu)) and the antialgal activity

decreased in the following order: Cu(sal-β ala)(quin)⋅H2O > Cu(sal-β-ala)

(im)>Cu(sal-β-ala)(pz)⋅2H2O>Cu(sal-β-ala)(py)>Cu(sal-β-ala)(ur)>Cu(sal-β-

ala)(tu). The applied amino acid strongly affected the antialgal activity of Cu(II)

cuprates against C. vulgaris and it decreased in the following order: β-alanine,

α-alanine, phenylalanine, valine. The IC50 values determined for compounds

NH4[Cu(sal-β-ala)(NCS)] and Na4[Cu2(sal-β-ala)2(NCS)2](SCN)2⋅4H2O (40.8 and

37.4μmoldm−3, respectively) were about two times lower than for K[Cu(sal-DL-α-

ala)(NCS)] and ve times lower than the corresponding IC50 value estimated for

K[Cu(sal-DL-α-ala)(NCO)] (82.5 and 198.3μmoldm−3, respectively) [216].

In M(II) nicotinamide complexes M(L)2(nia)2 (M=Cd or Zn) which contain coor-

dinated molecules of nicotinamide and anionic ligands L, where L is CH3COO−(ac)

and NCS−, nicotinamide molecule is coordinated to M(II) atom through the ring

nitrogen atom as a monodentate ligand and anionic ligands are also monodentate

coordinated to M(II) atom, whereby acetate anion acts as O-donor ligand and

NCS−anion as N-donor ligand [217]. The IC50 values related to PET inhibition

in spinach chloroplasts determined for Zn(NCS)2(nia)2, Cd(NCS)2(nia)2, and

Cd(ac)2(nia)2 (4.82, 4.44, and 6.64mmoldm−3, respectively) were greater than that

of CdCl2⋅2.5H2O (IC50=1mmoldm−3). The IC50 value for Zn(ac)2(nia)2 could not

be determined and treatment with 6.64mmoldm−3 of Zn(ac)2(nia)2 resulted only in

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25% PET inhibition with respect to the control (application of 5.26mmoldm−3 of

ZnCl2 led to 75% PET inhibition). It is evident that in comparison with acetate

ligands the NCS− ligands contributed to enhanced toxicity of M(L)2(nia)2 com-

plexes and compounds with Cd as the central atom were more toxic than those with

Zn compounds with the same ligands. Moreover, it could be assumed that during the

experiment the ligands L and nicotinamide probably were not replaced by the “bio-

logical ligands” (amino acid residues) occurring in the proteins of spinach chloro-

plasts. Similar results related to toxicity of above-mentioned Cd and Zn compounds

were obtained also in experiment focused on reduction of chlorophyll content in

C. vulgaris [217]. Application of 100μmoldm−3 of M(L)2(nia)2 resulted in reduced

dry mass of roots and shoots of hydroponically cultivated maize plants exposed to

metal compounds for 7 days related to the untreated control plants, whereby the root

dry mass was inhibited to a greater extent than the shoot dry mass with the excep-

tion of Zn(ac)2(nia)2 which did not affect shoot dry mass. The presence of NCS−

ligands caused enhanced toxicity of Zn(NCS)2(nia)2 complex comparing to the

effect of Zn(ac)2(nia)2. The toxicity of similar Cd(L)2(nia)2 complexes was greater

than that of complexes with Zn as central metal atom. In general, treatment with

M(ac)2(nia)2 resulted in better accumulation of metals in the individual plant organs.

However, while the structure of L ligands in Cd(L)2(nia)2 did not affect Cd content

in leaves, application of Zn(ac)2(nia)2 led to approximately four times higher Zn

content in maize leaves than application of Zn(NCS)2(nia)2. In maize plants treated

with Zn(L)2(nia)2 complexes besides estimation of Zn content in the plants also the

content of two essential metals Mn and Cu was determined and the more toxic effect

of NCS−ligands related to that of acetate ligands was reected also in signicantly

lowered amounts of Mn and Cu in roots, stems, and leaves of maize plants treated

with Zn(NCS)2(nia)2 compared with Zn(ac)2(nia)2 [217].

The IC50 values related to OER inhibition in C. vulgaris determined for ZnCl2⋅H2O

and a set of carboxylato and halogenocarboxylato zinc(II) compounds ranged from

0.112 to 1.362 mmol dm−3 and the inhibitory activity decreased in the following

order: Zn(BrCH2COO)2>ZnCl2⋅H2O> Zn(ICH2COO)2 > Zn(ClCH2CH2COO)2 >

Zn(ClCH2COO)2⋅2H2O > Zn(CH3CH2COO)2 > Zn(CH3CH2CH2COO)2 >

Zn(CH3COO)2⋅ 2.5H2O > Zn((CH3)2CHCOO)2 > Zn(HCOO)2 indicating a quasi-

parabolic course of OER- inhibiting activity on the lipophilicity of carboxylato

zinc(II) complexes [218]. It could be noted that the branching of the alkyl chain was

connected with decreased solubility of the compound and resulted in decreased bio-

logical activity, while the introduction of halogene substituent (Cl or I) into ligand

led to activity increase. The above-mentioned Zn(II) compounds also inhibited

growth of Sinapis alba roots, whereby the inhibitory activity of the Zn(II) com-

pounds containing RCOO− ligands was higher (IC50=0.033–2.147mmoldm−3) than

that of ZnCl2⋅H2O (IC50=2.971 mmoldm−3) and the halogenocarboxylato zinc(II)

complexes exhibited higher inhibitory activity than the corresponding carboxylato

zinc(II) compounds [218]. At application of 1mmoldm−3 of ZnCl2⋅H2O and Zn(II)

complexes Zn(CH3COO)2⋅2.5H2O, Zn(CH3CH2COO)2, Zn(ClCH2CH2COO)2 and

Zn(CH3CH(Cl)COO)2 on hydroponically cultivated Zea mays L. (c.v. Karolina)

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shoot length reduction was observed and two compounds, Zn(ClCH2CH2COO)2

(with Cl substituent bonded on β-carbon) and ZnCl2⋅H2O, caused also strong reduction

of root dry mass. The presence of Cl-substituent on α-carbon in Zn(CH3CH(Cl)

COO)2 led to signicant activity decrease. On the other hand, treatment with studied

compounds at concentration 1mmol dm−3 practically did not affect root dry mass

(with the exception of Zn(CH3CH(Cl)COO)2, which caused an increase of root dry

mass, comparing to the effect of equimolar ZnCl2⋅H2O). The bioaccumulated metal

content in roots, stems, and leaves of maize plants depended on the applied carboxy-

lato and halogenocarboxylato zinc(II) compounds, it increased with their increasing

concentrations and the effectiveness of metal translocation from roots into the shoots

was connected with the structure, as well as the applied external concentration of

Zn(II) compounds. In general, BAF values related to roots, stems, and leaves

determined for ZnCl2⋅H2O were higher than those determined for Zn(II) com-

plexes Zn(CH3COO)2⋅2.5H2O, Zn(CH3CH2COO)2, Zn(ClCH2CH2COO)2 and

Zn(CH3CH(Cl)COO)2 and it could be concluded that ionic compound ZnCl2⋅H2O

more easily penetrated into the roots and its translocation in maize plants is also

effective. Moreover, it could be assumed that free Zn(II) ions form complexes with

organic acids occurring in the plant, and Zn will be translocated into the above-

ground part of the plant in the form of Zn-malate, Zn-citrate, etc. [209].On the other

hand, TF values corresponding to the ratio of accumulated metal amount in shoots

and roots (which sharply decreased with increasing concentration of Zn(II) com-

pounds) estimated for the concentration range 10−5–10−3moldm−3 were as follows:

0.679–0.484 for ZnCl2⋅H2O, 0.606–0.102 for Zn(CH3COO)2⋅2.5H2O, 0.954–0.134

for Zn(CH3CH2COO)2, 1.195–0.311 for Zn(ClCH2CH2COO)2 and 1.193–0.069 for

Zn(CH3CH(Cl)COO)2 [218].

The IC50 values of the Ni(II) compounds of the type NiX2L4 (where X = Cl, NCS

or bromoacetate (Brac); L = nicotinamide (nia), 3-hydroxymethylpyridine (ron),

imidazole (iz) and 4-methylpyridine (4-pic)) concerning inhibition of mustard

(Sinapis alba L.) roots were estimated in the range from 0.414 to 0.944mmoldm−3

and increased in the following order: NiCl2⋅6H2O<Ni(NCS)2( nia)4<Ni(NCS)2(iz)4

< Ni(BrCH2COO)2(nia)4 < Ni(NCS)2(ron)4 < NiCl2(ron)4 < Ni(NCS)2(γ-pic)4.

The inhibitory activity of Ni(II) complexes markedly depended on the structure of

N-donor ligands L (ron, iz, nia, γ-pic), whereby the lowest inhibitory activity of

Ni(NCS)2(γ-pic)4 could be connected with the fact that the ligand 4-methylpyridine

do not form supplementary H-bonds with suitable “biological” ligand in the cell. On

the other hand, the formation of H-bonds between –CONH2 group of nicotinamide

ligand or NH group of imidazole ligand and suitable target sites of proteins is highly

probable and these interactions result in plant growth reduction. The root growth-

inhibiting activities of NiCl2(ron)4 and Ni(NCS)2(ron)4 were similar indicating

minor effect of X ligand on the biological activity. Lowered inhibitory activity of

Ni(BrCH2COO)2(nia)4 with respect to that of Ni(NCS)2(nia)4 could be connected

with more lipophilic X ligands (BrCH2COO−) causing decrease of aqueous solubil-

ity of this Ni(II) compound as well as its limited transport through hydrophilic

regions of the cell membranes [219]. Using EPR spectroscopy it was found that

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above-mentioned Ni(II) compounds inhibited PET in spinach chloroplasts due to

interactions with Z+/D+ intermediates, i.e., with tyrosine cation-radicals TyrZ and

TyrD situated in D1 and D2 proteins on the donor side of PS II and with the manga-

nese cluster in the oxygen evolving complex, as well [113]. The Ni(II) complexes

applied at higher concentration inhibited shoot growth of maize (cv. Karolina) more

than root growth; the highest reduction of dry mass of stems and leaves of maize

was observed for the treatment with two compounds with NCS− ligands, namely

Ni(NCS)2(nia)4 and Ni(NCS)2(iz)4. Based on the comparison of the effects of tested

Ni(II) compounds applied in concentrations 250 and 10μmoldm−3 on dry mass of

maize leaves of hydroponically cultivated plants, it could be concluded that the

most toxic effects were exhibited by Ni(NCS)2(nia)4 and Ni(NCS)2(iz)4 which

reduced dry mass of leaves already at relatively low concentration (10μmoldm−3).

On the other hand, application of 10 and 100μmoldm−3 NiBr2(nia)2 had stimulating

effect on growth of primary root of maize plants [220].

Using EPR spectroscopy effect of three iron compounds, FeCl3⋅6H2O,

[Fe(nia)3Cl3] and [Fe(nia)3(H2O)2](ClO4)3 on PET in spinach chloroplasts was stud-

ied and it was found that due to the interaction of these compounds with tyrosine

radicals TyrZ and TyrD located at the donor side of PS II, electron transport between

the photosynthetic centers PS II and PS I was interrupted and the treatment with

[Fe(nia)3(H2O)2](ClO4)3 resulted also in a release of Mn(II) ions from the oxygen-

evolving complex situated on the donor side of PS II [221]. An adverse effect of iron

stress on the photosynthetic electron transport was observed previously by several

researchers (e.g., [222, 223]). Application of 1mmoldm−3 of Fe(III) compounds

inhibited the primary root growth of maize much more than shoot growth and the

growth of adventitious roots as well as root hairs was suppressed [221]. However, at

this concentration expressive reduction of leaf dry mass was estimated,

[Fe(nia)3(H2O)2](ClO4)3 being the most toxic. In this compound the ClO42− anions

are not bound to the Fe atom by a coordination bond and it can be assumed that they

exhibit toxic effects, whereby it could be supposed that water molecules in the coor-

dination sphere of this complex can easily be substituted by another ligands such as

residues of amino acids in proteins. Treatment with [Fe(nia)3Cl3] resulted also

in signicantly higher Fe concentration in plant roots than application of

[Fe(nia)3(H2O)2](ClO4)3 and FeCl3⋅6H2O.

Previously, it was reported that also alga Scenedesmus quadricauda cells

accumulated 2.7–19.6 times higher Fe amounts in cells after treatment with iron

complexes compared to the inorganic salt FeCl3⋅6H2O [224]. On the other hand, leaf

Fe concentration of all three above-mentioned Fe(III) compounds was higher than

that in stems, therefore it could be assumed that they substitute their H2O ligands

and form complexes with organic amino acids occurring in the cell and these com-

plexes secure the mobility of Fe within the plant. The TF values estimated for the

treatment with 0.25, 0.5, and 1.0 mmol dm−3 decreased in the following order:

FeCl3⋅6H2O (0.277, 0.265, 0.143) > [Fe(nia)3(H2O)2](ClO4)3 (0.142, 0.189,

0.137)>[Fe(nia)3Cl3] (0.046, 0.038, 0.066) [221].

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7.3 Medicinal Plants

There are many hundreds of medicinal plants that can be grown in temperate climates

and there are probably a great deal more with properties as yet undiscovered.

Medicinal plants are thus potential plant factories for new natural drugs. Much more

research needs to be carried out on a whole range of medicinal plants in order to nd

safer, more holistic alternatives to the synthetic drugs so often used nowadays.

Moreover, medicinal species are widely used not only in pharmaceutical but also in

food and cosmetics industries, then higher content of toxic metals in their organs is

undesirable. Thus, it is inevitable to have sufciency information concerning the

effect of both bioelements and toxic substances on this important group of the plants.

7.3.1 Effect ofToxic Metals andBioelements

onMedicinal Plants

Because heavy metals may be introduced into medicinal plant products through

contaminated agricultural resources and/or poor production practices, it is neces-

sary to document levels of toxic metals in herbs extensively used in preparation of

products and standardized extracts and investigate whether some species can accu-

mulate toxic elements exceeding permissible levels proposed by the World Health

Organization and European Pharmacopoeia (e.g., [225–228]).

Some regions in Eastern Slovakia regions are traditionally used for commercial

chamomile cultivation in eld conditions, therefore continuous control of heavy

metals content in their shoots utilized for therapeutic purposes is indispensable.

Therefore, we collected and analyzed data concerning Cd accumulation in chamo-

mile plants cultivated in Eastern Slovakia regions in eld conditions [229]. From

data related to the period 1999−2001 that were obtained in four investigated locali-

ties (Streda nad Bodrogom, Košice, Michalovce, Nová Lubovňa), it was found that

the highest Cd level in anthodia (0.168±0.078mgkg−1) was estimated in plants

growing in the locality Streda nad Bodrogom where the lowest Cd content in the

soil was estimated (0.111 ± 0.042 mg kg−1) and the resulting BAF was 1.514.

Because for further three localities in which Cd concentration in soil was in the

range from 0.222 to 0.335mgkg−1 and the BAF values ranged from 0.676 to 0.234

it is evident that Cd translocation was more effective from less contaminated soil.

However, Cd content in anthodium of chamomile plants cultivated and collected in

different localities of Eastern Slovakia in the period 1995–2002 showed signicant

uctuations in individual years with extreme values in years 2001 (0.003mgCdkg−1)

and 2000 (0.505 mg Cd kg−1) suggesting considerable impact of actual climatic

relations on the metal uptake and accumulation. This was conrmed by the linear

increase of Cd content in chamomile anthodium dry mass with the increase of mean

hydrothermic coefcient of Seljaninov (HC; an integrated index of hydrothermic

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parameters taking into account the total amount of precipitations and thermal sum

of the mean daily temperatures exceeding 10°C in the investigated period) [230] for

the period April–June evaluated from the data measured in the meteorological sta-

tions Streda nad Bodrogom and Michalovce in years 1999−2002. In our pot experi-

ments using Cd concentration range in the soil 0–20mg kg−1, the Cd content in

shoots and roots of chamomile plants was higher than that in anthodia; however, in

both cases it increased linearly with increasing Cd concentration in the soil. The

decreased bioavailability of Cd bound to soil particles was reected in lower BAF

values compared to those determined in experiments in which plants were cultivated

hydroponically in the presence of soluble Cd salt (e.g., [231]). Moreover, Grejtovský

etal. [232] reported that Cd accumulation in anthodia of chamomile plants culti-

vated on naturally contaminated soil is lower than in those which were cultivated in

articially contaminated substrates: anthodia of plants cultivated on naturally con-

taminated soil contained only 17% Cd from the total accumulated Cd amount in the

plant, while anthodia of plants cultivated in pot experiments using additional Cd

contamination accumulated till 33.25%, whereby not even the highest applied Cd

dose (20.0mg kg−1) exhibited toxic effect on chamomile biomass or any visible

damage of plants. In general, it should be stressed that Cd bioavailability does not

depend only on plant species but it is inuenced by many factors, such as pH,

organic substances in soil, mechanical composition of soil, redox potential, cation

exchange capacity, etc. (in detail see [233, 234]).

Pavlovič etal. [231] investigated the effect of Cd on physiological and produc-

tion characteristics in two tetraploid cultivars of Matricaria recutita (cv. Goral and

cv. Lutea) in response to the uptake and accumulation of Cd under different cultiva-

tion conditions. The concentration gradient that was used in the experiments reects

Cd content in the soil from non-contaminated to highly contaminated sites [235].

Seeds were germinated and grown in soil in growth chamber under standard condi-

tions: 25°C, 80% relative humidity and 100μmolm−2s−1 PAR irradiance with day/

night photoperiod of 16h light/8h dark. Fourteen days after germination, 20 seed-

lings of each variant were washed to remove the soil adhering to the roots and pri-

mary root length was measured. The lter papers with the plants were coiled, put

into ask and 2cm submerged in Hoagland solution with the following Cd concen-

trations: 3, 6, 12, 24, 60μmol dm−3 Cd(NO3)2×4H2O p.a. The plants growing in

Hoagland solution without Cd served as control variant. The plants were cultivated

at 25°C, 80% relative humidity and 200μmolm−2s−1 PAR for 7 days. Primary root

length, root and shoot dry mass, as well as Cd content in plant organs were deter-

mined after 7 days of treatment. In another experiments the older plants were grown

in the greenhouse conditions in the soil for 7 weeks after germination. Then, the

plants were washed in water and transferred to hydroponic Hoagland solutions

(control) and Hoagland solution containing above-described Cd concentrations as

well as cultivation conditions. Root and shoot dry weight and Cd content in plants

organs were determined after 7 days of exposure. Signicant inhibition of root

growth was observed in both chamomile cultivars after Cd treatment. Fragility,

browning, and twisting of roots were also observed. In shoots leaf roll, chlorosis and

leaf growth inhibition occurred. During the root test chamomile plants cv. Goral

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formed the anthodia in all concentrations except control, despite the fact that the

plants were only 3 weeks old. From 4 to 5 weeks earlier blossoming under Cd

administration was also recorded for Cd hyperaccumulator Arabidopsis halleri [236].

The plants growing in paper rolls exceeded the limit of Cd hyperaccumulator

(100μgg−1) sensu Baker [237] at 60μmoldm−3 Cd in solution. Grejtovský and Pirč

[234] also found over 100μgg−1d.w. in shoots of plants growing in contaminated

substrates. Experimental plants growing in hydroponic solution exceeded the

threshold for Cd hyperaccumulator at 6μmoldm−3 Cd and accumulated up to ve

times more Cd in the shoots than the plants growing in paper rolls. It should

be stressed that 12μmoldm−3 cadmium concentration in hydroponic solution repre-

sents strong contaminated soil [235]; however, the Cd effect on plant was stronger

in comparison to the soil, because Cd ions are not bound to the soil particles and so

all ions are available for plant uptake. The measurements conrmed higher inhibi-

tion of photosynthesis in cv. Lutea (70% of control, P=0.01) although these plants

accumulated less Cd than the plants of cv. Goral (75% of control, P=0.05). Similar

decrease of shoot dry weight (72% of control) in both cultivars was also detected.

Decrease of net photosynthetic rate could be due to structural and functional disor-

ders in many different levels. In both chamomile cultivars shoot and root respiration

rates were not changed signicantly. Considering all found results it was concluded

that this plant species exhibited high tolerance to Cd treatment. Since Cd applica-

tion induced higher production of specic secondary metabolites in chamomile

plants such as α-bisabolol [238], polyacetylenes ene-yne-dicycloethers and sesqui-

terpene (E)-β-farnesene [239], herniarine, umbelliferone [240], these substances

could play a supplementary role in detoxication mechanism induced by Cd. This

additive defensive mechanism was also conrmed by Kráľová and Masarovičová

[241] for cadmium and Hypericum perforatum L. and secondary metabolites hyper-

icin, pseudohypericin (naphthodianthrone derivates) and quercetin that are pro-

duced by this medicinal plant. Similarly, Palivan etal. [242] observed formation of

copper complexes with hypericin.

Within investigation of metal impact on M. recutita plants, we focused our atten-

tion also on the effect of the stage of ontogenetical development on growth of cham-

omile plants and accumulation of supplied cadmium and copper in plant organs

[243]. Seeds of M. recutita L., cv. Bona were sown on the soil in two agrotechnical

date: March 30 (experiment A) and May 28 (experiment B). Plants from experiment

A were grown 12 weeks in the test room at daylight, while those of experiment B

were cultivated only 8 weeks. Then the plants were placed into hydroponics where

they were cultivated 7 days at controlled conditions: control and variants 6, 12, and

24μmoldm−3 CdSO4 or CuSO4, all in Hoagland solution and thereafter some pro-

duction characteristics and metal content in plant organs were estimated. Plants of

cv. Bona were found to be tolerant against application of CdSO4 and CuSO4 because

more expressive toxic effects were observed until at application of the highest metal

concentration (24μmoldm−3). However, it could be noted that while the shoot dry

mass of plants cultivated in both experiments was comparable, root dry mass of

plants (control as well as Cd (Cu) variants) from experiment A was approximately

twofold than that from experiment B. However, in this context it could be men-

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tioned that plants grown in soil in experiment B were exposed to higher levels of

irradiation.

Cd and Cu accumulation in both plant organs increased with increasing metal

concentration in hydroponics, whereby in plant roots it was considerably higher

than in the shoots. Elder, app. 3-month-old chamomile plants from experiment A,

originating from sowing in agrotechnical date, showed higher physiological activity

and accumulated in plant tissues higher Cd amounts than 2-month-old plants from

experiment B, which were sown later (May), thus after the agrotechnical date.

The BAF values related to Cd accumulation in roots were in the range 2177–

2752 for plants from experiment A and 1831–2171 for plants from experiment B,

whereas the BAF values related to Cu accumulation in roots ranged from 1124 to

1836 (experiment A) and from 1273 to 1578 (experiment B), respectively. The BAF

values related to Cd accumulation in shoots showed exponential decrease with

increasing metal (Cd or Cu) concentration.

In contrast to BAF values related to metal accumulation in roots which were only

slightly higher for Cd compared to Cu, the BAF values related to metal accumula-

tion in shoots estimated for Cd were approximately by one order higher than those

determined for Cu. In the case of Cd they ranged from 266.9 to 441.7 (experiment

A) and from 246.3 to 355.3 (experiment B), while the corresponding BAF ranges

for Cu were 13.2–45.6 (experiment A) and 13.6–35.8, respectively. These results

conrmed our previous ndings [244] concerning markedly higher Cd mobility in

chamomile plants. The values of translocation factor evaluated for Cd were in the

range 1.154–0.68 (experiment A) and 0.568–0.353 (experiment B), while for Cu the

TF values were estimated in the range 0.134–0.067 (experiment A) and 0.079–0.048

(experiment B), respectively. While at plants exposed to the lowest metal concentra-

tion (6μmoldm−3) within experiment A the fraction of accumulated metal allocated

in shoots related to the total amount of metal accumulated by plants was >50% Cd

and 11.9% Cu, in plants from experiment B the corresponding values were lower,

namely 36.2% Cd and 4.5% Cu, respectively. Signicantly higher Cd fraction in

shoots of plant cultivated within experiment A was observed also at application of

higher CdSO4 concentrations, while at application of higher CuSO4 concentrations

the differences were lesser. Thus, it can be concluded that physiologically more

active shoots exhibited higher attraction power due to more intensive transpiration

ow (e.g., [245]).

For plants from experiment B the ratio of root dry mass to shoot dry mass ranged

from 0.323 to 0.352, while the corresponding range for plants cultivated in experi-

ment A was only 0.142–0.171. This was reected probably in lower tolerance of

plants from experiment A against Cd-induced stress what led to enhanced extent of

Cd accumulation in both plant organs. High Cd fraction allocated in shoots related

to the total Cd amount accumulated by plants was signicantly affected by the fact

that in plants from experiment A shoot dry mass of Cd-treated plants was 5.86–

7.06-fold higher compared with root dry mass, while for plants from experiment B

the ratio of shoot dry mass to root dry mass was only 2.84–3.10.

Based on the above-discussed results it could be concluded that M. recutita, cv.

Bona is relatively tolerant against Cd-induced stress, however the translocation of this

metal from roots to shoots is signicantly affected by intrinsic factors (ontogenetic

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development, redistribution of assimilates between root and shoot) as well as external

factors such as date of outplanting or cultivation conditions (especially temperature

affecting the transpiration ow).

Interesting results were found by Masarovičová etal. [246] who investigated the

effect of cadmium (6–240μmolCddm−3) on root and shoot dry mass, length of root

and shoot and root respiration rate of 6 weeks old plants of Chamomilla recutita and

Hypericum perforatum. It was found that root dry mass of both studied medicinal

plants was the most inhibited parameter, whereas the length of main root was not

strongly affected. This fact could be explained with signicant reduction of lateral

roots and root hairs by cadmium treatment. Stress-induced higher respiration rate of

metal-treated plants correlated with root growth inhibition what was reected in

lower value of root dry mass. Relatively high Cd uptake into the root required

increased energy costs coming from root respiration rate. Both studied species accu-

mulated in the shoot high Cd concentrations without evident negative effect on their

growth and dry mass production. Based on found results it seems that both medici-

nal plants could have application in rehabilitation and recovering of cadmium con-

taminated sites. Similar ndings were earlier published by Kráľová etal. [247] for

older (6 months old) plant of Hypericum perforatum cultivated hydroponically:

control variant in 0.05% Ca(NO3)2 and Cd treated variant in 0.05% Ca(NO3)2 with

12μmoldm−3 Cd(NO3)2 (pH=5.5). The root dark respiration rate of the Cd-treated

plants was faster than in control plants. The highest Cd concentration was deter-

mined in the root (1792μgg−1d.w.) compared with the leaves (290μgg−1d.w.) or

stem (220μgg−1 d.w.). Moreover, Cd supported the release of some bioelements

(Mn, Fe, and Cu) from the membranes in both, the stem as well as root. Consequently,

ions of these bioelements were transported into the leaves where their higher con-

tent was estimated.

Kummerová etal. [248] studied effects of zinc (12–180μmoldm−3) alone and in

mixtures with 12μmoldm−3 Cd on metal accumulation, dry masses of roots and

shoots, root respiration rate, variable to maximum uorescence ratio (FV/FM), and

content of photosynthetic pigments in hydroponically cultivated chamomile

(Matricaria recutita) plants. The content of Zn in roots and shoots increased with

increasing external Zn concentration and its accumulation in the roots was higher

than that in the shoots. While at lower Zn concentrations (12 and 60μmoldm−3) the

presence of 12 μmol dm−3 Cd decreased Zn accumulation in the roots, treatment

with 120 and 180 μmol dm−3 Zn together with 12 μmol dm−3 Cd caused enhance-

ment of Zn content in the root. Presence of Zn (12–120 μmol dm−3) decreased Cd

accumulation in roots. On the other hand, Cd content in the shoots of plants treated

with Zn + Cd exceeded than in the plants treated only with 12 μmol dm−3 Cd. Only

higher Zn concentrations (120 and 180 μmol dm−3) and Zn + Cd mixtures nega-

tively inuenced dry mass, Chls and Cars content, FV/FM and root respiration rate.

Chl b was reduced to a higher extent than Chl a.

In the paper of Owen etal. [249], an “expected” range for 16 elements (Al, B, Ba,

Ca, Cd, Cr, Cu, Fe, Mg, Mn, Mo, Ni, Pt, Sr, Y, and Zn) in Hypericum perforatum dry

herb and processed preparations (tablets and capsules) was determined. The major

elemental constituents in the analyzed samples were Ca (300–199,000 μg g−1), Mg

(410–3530 μg g−1), Al (4.4–900 μg g−1), Fe (1.154–760 μg g−1), Mn (2.4–261 μg g−1),

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Sr (0.88–83.6 μg g−1), and Zn (7–64 μg g−1). The application of PCA (Principal

Component Analysis) to the elemental proles for the analyzed samples clearly dif-

ferentiated the dry herb samples from the processed samples with additional differ-

entiation between tablets and capsules. A reduction in the average concentration of B,

Ba, Cd, Ni, and Mn occurred post formulation and it has been postulated that this

could be due to factors such as the extraction process and/or powder dilution. Higher

levels of Ca and Mg found in processed products were identied as expected, but

higher levels of Cr, Y, and Sr were also found, which could be due to contamination

from metal alloys used in the manufacturing process. PCA model identied a 7-ele-

ment ngerprint (Ba, Ca, Cd, Fe, Ni, Sr, and Y) capable of differentiating between the

three categories of investigated products of H. perforatum. Results indicating sample

forms (i.e., herb, tablet, and capsule) were differentiated by a change in the elemental

prole contributed by excipient addition, dilution, and/or the extraction process.

7.3.2 Effect ofMetal Chelating Agents onMedicinal Plants

Mobilizing amendments such as chelating and desorbing agents increase the bio-

availability and mobility of metal(loid)s, whereby mobilizing agents can be used to

enhance the removal of heavy metal(loid)s through plant uptake and soil washing

[250] and thus chelator-enhanced phytoextraction of heavy metals is an emerging

technological approach for a non-destructive remediation of contaminated soils.

The presence of ethylenediaminotetraacetic acid (EDTA) in soil can alter the mobil-

ity and transport of toxic metals such as Zn, Cd, and Ni due to the formation of

water soluble chelates, thus increasing the potential for metal pollution of natural

waters. However, due to environmental persistence of EDTA in combination with

its strong chelating abilities in phytoextraction increasingly less aggressive alterna-

tive strategies such as the use of organic acids or more degradable aminopolycar-

boxylic acids are preferred [251, 252].

We investigated the effect of EDTA on bioaccumulation of Cu, Zn, and Cd with

2-month-old Matricaria recutita L. plants (cv. Goral) cultivated in hydroponics and

exposed for 7 days to 12, 24, and 60 μmol dm−3 of individual metals alone or

with addition equimolar EDTA concentration [244]. The application of tested com-

pounds without and with EDTA practically did affect neither the length nor the dry

matter of roots and shoots of M. recutita. The metal content in plant organs increased

with increasing metal concentration in the hydroponic solution, whereby accumula-

tion of all tested metals in roots was higher than in the shoots. Addition of equimolar

EDTA concentration resulted in signicant decrease of bioaccumulated Cu and Zn

amount in plant roots, whereas sharp increase of Cu shoot concentration was

observed, while Cd shoot concentration was elevated only slightly and Zn concen-

tration showed even a moderate decrease. For treatment with the lowest metal con-

centration (12μmoldm−3) the fraction of metal accumulated in the shoots from the

total metal amount accumulated by plant without and with EDTA addition increased

from 16.9 to 33.5% for Cd, from 44.06 to 65.04% for Zn, and from 12.46 to 34.41%

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for Cu. Independently on the concentration of applied metal, the addition of

equimolar EDTA concentration caused that more than 60% Zn and more than 30%

Cd accumulated by chamomile plants was allocated in the shoots, while the enhance-

ment of shoot metal contents in plants treated with higher Cu concentration due to

application of EDTA was dramatic, from 5.2 to 45.2% at 24μmoldm−3 and from 4.8

to 59.2% at 60μmoldm−3. The above-mentioned results indicate that the effect of

EDTA on metal bioaccumulation in roots and shoots of chamomile plants depends

on the applied metal (Cd, Zn, or Cu) and is closely connected with values of stabil-

ity constants of corresponding metal chelates with EDTA.Very efcient transloca-

tion of Cu into the shoots observed at the presence of EDTA could be connected

with the largest value of stability constant of Cu-EDTA chelate (logK1 = 18.8)

which is more than 100 times higher than the corresponding stability constants for

Cd-EDTA (logK1=16.36) and Zn-EDTA (logK1=16.5) chelates [253].

In an another experiment we compared the effect of ZnSO4 without and in the

presence of equimolar EDTA as well as zinc acetate applied in a wide concentration

range (12–120μmoldm−3) on zinc accumulation in 7-week-old chamomile plants

(cv. Goral) exposed for 7 days to Zn compounds. Within this concentration range

the fraction of Zn allocated in the shoots from total metal amount accumulated by

plant depended on the applied Zn form and was as follows: 30–45% for ZnSO4,

35.2–54.4% for Zn(CH3COO)2 and 62–70% for ZnSO4-EDTA application. Signi-

cantly lower effect of zinc acetate on Zn accumulation in the shoot is connected

with very low stability constant of Zn with acetic acid (logK1=1.03) [254]. Reduced

Zn accumulation in roots at EDTA application and at a much lower extent also at

Zn(CH3COO)2 application can be explained by suppressed transport of the formed

complexes through plasmalemma of root cells, whereby specic binding of these

complexes on cell walls does not occur like to binding of ZnSO4.

A further set of our experiments was focused on the study of effects of CuSO4,

copper salicylate tetrahydrate ([Cu(sal)2⋅(H2O)2]2⋅H2O) and seven copper(II)

chelates: copper N-pyruvidene-β-alaninate trihydrate ([Cu(pyr-β-ala)(H2O)2]H2O),

potassium salicylidene--glutamatecuprate(II) dehydrate (K+[Cu(sal-L-glu)]⋅2H2O)

as well as pyruvideneglycinatocopper(II) complexes with additional molecular

S-donor ligands, urea and ethylenethiourea (Cu(pyrgly)(urea)(H2O); Cu(pyrgly)

(ettu)(H2O)), N-donor ligands, pyridine and aniline (Cu(pyrgly)(py)(H2O)2;

Cu(pyrgly)(anil)(H2O)), and O-donor ligands, H2O ([Cu(pyrgly)(H2O)3]) applied at

concentration 24μmoldm−3 on dry mass and Cu accumulation in plant organs of 2

months old plants of Matricaria recutita (var. Goral) which were exposed to metal

compounds for 7 days [255]. At the applied concentration 24μmoldm−3 the studied

Cu compounds did not affect signicantly the length of roots and shoots of chamo-

mile plants. In general, Cu was allocated predominantly in chamomile roots, how-

ever application of Cu in the form of chelate led to more effective Cu translocation

into the shoots in comparison to CuSO4 treatment. The BAF values for CuSO4

related to roots and shoots were 3141 and 18.0, respectively, while for Cu chelates

the BAF values related to roots ranged from 737 (Cu(pyrgly)(py)(H2O)2) to 2524

([Cu(pyr-β-ala)(H2O)2]⋅H2O) and those related to shoots from 14.4 (Cu(pyrgly)

(urea)(H2O)) to 21.6 ([Cu(pyrgly)(H2O)3]). The fraction of Cu allocated in the

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shoots from total metal amount accumulated by chamomile plant decreased as

follows: Cu(pyrgly)(py)(H2O)2 (7.3%) > [Cu(pyrgly)(H2O)3](6.4%) > Cu(pyrgly)

(urea)(H2O)(5.3%) > Cu(pyrgly)(ettu)(H2O)(4.7%) > Cu(pyrgly)(anil)(H2O)

(4.3%) > [Cu(sal)2·(H2O)2]·2H2O(4.0%) > K+[Cu(sal--glu)]·2H2O(3.8%) >

CuSO4·5H2O(2.0%)>[Cu(pyr-β-ala)(H2O)2]·H2O (1.2%). Hence, with the excep-

tion of [Cu(pyr-β-ala)(H2O)2]·H2O the studied chelates accumulated in shoots more

Cu than CuSO4·5H2O, even though the stability constants (logK1) related to Cu

chelates with some amino acids, namely 19.2 for cysteine, 10.6 for histidine, 8.22

for glycine, 7.85 for glutamic acid, 7.13 for β-alanine and the corresponding logK1

value for salicylic, citric acid and pyruvic acid (10.6, 6.1, and 2.2, respectively) are

signicantly lower than logK1 for Cu-EDTA chelate (18.8) [253].

According to Nowack etal. [252], metal chelates with EDTA are taken up via the

apoplastic pathway and disruption of the Casparian band is necessary to achieve the

high metal concentrations in shoots. Therefore, adding chelators to a soil increases

not only the total dissolved metal concentration but also changes the primary route

of plant metal-uptake from the symplastic to the apoplastic pathway. The other

synthetic chelators and low molecular weight organic acids were also found to be

suitable for improving phytoremediation of metal-polluted soils (e.g., [256]).

For example, application of chelator ethylenediaminetriacetic acid on the Cd-

contaminated soil resulted in more than twofold increase of total Cd in Calendula

ofcinalis [257], while the combinative treatment using Cd (30mgkg−1)+humic

acid (2gkg−1)+EDTA (5mmolkg−1) caused maximum Cd-accumulation in root,

shoot and ower of C. ofcinalis up to the extent of 115.96, 56.65 and 13.85mgkg−1

and similar Cd contents in plant organs were achieved also with the treatment con-

sisting of Cd (15mgkg−1)+humic acid (2gkg−1)+ethylenediaminedisuccinic acid

(EDDS; 5mmolkg−1) [258].

7.3.3 Effects ofCadmium andZinc Compounds Containing Se

inDifferent Oxidation States onCrops andMedicinal

Plants

Selenium is an essential nutrient for animals, microorganisms and some other eukary-

otes and it has benecial effects on vascular plants. The ability of some plants to

accumulate and transform Se bioactive compounds has important implications for

human nutrition and health, and for the environment [259]. However, high Se concen-

trations are phytotoxic [260–262]. The major selenocompound estimated in cereal

grains, grassland legumes, and soybeans was found be selenomethionine [263].

The role of Se in stimulation and inhibition of plant growth in various agricul-

tural crops as well as biofortication of some crops with Se using agronomic and

genetic approaches is summarized in the review paper of Kaur etal. [264], while

Sieprawska etal. [265] comprehensively reported about involvement of Se in pro-

tective mechanisms of plants under environmental stress conditions. Feng etal. [266]

summarized the ndings concerning implication of Se in the regulation of ROS and

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antioxidants, the inhibition of uptake and translocation of heavy metals, changes in

the speciation of heavy metals, in rebuilding of the cell membrane and chloroplast

structures and recovery of the photosynthetic system and the researchers suggested

that Se could be involved also in regulation of the uptake and redistribution of ele-

ments essential in the antioxidative systems or in maintaining the ion balance and

structural integrity of the cell and it may interfere with electron transport by affect-

ing the assembly of the photosynthesis complexes. Increased total respiratory

activity in leaves and owers of Se-treated Brassica plants resulted in higher seed

production [267].

The increase in growth of hydroponically cultivated mungbean (Phaseolus

aureus Roxb.) plants due to treatment with Na2SeO4 resulted in signicant stimula-

tion of the activity of starch hydrolyzing enzymes-amylases and sucrose hydrolyz-

ing enzyme-invertase which was associated with elevation of activities of sucrose

synthesizing enzymes-sucrose synthase and sucrose phosphate synthase indicating

that upregulation of enzymes of carbohydrate metabolism provided energy sub-

strates for enhanced growth [268]. Owusu-Sekyere etal. [269] reported that carbo-

hydrate metabolism upregulated by Se via altered redox potential may have some

stimulatory effects on nodulation of alfalfa (Medicago sativa L.), a N2-xing

plant of high nutritive value, which is an important forage legume for sustainable

agriculture.

Although chemical properties of Se are similar to sulfur, and in plants they share

common metabolic pathways and compete in biochemical processes affecting

uptake, translocation, and assimilation pathways in plants, incorporation of Se

instead of S can result in altered tertiary structure and dysfunction of proteins and

enzymes, for example if selenocysteine is incorporated into proteins in place of

cysteine [270, 271]. Lyi etal. [272] reported that due to reduction of selenate and

selenite to selenide and subsequent coupling with O-acetylserine, selenoaminoacids

(Se-cysteine and Se-methionine) are formed which can be non-specically incorpo-

rated into proteins in place of cysteine and methionine and contribute to Se toxicity

in Se non-accumulator plants.

Higher plants take up Se preferentially as selenate via the high-afnity sulfate

permease [273]. Selenite is passively taken up into plants, while selenate enters

plant cells through a process of active transport mediated by sulfate transporters and

directly competes with uptake of sulfate [274–276]. Li etal. [277] found that the

phosphate transporter competitively carries selenite in wheat. Sulfate was found to

be involved in the root-to-shoot translocation of Se in B. napus supplied with sele-

nate, but not selenite [278].

Selenate is the predominant form of bioavailable Se in oxic soils and selenite is

more abundant in anoxic wetland conditions. The reduction of selenate to selenite

appears to be a rate-limiting step in the Se assimilation pathway, since most plants

supplied with selenate accumulate predominantly selenate, while plants supplied

with selenite accumulate organic Se [279]. Plants can also volatilize methylated Se.

While the enzyme ATP sulfurylase appears to be rate-limiting for the assimilation

of selenate to organic Se, cystathionine-γ-synthase is rate-limiting for dimethylsel-

enide volatilization [280]. Foliar application of selenite showed approximately 50%

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less efcient accumulation of Se in shoots compared with selenate and both

treatments exhibited a positive effect in particular on the level of reduced glutathi-

one, whereby selenate-treated plants exhibited higher content of phytochelatin 3in

red clover (Trifolium pratense L.) [281].

In Triticum aestivum and Indian mustard (Brassica juncea) plants grown on sel-

eniferous area of Punjab in India, the highest Se enrichment was estimated in the

upper plant parts what corresponds to the high uptake rate and mobility of selenate

within plants. Occurrence of dimethylselenide and methylselenocysteine in differ-

ent plant parts indicated that active detoxication takes place via methylation and/or

volatilization [282]. The organic Se raised Se concentrations in Brassica napus

plants much less effectively than the inorganic selenite [283].

Morlon etal. [284] reported that the inhibition of growth of Chlamydomonas

reinhardtii alga by selenite was linked to impairments observed at the subcellular

level, whereby chloroplasts were the rst target of SeO32− cytotoxicity, with effects

on the stroma, thylakoids and pyrenoids and Geoffroy etal. [285] demonstrated that

selenate disrupts the photosynthetic electron chain in C. reinhardtii alga causing

also ultrastructural damage (chloroplast alterations, loss of appressed domains).

Se can alleviate phytotoxic effects of heavy metals. Barrientos etal. [286] found

that it is possible to counterbalance negative effects of Cd concerning growth inhibi-

tion, decreased concentration levels of essential micronutrients and oxidative dam-

age by addition of Se. Addition of Se improved the dry weight of root and shoots,

photosynthesis and stomatal conductance in Cd-treated hydroponically cultivated

cucumber (Cucumis sativus) at early growth stage [287]. Treatment with Se or S

alleviated Cd-induced oxidative stress by increasing proline accumulation as a

result of increased activity of glutamyl kinase and decreased activity of proline

oxidase and reduced ethylene level, increased the activity of glutathione reductase

and glutathione peroxidase, reduced oxidative stress and improved photosynthesis

and growth of wheat [288]. Application of Se alleviated Cd toxicity in pepper

(Capsicum annuum L.) plants at the reproductive stage by restricting Cd accumula-

tion in fruits and enhancing their antioxidant activity what resulted in improvement

of the reproductive and stress tolerance parameters [289]. Benecial effects of Se on

different plants under Cd stress were also described by several other researchers

(e.g., [290–292]). Arsenic-induced oxidative stress in roots and shoots of Oryza

sativa L. was signicantly ameliorated by Se supplementation through modulation

of antioxidant enzymes and thiols [293] and exogenous Se application alleviated

chromium toxicity by preventing oxidative stress in cabbage (Brassica campestris

L. ssp Pekinensis) leaves [294].

Root elongation test is suitable to evaluate hazardous waste sites and to assess

toxicity of metals (e.g., [295–297]), including metal nanoparticles [298, 299]. Chen

etal. [300] who investigated roots of Brassica rapa under Se(IV) stress found that

Se inhibits root elongation by repressing the generation of endogenous hydrogen

sulde in root tips. Application of H2S donor NaHS resulted in the increase in

endogenous H2S and signicantly alleviated Se(IV)-induced ROS over-accumulation,

oxidative impairment, and cell death in root tips, which further resulted in the recov-

ery of root growth under Se(IV) stress.

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Lepidium sativum L. and Sinapis alba L. are sensitive test species widely used in

the phytotoxicity testing of toxic metals because they are rapidly growing species

and are cheap and easy to analyze (e.g., [19, 20]). Gomez-Ojeda etal. [301] inves-

tigated the effect of CdCl2 and Na2SO3 treatment on Lepidium sativum L. and found

that after exposure to both elements the changes in glyoxal and methylglyoxal con-

centrations were clearly attenuated as compared to a single stress or treatment and

possible in vivo formation of CdSe quantum dots was also suggested.

To compare phytotoxic effects of Se(IV) and Se(VI), the effect of selenium oxo-

acids and some of their salts of the type MSeO3 and MSeO4 (where X = Cd, Zn, and

Na2) on length of roots and shoots of rapeseed (Brassica napus L.) after 72h expo-

sure of seeds in the dark at mean air temperature (25 ± 0.5 °C) was evaluated

and the set of tested compounds was completed with three other Se-containing

compounds, H2SeO3, an adduct of H2SeO4 with nicotinamide (H2SeO4.nia), and

Cd(NCSe2)(nia)2 (in which oxidation state of Se is (–II), as well as with CdSO4,

ZnSO4, and Cd(NCS)2(nia)2) [261]. With the exception of H2SeO3 and Na2SeO3 root

growth of rapeseed seedlings was inhibited by tested compounds to a greater extent

than the shoot growth. The inhibitory activity of the compounds containing Se in

their molecules expressed by IC50 values varied in the range from 77 to 270μmoldm−3

for root growth inhibition and from 134 to 710μmoldm−3 for shoot growth inhibi-

tion, while that of their sulfur analogs (CdSO4, ZnSO4, and Cd(NCS)2(nia)2) was

several times lower. As the most effective inhibitor Cd(NCSe)2(nia)2 was estimated

with IC50 values 77μmoldm−3 for root and 134μmoldm−3 for shoot growth inhibi-

tion. This complex contains nicotinamide ligands which are unidentate coordinated

to Cd(II) atom and NCSe ligands (acting as N-donor ligands) which are also uniden-

tate coordinated to Cd(II) atom [302]. It could be noted that CdSeO4 exhibited com-

parable root growth inhibiting toxicity with Cd(NCSe)2(nia)2 (IC50: 78μmoldm−3)

but with lower effectiveness in shoot growth inhibition (IC50: 256μmoldm−3).

The phytotoxicity of tested compounds related to root growth inhibition

decreased in the following order: Cd(NCSe)2(nia)2~CdSeO4>CdSeO3>ZnSeO4

> H2SeO4. nia > Na2SeO3>H

2SeO3 > Na2SeO4 > ZnSeO3 > CdSO4 > ZnSO4 >

Cd(NCS)2(nia)2 and for shoot growth inhibition the estimated rank was similar:

Cd(NCSe)2(nia)2>Na2SeO3>H2SeO3>H2SeO4 . nia>CdSeO4>ZnSeO4~CdSe

O3>Na2SeO4>ZnSeO3>CdSO4>ZnSO4>Cd(NCS)2(nia)2. In general, the phy-

totoxicity of Se(IV) compounds in shoots was higher than that of Se(VI) compounds,

but the shoot growth inhibition by CdSeO4 (IC50: 0.256μmoldm−3) was more effec-

tive than with CdSeO3 (IC50: 364μmoldm−3). This could be explained as follows: in

contrast to selenate which enters plant cells through a process of active transport and

treatment with CdSeO4 results in higher mobility, selenite is passively taken up and

translocation of Cd into the shoots at application of CdSeO3 is lower. Consequently,

CdSeO3 exhibit less toxic effect in shoots than the more mobile Cd selenate.

In an another study in which at the same experimental conditions the root and

shoot growth of cress (Lepidium sativum L.) seedlings in the presence of some Se

compounds was estimated the IC50 values related to root inhibition varied in the

range from 8.1 to 91 μmol dm−3 and those for shoot inhibition from 19.5 to

130μmoldm−3, whereby the most toxic compound was Cd(NCSe)2(nia)2 and the

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lowest toxicity was exhibited by Na2SeO4 [260]. Thus, sensitivity of cress seedlings

to the treatment with Se containing compounds was signicantly higher than that of

Brassica napus [261]. However, it could be noted that in cress seedlings comparable

phytotoxic effect to that of Cd(NCSe)2(nia)2 exhibited ionic compound with SeCN−

anion (KSeCN) as well as Na2SeO3 with IC50 values 10 and 13.3μmoldm−3 for root

inhibition and 24.5 and 27.2μmoldm−3 for shoot inhibition.

Tested compounds Cd(NCS)2(nia)2 and Cd(NCSe)2(nia)2 contain biologically

active isothiocyanate and selenocynate ligands.

Linear triatomic thiocyanate and selenocyanate groups with 16 valence electrons

possess a great variety of bonding possibilities [303]. In general, isothiocyanates

(R–N=C=S) were found to be quite reactive, although less than the related isocya-

nates (R–N=C=O). The isothiocyanate anion is a resonance hybrid with greater

charge on the S [304], although charge can be localized on either the sulfur (−S–

C≡N) or the nitrogen (S=C=N−), depending on the environment [305]. Thiocyanate

(R–S–C≡N) is sometimes produced, particularly in members of the Alyssum,

Lepidium, and Thlaspi families [306]. Thiocyanate ion was found to inhibit shoot

and root growth of several plants [271, 307, 308] and at concentrations exceeding

2000ppm NH4SCN inhibited PET in isolated chloroplasts and inhibited the conver-

sion of glycine to sugars, while it had no effect on conversion of glycine to organic

acids in leaf tissue of cotton (Gossypium hirsutum L.) [309]. The relative toxicity of

SeCN− was comparable to that of selenate and selenite using the metalloid-resistant

bacterium LHVE as the test organism, whereby the reduction and methylation of

SeCN− was similar to that of selenate and selenite by other metalloid-resistant

bacteria [310]. Moreover, cultures of LHVE amended with SeCN− on agar plates

produced red, elemental selenium after 3 days [311].

In plants sulfur from thiocyanate may enter the sulfur assimilation pathway what

results in the production of other volatile sulfur gases, e.g., dimethylsulde [312]

and the same assimilation pathway was reported for assimilation of SeO42− in

Brassica juncea [313, 314]. Therefore, a pathway of SeCN− metabolism was pro-

posed in analogy to thiocyanate metabolism in plants by de Souza etal. [315].

In the phytotoxicity test in which mustard was used Cd(NCSe)2(nia)2 (IC50 for

root/shoot: 0.077/0.134mmoldm−3) showed much higher toxicity than its sulfur

analog Cd(NCS)2(nia)2 (IC50 for root/shoot: 1.292/1.031mmoldm−3). The great dif-

ferences in the toxicity (of about one order) could be connected with the differences

in overall stability constants of both compounds which is 602.56 for complex com-

pound Cd(NCS)2 and only 199.53 for Cd(NCSe)2 [316]. Due to three times lower

stability of the compound with NCSe− ligands, after their release from the complex

Cd could interact with suitable target groups on biomolecules and NCSe− anion

may exert its harmful effect, as well.

In further experiments, we investigated the effects of some Cd and Zn compounds

containing Se in their molecules on production and biochemical characteristics as

well as on accumulation of Cd(Zn) and Se in three crops Vigna radiate [317], Pisum

sativum [318] and Brassica juncea [319] as well as in three medicinal plants,

Hypericum perforatum [178], M. recutita, cv. Goral [320] and cv. Lutea [302], and

Salvia ofcinalis, cv. Primorska [321].

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In experiment with Vigna radiata L. 3 days old seedlings were exposed to CdSO4,

CdSeO4, and CdSeO3 (0.5, 1.0, 2.0, 3.0 a 4.0μmoldm−3) in hydroponic solution and

cultivated for 7 days [317]. Due to treatment with CdSO4 and CdSeO4 reduction of

root dry mass was comparable (approximately 25% at 4.0μmoldm−3), while reduc-

tion of shoot dry mass of young seedling was greater with CdSO4 (approximately

42%) as with CdSeO4 (approximately 27%). For all tested compounds Cd concen-

tration in shoots increased linearly with increasing metal concentration in hydro-

ponics, while the dependence of Cd concentration in roots showed polynomic

course indicating gradual saturation of root tissue with Cd. Substitution of sulfur

with Se resulted in signicant reduction of Cd accumulation in both plant organs.

The lowest bioaccumulation of Cd was observed after CdSeO3 treatment. The TF

signicantly depended on the applied compounds and they decreased as follows:

CdSO4>CdSeO4>CdSeO3. Whereas after treatment with higher CdSO4 concentra-

tions (3 and 4 μmol dm−3, respectively) the fraction of Cd from total amount of

metal accumulated by the plant found in its shoots represented more than 20%, for

CdSeO4 and CdSeO3 it was only 14.5% and 7–8%, respectively. These results indi-

cate that V. radiata plants in the early ontogenetic stage are sensitive to Cd-induced

stress what is in agreement with previous ndings of Šimonová etal. [6] and Wahid

and Ghani [322, 323]. The reduced toxicity of Cd selenate and Cd selenite com-

pared to that of CdSO4 is caused probably by formation of insoluble Cd-Se com-

pounds in roots [324, 325], whereby this effect was more pronounced with selenite.

According to Wahid and Ghani [322, 323] accumulated Cd exhibits toxic effects

mainly in mesophyll, presumably due to interference with essential nutrient uptake

what results in the reduction of growth in different phenological stages of V. radiata.

Cultivars of V. radiata which were tolerant to Cd induced stress exhibited higher

peroxidase and catalase activity than the sensitive cultivars [326]. Dhir etal. [327]

observed gradual increase of proline concentration with increasing Cd2+ concentra-

tion in V. radiata plants and Anjum etal. [328] demonstrated that Cd tolerant geno-

type of V. radiata has powerful antioxidant defense system securing sufcient

protection against Cd-induced oxidative stress.

In experiment with Pisum sativum L. 3 days old seedlings were exposed to

CdSeO4, CdSeO3 and Cd(NCSe)2(nia)2(3–60μmoldm−3) for 14 days in controlled

conditions [318]. Dry mass of roots and shoots of pea plants decreased with increas-

ing concentration of the studied compounds and treatment with 60 and 120μmoldm−3

resulted in desiccation of the shoots as well as in the damage of root cells by Cd

resulting in uncontrolled ion uptake what was documented with signicant increase

of the corresponding BAF values. Cd concentration in the roots reached higher lev-

els than in the shoots and accumulated Cd amount in plant organs increased with

increasing Cd concentration. In general, the BAF compounds related to Cd accumu-

lation in roots decreased in following order: Cd(NCSe)2(nia)2>CdSeO3>CdSeO4,

while BAF values related to Se accumulation in roots estimated for CdSeO4 were

signicantly lower than those for CdSeO3 and Cd(NCSe)2(nia)2, respectively. The

corresponding BAF values related to Cd bioaccumulation in pea shoots ranged from

7.3 to 50.0 for CdSeO4, from 5.6 to 35.6 for CdSeO3, and from 11.9 to 35.9 for

Cd(NCSe)2(nia)2, while the TF ranges related to Se bioaccumulation in shoots were

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49.4–173.8 for CdSeO4, 8.7–25.9 for CdSeO3, and 5.6–16.9 for Cd(NCSe)2(nia)2.

Application of 3 μmol dm−3 resulted in reduced Cd uptake for treatment with

CdSeO3 and reduced Se uptake for treatment with CdSeO4 indicating interactive

effects of Cd and Se. The results conrmed higher mobility of CdSeO4 within the

pea plants when 38% of Cd and 89% of Se from the total accumulated metal amount

by the plant was allocated in the shoots, while the corresponding fraction estimated

for CdSeO3 and Cd(NCSe)2(nia)2 reached only 18% for both Cd and Se. Similar

ndings were presented by Arvy [274], Shanker etal. [324, 329], Kráľová etal. [302],

and Lešíková etal. [320]. According to Whanger [325], the presumed protective

effect of Se against cadmium and mercury toxicity is through the diversion in their

binding from low-molecular-mass proteins to higher-molecular-mass ones.

Indian mustard (Brassica juncea L.) belongs to plants which are able to accumu-

late considerable concentrations of both Cd and Se in their shoots [130, 330] what

could be explained with rapid accumulation of phytochelatins in the roots where the

majority of the Cd is coordinated with sulfur ligands, probably as a Cd-S4 complex,

while in the xylem sap Cd is coordinated predominantly with O- or N-donor ligands

[130]. Signicant decrease of Cd uptake by selenite and selenate application was

reported by Shanker etal. [329], selenite being more effective and protective effect

of Se(IV) against Cd-induced DNA damage and chromosomal aberrations was esti-

mated, as well [28].

In experiment with Brassica juncea L., cv. Vitasso we exposed 3 weeks old

plants to 12, 24, and 60 μmol dm−3 CdSeO4, CdSeO3, Cd(NCSe)2(nia)2, and

Cd(NCS)2(nia)2 for 7 days [319]. The inhibition of B. juncea plant growth by the

studied compounds was reected in reduced length and dry mass of plant organs

and it increased with increasing compound concentration, dry mass of plant organs

being affected to greater extent. The toxicity of CdSeO3 and Cd(NCSe)2(nia)2 was

comparable and CdSeO4 was found to be less toxic. Addition of 60μmol dm−3

CdSeO3 and Cd(NCSe)2(nia)2 led to considerable decrease of leaf water content,

while lower loss was observed also at treatment with CdSeO4 and Cd(NCS)2(nia)2.

All tested compounds applied in tested concentration range caused strong reduction

of leaf Chl content. Se oxidation state strongly affected Cd and Se concentration in

plant organs. BAFs for roots as well as shoots related to Cd and Se decreased with

increasing compound concentration. The ranges of BAF values for shoots concern-

ing Cd(Se) were for individual tested compounds as follows: 170.5–93.7 (351.2–

204.1) for CdSeO4, 84.1–28.0 (45.8–13.4) for CdSeO3, 132.3–62.6 (91.8–46.3) for

Cd(NCSe)2(nia)2 and 191.8–81.8, for Cd(NCS)2(nia), while the corresponding

ranges of BAF values estimated for roots were several times higher. From these data

it is evident that (1) treatment with CdSeO4 led to the highest BAF values for both

plant organs, (2) CdSeO3 treatment resulted in the lowest Cd accumulation in both

roots and shoots and (3) application of Cd(NCSe)2(nia)2 caused higher Cd root and

lower Cd shoot concentration, in comparison with Cd(NCS)2(nia)2 addition.

Similarly, shoot Se concentrations decreased as follows CdSeO4>Cd(NCSe)2(nia)

2>CdSeO3. The fraction of Cd(Se) from total amount of elements accumulated by

the plant found in its shoots was >50% of Cd and >87% of Se at treatment with

CdSeO4, 39.6–46.6% Cd and 20.3–30.1% Se at treatment with CdSeO3 and 49.0–

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55% Cd after application of Cd(NCSe)2(nia)2. The results conrmed higher toxicity

of Se(IV) compared to Se(–II) as well as toxicity increase after substitution of sulfur

in Cd(NCS)2(nia)2 with Se(–II) and are in agreement with ndings of Ximenez-

Embun etal. [330], Zayed etal. [331], and De Souza etal. [315] who also used B.

juncea as a model plant.

In all experiments with medicinal plants 6 weeks old plants were exposed to the

same Cd compounds for 7 days as in above-discussed studies with crops, while

S. ofcinalis was exposed also to their zinc analogues.

Application of CdSO4, CdSeO4, CdSeO3, Cd(NCS)2(nia)2, and Cd(NCSe)2(nia)2

applied at concentrations 12, 24, and 60 μmol dm−3 reduced dry mass of plant

organs, water content of shoots as well as leaf Chl content in Hypericum perforatum

L. and these effects increased with increasing compound concentrations. Treatment

with 60μmoldm−3 of studied compounds caused leaf desiccation and leaf fall what

was reected in reduced shoot dry mass.

At this concentration the loss of leaf water content was the highest for treatment

with CdSeO4 and Cd(NCSe)2(nia)2 and it further decreased in the order

CdSO4>CdSeO3> Cd(NCS)2(nia)2 [178]. Toxic metals such as Cd affect plasma

membrane permeability what results in reduction of water content [177, 332]. The

adverse effect of tested compounds on leaf Chl concentration decreased as follows:

Cd(NCSe)2(nia)2>CdSeO3>CdSeO4≈Cd(NCS)2(nia)2≈CdSO4. While Cd affects

chlorophyll biosynthesis and inhibit protochlorophyll reductase and aminolevulinic

acid (ALA) synthesis [333], Se can inhibit Chl synthesis not only by acting on con-

stituent biosynthetic enzymes but also through lipoxygenase-mediated lipid perox-

ide levels and inhibition of antioxidant defense component [334]. Treatment with

Se(VI) reduced Chl concentration in lettuce (Lactuca sativa) [335] and inltration

of adult coffee plants leaves with Se(IV) also resulted in decrease of chlorophylls,

carotenoids, and xanthophylls [336].

While treatment with CdSeO4, CdSO4, and Cd(NCS)2(nia)2 resulted in linear

increase of Cd concentration accumulated in roots with increasing compound

concentration, at application of 60 μmol dm−3 of CdSO4 and Cd(NCSe)2(nia)2

consecutive saturation of the roots with Cd was observed [178]. Similarly, also Se

concentration in roots and shoots of H. perforatum plants increased linearly with

the applied compound concentration, only at the treatment with the highest

Cd(NCSe)2(nia)2 concentration saturation of the roots with Se was observed. With

regards to application of individual studied compounds root Se content decreased in

the following order: CdSeO3>Cd(NCSe)2(nia)2>CdSeO4, while for shoot Se con-

centration this sequence was opposite: CdSeO4 > Cd(NCSe)2(nia)2 > CdSeO3.

Because the most effective Cd accumulation in both plant organs was observed with

CdSO4, it is evident that reduced Cd accumulation obtained with Se-containing

compounds is due to Cd-Se interference. According to Shanker etal. [324] the less

mobile anion SeO32− after being reduced to selenide tends to form Cd–Se complex,

which appears to be unavailable for the plants, while the more mobile anion SeO42−

is available for Cd–Se formation only after following a more complicated redox

processes involving Se(VI) in selenate, Se(IV) in selenite and Se(0) species.

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Portion of Cd allocated in H. perforatum shoots related to the total Cd amount

accumulated by the plant was about 20% for treatment with CdSO4 and

Cd(NCS)2(nia)2 and about 12.8%, 10%, and 6% for treatment with Cd(NCSe)2(nia)2,

CdSeO4, and CdSeO3 what are considerably lower values than the corresponding Cd

portions estimated for Matricaria recutita, cv. Goral [320], Brassica juncea [319],

and also Pisum sativum plants [261]. Portion of Se allocated in shoots related to the

total Se amount accumulated by H. perforatum plants achieved approx.. 86%,

48.6%, and 45.9% after addition of CdSeO4, Cd(NCSe)2(nia)2, and CdSeO3, while

the corresponding Se portions were 91.5%, 27.8%, and 25.8% in M. recutita, cv.

Goral, 90.3%, 51.5%, and 26.4% in Brassica juncea [178] and 89%, 18%, and 18%

in Pisum sativum plants [318].

In experiments with M. recutita plants focused on the effects of four Cd

compounds (CdSeO4, CdSeO4, Cd(NCSe)2(nia)2 and Cd(NCS)2(nia)2) applied at

concentrations 12, 24, and 60 μmol dm−3 two chamomile cultivars, cv. Goral [320]

and cv. Lutea [302] were chosen and in the experiment with cv. Goral beside accu-

mulated Cd concentrations in plant organs also those of accumulated Se were

estimated.

In cv. Lutea application of Cd(NCS)2(nia)2 affected neither the length nor the dry

mass of roots and shoots, while other three compounds partially reduced dry mass

of plant organs already at application of 24mol dm−3, while in cv. Goral adverse

effects were observed only for treatments with 60μmoldm−3. The accumulated Cd

and Se content in roots and shoots of chamomile plants treated with Cd salts of Se

oxoacids increased with increasing compound concentration in hydroponic solution

and depended on the oxidation state of Se.

The ranges of BAF values for Cd related to roots in the concentration interval

12–60moldm−3 were after application of Cd compounds as follows: 582.3–209.6

(Goral) and 863–280 (Lutea) for CdSeO4; 581.9–505.9 (Goral) and 555–731(Lutea)

for CdSeO3; 518.9–646.4 (Goral) and 633–759 (Lutea) for Cd(NCSe)2(nia)2; 673.8–

335.5 (Goral) and 584–311 (Lutea) for Cd(NCS)2(nia)2. The corresponding ranges

of BAF values for Cd related to shoots were as follows: 158.0–60.4 (Goral) and

155.0–70.7 (Lutea) for CdSeO4; 100.1–29.4 (Goral) and 54.4–28.7 (Lutea) for

CdSeO3; 74.1–30.6 (cv. Goral) and 48.6–25.1 (Lutea) for Cd(NCSe)2(nia)2; 141.9–

91.8 (cv. Goral) and 90.5–60.3 for Cd(NCS)2(nia)2.

The ranges of TF values estimated for Cd were 1.014–1.316 (Goral) and 0.60–

0.93 (Lutea) for CdSeO4; 0.483–0.512 (Goral) and 0.200–0.105 (Lutea) for CdSeO3;

0.398–0.151 (Goral) and 0.27–0.12 (Lutea) for Cd(NCSe)2(nia)2; 0.737–0.905

(Goral) and 0.60–0.63 (Lutea) for Cd(NCS)2(nia)2.

In both cultivars the TF for Cd estimated with application of CdSeO4 and

Cd(NCS)2(nia)2 were more than two times higher than those found for CdSeO3 and

Cd(NCSe)2(nia)2, whereby the highest fraction of Cd accumulated in shoots was

observed for CdSeO4, while the lowest one for Cd(NCSe)2(nia)2. From both culti-

vars Goral was found to be more tolerant to the cadmium exposure compared to the

cultivar Lutea what is in agreement with the ndings of Pavlovič etal. [231].

While after application of CdSeO4 and Cd(NCS)2(nia)2 approximately 40% from

the total amount of Cd accumulated by cv. Lutea plants were allocated in shoots and

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for treatment with CdSeO4 in cv. Goral this portion exceeded 50%, due to treatment

with CdSeO3 and Cd(NCSe)2(nia)2 approximately 80% (or more) Cd from the total

amount of metal accumulated by the plants remained in roots.

Se speciation signicantly affected also bioaccumulated amount of Se in chamo-

mile (cv. Goral) plant organs [320]. Due to treatment of chamomile plants with

CdSeO4 more than 90% from the total uptaken Se by plants was allocated in shoots,

while this portion was about 30% for the treatment with Cd(NCSe)2(nia)2 and only

approximately 17% at treatment with 12 and 24μmoldm−3 CdSeO3. These results

conrmed higher mobility of CdSeO4 within the plant compared with CdSeO3 and

are in agreements with previous ndings (e.g., [178]).

In experiment with sage the effect of compounds of the type MSeO4, MSeO3,

M(NCSe)2(nia)2, MSO4, M(NCS)2(nia)2 with M = Cd or Zn on root and shoot dry

mass and Cd, Zn, and Se bioaccumulation in plant organs of hydroponically culti-

vated Salvia ofcinalis L., cv. Primorska plants was investigated [321]. For treat-

ments with Cd compounds concentration of 120μmoldm−3 was applied, while for

treatments with Zn compounds two concentrations (60 and 120μmol dm−3) were

used. These concentrations were toxic for the plants what was reected in wilting

and desiccation of plant leaves. BAFs related to Zn, Cd, and Se accumulation in the

roots were higher than those determined for the shoots. The highest BAF values

related to metal accumulation (Cd or Zn) in the shoots were observed for CdSO4,

ZnSO4, and Zn(NCS)2(nia)2 application. Oxidation state of Se in the studied Cd and

Zn compounds affected not only the toxic effect of these compounds, but also the

uptake and translocation of Cd and Zn into sage plants. More effective Cd transloca-

tion into the shoots was observed at application of CdSO4 and CdSeO4 compared to

CdSeO3, while the highest Zn mobility in S. ofcinalis plants was estimated at

application of 120 μmol dm−3 Zn(NCSe)2(nia)2 and Zn(NCS)2(nia)2. The lowest

Se translocation into the shoots exhibited both complex compounds of the type

M(NCSe)2(nia)2 (M = Cd or Zn).

Portion of Cd allocated in S. ofcinalis shoots related to the total Cd amount

accumulated by the plant was after application of Cd compounds (c=120μmoldm−3)

≈80% for CdSO4 and CdSeO4, 67.6% for CdSeO3 and 58–59% for Cd(NCSe)2(nia)2

and Cd(NCS)2(nia)2, respectively. The corresponding portion of accumulated Se

represented ≈80% for CdSeO4 and CdSeO3 and ≈47% for Cd(NCSe)2(nia)2. While

application of ZnSO4 resulted in relatively low portion of Zn allocated in S. ofci-

nalis shoots related to the total Zn amount accumulated by the plant (43%), the

highest values of this portion was obtained with application of Zn(NCSe)2(nia)2(73%),

whereby for Se accumulation it represented only 28.5%. At assessment of toxic

effects of complex compounds of the type M(NCSe)2(nia)2 (in which the oxidation

state of selenium is (–II)) also the overall stability constants estimated for Cd(NCSe)2

(199.53) and Zn(NCSe)2 (4.37) have to be considered. It is evident that the stability

constant for Zn(NCSe)2 is very low and therefore it can be assumed that after rapid

dissociation of the complex the released NCSe− ligands can interact with suitable

target groups on biomolecules.

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7.4 Trees: Characteristic Functions andUtilization

Wood of the trees, used by human societies for millennia, undoubtedly remains one

of the world’s most abundant raw materials for industrial products and renewable

energy. In general, wood-working industry represents app. 40% portion from total

technically utilizable potential of biomass. Besides this most important production

function trees possess also further, non-production functions. Trees (both forest and

fast growing trees, respectively) play a signicant role in reducing erosion and mod-

erating the climate. The last mentioned function of the trees is extraordinarily

important from the aspect of global atmosphere warming. Additionally, trees not

only remove carbon dioxide from the atmosphere and store large amount of carbon

in their tissues, but also release large part of oxygen into the atmosphere. Moreover,

enormous leaf biomass is after decomposition of the source of mineral nutrition in

the soil (cycle of the mineral nutrients) and can also serve as a lter against various

pollutants. Finally, trees and forests provide a habitat for many species of plants and

animals.

It should be also considered that the response of forest ecosystems to increased

atmospheric CO2 is constrained by nutrient availability. It is thus crucial to account

for nutrient limitation when studying the forest response to climate change. Jonard

etal. [337] in their last study described the nutritional status of the main European

tree species, to identify growth- limiting nutrients and to assess changes in tree nutri-

tion during the past two decades. These authors emphasized that increased tree pro-

ductivity, possibly resulting from high N deposition and from the global increase in

atmospheric CO2, has led to higher nutrient demand by trees. Based on found results

it was suggested that when evaluating forest carbon storage capacity and when plan-

ning to reduce CO2 emissions by increasing use of wood biomass for bioenergy, it

is crucial that nutrient limitations for forest growth should be considered.

It should be emphasized that large resources of biomass energy are related mainly

to forestry residues, forestry fuel wood, and fast growing woody plants (e.g., willow,

poplar, black locust, and European alder). In north European countries willow and

poplar have already great tradition for their plantation cultivation. However, new

biotechnological approach showed that energetic plants including the trees (such are

above-mentioned fast growing trees) have also signicant application for environ-

ment friendly management, mainly in phytoremediation technology. Phytoremediation

can be presented as a cleanup technology belonging to the cost-effective and environ-

ment-friendly biotechnology. Several types of phytoremediation technologies being

used today is briey outlined in the part of 7.5 of this chapter.

Trees such as poplar, willow, black locust, ash, or alder are not indeed fast grow-

ing species (for comparison see Masarovičová etal. [338]) but for their convenient

biological features these woody plants can be used to clean up substrates contami-

nated by both inorganic and organic pollutants. These plants have perennial charac-

ter, long lifespan, high transpiration rate, quick regeneration of removed above-ground

parts, and easy vegetative reproduction (in detail see Stomp etal. [339]). Moreover,

fast growing trees have an extensive and massive root system penetrating deeply into

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the soil and ensuring efcient uptake of water containing the pollutants from the

substrate. Mainly poplar and willow have been shown to be excellent species for

phytoremediation purposes because they can be cultivated at high rates of growth

and thus produce a large biomass. Leaves of this biomass have not only large tran-

spiration potential but they also can uptake large amount of contaminated water.

According to Chapell [340] advantages of, e.g., genus Populus in phytoremediation

are great number of species, fast growth up (3–5m/year), high transpiration rate

(100L/day) and not being a part of food chains. The use of plants producing large

biomass for metal extraction from soil was proposed as an alternative to hyperac-

cumulators (these plants are mostly characterized by low biomass production and

high tolerance against toxic substances) because high biomass production estab-

lishes to compensate moderate heavy metal concentrations in their shoots. Poplars

allow several cycles of decontamination, their leaves can be easily collected and the

contaminated biomass substantially reduced by incineration (cf. [341, 342]). It

should be stressed that precondition for effective utilization of woody plants in

phytoremediation technologies is their sufcient toxic metal tolerance. Therefore,

effects of toxic metals (especially Cd, Hg, Pb, and Cu) on structure and function of

trees are still intensively studied. As cadmium belongs to the most dangerous envi-

ronmental pollutants and has toxic and mutagenic effects on both the plants and

animals, our attention was focused predominantly on this toxic metal (see next

Sect. 7.4.1).

7.4.1 Effect ofBioelements andToxic Metals onWoody Plants

For correct experimental design both, the most important and the most difcult is to

prepare young individuals of woody plants from the cuttings to have good devel-

oped shoots as well as roots. Thus, in our earlier paper [343] we studied growth

parameters (including rooting and root growth) of six fast growing trees: Salix vimi-

nalis L., S. alba L., clone 21, S. purpurea L., S. cinerea L. and two poplar species—

Populus euroamericana cv. Gigant and Populus x euroamericana cv. Robusta

cultivated under two different conditions. Stem cuttings ca 18cm long from last

year shoots were cut in March before the beginning of growing season. The cuttings

were grown hydroponically in growth cabinet under the following conditions: air

temperature 25°C, relative air humidity app. 70%, 12h photoperiod with irradiance

100μmolm−2s−1 PAR.Since effect of both different cultivation conditions and the

cadmium were observed, rst part of cuttings (variant A) was directly rooted and

grown in control: 100μmoldm−3 Ca(NO3)2 and in Cd concentration: 10μmoldm−3

Cd(NO3)2 combined with 100μmoldm−3 Ca(NO3)2 treatment for 21 days. Second

part of cuttings (variant B) was rstly rooted in Knop nutrient solution for 10 days.

Then the plants were transferred into 100μmoldm−3 Ca(NO3)2 and after 3 days half

of them was placed into 10μmoldm−3 Cd(NO3)2 for 7 days. In the variant B the total

time of hydroponic cultivation was the same as in the variant A.The solutions

were changed every 3 days to prevent depletion of metals, nutrients, and oxygen.

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Twenty-one-day-old plants were washed in distilled water and used for experimental

evaluation. It was found that the roots responded to Cd treatment more sensitively

than the shoots. Cd treatment suppressed rooting and root growth (length and

biomass production) as well as its development in all tested species. Root system of

S. cinerea, S. alba, and Populus x euroamericana cv. Robusta were more tolerant to

Cd stress than the root system of the other studied species. Shoot growth parameters

of Salix species were signicantly reduced unlike Populus species, which were not

affected by Cd treatment.

Later [344] it was compared some physiological, production, and structural char-

acteristics of Salix alba L. and Populus x euroamericana cv. Robusta under two

variants of cultivation: rooting in Knop nutrient solution prior to Cd treatment and

direct cultivation in Cd. The measurement and equipment used have been described

in detail by Masarovičová and Kráľová [345]. Some production parameters of

S. alba roots (root cumulative length, number, and biomass production) and some

physiological characteristics of S. alba leaves (assimilation pigment content, net

photosynthetic rate, starch content, specic leaf mass) were positively inuenced by

pre-growing in Knop solution. Cd enhanced values of specic leaf mass in both spe-

cies and caused xeromorphic character of leaves—increased stomata density but

reduced stomata sizes. Assimilation pigment and starch contents, net photosynthetic

rate and specic leaf mass were positively inuenced by indirect treatment. Indirect

treatment lowered root Cd uptake in willow, Cd accumulation in cuttings of both

species and Cd accumulation in poplar shoot. Roots and shoots of P. euroamericana

cv. Robusta rooted in Knop nutrient solution were more sensitive to toxic effect of

Cd than plants cultivated directly in Cd treatment. Pre-growing in Knop nutrient

solution lowered root uptake of Cd in S. alba, accumulation of Cd in cuttings

of both species and translocation and accumulation of Cd into the shoots in

P. euroamericana cv. Robusta. Structural changes induced by Cd indicated better

adaptation of roots grown during the whole experimental period in Cd than of roots

formed in Knop solution and then transferred into Cd solution. The analyses of Cd

content in roots, cuttings, and shoots showed that Cd ions were accumulated mainly

in the roots. Barceló and Poschenrieder [346] summarized the main morphological

and structural effects of Cd on roots as follows: decrease of root elongation, root tip

damage, collapsing of root hairs or decrease of their number, decrease of root bio-

mass, increase or decrease of lateral root formation.

Similar results were published by Lunáčková etal. [347], but these authors addi-

tionally found that Cd impact increased root respiration rate of willow and poplar

plants. Higher values of this physiological parameter was caused by the fact that

toxic effect of Cd induced energy cost for increased metal ions uptake into the roots

and for repairing mechanisms as a consequence of metabolism damages.

Nikolič etal. [348] conrmed symptoms of Cd toxicity in Cd-treated hybrid pop-

lar plants (10–100 μmol dm−3 Cd): inhibited growth (plant height and biomass),

decreased root length and chlorosis of the leaves. The decreased photosynthetic

activity of treated plants may be connected with lower values of chlorophyll con-

tent. Gu etal. [349] investigated the effect of Cd2+ (10, 50 and 100μmoldm−3) on

the growth of four poplar cultivars. Root growth was signicantly inhibited at

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100 μmol dm−3. Cd accumulation increased signicantly with increasing Cd

concentration and with time in all organs of the Populus cultivars, whereby Cd was

accumulated mainly in the roots. Jensen etal. [350] studied growth performance

and heavy metal uptake by Salix viminalis in eld and growth chamber trials and

found that under eld conditions toxic metal uptakes were 2–10 times higher than

uptakes under growth chamber conditions. Vandecasteele etal. [351] investigated

the growth and metal uptake of two willow clones (Salix fragilis ‘Belgisch Rood’

and Salix viminalis ‘Aage’) cultivated in a greenhouse pot experiment using six

sediment-derived soils with increasing eld Cd levels (0.9–41.4mgkg−1). Willow

foliar Cd concentrations were strongly correlated with soil Cd concentrations. Both

clones exhibited high accumulation levels of Cd and Zn in the shoots. Celik etal.

[352] evaluated Robinia pseudoacacia L. leaves for biomonitoring of toxic metal

contamination in Denizli city, Turkey. Concentrations of Fe, Zn, Pb, Cu, Mn, and

Cd were determined in washed and unwashed leaves as well as in soils collected

from a wide range of sites with different degrees of metal pollution (industry, urban

roadside, suburban) and from a rural (control) site. All above-mentioned elements

were found to be at high levels in samples collected at industrial sites, except for

lead and copper which were found at high levels in samples collected from urban

roadsides that associated with the road trafc. The strong correlation between the

degree of contamination and concentrations in all plant leaves assessed display that

the leaves of R. pseudoacacia reect the environmental changes accurately.

At present [353] was published study concerning the effect of different Zn con-

centration on the ecophysiological response of four commercial Salix clones

(“1962”, “1968”, “Drago”, and “Levante”) selected for short rotation coppice, and

one natural clone (“Sacco”) obtained from a contaminated area. Physiological

parameters (net photosynthetic rate, Chla uorescence, Chl content, stomata con-

ductance) differed in dependence on the Zn concentration and clone. At the low Zn

concentration (300mgkg−1), the absence of any signicant reductions in parame-

ters investigated indicated an efcient plant homeostasis to maintain the metal con-

tent within phytotoxic limits. Stomatal limitation (observed at 750 and 1500mgkg−1,

which was found in all clones after 3 days of the treatment) might be caused by

indirect effects of Zn on guard cells. Commercial clone “Drago” was more sensitive

to Zn stress (showing inhibition of growth), while “1962” clone showed a down-

regulation of PS II photochemistry following the slowdown in the Calvin-Bensom

cycle. However, the natural Salix clone “Sacco” performed better, compared to the

other clones, due to activation of a photosynthetic compensatory mechanism.

In an other experiment we tested the sensitivity of Salix clone 102 against

CH3HgCl and HgCl2 [354] because mercury compounds represent severe risks,

often exert clastogenic effects in eukaryotes, especially by binding SH groups and

acting as spindle inhibitors, thereby causing c-mitosis and consequently aneuploidy

and/or polyploidy whereby from the aspect of genotoxicity methylmercury deriva-

tives and other ionizable organomercury compounds were found to be more active

in short-term tests than either non-ionizable mercury compounds (e.g., dimethyl-

mercury) or inorganic mercury salts (e.g., mercuric chloride) [355]. Approximately

25cm long stem cuttings of Salix clone 102 were cultivated hydroponically 20 days

7 Essential Elements andToxic Metals inSome Crops, Medicinal Plants, andTrees

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in growth chamber under controlled conditions (mean air temperature 25±0.5°C,

relative air humidity app. 70%; 12 h day and 12 h night; photosynthetic active

irradiance 80 μmol m−2s

−1) in either 100 μmol dm−3 Ca(NO3)2 (control), or

100μmoldm−3 Ca(NO3)2 with HgCl2 or CH3HgCl in concentrations 10.0, 1.0, and

0.1μmoldm−3 (pH=5.5) and then the length and dry mass of roots and shoots were

determined. Despite the presence of Ca(NO3)2 applied to secure better rooting

which is known to reduce toxic effects, in the presence of Hg compounds the cumu-

lative root length as well as root dry mass were more strongly affected than the

cumulative shoot length and shoot dry mass. At the concentration 10.0 and

1.0μmoldm−3 the toxic effect caused by CH3HgCl was pronouncedly higher than

that of HgCl2 applied in the same concentration: in the presence of organomercurial

compound the growth of roots was practically completely inhibited. For example,

related to the control, the inhibition of individual characteristics due to treatment

with 1μmoldm−3 CH3HgCl and HgCl2 were as follows: 97.5% and 29.4% (cumula-

tive root length), 62.5% and 21.7% (cumulative shoot length), 96.4% and 62.2%

and 36% (shoot dry mass). On the other hand, inhibition of root dry mass by

1μmoldm−3 CH3HgCl was 96.4% but no inhibition was observed due to treatment

with 1μmoldm−3 HgCl2. This is in agreement with previously described results that

the toxic effects of organomercurials are 1–2 orders higher than those of inorganic

Hg [356, 357]. It was reported that phenylmercuric acetate inhibits both Hill activity

and photophosphorylation [358] and Girault etal. [359] found that CH3Hg(II) interac-

tions with membrane phospholipids are electrostatic in nature and the phosphate moi-

ety is proposed as a potential binding site. The IC50 values related to PET inhibition in

spinach chloroplasts by some organomercurials were estimated as follows: 468 μmol

dm−3 for phenylmercuric borate, 657 μmol dm−3 for phenylmercuric acetate,

942μmoldm−3 for phenylmercuric citrate and 627μmoldm−3 for methylmercuric

chloride and using EPR spectroscopy as probable sites of action of organomercury

compounds in photosynthetic apparatus ferredoxin on the acceptor side of PS I and

the quinone electron acceptors QA or QB on the reducing side of PS II were suggested

[360]. According to Matorin etal. [361] increased toxic effect of methylmercury on

Chlorella vulgaris resulted from the decreased capacity of PS II for reparation and

damage of Chlamydomonas reinhardtii algal cells on the donor side of PS II and

impairment of the electron transfer from QA to QB after MeHg+ treatment was reported

by Kukarskikh etal. [362], while decreased photochemical activity of the PS II reac-

tion centers of diatom Thalassiosira weissogii after application of MeHg+ was

observed by Antal etal. [363]. It could be noted that considering the adverse effects of

organomercury compounds on the environment phytodetoxication of hazardous

organomercurials by genetically engineered plants was proposed (e.g., [364–366]).

7.4.2 Woody Trees asMedicinal Plants

Species of the genus Karwinskia from family Rhamnaceae (common names tulli-

dora or coyotillo) are medicinal woody plants (shrubs and trees) growing in the

subtropical and tropical areas of Mexico [367]. All parts of the plant produce

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secondary metabolites (toxins) characterized as anthracenones [368]. One of them

(T-514) was isolated and later named peroxisomicine A1 (PXM) with antitumor

effect on mammalian tumor cells [369]. Some isolated hydroxyanthracenones

belonging to the genus Karwinskia were found to possess also antimicrobial activ-

ity, particularly against Streptococcus pyogenes, Candida albicans, C. boidinii,

C. glabrata and Cryptococcus neoformans with minimal inhibitory concentrations

ranging between 16 and 2μg/mL [370]. Recently, Rojas-Flores etal. [371] isolated

from the dried fruits of Karwinskia parvifolia ve new “dimeric” napthopyranones,

karwinaphthopyranones, possessing a methoxy group at C-5′, some of which

possessed antiproliferative activities in representative human cancer cell lines, with

half-maximal growth inhibitory concentrations in the micromolar range. The capa-

bility of scavenging •OH radicals by phenolic metabolites of Karwinskia humbold-

tiana leaves was investigated, as well and it was found that for the antioxidant

effects are responsible metabolite such as (+)-epicatechin and avonol derivatives

quercetin, quercetin 3-O-glucoside (isoquercitrin), quercetin- 3- O-galactoside

(hyperosid), quercetin-3-O-arabinoside, quercetin 3-O-rutinoside (rutin), kaemp-

ferol 3-O-arabinoside, and kaempferol 3-O-rutinoside [372].

Saavedra etal. [373] found in yeast that PXM shows specic activity on peroxi-

somes. However, concentration of this substance in individual parts of the plant is

very variable and depends on environmental conditions, especially on drought, dif-

ference in the temperature and air humidity between the day and night or summer

and winter [374] and time of plant collection, too [375]. Since no information was

available as to how environmental variations affect CO2 exchange (photosynthesis

and respiration) as well as production of organic substances (important for synthesis

of secondary metabolites, including PXM) we estimated these parameters in the

leaves of Karwinskia parvifolia Zucc. grown under two temperature regimes:

day/night temperature of 35/20°C (summer temperature regime, SR) and 20/5°C

(winter temperature regime, WR). These temperature regimes were similar to aver-

age air temperature in SR and WR in natural areas of the studied species—Nuevo

León in Mexico. The other growth conditions were identical for both the SR and

WR: photoperiod 16h, irradiance 200μmolm−2s−1, day/night relative air humidity

85/50% (in detail see [376]). We estimated net photosynthetic rate (PN) and chemi-

cal composition (starch, reducing sugars, nitrogen, carbon, and hydrogen) of the

Karwinskia parvifolia leaves. On the basis of the CO2 curves of photosynthesis, the

values of PN increased in response to increasing CO2 concentration and irradiance.

Chemical composition of the leaves was different: the level of starch and the content

of reducing sugars were higher in the plants cultivated under summer temperature

regime (35/20°C), however content of the N and C was higher in the plants culti-

vated under winter temperature regime (20/5°C). Our study also provided the val-

ues of the other characteristics: dark respiration rate, Chl concentration, stomata

characteristics, and specic leaf area. It was conrmed that higher temperature in

SR signicantly inuences stomata apparatus properties such as stomata density,

stomata length, and width. However, there were no statistically signicant differ-

ences in dark respiration rate, Chl content per leaf area, and in specic leaf area

between plants cultivated in summer temperature regime and winter temperature

regime [376].

7 Essential Elements andToxic Metals inSome Crops, Medicinal Plants, andTrees

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According to Masarovičová and Lux [377], quantitative data on physiological

characteristics (mainly CO2 exchange) and production parameters (growth and bio-

mass formation) of the plants (e.g., Karwinskia species) can help to optimize con-

ditions for both their cultivation and the production of secondary metabolites

(including anthracenones) and also conrm the relationship between photosynthesis

(primary metabolism) and the synthesis of PXM (secondary metabolites). Since

nitrogen is often a major limiting factor for photosynthetic processes, Masarovičová

et al. [378] investigated photosynthetic characteristics, biomass partitioning and

PXM production of Karwinskia parvifolia Zucc. as well as Karwinskia humboldti-

ana Rose grown under controlled conditions at two different N supply. Two-year-

old plants in pot were watered by a nutrient supply system twice daily with 455mL

of modied nutrient solution containing high nitrate supply (HN) 2.0mmol dm−3

KNO3 and 1.5mmol dm−3 Ca(NO3)2 or low nitrate supply (LN) 198.8μmol dm−3

KNO3 and 150.8μmol dm−3 Ca(NO3)2, and for both nutrient solution, 270 μmol

dm−3 MgSO4, 190 μmol dm−3 KH2PO4, 41 μmol dm−3 Fe-EDTA, 20μmol dm−3

H3BO3, 2μmoldm−3 MnSO4, 0.9μmoldm−3 ZnSO4, 0.3μmoldm−3 Na2MoO4 and

0.2μmoldm−3 CuSO4. Pots were placed for 2 months (15 May–15 July) in a growth

cabinet (1600 SP, Weiss Bioclim, The Netherlands) with the following conditions:

14h day length, irradiance at mean plant height 400μmolm−2s−1, 25±0.5°C day

and night temperature, 80± 5% relative air humidity day and night. The above-

mentioned authors found signicant differences in growth, CO2 exchange (photo-

synthesis and respiration), Chl and nitrogen concentration between plants grown at

HN and LN.At HN, the plants of both species grew faster than those cultivated at

LN.Rates of photosynthesis, leaf respiration and root respiration, quantum yield,

the concentration of chlorophylls, specic leaf area, leaf mass ratio and PXM con-

centration were higher in plants grown at HN. K. parvifolia responded more strongly

to the nitrogen treatment than K. humboldtiana, in terms of growth, as well as with

respect to photosynthesis and PXM concentration. Based on found results it could

be concluded that it seems useful to modify biomass production through nutrients

(especially nitrogen) and in this manner also to inuence production of pharmaceu-

tically effective substances (mainly PXM) that occur in Karwinskia species. In K.

parvifolia the effect of nitrogen on fruit yield was also studied. It was found that

nitrogen-treated plants possessed higher dry mass of the fruit than control plants [379].

Zelko and Lux [380] investigated the effect of Cd(NO3)2 on growth, structure,

and development of roots of Karwinskia humboldtiana. Cadmium signicantly

reduced the growth of primary and lateral roots of K. humboldtiana plants cultivated

in hydroponics even at the lowest concentration of Cd(NO3)2 (1 μmol dm−3).

Sections of root tips revealed differences between control and cadmium-treated

roots, especially in dimensions and vacuolization of meristematic and cortical cells

and in sloughing off the root cap border cells. The authors found that development

of root endodermis was also affected by Cd and formation of the apoplastic bar-

rier—Casparian bands—in endodermis started closer to the root apex in Cd-treated

plants in comparison with control plants. Moreover, the appearance of Casparian

band can precede xylem elements formation and the second stage of endodermal

development—deposition of suberin lamellae was accelerated after Cd treatment.

E. Ma sarovičová and K. Kráľová

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In this context we should stress that secondary metabolites play important role in

defensive mechanism of plants (especially medicinal plants) against toxic metals,

because some of them can bind metal ions. As mentioned above, this additive defen-

sive mechanism was also conrmed by Kráľová and Masarovičová [241] for

cadmium and Hypericum perforatum L. and secondary metabolites hypericin, pseu-

dohypericin (naphthodianthrone derivates) and quercetin that are produced by this

medicinal plant. Defensive mechanisms of medicinal plants are thus connected with

strategies of these plants to tolerate negative effects of toxic metals or bioelements

in their non-physiological concentrations.

7.5 Phytoremediation, Phytofortication,

andNanoagrochemicals

7.5.1 Phytoremediation

One reason for interest in plant metal interaction has been the recent attention on the

use of plants either to remediate toxic metal-contaminated soils or increasing the

bioavailable concentrations of essential nutrients in edible portions of food crops

through agronomic intervention or genetic selection. In addition, since plants are

known to interact with different metals, they have been used for the “green biosyn-

thesis” of metal nanoparticles. Such bioinspired methods are dependable, environ-

mentally friendly and benign. In general, phytoremediation, phytofortication, and

metal nanoparticles biosynthesis are thus natural green biotechnologies with using

crops, medicinal plants, as well as trees.

Phytoremediation is environment-friendly and cost-effective green technology

for the removing of toxic substances from the environment using convenient plant

species. In general, several types of phytoremediation technologies are available for

clean-up of soils and water contaminated by organic or inorganic pollutants. The

most important of them are: phytoextraction (reduction of soil metal concentration

by cultivating plants with a high capacity for metal accumulation in the shoots),

rhizoltration (adsorption or precipitation of metals in the roots or absorption by the

roots of metal-tolerant aquatic plants), phytostabilization (immobilization of metals

in soils), rhizodegradation (decomposition of organic pollutants by rhizosphere

microorganisms), hydraulic control (absorption of large amounts of water by fast

growing plants and thus prevent expansion of contaminants into adjacent uncon-

taminated areas), and phytoresaturation (re-vegetation of barren area by fast grown

plants that cover soils and thus prevent the spreading of pollutants into environ-

ment) (e.g., [381, 382]).

The most frequently used phytoremediation technology is phytoextraction

involving the cultivation of metal-tolerant plants that concentrate soil contaminants

in their shoots. At the end of the growth period, plant biomass is harvested, dried

or incinerated, and the contaminant-enriched material is deposited in a special

dump, added into a smelter or the metals can be extracted from the ash [383].

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Besides metal hyperaccumulators metal-tolerant species (e.g., Hordeum vulgare,

Triticum aestivum, B. napus, B. juncea, Helianthus annuus) can accumulate high

concentration of some toxic metals in the shoots. Moreover, the fast growing (high

biomass producing) plant species, such as Salix spp. and Populus spp., can also be

used. These trees have lower shoot metal-bioaccumulating capacity, but their

efcient clean-up of contaminated substrates is connected with their high biomass

production. Within the Brassica genus, there also exist some other species which

show the tendency to accumulate high metal concentrations, and which can be char-

acterized as metal accumulators. Some of these species grow fast and produce a

high biomass. Besides already mentioned rapeseed (B. napus) and Indian mustard

(B. juncea) it is also eld mustard (B. rapa) [384].

7.5.2 Phytofortication

Phytofortication is the fortication (enrichment) of plants with essential nutrients,

vitamins, and metabolites during their growth and development to be more available

for human or animal consumption. As many of the metals that can be hyperaccumu-

lated are also essential nutrients, it is easy to see that food fortication and phytore-

mediation are two sides of the same coin [385]. Since plants are at the beginning of

food chain, improving the nutrients uptake from soil and enhancing their movement

and bioavailability in the edible parts of crops or in the feed will provide benets for

humans as well as animals. Phytofortication provides a feasible means of reaching

malnourished populations mainly in relatively remote rural areas, where there is

limited access to commercially marketed fortied foods. There are two main chal-

lenges ahead: (1) to develop crops that have an increased content of essential ele-

ments in the edible parts of plant but that at the same time exclude toxic elements

that exhibit similar chemical properties; (2) to avoid sequestration of bioelements in

the inedible parts of plants (e.g., in the roots). A breeding approach to produce nutri-

tionally improved food crops relies on genetic diversity in natural populations that

can be crossbred to introduce traits/genes from one variety or line into a new genetic

background [386]. Thus phytofortication could be divided into agronomic and

genetic phytofortication. The rst approach uses soil and spray fertilizers enriched

by individual essential elements (e.g., Fe, Zn, and Se). Agronomic phytofortica-

tion has been successfully adopted in Finland as a cost-effective method for enrich-

ment of crops by Se. It should be stressed that increasing Se content in wheat is a

food systems strategy that could increase the Se intake of whole populations.

Genetic phytofortication presents the possibility to enrich crops by selecting or

breeding crop varieties, which enhance bioelement accumulation (in detail see

[387]). According to Genc etal. [388] strategy utilizing plant breeding for higher

nutrient concentration together with agronomic biofortication (e.g., applying sel-

enate to cereal crops by spraying or adding to fertilizer) is likely to be the most

effective way to improve the nutrition of populations. Because selenium as an

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essential micronutrient for humans and animals is decient in at least a milliard

people worldwide, Se-accumulating plants are a source of genetic material that can

be used to alter selenium metabolism and tolerance to help develop food crops that

have enhanced levels of anticarcinogenic Se compounds [259]. Application of sel-

enate on soil could be used by food companies as a cost-effective method to produce

high-Se wheat products that contain most Se in the desirable selenomethionine

form. Increasing Se content in wheat is a food systems strategy that could increase

the Se intake of whole human population [389].

Based on found results it should be emphasized the interest in the potential

exploiting of hyperaccumulators of bioelements as a rich genetic resource to develop

engineered plants with enhanced nutritional value for improving public health [386].

Importance of phytofortication for the humans makes this an exciting line of future

research in the eld of hyperaccumulation of essential elements.

7.5.3 Nanoagrochemicals

In recent years many efforts were done to minimize negative effect of fertilizers

(especially synthetic fertilizers) on soil, water, and air by design of new improved

fertilizers. Nanotechnology opened novel applications in different elds of both

biotechnology and agriculture. Studies showed that the use of nanofertilizers causes

an increase in nutrients use efciency, reduces soil toxicity, minimizes the potential

negative effects associated with over dosage and reduces the frequency of the appli-

cation. Nanotechnology thus has a high potential for achieving sustainable agricul-

ture, mainly in developing countries [390].

Studies of nanomaterials in plant systems have demonstrated that dependent on

dose, nanotechnology can be leveraged in developing novel fertilizers to enhance

agricultural productivity. For instance, in wheat, ZnO nanoparticles caused the stim-

ulation of lateral roots and changed the root architecture, which could contribute in

the overall uptake of nutrients. In bean, a low dose (100mgkg−1) of ZnO nanopar-

ticles stimulated shoot growth, similar to ndings in chickpea and green pea. The

same ascertainment for further plant species and other metal nanoparticles were

also published by Masarovičová etal. [391]. However, these authors described also

negative effects of metal nanoparticles on plants depending on their concentration

as well as on experimental conditions. Dimkpa [392] reminded that the delivery of

mineral nutrients in nanoform is predicated on a variety of benecial features,

including timing of nutrients release, sustained release of nutrients, synchronization

or targeted environmental response, and directed nutrient delivery. Mineral nutri-

ents can be also encapsulated in nanopolymers that also could either be directly

absorbed by the plant, releasing the cognate nutrient in plants, or be engineered to

timely dissolve in the rhizosphere, releasing the encapsulated nutrients according to

plant’s need.

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Originally nanoagrochemicals (nanofertilizers, nanopesticides, plant growth-

stimulating nanosystems) were designed to increase solubility, enhance bioavail-

ability, targeted delivery of substances in both, soil and plants, controlled release

and/or protection against degradation resulting in the reduced amount of applied

active ingredients. These substances reduce the amount of applied active ingredients

by means of their enhanced bioavailability and protection against degradation and

thus decrease the dose-dependent toxicity for non-target organisms (in detail see

[393, 394]). These authors emphasized the further rapidly growing application of

nanotechnology such as detection of pathogens and contaminants by using nanosen-

sors and indicators, food packaging, food security, encapsulation of nutrients, and

development of new functional products. It was found that nanoscale food packing

materials may extend food life, may improve food safety, may alert consumers that

food is contaminated or spoiled and even may release preservatives to extend the

life of the food in the package. However, they also stressed that increased attention

must be devoted to the impact of risk factors associated with usage of nano-size

materials on the environment and possible adverse effects on non-target organisms,

especially humans. Similarly, Dimkpa [392] stressed that despite potential benets,

the application of nanotechnology in plant fertilization could come with risks for

environment: non-target plants, plant-benecial soil microbes and other life forms

could be affected if nanomaterials are misused.

7.6 Concluding Remarks

Global climatological changes, including “greenhouse effect”, induced water

deciency in the environment and thus “blue revolution” has been started after well-

known “green revolution” which appeared in the beginning of 60 years of the last

century. In comparison with “green revolution”, when excessive application of dif-

ferent fertilizers and pesticides was preferred, “blue revolution” will change the

approach to water conservation and management. Mainstream plant production thus

will be in the sense “blue revolution—more crops for every drop”. Water, as a criti-

cal issue, needs to move to the center stage of policy-making in the whole society,

as this is a time bomb ready to go off any time. Recently appeared a new approach

in transforming life sciences to technologies through the “converging Technologies”

that represents combination of nano-, bio-, information- and cognition technologies

knows as “NBIC technologies”. These “converging technologies” allow for totally

new combinations of biological and non-biological materials that will open new

possibilities to interfere with living organisms. However, intensive improvement of

biological sciences accompanied with many novel technologies promotes new sub-

stantial issues concerning ethics. For that reason, both scientists and politicians will

have to accept fundamental bioethical principles to ensure the sustainable develop-

ment of human society as well as essential protection of the environment and nature.

E. Ma sarovičová and K. Kráľová

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235

Acknowledgements This contribution was nancially supported by the Grant Agency VEGA,

grant No. 1/0218/14 and code ITMS 26240120004, funded by the ERDF.

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7 Essential Elements andToxic Metals inSome Crops, Medicinal Plants, andTrees

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Part III

Phytoremediation of Aquatic Ecosystems

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259© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_8

Chapter 8

Phytoremediation Using Aquatic Macrophytes

AmtulBariTabindaAkhtar, AbdullahYasar, RabiaAli, andRabiaIrfan

Abstract Phytoremediation is a plant-based technology that is also called green

technology. After the discovery of hyperaccumulating plants, this technology gained

increasing attention. These hyperaccumulating plants are having the ability to

uptake, store, transport, and focus on large quantity of specic poisonous elements

in their body parts such as aboveground parts and harvestable parts. Phytoremediation

has a number of processes that are phytoextraction, rhizoltration, phytovolatiliza-

tion, etc. Both type of plants (terrestrial and aquatic) have been tested, and these are

having characteristics to treat polluted soils and waters. A number of aquatic mac-

rophytes have been found that are used for the removal of toxic contaminants such

as arsenic, zinc, cadmium, copper, lead, chromium, and mercury. Some of these

aquatic macrophytes are water hyacinth, water spinach, water ferns, hydrilla, and

watercress. Metal uptake ability and mechanisms of many other macrophytes have been

studied or investigated. Many of these studies proved that aquatic macrophytes have

potential for phytoremediation. Phytoremediation is cost-effective, environment-

friendly, and has gained rising appreciation. More than 400 plant species have been

known that are having the ability to remediate soil and water. This chapter provides

a look into new developments in research and practical applications of phytoreme-

diation by using aquatic macrophytes.

Keywords Phytoremediation • Aquatic macrophytes • Hyperaccumulators • Heavy

metals • Phytoextraction • Phytostabilization • Rhizoltration • Phytovolatilization

and phytotransformation

A.B.T. Akhtar (*) • A. Yasar • R. Ali • R. Irfan

Sustainable Development Study Centre, GC University,

Katchery Road, Lahore 54000, Pakistan

e-mail: amtulbaritabinda@gcu.edu.pk; yasar.abdullah@gmail.com; rabia_gcu@hotmail.com;

rabia.irfan8@gmail.com

guarino@unisannio.it

260

8.1 Phytoremediation

In the past 30years, the quantity and density of toxic waste efuents have increased.

This is due to the rapid growth [1], uncontrolled disposal of waste, accidental spill-

age, sludge application to soils, heavy metals, and higher complexity of chemical

industries. All these factors contribute towards contamination of ecosystem [2].

Freshwater resources are changed by the human activities. Due to high deterioration

of water quality, this resource is deteriorating in many areas of the world [1].

Ultraltration, chemical precipitation, chemical oxidation, chemical reduction, reverse

osmosis, electrochemical treatment, coagulation, and occulation are traditional

technologies used for the treatment of pollutants, especially heavy metals [3, 4]. All

these technologies have some limitations and benets [5].

The word phytoremediation is made of two words. The rst word is Greek that is

phyto which means plant, and the second word is Latin that is remedian. It means to

eliminate an evil and restoring balance, in other words remediation [6]. This term

has been used since 1991. There are a number of technologies in phytoremediation

in which plants and some soil microbes are used. These plants and microbes are

used to decrease the quantity, movement, and toxicity of pollutants in different

mediums such as soil, groundwater, and other polluted media [7, 8]. In other words,

phytoremediation uses the plants and their natural, biological, physical [9], and

chemical activities and processes to eliminate, immobilize, and detoxify environ-

mental pollutants in a growth medium which could either be soil, sediments, and

water [10]. Phytoremediation is used to treat a variety of contaminants or pollutants

in small level eld and in laboratory [6, 11]. These contaminants or pollutants are

heavy metals, radionuclides, petroleum hydrocarbons, polychlorinated biphenyls,

chlorinated solvents, polyaromatic hydrocarbons, organophosphate insecticides,

explosives, and surfactants [12, 13] in air, water, and soils with the help of plants

(natural and genetically engineered) [6]. Phytoremediation is a novel, efcient, in

situ applicable, nonpolluting, and solar-driven remediation strategy that is effective

in its cost and is environment or eco-friendly [9, 10, 14, 15].

8.2 Mechanism ofPhytoremediation (Fig. 8.1)

Plants are having power to uptake, destroy, alter, and restore the pollutants. Plants

are having the ability to remediate contaminated sites [16]. Plants remediate the

contaminants and contaminated sites through several methods [6]. Generally,

these plants hold the pollutants without upsetting topsoil, as a result conserving its

value and richness. By adding organic matter in the soil, these plants improve soil

fertility.

Some of the plants break the dangerous contaminants from the ground and mean-

while their roots take water and nutrients from the polluted soil, remains, and under-

ground water. By adopting natural processes such as by storing the contaminants in

A.B.T. Akhtar et al.

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261

the roots, stems, and leaves, plants can purify the pollutants. As far as their roots can

get in touch converting harmful chemicals into vapors, which are released into the

air. Plant breaks sorbed pollutants to less harmful chemicals into their root zone.

Green plants have a huge capacity to take contaminants from the surroundings and

achieve their detoxication by different processes. If the contaminated plants are

left on spot for deterioration, the pollutants will be back to the soil. For total elimi-

nation of pollutants from an area, the plants must be cut and disposed of somewhere

else in a less polluting way. The time period, required number, and type of species

depend on the site characteristics and mostly the contaminant type. The most sig-

nicant factors that have to be taken into reection to a site where phytoremediation

is used are: kind of contaminants, plant species, levels of contamination [16].

8.3 Plants Used inPhytoremediation

Some of the important plant families which hyperaccumulate the pollutants are

Brassicaceae, Euphorbiaceae, Lamiaceae, Scrophulariaceae, etc. Indian mustard

(Brassicaceae juncea) is a plant having large biomass. The rapid growth of plants is

having the ability to accumulate nickle, lead, and cadmium in its shoots [7]. Some

of the plants such as corn, sorghum, and sunower are said to be good because their

growth rate is fast and produce large biomass [17, 18]. Alfalfa (Medicago sativa) is

tree roots take

in water and

contamination

from the ground

vapors

water enters tree

where contamination

is cleaned up

contaminated

soil

clean

soil

contaminated groundwater

water

table

Fig. 8.1 Mechanism of phytoremediation, Source: USEPA 2012 [7]

8 Phytoremediation Using Aquatic Macrophytes

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an impending source of biological materials for the exclusion and revival of heavy

metal ions [19]. Poplar trees, Forage kochia, Kentucky bluegrass, Scirpus species,

Coontail, American pondweed, and Arrowhead are some others. Some of the aquatic

plants are water hyacinth and duckweed [20]. Currently, a fern Pteris vitatta is

reported to store arsenic [21].

8.4 Types ofPhytoremediation

Phytoextraction (or phytoaccumulation), phytostabilization, rhizoltration, phytovola-

tilization, and phytotransformation are the main phytoremediation techniques [22].

8.4.1 Phytoextraction

All over the world, phytoextraction has been gaining popularity for the last 20years.

Phytoaccumulation, phytosequestration, and phytoabsorption are some of the other

names of phytoextraction. Phytoextraction involves plant roots that absorb pollut-

ants from soil or water and accumulate them in aboveground biomass such as plant

shoots [23]. Brassica juncea and Thlaspi caerulescens are commonly used plants

for phytoextraction [24].

8.4.1.1 Mechanism (Fig. 8.2)

The plants take up pollutants with the help of roots and accumulate them in the root

structures or then carry these into other or upper parts of the plants. A plant may

carry on this process until it is removed. After removal, small amount of the pollut-

ants stays behind in the soil, so the growth and removal cycle must regularly be

repeated through a number of crops to get a considerable cleaning. After this proce-

dure, the puried soil can sustain other plants. The essential time for removal is

dependent on the type and amount of metal pollution, the length of the growing

period, and the effectiveness of metal exclusion by plants [24].

8.4.1.2 Advantages andDisadvantages

This technology is appropriate for treating large areas of land, and the contamina-

tion level of these areas is low to moderate. Plant growth is not maintained in highly

contaminated soils. Soil metals should also be bioavailable and subjected to assimi-

late by plant roots. The two properties such as high accumulation of metals and high

production of biomass result in maximum removal of metals [24].

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8.4.2 Phytostabilization

Phytostabilization also called phytoimmobilization or phytorestoration is the use of

specic plants for stabilization of pollutants in polluted soils [26]. It is a remedia-

tion method that is based on plants. It stabilizes contaminants and prevents contact

through wind and water erosion. It provides hydraulic control, which decreases the

upright movement of pollutants. It decreases the mobility of pollutants physically or

chemically by root absorption [27, 28]. This method is used to slow down the move-

ment and availability of contaminants in the environment resultantly preventing

relocation and entry of these contaminants into groundwater and food chain [29].

The primary focus of phytostabilization is storage of pollutants in soil. In this tech-

nique, pollutants are not stored in plant tissues and become less available biologi-

cally so as a result there will be less exposure to livestock, wildlife, and humans.

Agrostis tenuis and Festuca rubra are commercially accessible for the treatment of

lead-, zinc-, and copper-contaminated soils.

8.4.2.1 Mechanism

This process decreases the movement of the pollutants, their relocation to the under-

ground water. This method can also be used to restore plant life at sites where natu-

ral plantation fails to stay alive due to high concentration of metals in supercial

1. Uptake of

bioavailable metals

2. Transfer from roots

to shoots

3. Accumulation in

shoots

4. Harvest

of shoots

Biomass can be:

- reduced (compaction,

composting, thermal

treatments) for disposal:

-used to metals recovery

(phytomining)

Fig. 8.2 Mechanism of phytoextraction, Source: Paulo etal. [25]

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soils or physical instability to surface materials. For restoring the vegetation of

polluted sites, metal-tolerant species are used. Due to which there will be less

chances of transfer of pollutants in soil and water through wind erosion and uncov-

ered soil surface. This technique is used for the treatment of lead, zinc, arsenic,

cadmium, copper, and chromium [30].

8.4.2.2 Advantages andDisadvantages

Phytostabilization has many advantages over other remediation methods because it

is cheap, environment-friendly, easy to apply or use, and add visual value [28]. It is

most successful at those places which are having smooth soils with organic sub-

stances but is appropriate for treatment of large variety of places where large areas

of surface pollution exist. Phytostabilization is not possible at highly polluted places

because plant growth and life is not possible there [27].

8.4.3 Rhizoltration

Rhizoltration is the method or technique in which plant root stake in toxic metals

from efuents accumulate and precipitate them. Water hyacinth (Eichhornia

crassipes), duckweed (Lemna minor), and pennywort (Hydrocotyle umbellata) are

the plants used for rhizoltration [31].

8.4.3.1 Mechanism (Fig. 8.3)

Plant roots excrete specic chemicals in the root environment that creates biogeo-

chemical conditions that resultantly precipitate the contaminants onto the roots or in

the water body. When the plant roots become ooded with the pollutants, then only

the roots or whole plants are cut off for dumping [31].

8.4.3.2 Advantages andDisadvantages

Rhizoltration is a cheaper technique that is applied for the treatment of consider-

able amount of metals (chromium, lead, and zinc) in surface water and groundwater.

But the purpose of this technology is more difcult to achieve and inclined to failure

than other methods of related cost. It requires a trained and skilled manpower for

making and preservation of hydroponically grown systems. The services and spe-

cialized equipment requirement can increase operating costs [33].

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8.4.4 Phytovolatilization

Phytovolatilization is the method in which plants take pollutants from its surround-

ings and then transpire the contaminants. The transpiration is the process in which

water moves in a plant from bottom root part to the upper part and then evaporated

in the leaf pores. Cultivated tobacco (Nicotiana tabacum), Swamp lily (Crinum

americanum), Spring wheat (Triticum aestivum), Mouse-ear cress (Arabidopsis

thaliana), Water hyssop (Bacopa monnieri), and White clover (Trifolium repens)

are commonly used plants for phytovolatilization [34–36].

8.4.4.1 Mechanism (Fig. 8.4)

Phytovolatilization is the process in which a pollutant is taken up by a plant and

undergoes transpiration process. Meanwhile, a pollutant or a modied form of a

pollutant is released by the plant into the atmosphere. Another process called

phytodegradation is a linked phytoremediation process that can occur along with

phytovolatilization.

8.4.4.2 Advantages andDisadvantages

Sites that use this technique of phytovolatilization may not need much supervision

after the plantation of these plants. This remediation method has the additional advan-

tages, for example, these sites are less disturbed, very less chances of erosion, and

plants used in this process need not to be disposed of. Phytovolatilization would not be

Fig. 8.3 Mechanism of rhizoltration, Source: IGECE [32]

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suitable for places which are near the highly populated sites or at some other places

with distinctive weather patterns that endorse the quick settlement of unstable com-

pounds [37]. Opposite to other treatment techniques, if pollutants have been removed

through phytovolatilization, there will be less control over their relocation to other

places. The use of this process is restricted because the pollutants completely remove

but only changes its place such as transferring from one segment of the environment

(soil) to another segment (atmosphere) and afterwards redeposited at some other place.

Phytovolatilization is very contentious among all techniques of phytoremediation [38].

8.4.5 Phytotransformation

Phytotransformation is the uptake of contaminants from soil and water that are

either organic or nutrient contaminants [39].

8.4.5.1 Advantages andDisadvantages

Phytotransformation is applicable on petrochemical sites and other storage areas,

e.g., agricultural-based chemicals, ammunition wastes, chlorinated solvents, fuel

leakages, landll leachates [39].

Fig. 8.4 Mechanism of phytovolatilization, Source: IGECE [32]

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8.5 Classication ofAquatic Macrophytes

Aquatic macrophytes are also called aquatic plants or water plants. These are usually

presented in seven plant divisions or classes which are Cyanobacteria, Bryophyta,

Pteridophyta, Chlorophyta, Rhodophyta, Xanthophyta, and Spermatophyta.

Depending upon the growth pattern, the aquatic plants are categorized into four

major groups such as:

• Group I is also called emergent macrophytes. These are the plants which are hav-

ing roots in soil and the plant growth is rising to considerable heights above the

water. Examples of these plants are Phragmites australis, Typha latifolia, etc.

• Group II commonly called oating macrophytes. The plants of this group are

mostly located on waterlogged sediments and are found at water depths of about

0.5–3.0m and it includes angiosperm plants. Examples of these are Potamogeton

pectinatus, etc.

• Group III consists of submerged macrophytes or plants. These are grown mostly

below the water surface. It includes mosses, angiosperms, charophytes, and

pteridophytes.

• Group IV includes free-oating plants. These are nonrooted to rock layer plants.

This group is highly diversied in its habitats and characteristics.

Aquatic macrophytes or aquatic plants in comparison with terrestrial plants are

more appropriate for the treatment of wastewater. Aquatic macrophytes are having

many distinct characteristics such as growth rate, large production of plant body,

pollutants uptake ability, and better distillation effects due to direct contact with

polluted water. These macrophytes perform important functions at structural and

functional levels of aquatic ecosystems. Some of the structural level functions are

changes in water movement, shelter to sh and other invertebrates of aquatic habitat,

and a good food source. At functional level, these macrophytes alter the quality of

water by balancing oxygen, nutrient cycle, and heavy metals accumulation [40].

Aquatic macrophytes have the ability to accumulate heavy metals. This characteris-

tic makes them attractive for research particularly for the treatment of industrial and

household waste water [9, 41, 42]. The potential of aquatic plants to phytoremedia-

tion is mainly dependent on: the acceptance of plant species and difference in

uptake or storage potential for the same heavy metal. In phytoremediation, some

environmental factors should be maintained like chemical species, initial concen-

tration of the metal, interface of different heavy metals, temperature, pH, redox

potential, and salinity. Another phytotechnology is using the oating macrophytes

for treating water in which different types of duck weed and water hyacinths have

been used. Root zone plants may also be used for the treatment of small volumes of

sewage water.

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8.5.1 Examples ofAquatic Macrophytes fortheRemoval

ofContaminants

8.5.1.1 Eichhornia crassipis (Water Hycianth) (Fig. 8.5)

Water hyacinth is a native tropical and subtropical aquatic plant. Among seven spe-

cies of water hyacinth, Eichhornia crassipis is mostly common, grows fast and

highly tolerant to pollution [43], and is used in treatment of wastewater due to its

high absorption capacity of heavy metals [44]. Arsenic removal capacity of

Eichhornia crassipes was larger than the other aquatic macrophytes due to its high

production of biomass and suitable climatic conditions [41]. Plant shows high

growth rate and huge vegetative reproduction [45], and it is the most troublesome

weed, found in a large amount throughout the year and is very efcient in absorption

of lead, zinc, manganese, cadmium, copper, and nickel by root or shoot system [26].

Water hyacinth can be the best option for the elimination of heavy metals [46].

Eichhornia crassipes has been used in treatment of wastewater and for improving

the quality of water. It does so by reducing the levels of organic and inorganic nutri-

ents [26]. Irfan [47] performed 1-month treatment of water at four different concen-

trations of chromium (Cr) and copper (Cu) by using Eichhornia crassipes. This

plant successfully removed these heavy metals without any sign of being affected

by it. Eichhornia crassipes removed 80.94% chromium (Cr) and 95.5% copper (Cu)

during 1-month experiment.

8.5.1.2 Azolla caroliniana (Mosquito Fern)

Azolla is highly efcient to accumulate toxic heavy metals and can remove pollut-

ants from wastewater [42].

Fig. 8.5 Eichhornia

crassipis

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8.5.1.3 Brassica juncea (Mustard Green) (Fig. 8.6)

Brassica juncea, Helianthus annuus, and Zea mays were mostly studied in phytore-

mediation from 1995 until 2009. Many researches showed that Brassica juncea is

very efcient for soil remediation and accumulates cadmium. It shows high removal

efciency of zinc because of more biomass production.

Removal efciency of zinc, copper, and lead was compared among three species

of Brassica that is Brassica oleracea, Brassica carinata, Brassica juneca, and

Brassica oleracea showed high removal of zinc and copper in its shoots than the oth-

ers zinc and lead accumulation was reported almost constant in all three species.

8.5.1.4 Pistia stratiotes (Water Lettuce)

Pistia stratiotes is an aquatic macrophyte that rapidly grows with large biomass.

It shows high removal efciency for removal of heavy metals due to extensive root

system. Dead Pistia stratiotes is found very efcient and low cost alternative for the

removal of diluted heavy metals like lead and cadmium for the treatment of indus-

trial efuents [26, 48]. Pistia stratoites removed 77.3% chromium (Cr) and 91.29%

copper (Cu) at four different concentrations of these heavy metals during 1-month

treatment of water [47].

8.5.1.5 Lemnoideae (Duckweeds) (Fig. 8.7)

Duckweed is a free-oating aquatic plant. It grows fast in many aquatic conditions.

Optimum temperature for the growth of the plant ranges from 5 to 35°C with a wide

range of pH from 3.5 to 10.5 [47].

Fig. 8.6 Brassica juncea

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Duckweeds are mostly found in ponds and wetlands. The plant (Lemna species)

shows high capacity for the exclusion of toxic metals from water. Lemnoideae

minor grows well from 6 to 9 pH and accumulates up to 90% of soluble lead from

water. Increased concentration of nitrate and ammonia inhibits the growth rate of

Lemnoideae minor [49].

8.5.1.6 Hydrilla verticillata (Hydrilla) (Fig. 8.8)

Hydrilla verticillata (hydrilla) is an aquatic weed that forms thick layer in the whole

water body. Whole plant can help in removal of contaminants. Denny and Wilkins [50]

reported that shoots are more efcient in heavy metals uptake instead of roots. Hydrilla

showed 98% uptake of lead when exposed to concentrated lead solution for 1week [51].

8.5.1.7 Spirodela intermedia (Duckweed)

Spirodela intermedia is a oating aquatic macrophyte, shows high growth rate even

under varied climatic conditions, and can accumulate cadmium, chromium, and

lead from water column [52]. Plants can diminish algal production by extending

itself all over the water surface and restrict the light penetration and ultimately pho-

tosynthesis [53].

Fig. 8.7 Lemnoideae

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8.5.1.8 Schoenoplectus californicus (Giant Bulrush)

Schoenoplectus californicus (giant bulrush) shows wide distribution geographi-

cally. It is a vascular plant that grows along the American continent below water

level and takes up nutrients from sediments through its roots. It is highly tolerant to

high metal concentration in streams and lakes [54].

8.5.1.9 Ricciocarpus natans (Fig. 8.9)

Ricciocarpus natans is a free-oating aquatic plant that can accumulate elements

directly from water [55]. Ricciocarpus natans lack ower, stems, roots, and vascu-

lar tissues, known as liverwort.

In a study, three aquatic macrophytes Eichhornia crassipes, Pistia stratiotes, and

Spirodela polyrrhiza were compared to check their removal efciency for heavy

metals. It was found that Eichhornia crassipes removes more metals than Pistia

stratiotes and Spirodela polyrrhiza [56]. External supplementation of ethylene

diamine tetra acetic acid (EDTA) was studied on Spirodela polyrhiza plant showed

high uptake of heavy metals such as arsenic (V) and arsenic (III).

8.5.1.10 Vallisneria spiralis (Fig. 8.10)

An experiment was performed for 21 days on Vallisneria spiralis to check its capa-

bility for removal of copper (Cu) and cadmium (Cd) with different concentrations

in a prepared pot containing sediment high accumulation was observed in roots and

shoots by decreasing chlorophyll.

Fig. 8.8 Hydrilla

verticillata

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A positive correlation was found between the level of metals in soil and plants

and/or between metals in water and plants. Salvinia natans showed high efciency

to accumulate different concentrations of Iron (Fe), Copper (Cu), and Zinc (Zn) at

different time periods [57].

8.6 Benets ofPhytoremediation

As a natural process, phytoremediation offers many benets.

• The process does not disturb the local environment and maintain the landscape.

• Most useful at shallow and low level contaminated sites.

• A wide variety of environmental contaminants can be treated.

• The idea is aesthetically good and has public acceptability. It is suitable for those

areas where other techniques are not applicable. It is cost-effective than the other

remediation techniques.

Fig. 8.9 Ricciocarpus

natans

Fig. 8.10 Vallisneria

spiralis

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• Phytoremediation has less maintenance and installation costs in comparison to

other techniques.

• The plantation on contaminated soils can prevent metal leaching and erosion.

Fast growth and large biomass producing plants can also be used for energy pro-

duction [58].

• Phytoremediation can help in reuse and recovery of valuable metals.

• This naturally occurring process is least harmful for the environment and

surrounding people.

8.7 Drawbacks or Limitations andChallenges

ofPhytoremediation

Even though phytoremediation is an environment-friendly process, it does have

negative aspects as well.

• It is comparatively a long-term remediation process.

• It can cause toxic effects to the food chain by transferring contaminants from

water or ground to foraging animals.

• The process has a less deep remediation zone ranging from 12in. to 15ft.

• High metal contamination can be harmful to the plants but some species are

highly efcient for the removal of toxicity.

• By the accumulation of toxic metals and contaminants, plants become harmful to

livestock and general public so there should be restricted access to the site.

• It is not suitable for highly contaminated areas because plants can accumulate

low to moderate level of contaminants from water and soil.

The development of phytoremediation as an eco-friendly process involves many

challenges in the future, e.g., development of local capacity and to establish effec-

tive regulatory policies. There is a lack of experience using phytoremediation less

available data, performance standards, and cost-benet analysis.

8.8 The Future ofPhytoremediation

Research is continuing in order to locate gene coding of plants which are having the

ability to hyperaccumulate some specic heavy metals in plants. With the help of

this different characteristics can be combined into a single plant species [58]. In

spite of a number of challenges, phytoremediation seems to be a green remediation

technology with a high potential. Unlike other physical and chemical methods for

removal of heavy metals and other contaminants, phytoremediation is low cost, eco-

friendly technology and does not destroy native soil microora and fauna. Screening

of local plants for phytoremediation and evaluation of the effects of different param-

eters during phytoremediation is in progress and for this purpose many

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interdisciplinary studies and researches are in process. Advance studies are identify-

ing a number of different proteins that are involved in transportation of pollutants

across membranes and vacuolar sequestration of these pollutants or heavy metals to

deeply understand the mechanism of phytoremediation. For phytoremediation and

phytomining of heavy metals, phytoextraction is expected to be a business

technology.

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A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_9

Chapter 9

Remediation ofPharmaceutical andPersonal

Care Products (PPCPs) inConstructed

Wetlands: Applicability andNew Perspectives

AnaRitaFerreira, AlexandraRibeiro, andNazaréCouto

Abstract Nowadays, wastewater treatment plants (WWTPs) considered not very

effective in removing all types of organic compounds, including pharmaceuticals

and personal care products (PPCPs). The efuent discharged containing PPCPs

shows negative impact on fresh/marine waters, even at vestigial concentrations. The

integration of constructed wetlands (CWs) as a biological treatment technology in

WWTPs may be an option to effective removal of PPCPs, which is crucial for water

bodies’ protection. On the other hand, if they arrive to water bodies it is important

to understand the self-restoration capacity of the system. This chapter makes an

overview (based on literature and experimental data) about the effectiveness of

CWs as a polishing step in WWTPs and the potential to remove contaminants if

they arrive to salt marsh areas. In both cases, there is a same principle behind. CWs

dened as articially engineered ecosystems designed and constructed to control

biological processes as in natural wetlands, but in a controlled natural

environment.

A case study highlights the remediation potential to remove target PPCPs in both

environments. Simulated CWs (spiked wastewater) planted with Spartina maritima

and light expanded clay aggregates (LECA) as substrate. Simulated salt marsh areas

(spiked elutriate soaked in sediment) were planted with the same plant but with

sediment as substrate. The presence of a physical support and/or S. maritima

decreased contaminant levels either in WWTPs or in estuarine simulated environ-

ment. Plant uptake, adsorption to plant roots/sediments and bio/rhizoremediation

are strong hypothesis to explain the decrease of contaminants either in CWs or in

salt marsh environment. The chapter also discusses the concept of energy produc-

tion in CWs as a way to increase the competitive advantages of CWs over other

treatment systems, by coupling an efcient removal together with a protable tech-

nology, which may decrease WWTP energetic costs.

Keywords WWTPs • PPCPs • CWs • Spartina maritima • LECA

A.R. Ferreira (*) • A. Ribeiro • N. Couto

CENSE, Departamento de Ciências e Engenharia do Ambiente, Faculdade de Ciências

e Tecnologia, Universidade Nova de Lisboa, Caparica 2829-516, Lisbon, Portugal

e-mail: arl.ferreira@campus.fct.unl.pt

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9.1 Introduction

Water pollution is a relevant problem as it compromises the quality of a resource

that is essential to life. In 2008, the production of hazardous chemicals (i.e., toxic

chemicals dened by Eurostat) was ca.200 million tons [1]. In 2011, the European

Environment Agency reported that hazardous substances, like pharmaceutical and

personal care products (PPCPs), have a detrimental effect in EU fresh and marine

waters [1]. PPCPs constitute a wide group of compounds largely consumed in mod-

ern societies aiming to improve the quality of daily life [2]. After utilization, e.g.,

pharmaceutical compounds are not completely metabolized in the body of humans

and animals and as a result, metabolites, conjugates, and their native forms are

excreted into the sewage system [3]. In addition, the unused and expired PPCPs are

usually disposed with normal household waste or discarded into sink or toilets [4].

Wastewater treatment plants (WWTPs) receive wastewaters that contain a lot of

different trace polluting compounds but are not specically designed to eliminate all

of these compounds [4–7]. Consequently, after WWTP treatment, various kinds of

PPCPs and their metabolites have been detected into surface water, ground water,

and even drinking water [8–12]. Upon entering the aquatic environment, and even

at trace levels, PPCPs and their metabolites became a potential risk to the health of

aquatic life and human beings. The available information on the ecotoxicology of

these compounds is scarce, and the potential risks to the water environment are still

under debate [2, 13–15]. However, it is clear that human pharmaceuticals cause

e.g., antibiotic resistance in microorganisms and will negatively impact aquatic

communities through feminization of male sh and affect kidneys, gills and liver in

sh [13, 16].

In WWTPs, different types of treatment technologies are applied aiming to

enhance organic contaminants, i.e., PPCPs removal. In fact, advanced oxidation

processes, activated carbon adsorption, membrane separation, and membrane biore-

actor are available to restore and maintain the chemical, physical, and biological

conditions of wastewaters [17]. However, advanced treatment processes involve

high capital and operational costs and selecting low-cost alternative treatments for

the removal of emerging contaminants seems to be a very promising option [3, 6, 13].

Therefore, the quest for green, cost-effective, and energy sustainable technologies is

a subject of debate today.

Constructed wetlands (CWs) represent an option that ts these purposes as they

represent a green treatment technology, cost-effective, with low operation and

maintenance requirements [18]. CWs are part of the tertiary treatment in WWTPs

and may be assumed as a polishing step before the discharge for the aquatic bodies.

CWs are dened as articially engineered ecosystems designed and constructed to

control biological processes as in natural wetlands, but in a controlled natural

environment. CWs has been widely used to treat various kinds of wastewaters [19],

such as domestic [20], agricultural [21], and industrial wastewater [22] but also

storm water and acid mine drainage [23]. However, removal rate in CWs (affecting

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279

the residence time) and the effect/area of inuence from the plant have been reported

as limitations to this technology [6].

This chapter is an overview about the existent practices concerning PPCPs

removal using CWs. The capability of a CW or a simulated salt marsh area (both

planted with Spartina maritima) to promote the removal of two PPCPs with differ-

ent physico-chemical properties, either in the presence or absence of a support

matrix will also be discussed. At the end of the chapter, insights about the integra-

tion of energy production in CWs will discussed. The main aim of this concept is to

increase the competitive advantages of CWs over other treatment systems, by cou-

pling an efcient removal together with a protable technology, which may decrease

WWTP energetic costs.

9.2 Phytoremediation

9.2.1 General Aspects

Phytoremediation is an environmentally friendly technology that uses plants for the

degradation, removal, and detoxication of contaminants from soils, sediments, or

waters [24]. Different mechanisms can be used to immobilize, sequester, degrade,

or metabolize in place (either inside or outside the plant) depending on the type of

contaminant, the site conditions, the level of cleanup required, and the type of plant [25].

The phytoremediation of organic contaminants, such as PPCPs, is complex and car-

ried out through different approaches. The contaminant absorbed by the plant and

then metabolized into nontoxic metabolites (phytodegradation). The capacity to

enter into the plant depends on the lipophilicity of the pollutant. It is accepted that a

Log Kow between 0.5 and 3 is adequate for this purpose [26]. However, contaminants

can remain outside the plant. In rhizosphere, organic contaminants may be biode-

graded by microorganisms that spur from root exudates (e.g., carboxylic acids,

amino acids) in a synergistic action between plant and microorganisms [27]. The

evolution of phytoremediation-related literature and from this, the relation with

organic contaminants assessed to understand the present research tendency regard-

ing this topic. Figure 9.1 shows the number of publications containing for the word

“Phytoremediation” and then “Phytoremediation AND organic contaminants.” The

data was obtained from the Scopus database with the search eld text =

(Phytoremediation AND Organic contaminants) from 2000 to 2014. The results

were rened based on: type of Literature = (Article OR Review) and subject

area=(Life Sciences). The results show that the phytoremediation is intensively

studied but literature regarding phytoremediation of organic contaminants repre-

sents a small percentage (between 32% in 2000 and 16% in 2014). The interest

regarding phytoremediation of organic contaminants is less than the researches in

phytoremediation of other organic compounds. Nevertheless, there is a growing

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tendency regarding the studies with phytoremediation and organic contaminants

(including PPCPs) (passing from 24 studies in 2000 to a maximum of 98in 2013).

9.3 Constructed Wetlands

9.3.1 General Aspects

CWs have been widely employed since its rst full-scale application in the late

1960s. During the last ve decades, CWs have evolved from empirical research into

success, increasingly more popular applications, e.g., habitat restoration for native

and migratory wildlife, anthropogenic discharge for wastewater, storm water runoff,

sewage treatment, land reclamation following mining or reneries [28]. The CWs,

also known as engineered wetlands, are designed to mimic the process involved in

natural wetland systems but within a more controlled environment [18]. Physico-

chemical properties of wetlands provide many positive attributes for contaminants

remediation [29]. In sequence, CWs have also demonstrated to be a sustainable and

operational technology to include in conventional WWTPs aiming for an efcient

decrease of total suspended solids, biochemical oxygen demand (BOD), or elimina-

tion/decrease of various pollutants including nitrogen, phosphorus and heavy met-

als [30]. In recent years, the applicability of CWs for the remediation of PPCPs has

been increasingly explored and proved to be successful for a variety of compounds

with a simultaneous improvement of water quality [31–36].

2000

56 90 131 118

192 234 249 293 310 365 350

501 468 516 582

28 31

33 39

48 47

66 67

100

200

300

400

500

600

700

800

Phytoremediation Phytoremediation AND organic contaminants

2001 2002 2003 2004 2005 2006 2007

Year

Number of publications

2008 2009 2010 2011 2012 2013 2014

Fig. 9.1 Number of articles or reviews published on the phytoremediation area from 2000 to 2014

(Source: online version of Scopus database accessed in 26.11.2015; search eld: Phytoremediation

AND Organic contaminants)

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CWs can be classied according to their hydrology (free water surface, subsurface

ow, and hybrid), ow path (horizontal or vertical), and types of macrophyte (free-

oating, emergent, and submerged) [6, 37]. According to the review of application

of CWs for wastewater treatment in developing countries performed by Zhang etal.

(2014) [38] horizontal subsurface ow (HSSF) CWs have been the most frequently

employed aquatic plant-based systems to remove pharmaceutical compounds

although vertical subsurface ow (VSSF) CWs and hybrid CWs have also shown

good removal efciencies for pharmaceuticals. The treatment performance in CWs

is critically dependent on the optimal operating parameters and includes water

depth, hydraulic load, hydraulic retention time, and feeding mode related to the

sustainable operation for wastewater treatments [18]. The contaminants removal in

wastewater involves a set of abiotic and biotic processes inuenced by plants, sub-

strate, and associated microbial assemblages, which assist in integral contaminant

removal, while the more homogeneous conditions in WWTPs (without these

dynamic interactions) induce fewer degradation pathways [30]. The physico-chem-

ical processes contributing to contaminants degradation in CWs have not been thor-

oughly described [39] and it is imperative to understand the transformation processes

that driven PPCPs removal, aiming to optimize CWs design for an effective con-

taminants removal.

CWs have advantages over the natural wetlands but also have some limiting fac-

tors. Land requirement is a limiting factor for their broader application, especially

in regions where land resources are scarce and population density is high. In addi-

tion, the biological components can be sensitive to toxic chemicals (e.g., ammonia

and pesticides) and peaks of contaminants in water ow may temporarily reduce

treatment effectiveness. Another point is the possible re-entry of contaminants after

the death of plants, which may result in a poor removal performance of CWs. To

prevent this, it is necessary to develop an appropriate plant harvest strategy, with a

focus on the reclamation and recycling of plant resources in CWs.

9.3.2 Salt Marsh Plants

The role of plants in CWs has been frequently discussed and several studies state

their crucial role, being considered the essential component of the design of CW

treatments [38]. The roots maintain the hydraulic properties of the substrate, and the

shoots protect the surface from erosion while shading prevents algae growth.

Besides, plants play another important role in stimulating the development and

activities of microbial populations, which are supported by the rhizodeposition

products (i.e., exudates) promoting the occurrence of various biological processes

in the rhizosphere (e.g., transformation and mineralization of nutrients and organic

pollutants) [40]. Not all plants are suitable for waste treatment since plants must be

able to tolerate the combination of continuous ooding and exposure to waste

streams containing relatively high and often variable concentrations of

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contaminants [41]. Therefore, the study of plant species is crucial to obtain better

treatment efciency in CWs.

Salt marsh plant species are morphologically adapted to cope with environ-

mental stress, such as, high concentrations of salt and/or insufcient water condi-

tions. In wetlands, these types of plants have been reported to be one of the main

factors inuencing water quality by their capability of utilizing nitrogen, phos-

phorous, and other nutrients [18]. Salt marsh plants also have shown potential to

remediate inorganic [42] and organic [43] contaminants. Table 9.1 summarizes

studies using salt marsh plants for PPCPs removal in aquatic media simulation.

The most popular salt marsh plants are Phragmites australis, Typha spp., includ-

ing Typha angustifolia.

9.3.3 Substrates

Substrate or support matrix is considered as an important component of CWs that

provides a suitable growth medium for plant and microorganisms together with a

successful movement of wastewater [49]. The frequently used substrates include

natural (sand, gravel, clay), articial (light weight aggregates, activated carbon),

and industrial (slag) materials [18]. Substrates can remove contaminants from

wastewater by exchange, adsorption, precipitation, and complexation [36]. For this

reason, the chosen materials are extremely important when designing CWs as, e.g.,

a material with high sorption capacity will improve contaminants removal [50].

Calheiros etal. [49] studied the treatment of tannery wastewater by Typha latifolia

in CWs established with three different substrates. The tested substrates proved to

be adequate for T. latifolia development with higher organic removal for the two

Table 9.1 Salt marsh species reported for PPCPs removal from aquatic medium

Plants PPCPs References

Typha spp. Carbamazepine, clobric acid, and ibuprofen [36]

Typha angustifolia Triclosan [44]

Ibuprofen, diclofenac, caffeine, and methyl

dihydrojasmonate

[39]

Scirpus spp. Carbamazepine, ibuprofen, naproxen, tramadol [8]

Scirpus validus Caffeine [45]

Carbamazepine [46]

Phragmites

australis

Enrooxacin, ceftiofur, and tetracycline [47]

Ibuprofen, naproxen, diclofenac, tonalide, and bisphenol A [48]

Ibuprofen, diclofenac, caffeine, and methyl

dihydrojasmonate

[39]

Typha and

Phragmites

Clobric acid, carbamazepine, caffeine, methyl

dihydrojasmonate, galaxolide, tonalide, ibuprofen,

naproxen, ketoprofen, and diclofenac

[33]

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expanded clay aggregates when compared to the ne gravel. Dordio etal. [51]

showed in laboratorial batch experiments that light expanded clay aggregate (LECA)

is considered a good sorbent for acidic (e.g., clobric acid and ibuprofen) and neu-

tral pharmaceutical compounds (carbamazepine) with removal efciencies between

75% and 97%. Recently, biosorbents such as rice husk, pine bark, and granulated

cork have also been considered as interesting alternatives to the common substrate

materials in CWs due to their low cost, economical value of reuse, and easy disposal

by incineration certain [6].

9.4 Case Study

The aiming of this study was to understand, in the tested conditions, the remedia-

tion potential of the different components of the system (plant, substrate) after a

(simulated) PPCPs contamination before (CWs) and after efuent discharge (salt

marsh area). In addition, the capacity of planted CWs and LECA as a support

medium to remove contaminants were also evaluated. The rst study tested the

potential of CWs for PPCPs removal and the second simulates the self-restoration

capacity of the salt marsh area affected by PPCPs load. In both cases, S. maritima

was the chosen plant species. This plant species is frequently found in Portuguese

estuaries and may potentially be used in CWs. Two PPCPs with different physico-

chemical properties were chosen: caffeine (CAF) and oxybenzone (HMB). CAF

has a Log Kow of −0.77, pKa of 10.4, and solubility of 2.16*104 mg L−1 at

25°C. CAF is one of the most consumed stimulant of central nervous system

worldwide [45]. HMB is a UV lter increasingly used in personal care products,

in particular as light-lters to protect the human skin from harmful exposure to

UV irradiation [52]. HMB has a Log Kow of 3.8, pKa of 7.6, and solubility of

69mg L−1.

9.4.1 Methodology

The work was divided into two different parts: [53, 54]. For sake of clarity, a com-

parative assessment between both is carried out. Experimental design of the work is

shown in Fig. 9.2.

CWs were prepared with LECA as substrate and with continuous entry of

contaminants, simulating real operating parameters (residence time; Assay 1).

Simulation of the salt marsh environment was carried out with sediment soaked in

the respective elutriate, allowing simulation of nutrients and contaminants exchange

among plants, solution, and sediment, as occurs in the natural environment (Assay 2).

The efuent was collected after a secondary treatment stage in a WWTP from Águas

de Lisboa e Vale do Tejo located in Quinta do Conde, Sesimbra, Portugal

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(38°34’13”N, 9°2’7” W). The plants, water, and colonized sediment were collected

at low tide from a salt marsh, located in the Tagus River Estuary, Portugal

(38o36’59.39”N; 9o02’33.41”W).

All the microcosms where wrapped in aluminum foil to protect of the sunlight

and simulate real light penetration conditions. Groups of S. maritima were homoge-

neously distributed (9.0±1.0g) by different treatments and exposed to the medium

(wastewater and elutriate). Plant roots were disinfected before the experiments to

stop bacterial activity. The experiments to simulate CWs were carried out for 7days,

but there was three spiking periods (at days 0, 3, and 6) making the concentrations

range from 0.5mg L−1 to 1.5mg L−1. The experiments to simulate the salt marsh

area were carried out for 10days, and the system was spiked with 1mg L−1 of each

contaminant. The purpose of different spiking periods is to simulate a successive

load of contaminants in CWs and a lower contaminant load in estuarine systems.

Three types of controls carried out in parallel (spiked matrix with isolated presence

of substrate or plant and non-spiked matrix with the presence of plant to evaluate

plant vitality). Photosynthetic pigments used to evaluate plant vitality when exposed

to contamination. High performance liquid chromatography (HPLC) used to quan-

tify the levels of different contaminants in the studied matrices.

Fig. 9.2 Experimental design of the work

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9.4.2 Results andDiscussion

The presence of contaminants may inuence the functions of plants and associated

efciency for contaminants removal. The evaluation of chlorophylls (a and b) and

carotenoids (μg g−1) of S. maritima exposed to PPCPs showed that this plant toler-

ates up to 1.5mg L−1 of CAF and HMB.Table 9.2 shows the remediation potential

of the system components (plant and substrates: LECA and sediment) in each simu-

lation environment (CWs and salt marsh areas) compared with respective controls.

In the CW, the presence of S. maritima only increased HMB remediation by 10%

but did not have any effect on CAF.Also, in the simulated salt marsh area S. mari-

tima had no effect on CAF remediation but promoted a decrease of 60% in HMB.

S. maritima promoted CAF remediation in about 20% with the presence of LECA

(CWs) or sediment (salt marsh area). HMB presented a different remediation behav-

ior. The remediation was neglectable in CWs (plant, wastewater, and LECA) and

was almost 40% in simulated salt marsh area (plant, elutriate and sediment).

Regarding the susbtrates, the presence of sediment enhanced the remediation of

HMB by 60% and of CAF by 17% in salt marsh simulation. In wastewater, LECA

presented 10% of HMB remediation, but negatively affected CAF remediation.

The uptake by plants is more probable for compounds with Log Kow values of

0.5–3 [26]. Recent studies show that compounds with other Log Kow values may

also enter the plant. Wu etal. (2013) [55] detected PPCPs with a detection fre-

quency of 64%, and concentrations range of 0.01–3.87ng g−1 (dry weight) in veg-

etables. Triclocarban, triclosan, and uoxetine (Log Kow>3) accumulated in roots

at levels higher than the other PPCPs, while translocation to leaves/stems was for

compounds with Log Kow<3, e.g., carbamazepine. Also, (ab)/adsorption to plant

roots and (bio)/rhizoremediation in liquid phase or substrate may be strong hypoth-

esis to the enhanced remediation in the tested conditions. The higher removal of

HMB, compared to CAF, explained by their octanol water partition coefcient (Log

Kow>3) and solubility, which promotes their retention by adsorption of the solid

matrices (bioconcentration in the roots or in the sediment through adsorption

processes, which is higher for hydrophobic contaminants). CAF has a very high

Table 9.2 Potential of remediation of plant and substrate

CW Salt marsh area

Plant Substrate Plant Substrate

Contaminant S. maritima

vs. controla

Planted vs.

unplanted

Unplanted

LECA bed

vs. controla

S. maritima

vs. controlb

Planted vs.

unplanted

Unplanted

vs. controlb

CAF (=) 0% (+) 20% (−) 40% (=) 0% (+) 19% (+) 17%

HMB (+) 10% (=) 0% (+) 10% (+) 60% (+) 38% (+) 60%

Note: (+), (−) or (=) means the potential of the plant or substrate comparing (vs.) with controls

aControl only with wastewater

bControl only with elutriate

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solubility and tends to remain in the liquid phase. Therefore, the presence of

microorganisms (either in simulated salt marsh area or CW/liquid or solid phase)

appears to favor biodegradation. The studied compounds are reported as biodegrad-

able, being indicated as readily biodegradable, mainly HMB [56, 57].

9.5 CWs Coupling Plant Microbial Fuel Cells

The combined/integrated treatment systems present a novel pathway to improving

CWs functions. The improvement of wastewater quality with simultaneous energy

recovery has garnered much attention in recent years [58]. Plant microbial fuel cell

(Plant-MFC) is an emerging technology, which consists in the conversion of solar

energy to bioelectricity. It was patented in 2007, and the proof-of-principle was

published in 2008 (e.g., [59]) and developed in an EU project 2009–2012 resulting

from a spin-off company Plant-e. Plant-MFC may represent an add-in value to

CWs. 50% of photosynthetic organic matter goes to soil where naturally occurring

bacteria oxidize it and transfer energy rich electrons to the anode of the fuel cell.

The energy can be used as electrical energy [60]. In addition, plants transfer oxygen

to the rhizosphere through the root system and enhance the aerobic degradation of

unutilized organic matter, nitrication and mineralization of aromatic amines [61].

Figure 9.3a presents a model of the plant-MFC. The maximum and long-term

(2weeks) power output of best performing Plant-MFC reached 0.44 and 0.222W

m−2 [60], a value comparable with conventional biomass–electricity chains, with

potential to cover energy consumption. The technology has been scaled up to 25m2

in a “green electricity roof” and has a potential to be applied in wetlands [62]. In the

case of CWs the “traditional” approaches, the anodic chamber is in the bottom

region of the system (Fig. 9.3b).

In this region, microbes oxidize the organic matter and promote denitrication

thus generating electrons (e−), protons (H+), and carbon dioxide. Electrical current

is generated when the electrons migrate to the cathode. The voltage difference

between the anode and cathode, together with the electron ow in the outer circuit,

generate electrical power [63]. The electrons from the anode also react with oxygen

(or other electron acceptors) at the cathode to produce water and other reduced

compound. Different electrode materials can be used for the process (e.g., stainless

steel mess, platinum, carbon paper, and granular active carbon). Carbon and graph-

ite are commonly used as anode and cathode electrode materials because they offer

high electrical conductivity and non-oxidative nature thus offering a good medium

for the attachment and growth of microbial communities [64]. It is important to note

that various operation parameters and designs have been developed lately by cou-

pling MFC into other wastewater treatment process in an attempt to maximize the

aerobic and anaerobic conditions. Different congurations can be found in the fol-

lowing references: [58, 65–67]. The study of the Plant-MFC concept extensively

explored while the integration of CW and MFC is still in the beginning. Combining

CW and MFC seems a promising green technology to be incorporated in WWTPs

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allowing a cost-effective process to produce electricity. Several works using this

technology to remove organic contaminants from wastewater with simultaneous

energy production have been reported. Some examples are given below.

Figure 9.4 shows the increased number of publications with the integration of

CWs and MFC, retrieved from Scopus (26 November 2015). In the online version

of Scopus database, the search terms text = (Constructed wetlands AND Microbial

Fig. 9.3 Schematic diagram of (a) model of a plant microbial fuel cell producing electricity and

driving a light source (adapted from [59]); (b) model of constructed wetland including the concept

of microbial fuel cell (adapted from [67])

2004

2006 2008 2010

Year

Number of Publications

2012 2014

2016

Fig. 9.4 Publications from 2004 to 2015 (Source: online version of Scopus database accessed in

26.11.2015; Search terms: Constructed wetlands AND Microbial fuel cells; Search eld: Article

Title, Abstract, Keywords; Document type: Article or Review)

9 Remediation ofPharmaceutical andPersonal Care Products (PPCPs) inConstructed…

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fuel cells) with a search eld: Article Title, Abstract, Keywords in a period between

2004 and 2015, were used. The results were further rened based on: type of

Literature = (Article OR Review) and subject area=(Life Sciences). The study car-

ried out by Villasenor etal. [68] operated into a HFCW-MFC using a bentonite layer

to separate the lower anaerobic anode compartment and the upper aerobic cathode

compartment. It was reported 95% of COD removal (mean inuent concentration of

560mg L−1) and a power density of 20.76mW m−2in the CW.These authors reported

that several factors inuenced the electricity generation, such as the aerobic envi-

ronment in the upper wetland zone, which in part, depends on the aeration potential

of the plants. In general, the aeration potential of macrophytes is rather low com-

pared with the conventional aeration systems in wastewater treatment plants.

The authors Zhao etal. [58] studied CW-MFC to treat swine wastewater oper-

ated in batch mode, in continuous, without and with air diffusion heads to aerate

the cathode region. 71.5% of COD was removed (with initial concentration of

3190–7080mg L−1) and a peak power density of 12.83mW m−2was produced.

The aeration in the cathode region signicantly enhanced the performance of the

CW-MFC, with the continuous mode demonstrating an average of 76.5% COD

removal (average inuent COD concentration of 1058.45±420.89mg L−1) and a

peak power density of 9.4mW m−2. Doherty etal. [66] studied the ability of the

alum-sludge- based CW-MFC to remove organics from wastewater while produc-

ing electricity with different ow directions on the CW-MFC performance. They

concluded that the ow direction inuenced the efciency of the system. The

authors say that the simultaneous upow–downow CW-MFC combats the two

major bottlenecks of CW-MFC power output: reducing the separation between the

electrodes and maintaining anoxic conditions at the anode and aerobic conditions

at the cathode.

Fang etal. [69] applied a vertical CW-MFC system to treat azo dye wastewater

(aromatic compounds) and simultaneously produced electricity. The system

achieved 91% of decolorization rate and a voltage output of about 610mV.The

results obtained by these authors showed that plants grown in cathode region had

potential to enhance the voltage output and slightly promoted dye decolorization

efciency. Villaseñor etal. [68] reported the inuence of plants in voltage, stated

that photosynthetic activity affected the redox conditions in the cathode compart-

ment, as the deposition of organic matter and O2 in the rhizosphere increased.

During the night, the voltage dropped to approximately 200mV in the horizontal

ow CW-MFC, planted with Phragmites australis, and gradually increased to maxi-

mum values during daylight. Liu etal. [61] have also shown the importance of

plants in power density and nutrient removal of CW-MFC.The authors incorporated

the root exudates of Ipomoea aquatica as part of fuel into the anode section of the

CW-MFC and produced a power density 142% higher than that of 5.13mW obtained

from the unplanted systems. They also promoted the reduction of internal resis-

tance. The planted CW-MFC removed 95% of COD whereas 92% of removal

achieved in the unplanted CW-MFC. The average nitrogen removal efciencies

were 54% and 91% in the unplanted and planted systems, respectively. The concept

of CWs coupled to MFC systems was tested with Typha latifolia [67]. Electricity

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was generated with maximum power density of 6.12mW m−2 and contaminant

removal was enhanced during wastewater treatment. The removal efciencies of

COD, NO−3, NH+4 were of 100%, 40%, and 91%, respectively. Despite the several

studies, the combination of CW-PMFC is an emerging technology and more

research is required to increase the power output (as nowadays it is too low to be

directly utilized) [70].

9.6 Conclusions

Population growth implies higher and faster generation of WWTP waste streams as

well as higher consumption of PPCPs. These compounds are not efciently removed

in WWTP treatment methodologies and the efuent discharge into water bodies

may lead to environmental and human risks. There is a need to nd sustainable solu-

tions to prevent this situation in future (by acting in WWTPs) or to remediation

areas that have been contaminated throughout the times (salt marsh areas). In both

environments, it is important to study the importance of “key-components” in the

system, i.e., matrix, plant species and substrates. The remediation capacity of the

system results from a dynamic interaction between matrix-plant-substrate compo-

nents and physico-chemical properties of the PPCPs, which will promote their dis-

persion/dilution in liquid fraction, adsorption to solid fraction, or bio−/

rhizoremediation. The concept of CWs as a green technology to remediate organic

contaminants matches the purpose of Plant-MFC with the associated benet of

electricity.

Acknowledgements The authors would like to thank AdP, Dr. Cristina Santos for providing the

samples, Eng. Olga Paredes, and the Laboratory of Control and Processes for the assistance.

Financial support was provided by UID/AMB/04085/2013 and 4KET4Reuse (SOE1/P1/E0253),

co- nanced by the European Regional Development Fund (FEDER). The Associação Nacional

de Farmácias (ANF), Portugal, is also acknowledged for giving the HPLC. N. Couto acknowl-

edges Fundação para a Ciência e a Tecnologia for her postdoc fellowship (SFRH/BPD/81122/2011).

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Chapter 10

Floating Wetlands fortheImprovement

ofWater Quality andProvision ofEcosystem

Services inUrban Eutrophic Lakes

EugeniaJ.Olguín andGloriaSánchez-Galván

Abstract The occurrence of eutrophic urban water bodies is widely spread globally

especially in countries where sanitary infrastructure is decient in medium and small

cities. Floating Wetlands also known as Treatment Floating Wetlands or Floating Islands

are a suitable option for the treatment or improvement of the water quality in urban

water bodies since they show several advantages over other systems, especially that they

can operate in situ and no additional surface of land is required. They have been applied

for the treatment of various types of water/wastewater ranging from low nutrient to high

nutrient content. Their efciency at removal of nutrients and other type of pollutants

depends on several factors being the most important ones the initial concentration of

pollutants, the environmental conditions, and the characteristic of the utilized plants.

Emphasis is given in the need of research at large-scale applications in situ and also in

the study of the potential of FW for the provision of ecosystem services. There are very

few studies oriented towards this latter issue, which is currently a very important one for

understanding their benets to the urban human communities.

Keywords Eutrophication • Eutrophic lakes • Water pollution • Phytoremediation

• Emergent plants • Articial wetlands • Ecosystems services

10.1 Introduction

It has been widely recognized that the excessive anthropogenic nutrient loading

(especially nitrogen and phosphorus) promotes eutrophication. This is one of the

most serious environmental problems affecting the water quality in fresh water bod-

ies. Alterations in aquatic communities, sh mortality, reduced oxygen levels,

excessive growth of algae (harmful algal blooms), and the increase of suspended

solid materials are the main effects of eutrophication [1]. At high densities, harmful

E.J. Olguín (*) • G. Sánchez-Galván

Institute of Ecology, Environmental Biotechnology Group,

Carretera Antigua a Coatepec # 351, El Haya, Xalapa, Veracruz 91070, Mexico

e-mail: eugenia.olguin@inecol.mx; gloria.sanchez@inecol.mx

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_10

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294

algal blooms produce odor problems and kill aquatic biota due to the release of

toxins, low dissolved oxygen, and high ammonia concentrations, which are associ-

ated with their proliferation and senescence [2]. The discharge of untreated or par-

tially treated domestic wastewater to the aquatic environment threatens public

health and aquatic ecosystems. In Indonesia, the national average of access to

wastewater facilities was 56% in 2010, from which only the 1% is related to sewage

plants treatment systems since septic tanks are the most common wastewater infra-

structure [3]. In Mexico, approximately 91% of inhabitants have access to improved

sanitation; however, only 50% of all collected sewage is treated in wastewater treat-

ment facilities [4], especially in big and most important cities, leaving behind the

small and scattered communities which are mainly in rural areas.

In China, the treatment rate of urban domestic wastewater was 72.9% in 2010 [5]

although in most small towns, untreated domestic wastewaters are discharged into

water bodies [6].

The water pollution in urban lakes and rivers, due to the excessive input of nutri-

ents from untreated domestic wastewaters, has been extensively reported in devel-

oping countries. Sánchez etal. [7] found that the urban lagoon called “La Pólvora,”

in the Grijalva River basin in Mexico, had a hyper-eutrophication level affecting the

biodiversity of crustaceans and mussels. Olguín etal. [8] analyzed the water quality

of the Sub-basin of Sordo River that passes through Xalapa, México.

Authors found that the concentrations of diverse parameters (nitrates, phosphates,

fecal coliforms, BOD, dissolved oxygen, pH, dissolved suspended total solids), in

different monitoring points, were above the permissible levels for the development of

aquatic life (US-EPA). Another urban lake called “Lago de Guadalupe” located in

Mexico City has been reported as a hyper-eutrophic lake with an anoxic environment

[9]. Recently, Olguín etal. [10] found in the lake system “Los Lagos del Dique” in

Xalapa, Mexico, a high content of nutrients, organic matter, and pathogens that

exceeds the values established in the Norm of EPA for aquatic life protection and the

presence of an excessive growth of algae, especially during the spring. In this case,

nonpoint source pollution from storm water runoff from the surrounding streets is one

of the major sources, especially during the rainy season. In China, 130 major lakes

were analyzed and the results showed a high level of eutrophication in 43.5% of them

while 45% had an intermediate status. All lakes are located around cities and receive

large amounts of municipal sewage without an appropriate treatment [11].

10.2 Need ofEnvironmentally Friendly Technologies

Effective solutions for the remediation of eutrophic urban lakes, the control of

harmful algal blooms, and the prevention of eutrophication are a matter of urgent

concern in those countries in which this problem is still prevalent in small and

medium size cities. The conventional wastewater treatment processes are not suit-

able for eutrophic lakes since in the latter case, the nutrient concentration is lower

and the water volume is huge, and land is not available for establishing treatment

plants. Thus, the implementation of environmentally friendly technologies with

high efciency and low cost is required.

E.J. Olguín and G. Sánchez-Galván

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Phytoremediation offers various alternatives to treat polluted water, being the use of

aquatic oating plants in phytoltration lagoons one of the most widely applied for treat-

ing poultry efuents [12], municipal [13], and aquaculture wastewaters [14]. On the

other hand, the use of emergent plants in constructed wetlands has also being applied for

the treatment of domestic wastewater [15, 16] and agroindustrial wastewater [17–19].

More recently, another type of environmentally friendly technology using emer-

gent plants adapted to the water environment has been usefully applied for the treat-

ment in situ of polluted water. Such type of technology has received various names

such as Floating Treatment Wetlands, FTW [20, 22], Floating Islands [21], and

Floating Wetlands [10]. For practical reasons, in the foregoing text, this type of

technology will be referred as Floating Wetlands (FW).

In general terms, the FW present several advantages over the other type of arti-

cial wetlands: (a) they are applied in situ, avoiding the need of large land areas for

treatment; (b) they can be built at a low cost using plastic containers supported by

empty bottles for providing buoyancy [22] or oating matrices of various types [20,

23, 24]; (c) they provide a support for the upper parts of the plants (leaves and

stems) and allow the roots to be hanging in the column water enhancing the plant

uptake and the support for rhizospheric microorganisms and their biolms with

degrading capabilities; (d) they are able to remove nutrients [25, 26]; (e) in some

cases, they can provide ecosystem services such as an increase in the dissolved

oxygen of the rhizospheric zone and removal of pathogens [22, 27].

The efciency of FW at the removal of organic matter and nutrients such as N

and P varies widely according to the initial concentration of pollutants in the water,

the type of plants utilized, and the environmental conditions. The initial concentra-

tion of N and P in storm water and reservoirs is much lower than in eutrophic lakes

and rivers. There have been several reports dealing with this type of polluted water,

Table 10.1 Efciency of FW at removal of organic matter and nutrients in stormwater reservoirs

Plant(s) used

% Removal

averages

Environmental

conditions Highlights References

Canna indica,

Thalia dealbata

and Lythrum

salicaria

COD: 71.2% Urban

stormwater

runoff sewage,

greenhouse

T. dealbata outperformed

C. indica and L. salicaria

in nutrients removal

Ge et al.

(2016) [46]

TN: 70.0%

TP: 82.4%

Scirpus

californicus

TP: 50% Controlled

conditions

with

stormwater

(mesocosms)

The application of oating

islands as a stormwater

technology can remove

nutrients through plant

uptake and biological

activity

Chang et al.

(2012) [25]

P-PO4: 45%

TN: 25%

HRT = 30 days

N-NO3: 60%

N-NH4: 55%

Chrysopogon

zizanoides

(Vetiver grass),

Typha

angustifolia and

Polygonum

barbatum

TP:

19.1–46.0%

Water from a

reservoir, in

situ

Estimates of nutrient

uptake rates showed that

Typha achieved the highest

uptake rates, compared to

Vetiver and Polygonum

Chua et al.

(2012) [40]

TN:

7.8–67.5%

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Table 10.2 Efciency of FW at removal of organic matter and nutrients in efuents from wastewater treatment plants and agroindustrial activities

Plant(s) used % Removal averages

Environmental

conditions Highlights References

18 local species from Tampa,

Florida, USA

Oxidized nitrogen

N–NO2 + N–NO3: <72%

Wastewater efuent

from treatment plant

An incremental TN removal rate for the

Floating Wetland Islands was calculated

to be 4.2kg N/m2 FWI per year

Vázquez- Burney

etal. [41]

TN mass removal efciency: 61%

Calamagrostis epigejos,

Phragmites australis,

T.latifolia, Juncus maritimus,

P.palustris and E. atherica

TN uptake: 10–15 g/m2Manure and biomass

digestate liquid

fraction

P. palustris and E. atherica present the

highest potential to be used to treat

digestate liquid fraction in oating

wetlands

Pavan etal. [42]

TP uptake: 1–3 g/m2

Ipomoea aquatica Forsskal TN: 11.2% Aquaculture

wastewater, in situ

The removal mechanism of nitrogen

compounds was due mostly by

microorganisms, whereas the removal

mechanism of phosphorus was mainly

by plant absorption

Zhang etal. [43]

N–NH4: 60.0%

N–NO2: 60.2%

TP: 27.3%

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and the removal efciency has been observed to vary widely and it has been reported

that plant uptake plays a very important roll (Table10.1).

On the other hand, FW have also been tested for the treatment of efuents from

wastewater treatment plants and from agroindustrial activities (Table 10.2),

observing also a wide variation in performance depending on the type of plant

utilized and other operation parameters.

10.3 Applications ofFW fortheImprovement ofWater

Quality inEutrophic Lakes

FWs have shown to be very efcient and convenient for the removal of nutrients in

eutrophic water bodies (Table10.3). Such efciency varies widely with the season,

the initial concentration of nutrients, and the plants utilized. There is no single opti-

mal combination of plants; it depends on the geographical location and the local

availability of each particular species. There are only very scarce studies in which

the removal of pathogens (such as fecal coliforms, FC) has been quantied. In the

case of the Mexican urban lakes [22], such efciency was found to be related to the

hottest season (summer).

Among the various designs that have been developed for using Floating Wetlands

for treating polluted water from rivers, there is a novel design, which is called hybrid

oating treatment bed (HFTB). It is a combination of a oating treatment bed (FTB)

with structures to promote the attachment of periphyton to improve the nutrient

removal capacity of the whole system. Periphyton is a complex assemblage of

aquatic organisms such as microalgae, bacteria, protozoa, and other organisms with

attaching capacity to surfaces. The combination of both, the oating plants, and

their rhizosphere and the periphyton maintained the TN and TP of the river at less

than 2.0 and 0.02 mg/L, respectively [28].

FWs have also demonstrated to remove toxic compounds generated by the algal

blooms present in eutrophic water bodies. Bao [29] assessed the removal efciency

of microcystin-LR (MC-LR is the main variant of microcystin) in FWs planted with

Oenanthe javanica during the treatment of water collected from a eutrophic river.

The removal ranges were 38.3–48.4, 42.5–56.5, 37.5–46.9, and 30.2–42.8% for

ow rates of 0.5, 2, 4, and 10 cm/s, respectively. However, the changes in ow rates

did not affect the absorption of MC-LR by the plant, which would suggest the pres-

ence of an additional mechanism of removal such as bacteria degradation. MC-LR

plant uptake was higher in roots than in stems and leaves at the inlet (average: 1.75

vs. 0.9 μg/kg fw) and outlet (average 1.2 vs. 0.6 μg/kg fw) of the FW in all different

ow rates tested in this study. The exposition of O. javanica to MC from a polluted

river did not inhibit its growth.

Few are the reports about the use of FWs, at large scale, to improve water quality

in eutrophic water bodies. Wang etal. [30] assessed the growth of Pontederia cordata

L. and Schoenoplectus tabernaemontani planted in FW established in a eutrophic

urban lagoon. The macrophytes adapted successfully to the stress due to the low

dissolved oxygen (1.2mg DO/L) during the summer and the low nutrient concentra-

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Table 10.3 Efciency of FW at removal of organic matter and nutrients in polluted rivers, lakes, and ponds

Plant(s) used % Removal averages

Environmental

conditions Highlights References

Phragmites australis and Canna indica N–NH4: 72.0% Polluted river

water,

experiments at

pilot plant level

The performance of the FW was critically

inuenced by the maturity of the roots,

during sudden increments of the hydraulic

and pollutants loads

Saeed etal. [43]

N-NO3: −35.0%

TP: 57.1%

BOD5: 8.2%

COD: 7.7%

TS: 23.9%

Typha orientalis Presl, Eleocharis dulcis,

and Juncus effuses

TN: 66.7% Urban lake, in

situ

The experimental units can quickly

enhance water quality; it is worth

promoting its application for water

landscapes and environmental conservation

Lu etal. [24]

TP: 74.4%

N–NH4: 99.9%

N–NO3: 93.7%

N–NO2: 98.9%

P–PO4: 63.5%

COD: 91.7%

Pontederia cordata L., Schoenoplectus

tabernaemontani

P removal: Urban pond, in

situ

P. cordata produced more biomass and

demonstrated higher P removal

performance than S. tabernaemontani

Wang etal. [30]

7.58mg P/plant of P.

cordata

1.62mg P/plant of S.

tabernaemontani

Pontederia sagittata and Cyperus

papyrus

WQI (scale from 1 to

100) was upgraded

from “polluted” (46)

to “moderately

polluted” (53) during

the summer

Eutrophic urban

lake, in situ

The FWs increased the DO concentration

(18–25%) and decreased the fecal

coliforms (85%). The plants’ growth and

the high productivity observed (4.87 kg/m2

month) indicated the nutrient uptake from

the water column

Olguín etal. [22]

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299

Pontederia sagittata and Cyperus papyrus COD: 71.7% Eutrophic urban

lake, in situ

The productivity of a mixture of both

plants after 3 months was higher in autumn

season than in winter season (4.14 vs.

2.33kg dw/m2month)

Olguín etal. [10]

N–NH4: 27.3%

N–NO3: 61.42%

TDS: 39.8%

Canna indica, Accords calamus, Cyperus

alternifolius, and Vetiveria zizanioides

COD: 15.3–38.4% Eutrophic river

water, in situ

Canna indica exhibited better growth,

higher DO levels, and a higher percentage

of nutrient removal attributable to plant

uptake

Bu and Xu [27]

TN: 25.4–48.4%

TP: 16.1–42.1%

Chl-a: 29.9–88.1%

Phragmites australis, Carex elata, Juncus

effusus, Typha latifolia, Sparganium

erectum, and Dactylis glomerata

COD: 29–66% River water, in

situ

The oating systems were easily installed

and required few maintenance operations

De Stefani etal.

[26]

BOD:41–52%

TP: 15–65%

TN: 13–29%

N–NO3: 12–14%

Phragmites karka TS: 55–60% River water, in

situ

The system is recommended as an

eco-friendly river water treatment for small

shallow, slow owing (or slightly stagnant)

water bodies

Billore etal. [45]

N–NH4: 45–55%

N–NO3: 33–45%

TKN: 45–50%

BOD: 40–50%

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300

tions during the winter (0.15 mg P/L and 1.15 mg TN/L). P. cordata was more

efcient to remove P than S. tabernaemontani (164.4 vs. 35.1 mg P/m2). The

changes in the biomass productivity and the P concentration pointed out that the P

was accumulated in stems during the summer and mobilized, during the autumn, to

the inferior parts of the plants (rhizomes and roots). In Mexico, two linear FWs

(17.5 and 33m2) have been established by Olguín etal. [22] in a eutrophic urban

lake in the city of Xalapa, Mexico (Fig.10. 1).

FWs were built with low-cost materials and planted with a combination of

Pontederia sagittata and Cyperus papyrus (Fig.10. 2). A water quality index (WQI)

was calculated with the data of selected parameters, according to Olguín etal. [8].

The results showed that FWs contributed to improve the WQI, especially during the

August and October’s monitoring period (summer and autumn). The WQI improved

from polluted (27–50) to moderately polluted (51–70), especially during summer.

10.4 Ecosystem Services Provision

An ecosystem service (ES) is a direct or indirect contribution of ecosystems to the

human well-being, which generates a link between the biophysical aspects of the

ecosystems and the human well-being. They have been classied as provisioning,

Fig. 10.1 Two large Floating Wetlands have been established as an efcient aquatic phytobarrier

or phytolter at the entrance of Lake 1in the Lakes System known as “El Dique” en Xalapa,

Veracruz, México [22]

E.J. Olguín and G. Sánchez-Galván

guarino@unisannio.it

301

regulating, supporting, and cultural, according to the type of service they provide

[31]. Wetlands provide all kinds of ES such as provisioning (i.e., food, fuel, wood,

and freshwater), regulating (i.e., ood control, climate, water quality, water supply),

supporting (nutrient recycling, soil formation, primary production, biodiversity)

and cultural (i.e., recreational, aesthetic, educational) [32]. The ES assessment of

wetlands can be carried out in two steps: (1) physical dimension measurement

including plant harvest, removal rates of N/P, carbon xation, and oxygen release

and (2) monetary evaluation in which the SE value, in the market, is estimated [33].

Research on ES provided by wetlands has been mainly focused on large wetlands

and few are the reports about small wetlands (<1 ha). However, often small wet-

lands have a high performance in the improvement of water quality (in terms of

area), compared to large wetlands [34].

The ecosystem services (ESs) concept has been extended to the design and man-

agement of landscapes. Constructed wetlands (CW) for the improvement of the

quality of wastewater, before being discharged to water bodies, is an example. CW

can directly or indirectly support human well-being by providing several ES such

as a clean water supply, habitats, food, aesthetic, education, and recreational ben-

ets [35].

FW can provide other ESs different to water quality improvement. They can

directly or indirectly support the human well-being providing ESs such as regula-

tion including ood reduction and atmospheric CO2 capture through vegetation,

improvement of the air quality through ltration and/or particle absorption, NOx,

Fig. 10.2 The Floating Wetlands established in Xalapa, Veracruz, México, have a combination of

Cyperus papyrus and Pontederia sagittata and provide various ecosystem services, including habi-

tat for birds [22]

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and other contaminants. ESs of provision include food and the harvested biomass

can be used as raw material for composting, ornamental purposes, etc. FW can also

contribute to biological and genetic diversity providing habitat for plants, microor-

ganisms, invertebrates, and vertebrates. Cultural ES include esthetic, educational,

and recreational benets [36, 37].

In the specic case of oating wetlands established in eutrophic lakes, they have

demonstrated that provide at least three ES such as water quality improvement,

nutrient recycling, and oxygen release. Dunne etal. [38] assessed the P removal by

a FW at large scale to treat water from a eutrophic lake. The cost of the FW perfor-

mance was also evaluated. FW removed 2.6 metric tons P/year mainly in particu-

late form. The highest removal was observed in the cold season but the performance

costs were constant in all seasons. The authors concluded that an operational

regime with low ow rates, in hot season, can increase the economic viability of

the process.

Bu and Xu [27] compared the performance of four different emergent plants

(Canna indica, Accords calamus, Cyperus alternifolius, and Vetiveria zizanioides)

in FW treating the water from a eutrophic river. They found that Canna was the

plant that exhibited better DO gradient distribution and higher DO levels compared

to the other three plants. Furthermore, as a result of creating a higher number of

aerobic microsites around their roots, this plant was also the one that showed a

higher percentage of nutrient removal. The harvested biomass from FWs can be

used as raw material to generate biogas through anaerobic digestion, which is an ES

that can be quantied. Biogas has been produced from Acorus calamus, Typha ori-

entalis, Pontederia cordata, Canna indica, Colocasia tonoimo, Thalia dealbata,

and Hydrocotyle vulgaris [39].

10.5 Final Remarks

Floating Wetlands have been applied to very different types of water/wastewater

ranging from storm water (low nutrient load), eutrophic lakes (medium to high

nutrient content) to efuents from agroindustrial activities (high nutrient load).

Their efciency depends on several factors, the following being the most

important:

• Quality of the water (initial concentration of pollutants)

• Emergent plant utilized and its rhizospheric microorganisms

• Irradiance and temperature

• Additional factors such as matrix and combination of plants

Currently, research could focus in experimental work carried out at pilot plant

level in situ, especially in the case of applications for the improvement of water

quality in eutrophic lakes. Likewise, there is the need to study more in depth the

provision of ecosystem services provided by Floating Wetlands.

E.J. Olguín and G. Sánchez-Galván

guarino@unisannio.it

303

Acknowledgments Authors acknowledge the technical assistance from Erik Gonzalez-Portela

and Francisco Javier Melo. They also thank the nancial support from the National Council of

Science and Technology (CONACYT) through the Project 215148, entitled “Improvement of the

water quality in polluted urban water bodies using Floating Wetlands.”

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10 Floating Wetlands fortheImprovement ofWater Quality andProvision…

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A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_11

Chapter 11

Green Aquaculture: Designing andDeveloping

Aquaculture Systems Integrated

withPhytoremediation Treatment Options

GuyR.Lanza, KeithM.Wilda, SusheraBunluesin,

andThanawanPanich-Pat

Abstract An increase in aquaculture for global food production has been one

response to the sharp reductions of the stocks of aquatic species used as a source for

traditional shing methods. Phytoremediation offers an environmentally compati-

ble approach that can be quickly integrated into existing aquaculture systems to

provide management of contaminants. The scenarios of Integrated Aquaculture–

Phytoremediation systems (IAPS) provided in this chapter are not intended to be all

inclusive but rather serve as selected examples of potential applications. Appropriate

IAPS will be highly site specic and will depend on local conditions including geo-

morphology, water sources, levels of ambient soil and water contamination, the

aquatic species under aquaculture, and the type of culture system used. The IAPS

design must provide a good balance that insures both the removal of excess nutrients

and other contaminants and an adequate supply of nutrients to support the growth of

“Can you help us get clean water?” (Woman in TayPhong village, Vietnam following the loss of

all of the sh in the village aquaculture pens in 2006 following a toxic spill in the Song Lan River.)

G.R. Lanza (*)

College of Environmental Science and Forestry, State University of New York,

Syracuse, NY 13210, USA

e-mail: glanza@esf.edu

K.M. Wilda

Blue Stream Aquaculture, Island Grown Initiative, WE Aquaculture, Edgartown, MA, USA

S. Bunluesin

WHO Country Ofce for Thailand, Permanent Secretary Building 3, Ministry of Public

Health, Nonthaburi 11000, Thailand

T. Panich-Pat

Faculty of Liberal Arts and Science, Department of Science, Cluster of Environmental

Science and Technology, Kasetsart University, Nakorn Pathom, Thailand

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the aquaculture products. IAPS can greatly enhance the global production of plant

and animal food particularly in developing countries with warmer climates and

highly diverse plant communities. IAPS that effectively removes snail-vectored

parasites (e.g., sh-borne zoonotic trematodes) are especially desirable because

snails are often cultured for food in aquaculture systems along with sh. The inclu-

sion of carnivorous plants (e.g., Utricularia sp.) in IAPS may offer one solution.

Utricularia sp. inhabiting wet soils and water are known to actively trap and

consume aquatic animals, and it may be possible to use carnivorous plants to remove

immature snails, snail eggs, miricidia, and cercariae as a treatment option in IAPS.

Keywords Phytoremediation • Aquaculture • Food security • Soil and water con-

taminants • Global food production • Snail-vectored diseases • S.E. Asia Aquaculture

11.1 Introduction

Increased demand on traditional global sheries coupled with ocean contamination,

altered patterns of aquatic species distribution due to climate change, El Nino

effects, and altered predator prey relationships has raised international concern

about sharp reductions in the stocks of available sh and other aquatic food species.

One response has been the increased development of aquaculture systems to pro-

duce animals and plants for food. A major challenge confronting the growth of

aquaculture activities is the increase in biological and chemical contamination asso-

ciated with aquaculture to produce food.

The primary aim of this chapter is to encourage the development and use of phy-

toremediation options appropriate for treating and improving the soil, sediment, and

water quality associated with aquaculture systems. Phytoremediation offers an

environmentally compatible approach that can be quickly integrated into existing

aquaculture systems and provide effective management of contaminants during

aquaculture operations.

Phytoremediation offers an excellent array of plant–microbe choices that can be

matched to a site-specic water quality problem in aquaculture. Matching the

appropriate plant or plant community to chemical and biological contaminants can

play a major role in conserving and protecting soil and water. Integrating phytore-

mediation options with various aquaculture systems can serve as a major tool to

achieve cost-effective, low energy treatments that can support sustainable aquacul-

ture production on a global scale.

The high diversity of plant species and their associated rhizoora, and the favor-

able climate and long growing seasons typical of semitropical and tropical regions

make phytoremediation an attractive and practical option in many developing

countries. Although this chapter will focus on sh aquaculture in tropical systems,

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many aspects of the basic design features presented here could be applied to facilities

in other climates and to the production of other organisms including crustaceans,

molluscs, reptiles, and plants.

The approach and selected scenarios of integrated aquaculture–phytoremediation

systems (IAPS) provided in this chapter are not intended to be all inclusive but

rather serve as selected examples of potential applications. Appropriate IAPS will

be highly site specic and will depend on local conditions including geomorphol-

ogy, water sources, levels of ambient soil and water contamination, the aquatic spe-

cies under aquaculture, and the type of culture system used. Sustainable IAPS will

require interdisciplinary collaboration between local farmers, agriculture and sher-

ies scientists, engineers, and government ofcials.

11.2 The Global Aquaculture Industry

Current aquaculture production is on the increase representing the fastest growing

sector in global livestock production [1]. The global aquaculture industry contrib-

uted 43% of all aquatic animal food for human consumption in 2007 (e.g., sh,

crustaceans, and molluscs, not including mammals, reptiles, and aquatic plants) and

is expected to grow further to meet the future demand. Freshwater ponds and tanks

were the source for 60% of the world aquaculture production in 2008 (56% by

value), despite only using 3% of the planet’s water. The rapid growth noted in the

production of carnivorous species including salmon, shrimp, and catsh has been

driven by globalizing trade initiatives and the positive economics of larger scale

intensive farming approaches [2].

The impact of climate change on future food supplies and global food security is

uncertain and sheries activity will undoubtedly face some effects inuencing the

sources of protein from sh and other aquatic species. Fish are an important source

of protein for a substantial proportion of the world’s population [3]. A portion of

150g of sh can provide about 50–60% of an adult’s daily protein requirements. In

2010, sh accounted for 16.7% of the global population’s intake of animal protein

and 6.5% of all protein consumed. Moreover, sh provided more than 2.9 billion

people with almost 20% of their intake of animal protein, and 4.3 billion people

with about 15% of such protein. Fish proteins can represent a crucial nutritional

component in some densely populated countries where total protein intake levels

may be low [4].

The current trend towards enhanced intensive systems with key monocultures

remains strong and, at least for the foreseeable future, will be a signicant contributor

to future supplies. Dependence on external feeds (including sh), water and energy

are key issues. Some new species will enter production and policies that support the

reduction of resource footprints and improve integration could lead to new develop-

ments as well as reversing the decline evident in some more traditional systems [2].

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11.3 Aquaculture inSoutheast Asia

Aquaculture in Asia represents over 80% of total global production [4], and the total

quantity of sh is projected to reach more than 95.6 million tons by 20,130 [5].

Southeast Asia has seen a signicant increase in aquaculture beginning in the 1990s [6].

Aquaculture represents a major component of the food security and overall econo-

mies of several countries in South East Asia [7] including Thailand, Vietnam, and

the Philippines. Thailand and Vietnam in particular are increasing their aquaculture

activities in response to both an increasing global market and local demand accom-

panied by a leveling off of the yield from capture sheries. The development of

freshwater aquaculture in the Philippines, associated environmental impacts, and

relevant environmental regulations and regulatory bodies was recently reviewed by

Legaspi etal. [8]. They described the complex relationship between aquaculture

and water quality and provided data from studies on Lake Mohicap to illustrate the

potential role of paleolimnology as a tool to help achieve a more ecologically sus-

tainable lake-based aquaculture in the Philippines.

Freshwater aquaculture in Thailand and Vietnam is mainly for domestic con-

sumption and provides a good protein source for local use and also bolsters local

food security. Small-scale freshwater aquaculture is currently providing the rural

poor with high quality protein food for local consumption. Brackish water aquacul-

ture can produce protable products for both in-country use and export from both

countries [9].

Freshwater aquaculture, mainly pond and rice-eld culture, has been practiced in

Thailand for more than 80 years. In 2003, total production from freshwater and

brackish water aquaculture in Thailand was approximately 320,000 and 450,000

tons, respectively. The main freshwater species cultured were Nile tilapia (Oreo-

chromis niloticus), hybrid catsh (Clarias macrocephalus X C. gariepinus), silver

barb (Barbodes gonionotus), giant river prawn (Macrobrachium rosenbergii), and

snakeskin gourami (Trichogaster pectoralis). The main brackish water-cultured

species were giant tiger prawn (Peneaus monodon ), whiteleg shrimp (Penaeus

vanamei), green mussel (Perna viridis), blood cockle (Anadara spp.), and oyster

(Crassostrea commercialis). At present, more than 50 freshwater aquatic species

have been cultured [9].

Nhan etal. [6] provides a detailed description of the general operation and basic

economics of integrated freshwater aquaculture, crop, and livestock production using

Integrated Agriculture–Aquaculture (IAA) farming systems in Vietnam. Madsen

etal. [10] has studied the freshwater snail populations in Asia and provides an excel-

lent description of the snail disease vectors including those found in integrated sh-

livestock ponds in common use by families in Vietnam [10–12]. The integrated

systems described by Dung etal. [12] consist of a garden (Vaun), a sh pond (Ao Ca),

and a cattle shed (Chuong) and are referred to as family VAC ponds (Fig. 11.1).

Recognizing the potential of aquaculture, since 1999, the Vietnamese govern-

ment promoted diversication in agriculture, aiming to reduce the share of rice to

the total agricultural output value while increasing the contribution of aquaculture

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to economic growth and poverty reduction [13–15]. In this context, stimulating

integration between sh, shrimp/prawn, fruit, livestock, and rice production on the

same farm, further referred to as integrated agriculture–aquaculture (IAA) systems,

is expected to contribute to agricultural diversication and enhance its sustainability. In

Vietnam, IAA-farming has been promoted through mass organizations such as the

Vietnam Gardening Association and Government Agricultural Extension Agencies [6].

An important characteristic of IAA-farming is the recycling of nutrients between

farm components [16, 17]. Through nutrient recycling, IAA-farming allows intensi-

cation of production and income, while reducing environmental impacts [18–20].

Intensive export-oriented Pangasius sp. culture in both cages and ponds is charac-

terized by large nutrient ows supported by the use of off-farm feeds and water

exchange making local nutrient recycling problematic [21–23]. Moreover, the

industrial scale of the business and its sensitivity to uctuations in global trade make

it risky and the domain of the resource-rich [24]. IAA-farming in contrast appears

to be a realizable approach for diversication of rice production whereby synergism

between on-farm components can be realized and whole system productivity opti-

mized rather than that of individual enterprises [18, 25].

The potential integration of farm components and attainable intensication lev-

els of IAA-systems are in part determined by the biophysical setting and the farm-

er’s aspirations and decisions [26, 27]. In Vietnam, the benets of traditional VAC

Fig. 11.1 Schematic representation of a VAC pond (Dung etal. [12], with permission of Elsevier)

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(garden-pond-livestock)-integrated systems [13–15] have been widely reported

and recent studies have investigated IAA commercial orchard and sh production

systems [6].

In the Mekong delta, freshwater IAA-farming is commonly practiced in the cen-

tral region, where soil and hydrological conditions are favorable for aquaculture.

Development agencies have tended to promote a rather standardized IAA-system

for the region in a “conventional, linear” approach (cited in [28]). Within the central

zone of the delta, however, different agro-ecologies exist and market opportunities

for farming inputs and outputs differ. In particular differences between rural and

peri-urban areas are likely and might be expected to have an impact on optimal

forms of IAA.In Northeast Thailand, Demaine etal. [29] found that location rela-

tive to urban centers was more important than agro-ecology in determining farmer

attitudes and any likelihood of intensication. Better market accessibility in peri-

urban areas and access to nutrients often stimulates intensication of aquaculture

compared with more rural areas [30], allowing IAA-farming to raise income and to

produce cheap food for urban consumers [18].

11.4 Potential Designs ofIntegrated Aquaculture–

Phytoremediation Systems

11.4.1 Chemical andBiological Contaminants

Sources of water used to supply aquaculture systems are often contaminated

with organic and inorganic contaminants and disease causing microorganisms. For

example, one major environmental challenge evident in many aquaculture systems

in Southeast Asia and other areas of the world is the presence of freshwater snails

that vector human and livestock diseases. Some aquaculture systems follow a poly-

culture approach that simultaneously produces sh, snails, and other aquatic food

species for human consumption. Snails that are intermediate hosts of sh-borne

zoonotic trematodes are of special concern in VAC ponds and other types of aqua-

culture systems. Dung etal. [12] provided an excellent description of the distribu-

tion of freshwater snails in family-based VAC ponds and associated water bodies

with special reference to the intermediate hosts of sh-borne zoonotic trematodes in

Nam Dinh Province, Vietnam.

In addition to the human and livestock health threats from aquatic species

infected with chemical contaminants, parasites, and other pathogens, reduced

marketability of shes, snails, and other aquaculture products harboring disease

organisms warrants control efforts to reduce contamination in aquaculture

systems.

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11.5 Aquaculture Systems inVietnam

Figure 11.1 provides a schematic representation of a VAC pond in Vietnam [12].

High levels of nutrient pollutants (e.g., nitrogen and phosphorus) from both external

and internal sources are of particular concern. Standard operations used to culture

aquatic plants and animals for human and livestock consumption typically contrib-

ute additional contaminants and pathogens as waste products through the use of

water contaminated with biological and chemical contaminants. These contami-

nants enter the aquatic food web supporting aquaculture operations and can accu-

mulate in sh, snails, and other aquatic food sources. The lower water quality that

results from the contamination contributes to both reduced overall yield of product

and increases the risk of contaminated product unt as human and livestock food

sources.

Figure 11.2 presents pond nutrient ows in an integrated aquaculture, crop, and

livestock system (IAA-farming) in the Mekong Delta of Vietnam [6]. Low, medium,

and high pig waste nutrient input to sh farming—fruit production systems are pre-

sented as examples. The main motivations for practicing IAA-farming included

increased income and food for home consumption from available farm resources

while reducing environmental impacts. Fear of conicts from the use of pesticides

was given as one reason that some farmers chose to not use aquaculture.

11.6 Protocol forIntegrating Phytoremediation

withAquaculture Systems

11.6.1 Assessing Water Quality

The rst step in the development of an effective integrated aquaculture–phytoreme-

diation system is a local water quality assessment. Major water quality problems

resulting from typical freshwater pond and stream aquaculture systems are listed in

Table 11.1. Increased total suspended and dissolved substances, increased biochem-

ical oxygen demand, dissolved oxygen depletion, and increased and excessive phy-

toplankton which can include toxic blooms are of particular concern. Table 11.1

also lists specic contaminants in typical aquaculture systems including unionized

ammonia, nitrates and nitrites, heavy metals/metalloids, (e.g., As, Al, Cd, Cu, Pb,

Zn), organics (malachite green, pesticides, algicides, herbicides, petroleum hydro-

carbons), and microbial pathogens and parasites (bacteria, viruses, protozoa, trema-

todes, cestodes, nematodes). Some of the parameters in Table 11.1 can be estimated

on-site using portable eld-testing kits while others will require lab testing off-site.

Paleolimnology can be used to characterize the water quality history of lakes and

medium to large ponds serving as aquaculture systems [8].

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11.7 Selecting Plant Species

Phytoremediation options designed to control and treat the identied contaminants

in aquaculture systems will also be very site specic and must be carefully planned

to accommodate the individual characteristics of a particular aquatic system. Design

parameters must allow for the integration of phytoremediation processes with the

basic operational schemes of common aquaculture systems. Native plants with a

relatively rapid growth rate and high biomass production are the most effective

Fig. 11.2 Pond nutrient ows in an integrated aquaculture, crop, and livestock system (IAA-

farming) in the Mekong Delta of Vietnam (Nhan etal. [6], with permission of Elsevier)

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candidates for the phytoextraction and phytostabilization of specic contaminants

common in aquaculture operations. Basic knowledge about plants and water quality

characteristics may be available from farmers and other local residents involved in

IAA activities.

Care must be taken to avoid competition between the plant and microbe com-

munities used to treat/remove contaminants and the processes required for cost-

efcient aquaculture operations. For example, livestock and crop wastes are

typically directed to aquaculture systems to fertilize the biological community

that provides food for sh, snails, and other aquatic herbivores under culture (see

Figs. 11.1 and 11.2). The integrated aquaculture–phytoremediation design must

provide a good balance that insures both the removal of excess nutrients and an

adequate supply of nutrients to support the growth of the aquaculture products.

Major considerations for an integrated and sustainable phytoremediation–aquaculture

system include the specic locale of the facility pond or river, climate/local weather

patterns, hydrology, general land use patterns in the surrounding area, and the

sources and types of major biological and chemical contaminants entering the water

and sediments.

Phytoremediation research in Southeast Asia has expanded considerably during

the past 15years beginning with the initial research completed at Mahidol University

and later at Burapha University in Bangkok and Bangsaen Thailand. There is a good

database of plant species available in the published literature describing plants used

for various applications of phytoremediation in developing countries (see for exam-

ple [31]).The database can be one good source of appropriate plant species for use

in designing integrated aquaculture–phytoremediation systems. Several basic fac-

tors should be considered in the process of plant selection. For example, plants used

in shoreline and inow/outow areas should be chosen on the basis of their growth

characteristics in different soils and sediments and their compatibility with other

plants in the community.

Table 11.1 Water quality problems and selected contaminants in aquaculture systems

Water quality problem Selected contaminants

Suspended and dissolved substances Inorganic and organic materials—TSS, TDS

Nutrient loadings Nitrates, nitrites, phosphorus, unionized ammonia

Oxygen depletion—Biochemical Oxygen

Demand (BOD)

Dissolved organics, sediment oxygen demand

Increased phytoplankton and toxic

blooms

Oxygen depletion, organic contaminants, microbial

toxins

Increased inorganic contaminants As, Al, Cd, Cu, Pb, Zn)

Increased organic contaminants Malachite green, pesticides, algicides, herbicides,

petroleum hydrocarbons

Microbial pathogens and parasites Bacteria, viruses, protozoa, trematodes, cestodes,

nematodes

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11.8 Design Parameters forIntegrated Aquaculture–

Phytoremediation Applications

The pond and river areas available for the application of phytoremediation options

to control and treat contaminants in aquaculture systems include (1) water supply-

ing the ponds through direct inputs from inow channels/canals and indirect inputs

from non-point source runoff, (2) sediments in the ponds and rivers, (3) bank areas

immediately surrounding the ponds and rivers, and (4) water exiting the pond

through outow channels/canals or downstream ow in rivers. Table 11.2 provides

selected examples of potential phytoremediation treatment options for aquaculture

systems experiencing common contaminants. Food security and water pollution are

of increasing concern, especially in developing countries. Biomass removed from

the aquaculture pond or river can be composted, used as fuel, or as food for humans

and livestock if the concentration of toxic contaminants is low enough.

11.9 Vegetative Filter Strips andNatural andConstructed

Wetlands

Vegetative lter strips (VFS) can be applied to areas immediately surrounding IAA/

VAC pond shoreline areas (see Figs. 11.1 and 11.2) and to the inow and outow

areas of the facility. The VFS plant community can be constructed using compatible

native plants that are known to be effective in the treatment of specic organic and

inorganic contaminant mixtures. In many cases, decorative plants including blue

ag iris and marigolds (e.g., Iris sp., Tagetes sp.) with good phytoremediation

potential can provide value-added benets to farmers as products sold to oral dealers.

Table 11.2 Examples of phytoremediation treatment options for aquaculture systems using

constructed communities of plants, algae, and bacteria

Contaminant removal system Plants

Vegetative lter strips (VFS) [32–35]

Pesticides Iris versicolor, Trypsacumdactyloides,

Andropogongerardii, Salix nigra

Petroleum hydrocarbons (TPH) Trifolium sp., Festua sp., Cynodon sp.

Heavy metals/metalloids Vetiveria sp., Chrysopogon sp., Chromolaena sp.,

Typha sp., Leersia sp., Tagetes sp., Acidosasa sp.

Natural and constructed wetlands [32, 36, 37]

BOD, TSS, nutrients, heavy metals/

metalloids, organics/malachite green,

coliform bacteria, parasites

Carex sp., Cyperus sp., Typha sp., Phragmites sp.,

Juncus sp., Rhizophora sp., Panicum sp., Leersia

sp.

Limnocorrals/cages/net pens/hydroponic rafts [38–44]

BOD, TSS, nutrients, organics/malachite

green, metals/metalloid

Lemna sp., Eichornia sp., Hydrilla sp.,

Ceratophyllum sp., C. indica

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Average phytoextraction coefficients for Typha sp.

total biomass

Values in parentheses indicate the mg of metal contaminant/kg dry weight of soil.

Values above the columns indicate the mg of metal contaminant/kg dry weight of plant

total biomass.

2.0

1.5

1.0

Phytoextraction coefficient

0.5

0.0 Cd (37.24) Cu (562.7)

(78.6)

(19.4)

(76.1)

(399)

coeff.value

Metal concentration (mg/kg)

Pb (382.1) Zn (686)

Fig. 11.3 Average phytoextraction coefcients for Typha sp. total biomass

Average phytoextraction coefficients for Leersia

sp. total biomass

Values in parentheses indicate the mg of metal contaminant/kg dry weight of soil.

Values above the columns indicate the mg of metal contaminant/kg dry weight

of plant total biomass.

Phytoextraction coefficient

Cd (37.24) Cu (562.7)

(381)

(101)

(95)

(2770)

coeff.value

Metal concentration (mg/kg)

Pb (382.1) Zn (686)

Fig. 11.4 Average phytoextraction coefcients for Leersia sp. total biomass

Natural and constructed wetlands can be used to compliment VFS communities

especially at the inow and outow areas of an aquaculture pond. Plants used in

VFS and/or constructed wetlands should be matched to soil or sediment types simi-

lar to their normal habitat. For example, Typha sp. grows best in wet, saturated soils

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while Leersia sp. favors moist to dry soils. Although erratic phytoextraction patterns

may occur over time, both plants can effectively remove small to moderate amounts

of heavy metals/metalloids, thus preventing the contaminants from entering the

aquaculture system and its food web.

Figures 11.3 and 11.4 provide examples of typical phytoextraction coefcients

seen with Typha sp. and Leersia sp. from an industrial area in the USA heavily con-

taminated with TPH, PCB, and several heavy metals [32]. Table 11.3 displays root

and shoot contaminant removal and Shoot/Root Quotients of Leersia sp. in test sedi-

ments and soils with varying organic content [32, 45]. Total or partial plant removal

can eliminate some of the contaminants using successive plantings over time.

11.10 Limnocorrals, Cages, Net Pens, andHydroponic Rafts

Treatment of contaminants in the pond, river basin, or canals can be accomplished

with plants housed in containment structures including limnocorrals, cages, net

pens, and hydroponic rafts. The site-specic characteristics of the aquaculture oper-

ation will determine which type or combination of containment structures is best

suited for integration with the aquaculture process. The interaction of different

contaminants (e.g., cadmium and zinc) and humic substances are important deter-

minants of contaminant behavior and removal [38] and should be considered in

designing a system. The specic absorption/adsorption characteristics of the plant

are also important considerations in the planning and design of integrated aquacul-

ture–phytoremediation systems.

In some cases, more than one type of containment structure can be used over

time. Caged oating plants could be used simultaneously with hydroponic rafts.

Floating Treatment Wetlands (FTW) with C. indica [39] could be used along with

caged Eichornia sp. or Lemna sp. Since all aquatic plants absorb contaminants and

then release them back to the environment when they die and decompose, the con-

tainment structure must be periodically removed, cleaned out, and repacked with

Table 11.3 Inorganic contaminant uptake by Leersia oryzoides in different soils

Contaminant (mg/kg) Soil type 1,2,3aRoots Shoots SRQ**

As 1 30 10 0.30

AS 2 140 140 1.00

Cd 3 86 16 0.18

Cu 3 370 11 0.30

PB 3 91 4 0.04

Zn 3 1770 1000 0.60

All soils had percent organic content by weight between 1 and 6. Shoot/Root Quotient Data from

Lanza [32], Amphia-Bonney etal. [45]

aSoil types 1=Standard loam, 2=Potting soil, 3=Wetland sediments

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fresh plants. For example, water hyacinth (Eichornia sp.) and duckweed (Lemna

sp.) listed in Table 11.2 absorb nutrients, heavy metals/metalloids, and other con-

taminants from water. In addition to inhibiting phytoplankton growth by competing

for nutrients, the plants remove toxic contaminants such as cadmium. Studies using

Lemna sp. collected from Rice City Pond (RCP) a cadmium-contaminated pond in

the USA (see Fig. 11.5) and an USEPA reference culture of Lemna sp. cultivated in

synthetic water (DNS) showed very good removal of cadmium after 2weeks of

culture. Using a “put and take” approach with the containment structures will pre-

vent the return of the contaminants to the pond or river by removing biomass before

death and decomposition.

11.11 Summary andFuture Research Needs

Food security and water pollution are of increasing global concern, especially in

developing countries. Methods to simultaneously augment food production and

decrease water pollution can be valuable additions to current aquaculture opera-

tions. IAPS offers a new approach to create sustainable aquaculture systems that can

provide green, low energy-low technology solutions in developing countries. If the

concentration of toxic contaminants is low enough, biomass removed from aqua-

culture ponds or rivers can be composted, used as fuel, or as food for humans and

livestock. Composting biomass can signicantly reduce the volume of plant

Cd uptake by Lemna in DNS (50

g/ICd)

* RCP

Plants

Day 1

P - Value Significant Difference

YES

0.478

0.016Day 15

Cd concentration (ug/g)

Mean Cd Concentration in Biomass

EPA plants

RCP plants

0.02

Control Stock Day

ay 15

1.30 1.32

18.49

44.66

Fig. 11.5 Cadmium uptake by Lemna sp.

11 Green Aquaculture: Designing andDeveloping Aquaculture Systems Integrated…

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320

material, but contaminated biomass would have to be safely disposed of in landlls

or other appropriate storage areas [46].

Typical biomass from plant material contains varying amounts of stored energy

as oxygenated hydrocarbons biomass. As a result, it can serve as a reliable source

of fuel if the amount produced merits collection and storage. The dry weight of

Brassica juncea used to phytoextract lead from soil produced 6 tons biomass per

hectare with 10–15,000mg/kg lead content [47]. The use of biomass for fuel may

be feasible as an augmentation to traditional solid fuels combusted under con-

trolled conditions that do not release contaminants to the atmosphere. In the case

of biomass use as food for livestock and humans, studies of nutrient removal by

Lemna sp. from two ponds in Brazil indicated that the ponds together produced

over 13 tons of biomass (68t/ha year of dry biomass), with 35% crude protein

content [36].

Using Integrated Aquaculture–Phytoremediation Systems (IAPS) can greatly

enhance the global production of plant and animal food particularly in developing

countries with warmer climates and highly diverse plant communities. Although

IAPS will remain site specic, additional research can clarify the most efcient

plant communities for many common types of aquaculture systems based on gen-

eral similarities in the aquatic products grown and the waste types and loadings

typically used to fertilize the food web supporting production. Information on IAPS

that create a balance between the livestock fertilization supporting good aquatic

product growth and excess fertilization that leads to undesirable water quality that

impedes good aquatic product growth will be very useful.

The presence of various microbial pathogens and parasites present a major chal-

lenge to sustainable IAPS and aquaculture systems in general. Additional research

is needed to develop IAPS that provide the effective removal of disease causing

organisms common in aquatic systems used for aquaculture. One good example is

provided by sh-borne zoonotic trematodes (FZT). Current research indicates that

sh-borne zoonotic trematodes FZT such as Clonorchis sinensis, Opistorchis

viverini (Opisthorchiidae), and intestinal trematodes of the family Heterophyidae,

constitute a public health hazard in Vietnam. These parasites have been linked to

consumption of raw or undercooked sh from aquaculture [11]. The FZT transmis-

sion pathways, however, are more complicated than just the presence of intermedi-

ate snail hosts in aquaculture ponds as ponds may exchange water with surrounding

habitats such as rice elds and irrigation canals (see Fig. 11.1), and these surround-

ing habitats may be a source of snails and cercariae and contribute to FZT infection

in cultured sh [11].

The fact that snails are often harvested as food from aquaculture ponds and rivers

complicates the problem of FZT.Research is needed to clarify the possible inclu-

sion of carnivorous plants in phytoremediation communities used in IAPS (see

Table 11.2). Plants in the bladderwort group (e.g., Utricularia sp.) inhabiting wet

soils and water are known to actively trap and consume aquatic animals including

mosquito larvae and tadpoles [48]. It may be possible to use bladderworts to remove

immature snails, snail eggs, miricidia, and cercariae as a treatment option in IAPS.

IAPS may contribute to providing a holistic approach to deal with all stages of the

FZT transmission cycle.

G.R. Lanza et al.

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Special Applications of Phytoremediation

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Chapter 12

Modelling Phytoremediation: Concepts,

Models, andApproaches

EditaBaltrėnaitė, PranasBaltrėnas, andArvydasLietuvninkas

Abstract In the phytoremediation modelling stage, which is specic due to

unavoidable assumptions and limitations, the complicated nature of natural pro-

cesses, and different qualications of model developers result in the variety of

phytoremediation- oriented models that differs in complication and the extent of

applicability. The variety of phytoremediation models is not only naturally under-

standable, but also serves specicity of model application. In other words, the

choice of a model and the need for detailed result depend on the prospects of the

model use, e.g., for preliminary assessment of the phytoremediation effect, phytore-

mediation cost estimation or contaminant distribution among the plant compart-

ments. This chapter discusses the prospects of application of the phytoremediation

assessment tools, such as Phyto-DSS, BALANS, Dynamic factor method, and Hung

and Mackay model used for simulating the contaminant transfer processes in the

soil–plant–atmosphere system.

Keywords Phytoremediation • Modelling • Dynamic factors • BALANS • Phyto- DSS

• Uptake • Chemical element

12.1 Introduction

The topic of phytoremediation, though rather new, e.g., interest in it peaked in the

EU in 2002–2007, has already reached the stage of modelling of biogeochemical

processes for this purpose. In the modelling stage, which is specic due to unavoid-

able assumptions and limitations, the complicated nature of natural processes, and

different qualications of model developers result in the variety of phytoremediation-

oriented models that differs in complication and the extent of applicability. On the

other hand, the variety of models is not only naturally understandable, but also

serves specicity of model application. In other words, the choice of a model and

the need for detailed result depend on the prospects of the model use, e.g., for pre-

liminary assessment of the phytoremediation effect, phytoremediation cost

E. Baltrėnaitė (*) • P. Baltrėnas • A. Lietuvninkas

Vilnius Gediminas Technical University, Saulėtekio al. 11, Vilnius 10223, Lithuania

e-mail: baltrenaite@aol.com

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_12

guarino@unisannio.it

328

estimation or contaminant distribution among the plant compartments. This chapter

discusses the prospects of application of models used for modelling the metal transfer

processes in the soil–plant–atmosphere system.

12.2 Different Prospects ofUsing Phytoremediation Models

The selection of a model for phytoremediation depends on the objectives of the

operator, the tasks for applying the model, the advantages and limitations of the

model, and the comprehension level of the result. Different models can be used by

operators that represent particular professions, e.g., environmental researcher, envi-

ronmental ofcer, environmental consultant, and owner of a contaminated site.

There are different prospects for use of a particular phytoremediation model. For

instance, an environmental researcher might be focused on chemical element con-

centration in different plant compartments using a model that involves a detailed

mechanism of chemical element uptake, translocation, and bioaccumulation mecha-

nisms. Changes in the phytoremediation-related processes after soil treatment in

comparison to the background situation are an important feature in understanding of

process mechanisms. A food safety specialist might need more detailed simulation

mechanisms to forecast contaminant concentration in plant compartments used for

forage and the risk of contaminant entering the food chain. A soil researcher might

be interested in the effect of soil chemical element concentration changes on the

chemical element concentration in the plant [1]. An environmental ofcer might be

looking for a model that would assist in assessment of environmental risk posed by

contaminated sites before and after the use of phytotechnologies. A consultant on

environmental matters might need to estimate the time required to extract chemical

elements till the target concentrations in soil to evaluate the phytoremediation

techniques compared to a range of other soil restoration options or to estimate the

costs of phytomediation activities. An owner of a contaminated site is usually inter-

ested in decontamination costs and the efciency of the technique based on a more

spatial/regional level, and for this reason the model must be robust and estimating the

macro-effects of the site. Models that integrate a soil amendment effect on the

contaminant background level in soil, the effect of uptake of a contaminant by plants,

and its removal by leaching in soil can assist in dening the critical loads of contaminants

with soil amendments and justifying the environmental regulations.

12.3 Detailed Models vs. Robust Models

In selecting a phytoremediation model, it is important to make a clear distinction

between steady-state and dynamic models. The steady-state models provide a prog-

nosis for innite time, while dynamic methods are more useful in predicting a time

period before the modelled variable reaches the target value. Depending on the aim,

E. Baltrėnaitė et al.

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one must choose a model with an appropriate level of detail in describing the

processes. A disadvantage of relatively complex mechanical models is that input

data for their application on a regional level is generally incomplete and values can

only be roughly estimated. Even if the model structure is correct (or at least ade-

quately represents current knowledge), the uncertainty of the output of complex

models may thus still be large because of the uncertainty of input data. Simpler

empirical models have the advantage of a smaller need for input data, but the theo-

retical basis, that is needed to establish condence in predictions, is small, which

limits the application of such models for different situations. Therefore, there is a

trade-off between the model complexity (reliability) and regional application [2].

When the aim of a model is to evaluate the phytoremediation on the regional scale

consisting of the receptors with different properties, it seems most rational to use a

relatively simple model with an aggregated description of processes in the total con-

sidered compartment. In choosing a model, one should be aware of the consequences

of simplications, such as ignoring certain processes (complexation, chemical

element cycling, etc.) and making certain assumptions (steady-state, homogenous

mixing, equilibrium partitioning) [2]. In order to gain insight into the consequences

of the choice of a certain model and limits, one could perform modelling with differ-

ent models using various limits and compare the results. In such a way, one also gets

insight into (1) the differences in vulnerability of various environmental compart-

ments and (2) the relevance of the different processes in the systems and of the

different ways of parameterizing certain processes [2].

12.4 Examples ofChemical Element Uptake Simulation

Methods: Advantages andLimitations

A group of models that are described below are the examples of different modelling

approaches towards chemical element uptake by plants in relation to the phytoreme-

diation effect. The scope and a principal process of modelling, as well as advantages

and limitations of the models are discussed.

12.4.1 Plant-Oriented Models

This program was named as a decision support system (DSS) and was intended for

predicting the effect of phytoextraction (phytoremediation) on soil metal concentra-

tion and distribution, as well as the economic feasibility of the process in compari-

son to other ways of land reclamation, i.e., natural attenuation (or inaction) or the

best alternative technology. The model is based on the transfer of metals through

transpiration ow, and therefore water transfer through a plant and the solubility of

metals in water is a basis for the algorithm for the model. Within the model, the

focus is given to the soil metal concentration and the changes of it are predicted by

12 Modelling Phytoremediation: Concepts, Models, andApproaches

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modelling the plant water use, taking into account the metal concentration in soil

solution, soil density, plant root distribution, and introduced root-absorption factor.

The root-absorption factor represents the metal concentration quotient within the

system “root xylem/soil solution” and lumps up the number of complex process

factors that inuence the uptake of metals by plant roots [3].

The model accepts a precondition that the potential uptake of chemical substance

i through the plant of certain species n depends on the density of the root system. To

enter the plant, the concentration (mg/kg) of chemical substance i at the depth d of

soil changes, and this variation is calculated according to Eq. 12.1:

Drj

TC dt

z,=ò

(12.1)

where ΔMz is variations in the concentration (mg/kg) of chemical substance i at the

depth of soil z; ρz is bulk density (g/cm2) of soil at depth z; t is time, days; Rz is a part

of the mass of plant roots (mass of the root system at the depth of soil z divided by

the general mass of the root system).

Among the model limitations, there are several ones including the following.

Environmental conditions, such as drought, that may limit the plant growth, are not

considered in the model. Root absorption factor values may vary because the metal

uptake by a particular species would be different in different soil types. The model

does not consider the fact that metals from the aerial parts may be further translo-

cated within the phloem back to the below-ground compartments [4]. The model

considers the uniform moisture distribution in the soil, and these conditions are not

typical in a real situation. Moreover, the depth-related distribution of ne roots is

not equal to the total uptake of metals by roots. This causes the remediation time to

be shorter than actual. Using Phyto-DSS model, the precise initial data is required

to forecast future biomass growth and metal accumulation, such as root develop-

ment, metal and water uptake rates. The climate conditions must be considered very

precisely because the model is sensitive to changes in water regime (rainfall, evapo-

transpiration). The model mainly considers metal accumulation in the above-ground

plant biomass as the principal feature for phytoremediation. It also provides no

possibility to simulate mixed plantations, which is usually a practical consideration

in many phytoremediation projects.

Later Liang et al. [5] has modied the phytoremediation evaluation approach

(Eq. 12.1) by incorporating iterative concepts (Eq. 12.2):

tMM

dAhC C

×=×××× -

()

=¼

1000 012 1

sf ,,,, ,

(12.2)

where t is the cycle number needed for phytoremediation (cycles), i is the cycle

number used for phytoremediation, Mi is the initial metal concentration in soil (g/ha),

Mf is the target metal concentration in soil (g/ha), Mp is the biomass production of

plant species (tons/ha/cycle), ds is the soil density (kg/m3), A is the contamination

E. Baltrėnaitė et al.

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area (m2), h is the soil depth with the metal content (m), Csi is the metal concentration

in soil after the ith harvest cycle (mg/kg), and Csf is the nal metal concentration in

soil (mg/kg).

The transfer of contaminants in the “soil–plant–atmosphere” system was also

modelled by Trapp and McFarlane [6], who used the PLANTX model for this pur-

pose. This model considers: (a) the dynamic transfer of contaminants to plants

from the soil solution and air; (b) metabolism of anthropogenic contaminants and

their accumulation in the roots, stem, leaves, and fruits of the plant. The model is

based on the processes of contaminants’ diffusion in the soil solution and the soil

pores lled with air and roots as well as their transfer to roots by the transpiration

ow, the exchange between the ambient air and leaves through their stomata due to

diffusion, and metabolism and distribution of contaminants due to the growth of

plants. The model can be used to predict contaminants’ concentration in plants;

however, it is intended for considering organic contaminants. Boersma etal. [7]

integrated the processes of contaminants’ transport and transformation into one

mathematical model, CTSPAC, combining two submodels representing soil and

plant, while Ouyang [8] discussed the problem of applying the model to 1,4-diox-

ane transfer to poplars.

Guala etal. [9] focused on the model of metal transfer from soil to plants based

on the mechanism of physiological absorption of plants, which aims to assess the

effect produced by it on their growth. This model is limited to the evaluation of Cd

and Ni concentrations in the soil and their effect on two cereal crops (rice—Lolium-

perenne L. and oats—Avenasativa L.). The model is based on the physiological

mechanisms of metals’ transfer and physiological characteristics of plants described

by Moreno etal. [10]. The model refers to the main relationship between the bio-

available form of metal in soil (A) and metal concentration in a tree (S):

AcfS

eS S=-+

()

÷×

(12.3)

where A is the available concentration of metal Mn+ (mg/L) in the soil solution; S is

the metal concentration in trees (mg/kg); α is the coefcient of absorption (L/kg/

year); and c/α, f/α, and e can be tted by experimental results in order to establish

the relationship between A and S.

The model does not involve aerogenic uptake nor the ux of aerogenic contami-

nants and is focused on inorganic contaminants. As mentioned above, it requires

experimental data to dene the values of coefcients c/α, f/α, and e.

The translocation of metals from the soil to a tree embraces various chemical,

physical, and biological processes (e.g., diffusion of metals in the soil and trees as

well as in the area of the roots and the tree itself, adsorption and absorption in the

tree and soil, the growth of the tree, and transpiration). Due to the complicated char-

acter of these processes and the action of the external factors (e.g., the climatic

conditions, type of the substrate), it is hardly possible to accurately describe them

by mathematical equations.

12 Modelling Phytoremediation: Concepts, Models, andApproaches

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Therefore, the models which can be found in the literature describing the transfer

of contaminants from soil to trees are rather simplied descriptions, usually based

on a few transfer coefcients and the evaluation of the transfer of contaminants only

to the whole tree organism. To create more complicated models of contaminants’

translocation, huge computing resources, as well as great amounts of experimental

data, allowing for validating the accuracy of the developed models, are required.

The other model, representing the uptake of chemical elements from sludge

amended soil which is based on the simplied model of the contaminant’s transfer

from soil to plants offered by Hung and Muckay [11], is described. This model can

evaluate the transfer of contaminants from soil to trees through the stem, then from

the stem to leaves, from the leaves to the air, from the air to the leaves, from the

leaves to the stem, and from the stem to the tree roots. In this model, uptake of con-

taminants from soil to plants is based on the equilibrium factors, describing the

distribution of contaminants in various media (e.g., soil, water, transpiration ow),

as well as on the rate of metabolism and time of diffusion. The time of the growth

of the main plant organs is also evaluated.

Modelling the uptake of metals to plants is a more complicated problem than

that associated with the transfer of organic contaminants (e.g., benzopyrene,

naphthalene) because, unlike the organic contaminants, metals are signicant for

physiological processes occurring in plants, such as their growth, metabolism,

and fermentation [12, 13].

Therefore, not only the factor of distribution of metals’ concentrations between

octanol and water (Kow) is important for the accumulation of metals in plants. In this

case, the coefcients of biological uptake of metals by plants and their translocation

are also required for modelling.

The main Eqs. 12.4–12.6 providing for the concentrations of contaminants in

leaves (Ci

l), stem ( Cst

i), and roots ( Ci

r), which are used in the present model, are

provided below.

CBBBK C

BKC

ier rs sl lw

al lw

=××××

+××

çö

÷×

(12.4)

CBBKC

BBKC

ier rs stw

al ls stw

aw st

=×× ×+×× ×

çö

÷×

(12.5)

CBK C

BBBK C

ier rw

al ls sr rw

=××

+××××

çö

÷×

(12.6)

where Ber is uptake ratio of the contaminant from soil to roots; Brs is uptake ratio of

the contaminant from roots to stem; Bsl is uptake ratio of the contaminant from stem

to leaves; Bal is uptake ratio of the contaminant from air to leaves; Bls is uptake ratio

of the contaminant from leaves to stem; Bsr is uptake ratio of the contaminant from

stem to roots; Сs is concentration of the contaminant in the soil; Сa is concentration

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of the contaminant in the air; Kew is equilibrium partition coefcient of the contaminant

between soil and water; Klw is equilibrium partition coefcient of the contaminant

between leaves and water; Kstw is equilibrium partition coefcient of the contami-

nant between stem and water; Kaw is equilibrium partition coefcient of the con-

taminant between air and water; Krw is equilibrium partition coefcient of the

contaminant between roots and water; Mi is molecular weight of the contaminant;

ρl, ρst, ρr are density of plant leaves, stem, and roots, respectively.

The parameters involved in the model by Hung and Mackay [11] are grouped in

several groups:

• Growth and metabolism: growth rate of leaves, stem, and roots; metabolism rate

of leaves, stem, and roots; duration of plant exposure to metals

• Morphological data of plant compartment: leaf surface area

• Metal characteristics: molar mass of metals, aerogenic concentration of metals,

metal uptake/translocation factors within environmental and plant compart-

ments: air/leaves, leaves/stem, leaves/stem, stem/roots, roots/stem, soil/roots;

total metal concentration in soil

• Physico-chemical data of plant: transpiration rate, density of leaves, stem, and roots

The processes considered in the presented model are not associated with the

specic features of contaminants. For example, the solution of a contaminant in the

plant has not been modelled; therefore, this model can be adapted to modelling

translocation of contaminants from the soil to a tree. In order to adapt the considered

model to modelling the translocation of metals, it was extended to include [12]:

• The factor characterizing the distribution of metal concentrations between the

octanol and water (Kow)

• The factor characterizing the distribution of metal concentrations between the

soil and water (Kd), which depends on the soil pH and the amount of the organic

material

• The coefcient of metal concentration in water (KT)

• The correction factors

12.4.2 Soil-Oriented Model

The model BALANS was developed in the Tomsk State University (Russia) by Prof.

Arvydas Lietuvninkas [14]. The model was developed to simulate the self- purication

of soil from metals that enter the soil in different ways (aerogenic metal deposition,

edaphic metals, and metal entering the soil with amendment). The considered

self-purication processes of the soil include: (1) metal uptake and removal with the

harvest and (2) the removal of metals from the soil as the result (consequence) of their

natural physical–chemical migration (Fig. 12.1) [15].

The deposition of aerogenic metals on the investigated territory is estimated

based on the values of the total load of aerogenic aerosols in the form of snow

12 Modelling Phytoremediation: Concepts, Models, andApproaches

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mineral dust and metal concentration in them [16]. The amount of metals bioaccu-

mulated in the harvested crops is estimated based on biomass removed with crops

and metal concentration in it. The natural removal of metals as a result of their

physical–chemical leaching in the soil is estimated using empirical annual migra-

tion coefcients, corresponding to the respective geographical climatic zone and

type of soil. The annual coefcient values of soil metal leaching lump up the com-

plex soil mechanism affecting metal dissolution, sorption, complexation, migration,

precipitation, occlusion, diffusion into minerals, binding by organic substances,

absorption and sorption by microbiota, and volatilization. This largely facilitates

simulation process being more practically attractive. The program uses annual metal

leaching coefcients for Cu, Zn, Pb, Co, Ni, Cr, V, and Mo chosen by default from

numerous literature sources, which correspond to the types of soil, such as

Albeluvisols and Podzols, characteristic of southern taiga and mixed forest zones

[17]. The model provides an option for the operator to change the coefcient values

based on changed simulation conditions.

The advantages of the model include: (a) differentiation between the removal of

metals by the process of natural physical–chemical migration and together with

harvested crops; (b) determining the balance of metals in the “soil–plant–atmo-

sphere” system in the long and short term, which is of particular importance in the

Fig. 12.1 The model BALANS is based on the systemic concept of getting of metals into the soil

of the area sown with wheat together with amendments and their removal from it. CAE

i is the

concentration of metal i in mineral dust of aerogenic origin; Ci

is the concentration of metal i in

the amendment; Ci

is the amount of metal i removed due to natural migration; Ci

is the remain-

ing amount of metal i in the amended soil; Ci

g is the concentration of metal i in the crop grains;

Cst

i is the concentration of metal i in the crop straws

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context of sustainable development; (c) evaluating the effect of various ways of

metals’ uptake by the soil (through aerogenic sources and with amendments) and

their removal (due to natural migration and with the harvested crops) on the balance

of metals in the soil; (d) the integration of the principles of acropetal and basipetal

distribution of metals in the plant into the model of their translocation in the “soil–

plant–atmosphere” system; (e) the inuence of complicated processes affecting the

system is expressed by the coefcients of annual contaminants’ transfer, thereby

creating conditions for effective practical use of the model.

The method takes into account the microrelief of the site and the load of aero-

genic contaminants, on which the removal and accumulation of metals in the soil

depend [14]. A schematic diagram of self-purication of the soil from metals and

their accumulation in the soils, depending on the elds’ microrelief, is provided in

Fig. 12.2.

The aerogenic load of mineral dust is assumed to be invariable (straight line 1).

On the autonomous geochemical landscape basis, the soil metal leaching is highly

controlled by the type of relief. The metal leaching is much more intense in the

case of autonomous geochemical landscape (a positive type of microrelief, i.e., a

strongly pronounced hill) and less intense in the case of superaquatic geochemical

landscape (a negative type of microrelief, i.e., strongly pronounced lowland) as

expressed by curve 2 (not taking into account the extent of the process) (Fig. 12.3).

Nonuniform removal of metals and the polluted soil manifests itself in their different

concentration. Thus, the soil of the positive microrelief type accumulates a consid-

erably smaller amount of these materials than the soil of the negative microrelief

type (Fig. 12.3, curve 3), and the self-purication of soil is more intense in this case.

In real conditions, the above difference is even more affected by the geochemical

characteristics of the particular metal and conditions of the site of investigation

(e.g., edaphic conditions, including the type and properties of the soil and the metal

pollution load).

abcde

Microrelief

Fig. 12.2 A schematic diagram of the BALANS model describing the removal of metals from the

soil (2) and their concentrations in it (3) at the constant load of mineral dust (1): I is the level of

these parameters; the type of the eld’s microrelief: a is a strongly pronounced hill; b is a weakly

pronounced hill; c is a at surface; d is weakly pronounced lowland; e is strongly pronounced

lowland. Note: the cases b and d are not included in the analysis of the investigation results

12 Modelling Phytoremediation: Concepts, Models, andApproaches

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12.4.3 Uptake-Process-Oriented Models

A widely known biological absorption of chemical element is based on the relation-

ship between a plant and the soil. Consequently, it is expressed as the relationship

between the concentration of chemical elements in a plant and the soil, in which it

grows. This relationship is used to determine the ability of plants to accumulate

chemical elements (when its value is more than one, plants are called the accumula-

tors, meanwhile when the relationship value is equal to or is about one, they are

referred to as indicators, and when this value is smaller than one, the plants are

called “excluders”). The considered relationship is also used for identifying the

harmful effect of metals and for determining a risk posed by them to the biota [18].

Mingorance etal. [19] used the term “enrichment factor” to compare the concentra-

tions of metals and other chemical elements in the investigated soil or plant to those

in the control objects.

However, the above-mentioned factors/coefcients, expressing the concentration

of chemical elements in plants compared to that in the soil, have some drawbacks.

From biogeochemical perspective, they reect the comparison of chemical element

concentration in various media (a plant and soil), but this refers only to a particular

area and to particular environmental conditions characteristic of this area at a par-

ticular period of time (e.g., 10 years after the sludge was spread over the soil) [20].

First, from the biogeochemical point of view, the comparison of various plants

based on the considered factors/coefcients could hardly be accurate because these

plants could have been growing in different conditions, different types of soil, and

elementary landscape, which could result in different mobility and accumulation of

chemical elements in them. Second, it is required to compare not only the concen-

trations of chemical elements in plants to their concentrations in soil or the control

plant, but to compare the differences in the process of chemical element uptake and

its intensity with respect to the control case. The more so because in evaluating an

uptake process we should compare processes rather than concentrations. Third, the

numerical value of the relationship between the uptake of chemical elements to the

Fig. 12.3 Simulation approaches and their application for sites on two consideration levels, for a

site with existing contamination, and for a site with potential contamination

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investigated plant and to the control one, which could facilitate the evaluating of the

variation of chemical element transfer, is lacking. However, a method based on

dynamic factors can provide it. Fourth, it is important that the effect of natural

processes, inuencing the uptake of chemical elements, should be integrated into

the estimate. These purposes could be achieved by introducing the higher order fac-

tors. They are calculated by comparing the value of chemical element uptake factor

obtained for the investigated territory to the respective value for the control territory.

The authors suggested these factors for describing ve types of chemical element

behavior in the soil–plant system, depending on the changes taking place in the soil

(Eqs. 12.7–12.11). They are referred to as dynamic factors because of their sensitivity

to the changes in variables involved in calculations [13, 21, 22].

The dynamic factor of bioaccumulation

dyn

(12.7)

where Ci

_ is the concentration of chemical element i in the plant ashes on the

treated territory, mg/kg; Ci

_ is the concentration of chemical element i in the treated

soil, mg/kg DW; Ci

_ is the concentration of chemical element i in the control soil,

mg/kg DW; Ci

_ is the concentration of chemical element i in the plant ashes on the

control territory, mg/kg.

The dynamic factor of biophilicity

BF C

dyn

(12.8)

where Ci

_ is the concentrations of chemical element i in the plant biomass on the

treated site, mg/kg DW; Ci

_ is the concentration of chemical element i in the plant

biomass on the control site, mg/kg DW.

The dynamic factor of translocation

dyn

(12.9)

where Ci

_ is the concentration of chemical element i in the vegetative organs of

the plant growing on the treated territory, mg/kg DW; Ci

_ is the concentration of

chemical element i in the roots of the plant growing on the treated territory, mg/kg

DW; Ci

_ is the concentration of chemical element i in the roots of the plant growing

on the control territory, mg/kg DW; Ci

_ is the concentration of chemical element i

in the vegetative organs of the plant growing on the control territory, mg/kg DW.

The dynamic phytoremediation factor

FR CBC

CBC

dyn

=´´ ´

´´ ´

bt ts

bc cs

(12.10)

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where Bt and Bc denote the annual growth of a tree on the control and investigated

territories, kg/ha; ρc and ρt are the soil density on the control and the investigated

territory, respectively, g/cm3.

The dynamic factor bioavailability

BIO

dyn

tot

bioavt c

tbioav c

(12.11)

where Ci

bioa

_ is the concentration of chemical element i in the soil solution on the

treated site, mg/kg; Ctot

_t is the total concentration of chemical element i in the soil

on the treated site, mg/kg; Ci

bioa

_ is the concentration of chemical element i in the

soil solution on the control site, mg/kg; Ctot

_c is the total concentration of chemical

element i in the soil on the control site, mg/kg.

In addition to their biogeochemical signicance, the dynamic factors have a

number of advantages in practical application:

(a) They integrate the information of four various types by combining the data

about the amount of chemical elements in two media (or the plant organs) and

the data on the control and the polluted (treated) territory into a single value,

thereby facilitating the evaluation of chemical element transfer.

(b) They are nondimensional and, therefore, easy to compare.

the reliability of the obtained results and the quality of evaluation.

The principal input and output data used in the selected models are provided in

Table 12.1.

12.5 Application Types ofPhytoremediation Models

The quantitative phytoremediation methods could be classied according to their

application eld and characteristics. According to the element uptake evaluation

level, models can be classied as those providing preliminary or screening results

(e.g., method of dynamic factors) or more detailed information (e.g., Hung and Mackay

Table 12.1 Minimal input data requested by a particular model

Model Input data Output information

Hung and

Mackay

• Molar mass of contaminant • Concentration of contaminant in a

particular plant compartment (leaves,

stem, roots) after a specied period

of time (years) (e.g., Baltrėnaitė and

Butkus [12])

• Contaminant concentration in soil

solution

• Total concentration of contaminant in

soil

• Soil density

(continued)

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Table 12.1 (continued)

Model Input data Output information

BALANS • Metal concentration in soil,

aerogenic dust, soil amendment

(if applicable), grains, and straw

(if crop is chosen as plant)

• Amount of metals leached down to

soil

• Area of a site of investigation • Amount of metals removed from soil

with crops

• Soil density • The half-life period of metals in soil

• Amount of amendment (if applicable) • See example at Baltrėnaitė etal. in

review [15]

• Annual plant yield

• Annual biomass increment (in case of

crops, annual biomass of straw and a

portion of straw mass removed from

the site are entered)

• Type of relief

• Soil metal leaching coefcients (soil

type is taken into account when

choosing)

Phyto-

DSS

• Climate data (amount of rainfall and

evapotranspiration)

• Content of metals in plant biomass

after the selected time period of

phytoremediation

• Area of the site • Increment of plant biomass after the

simulated period

• Soil type, density, and soil moisture

content

• Costs of phytoremediation

• Soil rain inltration rate

• Total and bioavailable metal

concentration (background and

maximum contamination) in soil

• Metal concentration in plant biomass

• Root absorption factor

• Metal threshold concentration for

plants

Regular

and

dynamic

factors

• Total and bioavailable soil

contaminant concentration in soil of

both control and treated sites

• Factors indicating the changes in

contaminant bioaccumulation,

biophilicity, translocation,

bioavailability in soil,

phytoremediation process changes

(e.g., Baltrėnaitė etal. [20])

• Total concentration of contaminant in

plant (at least in roots and shoots) in

both control and treated sites

• Soil density

• Annual increment biomass of plants

in both control and treated sites

model, Phyto-DSS). Such differentiation of models also denes the complexity of a

model and, thus, the usability. Easy-to-use models are more practically attractive

and in most cases are much welcome by the ofcers of environmental protection,

while more complex models are better tools for researchers or others specialized in

12 Modelling Phytoremediation: Concepts, Models, andApproaches

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the eld. In regard to the spatial/regional application of the modelled results, models

can be focused on providing the results of phytoremediation effect for a single plant

(e.g., Hung and Mackay model) or a simulation might be already developed to

produce up-scaled results (e.g., BALANS).

Phytoremediation effect itself can be simulated of a different scope, e.g., generating

information on the amount of the removed contaminants (e.g., Phyto-DSS), residual

portion of contaminants in soil (e.g., BALANS) or assessment of contaminant

uptake process changes (e.g., dynamic factors). The variety of model scopes denes

the uptake mechanism involved in the model. For example, simulation of the

transpiration ow in plant is the principal mechanism of Phyto-DSS that is more

oriented to the plant organism, while the BALANS model is based on the estimation

of mass balance of contaminant in soil. Sometimes, the scope of a model may refer

to the contaminant uptake estimation of the present situation or dening the plausible

situation in the future. In such a way, models can be used for evaluation of the

contaminant uptake in the site with existing contamination or used to predict the

phytoremediation effect when potential contamination occurs (Fig. 12.3).

12.6 Conclusions

A model is bad unless it is validated. Each model has a particular range of applica-

tions and is oriented to solving a special task. Models for phytoremediation process

simulation vary depending on the scope, the prevailing simulation mechanism, spa-

tial applicability of simulation results and the level of complexity. Within the plant–

soil–atmosphere system, the models are classied as plant-oriented, soil-oriented or

uptake-process-oriented. In regard to the mechanisms involved in simulation, the

models are based on contaminant transport with a transpiration ow, mass balance

of the contaminants in the soil–plant–atmosphere system.

References

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12 Modelling Phytoremediation: Concepts, Models, andApproaches

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343© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_13

Chapter 13

Genetic Control ofMetal Sequestration

inHyper-Accumulator Plants

ShahidaShaheen, QaisarMahmood, MahnoorAsif, andRaqAhmad

Abstract Heavy metal contamination is an emergent environmental dilemma all

over the world, posing serious threat to environment as well as human being by

disturbing the ecological balance. There are a number of physical, chemical, and

biological techniques applicable worldwide for wastewater treatment, but the phy-

toremediation techniques are the green, sustainable, and promising solutions to

problem of environmental contamination. Studies revealed that there are certain

hyper-accumulator genes present in plants, which make them more metal tolerant

than non-hyper-accumulator plants species where those genes are absent. In addition,

hyper-accumulator plants tackle with heavy metals by activating their responsive

genes for chelation, trafcking, and sequestration. Therefore, studying such hyper-

accumulator genes opens a gateway for the thorough understanding of phytoreme-

diation techniques.

Keywords Hyper-accumulator • Non-hyper-accumulator • Phytoremediation •

Contamination • Tolerant

Abbreviations

Al Aluminum

BjMT Brassica juncea metallothioneins

Ca Calcium

CaM Calmodulin

CBL Calcineurin B-like protein

CIPK Calcium-interacting protein kinase

CRKs Cysteine-rich receptor-like kinases

DHAR Dehydroascorbate reductase

DNA Deoxyribonucleic acid

GR Glutathione reductase

S. Shaheen • Q. Mahmood (*) • M. Asif • R. Ahmad

Department of Environmental Sciences, COMSATS Institute of Information Technology,

Abbottabad 22060, Pakistan

e-mail: mahmoodzju@gmail.com

guarino@unisannio.it

344

GSH Glutathione

H2O2 Hydrogen peroxide

K Kalium (potassium)

MAPK Mitogen-activated protein kinase

MDHAR Monodehydroascorbate reductase

MV Methyl viologen

MTs Metallothioneins

OSMT Oryza sativa metallothioneins

PCs Phytochelatins

RLKs Receptor-like kinases

ROS Reactive oxygen species

SOD Superoxide dismutase

tApx Tobacco ascorbate peroxidase

13.1 Introduction

Wastewater released from industries makes human lives easier but brings heavy

metals menace, which is disturbing the ecological balance. Heavy metals are non-

biodegradable chemical species which may accumulate in different plants parts and

therefore cause threats to plants and human health [1]. There are a number of tech-

niques developed over times for remediation of heavy metals, but natural treatment

systems are more effective compared to a conventional treatment system.

Phytoremediation or the use of living plants to remove heavy metals from soils and

water bodies and is proposed as a cost-effective and environment-friendly way to

clean up the contaminants [2, 3].

In nature, plants are tolerant towards some heavy metals to some extent and

assimilate these as essential nutrients. Green plants can be categorized on the basis

of plant-metal interaction as hyper-accumulating and non-accumulating plants. On

the basis of adaptations against heavy metals exposure, plants are divided into four

main categories, metal-tolerant species, metal-resistant species, metal-tolerant non-

hyper-accumulator species, metal hyper-tolerant hyper-accumulator plants species

[4]. Hyper-accumulators are plant species which are able to uptake, translocate, and

accumulate metals in aboveground plant tissues. A hyper-accumulator should have

an intensive root uptake system and faster root-to-shoot translocation. Roots uptake

metal from the soil and transport them to the stems and into the leaves. As low

concentration of trace metals are present in soil so high afnity transport system is

used to accumulate metal ions. A number of transporter genes are involved in this

process of metal transport [5].

In most of the plants, heavy metals interaction produces oxidative stress in the

chloroplast and mitochondrial membranes. This oxidative stress produce of ROS

species causes disruption of intercellular and extracellular membranous organelles,

ion leakage, lipid peroxidation, and DNA strand cleavage [6–8]. Most of the heavy

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metals are recalcitrant in nature thus causing serious damage to the environment.

These are nonbiodegradable in nature but biologically can be transformed from

more toxic to less toxic condition by their transformation of oxidation state and their

conversion from more complex to simplest forms [9].

In nature, plants are capable of self-protection by the production of less toxic

reactive compounds or by controlling metals transportation, accumulation, and

metal binding with cell wall and vacuole [10, 11]. Many plants when expose to toxic

concentration of metal ions try to avoid or decrease its uptake into root cells by

limiting the metal ions to the apoplast, binding them to the cell wall or cellular exu-

dates, or by reducing their long distance transportation. If this does not happened,

then metals already in the cell adopted storage and detoxication strategies, along

with metal transportation, chelation, trafcking, and sequestration into the vacuole.

When these actions were completed, then plants trigger oxidative stress defense

mechanism and synthesis of stress-related proteins and signaling molecules, such as

heat-shock proteins, hormones, and reactive oxygen species [12]. This review has

attempted a comprehensive description of plants mechanisms against heavy metals

avoidance, transportation, accumulation, and detoxication of heavy metals

contamination, and exploring the genetically based defense strategies adopted by

plants against trace element excess.

13.2 Avoidance Strategy inPlants

13.2.1 Extracellular Defense Strategy ofPlants AgainstHeavy

Metals

Plants possess different intrinsic and extrinsic defense strategies for tolerance or

detoxication whenever faces the stressful conditions due to the high concentrations

of heavy metals. Initially, regarding metal intoxication, plants implement avoidance

strategy to prevent the arrival of stress via restricting metal removal from soil or elimi-

nating it, and control metal entry into plant roots [13]. This can be attained by some

mechanisms such as restriction of metals by mycorrhizal association, metal sequestra-

tion, or complication by releasing organic compound from root [14, 15].

For heavy metals prevention or reduction of its toxicity impacts, plants develop

avoidance approach against HMs entrance. Plants adapted two main pathways by tak-

ing part in enhancing its complexity in roots vicinity. For the reduction of heavy metal

toxicity, plants enhance the pH of rhizosphere which released anions of phosphate.

Studies revealed that South American maize variety 3 released phosphate ions without

toxicity while sensitive maize variety 5 showed toxicity symptoms under Al stress

[16]. Studies revealed that under Cd stress, malate is secreted from sorghum (Sorghum

bicolor L.) roots, and citrate is secreted from maize roots [17]. Studies accomplished

the fact that root exudates in plants rhizosphere decrease the level of toxicity by

activating HM-binding proteins which inhibit the HM uptake [18].

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Similarly, oxalate released from the root apex facilitates the prevention of

Cd from entering into tomato (Lycopersicon esculentum L.) roots, thus in the

Cd-resistant tomato cultivar (Micro-Tom) these exudates promote Cd resistance.

Genetic studies revealed that under Al exposure Al-tolerant higher plants produces

more malic acid than sensitive genotypic plants species [19]. Thus, it is concluded

that the tolerant plant species may have adopted precipitation as an avoidance mech-

anism for the prohibition of the HM.

13.3 Signaling Strategy inPlant

13.3.1 Signals Transduction inPlants

In all plants, reaction towards heavy metal stress involves a complex signal transduc-

tion system that is trigger by sensing the heavy metal and is characterize by the pro-

duction of stress-related proteins and signaling molecules, and nally the transcriptional

activation of particular metal-responsive genes to neutralize the stress [20].

The most signicant signal transduction processes consist of the Cacalmodulin

system, hormones, ROS signaling, and the mitogen-activated protein kinase

(MAPK) phosphorylation ow, which activates stress-related genes [21]. There are

two main types of plants signaling, i.e., extracellular signaling and intracellular

signaling.

13.3.2 Signaling Networks

The ROS network is highly dynamic for plants growth, development, and stress

states thus producing ROS-signaling response effectively by ROS-scavenging and

ROS-producing protein [22]. The production of ROS physiologically occurs as a

by-product of biological reactions. During ROS production, P-450 and other cellu-

lar elements are released as a by-product [23]. Under chemical toxicity, ROS genes

network is being regulated by cytochromes P-450 which slow down the ROS level

in plants cells. For instance, the ROS gene network of Arabidopsis thaliana contains

more than 150 genes for the maintenance of ROS level in plants [24]. Calcium-

signaling network regulates the transmission of calcium signals through channels,

pumps, and carriers that between cellular, subcellular, and extracellular parts of

plants. Ca2+-binding proteins decoded and transmitted the information provided

by calcium signaling for transcription by Ca2+-responsive promoter elements that

ultimately regulate proteins phosphorylation [25].

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13.3.3 Extracellular Signaling

When plants exposed to multiple abiotic stress stimuli, it rapidly activates signaling

proteins MAPKs. An extracellular signaling ROS system has been named as “the

ROS wave” that covers about 8cm/min distance. The concept of “ROS wave” is

concerned with the perception or signaling of ROS produced in the plants [26]. In

many signaling network, the most important thing is the presence of transmembrane

proteins that act as receptor-like kinases (RLKs) and recognize signals with their

extracellular kinase containing parts then transmit them through the intracellular

kinase containing parts. RLKs manage developmental and hormone responses, sto-

mata closing and opening and stress response, and resistance against bacterial and

fungal pathogens [27, 28].

About 600 members of RLK gene family has been reported in Arabidopsis. In

extracellular parts of plants, the RLK groups like CYSTEINE-RICH RECEPTOR-

LIKE KINASES (CRKs) have two preserved cysteine domains (C-2x-C-8x-C;

DUF26 domain). Various studies revealed that on the basis of transcriptional and

phenotypic analysis of CRK mutants like their extracellular domain structure,

phenotype, and genotypic expression, these could be concerned in apoplastic ROS

signaling [29–32].

13.3.4 Intercellular Signaling

During intercellular signaling plants, information is transmitted in the form of

mobile signals, including transcription factors and membrane-associated proteins.

Generally, membrane-associated proteins are signicant in transcription, as small

RNAs and revealed intercellular movement through mobile peptides [33]. ROS are

important mediators of developmental procedures in different organisms like pro-

karyotes, fungi, plants, and animals through redox-sensitive transcriptional regula-

tor genes expression. In plants, regulation of peroxidase genes is possible by novel

ROS-sensitive transcription factor, UPB1 [34]. Various organelles within the cell

like chloroplasts, peroxisome, and mitochondria can generate reactive oxygen spe-

cies under stress situations and donate to plant stress tolerance [35]. In case of cal-

cium intercellular signaling, calcium signatures transformed the cellular levels of

calcium [36]. Cell organelles like vacuoles, endoplasmic reticulum, mitochondria,

and cell wall are the store houses of Ca2+ from where these are released when it is

necessary by the plant cells [37]. Similarly, cell organelles surrounded by double

membrane (e.g., mitochondria, chloroplasts, and nuclei) can generate Ca2+ signals

whenever posed to stress conditions [38].

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13.3.5 Calcium Signaling inPlants

When plants come across a number of physiological stimuli or stress, like low

temperature, drought, salinity stress, and pathogen or herbivorous attacks, then free

calcium ions of cytoplasm enormously activated [39]. The Ca2+ are later on tran-

scripted by intercellular reactions, mainly by Ca2+ sensor proteins which have been

preserved in all eukaryotic organisms, so that by activating complex downstream

signals in reaction of developmental and environmental stimuli. The physical

changes of Ca2+ binding is measured by structural changes of sensor proteins in Ca2+

relying proteins [40, 41]. In response to abiotic stress, calcium signaling is produced

with the regulation of cell cycle. The equilibrium of Ca2+ ions depends on the Ca2+

deciency, Ca2+ transporters, efux pumps, Ca2+/H+ antiporters, Ca2+ signatures,

Ca2+ memory, Ca2+ sensor, and transducer proteins [42] (Fig. 13.1).

Fig. 13.1 Representation of Ca2+ signals under diverse abiotic stresses regulated by CAM,

CRKS, CPKS, and CCaMK for the activation of regulatory genes

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13.3.6 Genes Involved inCalcium Signaling

Calcium and protein kinases signicantly take part in signaling pathways against

environmental stress in plants. The rst calcium-dependent, calmodulin-

independent protein kinase activities were reported in pea (Pisum sativum) extracts

20 years ago [43]. Various studies have shown that Ca-regulated proteins and

kinases [44]. Such as CaM protein [45] neurin B-like 218 (CBL) 8 proteins, CDPK

genes [46], and CBL-interacting protein kinase (CIPK) [47] are associated to abi-

otic stress response in plants.

By the activation of protein kinesis, calcium sensors or Ca2+-binding proteins detect

the high concentration of Ca2+. In response to the genes, expression of regulatory pro-

teins produced by protein kinesis can enhance transcription factors by phosphorylation

and by changing the metabolism that ultimately results in producing phenotypic

responses for enhancing stress tolerance [37]. Because of the rise of calcium level two

types of genes, specically up regulating or down regulating are overexpressed and

produces stunt growth or death of plant cells. These consist of known quick stress-

responsive genes in addition to genes of nonspecic function [48]. Recent studies show

that calcium signaling plays a signicant role in some pathways, for example, in

Arabidopsis, Ca2+-signaling pathway can also control a K+ channel for low-K response

in the presence of a blue light receptor phototropins, which successfully promotes

growth and plant development [49]. Calcium signaling is correlated with the sucrose-

signaling pathway that is an essential source of fructan synthesis [50]. Under abiotic

stress, calcium signaling controls the cell cycle progression.

13.3.7 ROS Signaling inPlants

Plants can sense, transduce, and translate ROS signal into suitable cellular response

with the assistance of redox-sensitive proteins, calcium mobilization, protein phos-

phorylation, and gene expression. ROS can be sense directly by some important

signaling proteins such as a tyrosine phosphatase through oxidation of conserved

cysteine residues [51]. ROS can also regulate many enzymes in signaling, such as

protein phosphatases, protein kinases, and transcription factors [52], and transmit to

other signal molecules and the ways forming part of the signaling network that

regulate the response downstream of ROS [53]. Usually, the power, lifetime, and

size of the ROS-signaling pool rely on balance between oxidant production and

removal by the antioxidant. By using mutants that lack in key ROS-scavenging

enzymes, Miller and coworkers determined a signaling pathway that is operated in

cells in response to ROS accumulation [54]. In tomato leaves, ROS produced in cell

walls of vascular bundle cells, as a result of wounding and produce H2O2 from

wound-inducible polygalacturonase which is acted as a second messenger for the

activation of defense genes in mesophyll cells, but not for signaling pathway genes

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in vascular bundle cells [55]. Tracing or detoxication of the unnecessary ROS is

achieved by well-organized antioxidative system that contains the nonenzymatic as

well as enzymatic antioxidants [56].

13.3.8 Genes Involved inROS Signaling

ROS are a sort of free radicals, reactive molecules, and ions that are obtained from O2.

It has been evaluated that about 1% of O2 used by plants is transformed to ROS [57].

The most frequently existing ROS are O2, O2•−, H2O2, and •OH.As O2 is a completely

nontoxic molecule as in its ground state it has two unpaired electrons with equidistant

spin which form it paramagnetic and, therefore not likely to take part in reactions with

organic molecules unless it is activated [58]. Naturally, in living organisms the produc-

tion of ROS is responsible for the intracellular communication system that regulated

the response to environmental stresses [59]. When the plants are under salinity stress,

the target of the ROS is regulated by vesicle trafcking complexes [60].

At low concentration, ROS have been concerned as second messengers in intra-

cellular signaling cascades that mediate several plant responses in plant cells with

closing of stomata [53, 61, 62], automatic cell death [63, 64], gravitropism [65], and

achievement of tolerance to both biotic and abiotic stresses [54, 66] (Fig. 13.2).

13.3.9 Genetic Control ofROS Production

Studies revealed that under abiotic conditions signaling is produced due to oxidative

stress, which result in activation of defense genes giving out specic adaptive

responses [67]. The system involved in up regulation of mRNA due to ROS produc-

tion can occur by redox-sensitive second messenger systems (e.g., MAP kinase acti-

vation) [68]. Plants genetics analysis showed that ROS signaling in Arabidopsis

plants enhances the antioxidative defense by rising the antioxidative genes expres-

sion and activation of the genes of inducible stress proteins [69]. The specic effect

of ROS-mediated signaling is related with the conrmation of denite genes expres-

sion. Some specic promoters and transcription factors have been recognized as a

producer of oxidative stress-responsive elements [70]. Under chilling and salt stress,

overexpression of a cytosolic APX-gene taken from pea (Pisum sativum L.) in trans-

genic tomato plants (Lycopersicon esculentum L.) improves the oxidative injury [71].

Likewise, tolerance against oxidative stress can also be improved by overexpression

of the tApx genes in tobacco or in Arabidopsis [72]. Several studies revealed those

in plants, under environmental stresses MDHAR show overexpression [73]. Gene

expression analysis of wheat showed two varieties of tolerant wheat bHLHs

(bHLH2: CA599618 and bHLH3: CJ685625) that have been affected by salinity [74].

Another gene family WRKY plays effective regulatory role in plants under biotic

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and abiotic stress conditions like cold, drought, wounding, salinity, UV, H2O2,

salicylic acid (SA), viral and bacterial attack [75, 76].

Overexpression of Arabidopsis MDHAR gene in tobacco enhanced salt toler-

ance and mitigated polyethylene glycol stress [77]. Tomato chloroplastic MDHAR

overexpressed in transgenic Arabidopsis improved its tolerance to temperature and

methyl viologen-related oxidative stresses [78]. Likewise, regulation of the gene-

encoding cytosolic DHAR was observed in L. japonicas, which was proved to be

more tolerant to salt stress than other legumes. This increase of DHAR was associ-

ated with its action in AsA recovery in the apoplast [79]. Transgenic potato

overexpressing Arabidopsis cytosolic AtDHAR1 proved higher tolerance to herbi-

cide, drought, and salt stresses [80].

Kwon etal. [81] veried that simultaneous expression of Cu/Zn-SOD and APX

genes in tobacco chloroplasts increased tolerance to methyl viologen (MV)

stress relatively to expression of either of these genes alone. Similarly, improved

Fig. 13.2 Representation of the transduction pathways involved in ROS signaling in response to

abiotic stresses. External stimuli produce ROS in chloroplast, mitochondria, and peroxisomes;

these activated MAPK cascade. These mitogen-activated protein kinesis that regulate transcription

factor and synthesis of genes (SFR, WRKY, bHLH, etc.) to overcome the negative effects of stress

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tolerance to multiple environmental stresses has been produced by instantaneous

overexpression of the genes of SOD and APX in the chloroplasts [82–84] SOD and

CAT in cytosol [85] and SOD and GR in cytosol [86]. Moreover, the instantaneous

expression of numerous antioxidant enzymes, such as Cu/Zn-SOD, APX, and

DHAR, in chloroplasts has shown to be more efcient than single or double expres-

sion for developing transgenic plants with improved tolerance to various environ-

mental stresses [87]. So, in order to attain tolerance to multiple environmental

stresses, increased importance is now given to produce transgenic plants overex-

pressing multiple antioxidants.

13.4 Detoxication Strategy inPlants

13.4.1 Detoxication Mechanism inPlants

Main detoxication mechanisms in plants are

1. Transportation to storage parts

2. Chelate formation

3. Compartmentalization in subcellular parts

4. Removal from the plant body [88].

13.4.2 Metal Transporters

Transport system of metal ions is very complex and miscellaneous. Metal transport

system is involved in uptake of metal, its translocation to various plant organs and

metal liberation in subcellular parts together with metal storage in vacuoles [89].

For long distance or intercellular transport in plants and subcellular compartmental-

ization of metals, low-molecular-weight chelators, such as glutathione, phytoche-

latins, histidine, or citrate, play a crucial role. By selective metal chelation and

trafcking or by internal transporter selectivity, there is a requirement to make a

distinction between metal cations of different elements.

A wide variety of transport proteins occurs that belongs to different families

including

• The zinc-regulated transporter, iron-regulated transporter protein (ZIP) family

• The cation diffusion facilitator (CDF) family

• The P1B-type subfamily of P-type ATPases

• The natural resistance-associated macrophage protein (NRAMP) family

• The yellow-stripe 1-like (YSL) subfamily of the oligopeptide transporter (OPT)

superfamily

• The copper transporter (COPT) family

• The Ca2+-sensitive cross complementer 1 (CCC1) family

• The iron-regulated protein (IREG) family

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Other membrane protein families that have been found to involve in transition

metal transport are the cation exchanger (CAX) family and three subfamilies of

ATP-binding cassette (ABC) transporters, the multidrug resistance-associated pro-

teins (MRP), the ABC transporters of the mitochondria (ATM), and the pleiotropic

drug resistance (PDR) transporters [90, 91]. A metal transporter is involved in metal

detoxication and metal hyper-accumulation [89].

Metal transporters are involved in:

1. Uptake from the soil to root

2. Translocation from the root to the shoot

3. Detoxication by storage in the vacuoles (Fig. 13.3)

13.4.3 Uptake fromtheSoil toRoot

Transporter genes involved in cellular uptake of metals from soil have been identi-

ed by researchers. A number of ZIP transporters are found to be involved in Zn

uptake across plasma membrane [89]. Fifteen potential ZIP genes may be identied

in the Arabidopsis thaliana genome [92].

Fig. 13.3 Mechanisms involved in heavy metal transport

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Transport properties of plant metal transporters that mediate metal entry into the

cytoplasm, for example, the Zn transporter AtZIP1 have been analyzed upon heterolo-

gous expression in yeast by measuring metal uptake into yeast cells [93]. IRT1 is one

of the most important members of ZIP family, involved in iron uptake from the soil

[93]. Puig etal. [94] found that AtZIP2 and AtZIP4 are involved in cellular accumula-

tion of zinc and copper. OsIRT1 and OsIRT2 have been proposed to contribute in Cd

uptake [95]. Milner etal. [96] found that expression of AtZIP1 is localized to the root

stele and is a vacuolar transporter while AtZIP1 expression was also found in the leaf

vasculature and is localized to the plasma membrane. Functional studies with

Arabidopsis AtZIP1 and AtZIP2 T-DNA knockout lines suggest that both transporters

play a role in Mn (and possibly Zn) translocation from the root to the shoot. AtZIP1

may play a role in remobilizing Mn from the vacuole to the cytoplasm in root stellar

cells and may contribute to radial movement to the xylem parenchyma. AtZIP2, on the

other hand, may mediate Mn (and possibly Zn) uptake into root stellar cells, and thus

also may contribute to Mn/Zn movement in the stele to the xylem parenchyma, for

subsequent xylem loading and transport to the shoot [96].

13.4.4 Translocation fromtheRoot totheShoot

In phytoextraction, transport of metals from root to shoot is of utmost important.

The study of Zn and Cd hyper-accumulators provides the evidence of involving

P-ATPase also called HMA (Heavy Metal transporting ATPase) as an important fac-

tor in their transport from the cystol of root cells into vascular tissues [97]. The

HMAs divide into two groups: those transporting monovalent cations (Cu, Ag) and

those transporting divalent cations (Pb, Cd) group [98]. Eight HMAs has been iden-

tied in Arabidopsis and Oryza sativa [99]. Analysis of the complete genomic

sequence in Arabidopsis shows the division of these eight HMAs in two groups:

HMA1–4 for the transport of Zn/Co/Cd/Pb and HMA5–8 for the transport of Cu/Ag

[100], while HMA2 and HMA4, are involved, in the transport of Zn and Cd. HMA4

confers increased Cd tolerance when expressed in yeast [99].

Transport of cadmium from root to shoot is a control process and most of the Cd

is stored in roots. To increase root-to-shoot translocation of Cd, transformation with

the genes of high biomass responsible for high root-to-shoot translocation such as

HMA4 can be done. In A. thaliana, HMA2 and HMA4 genes are involved in the

transport of Zn and also nonessential Cd to the shoots [101, 102]. Studies show that

both proteins encoding for Cd loading in xylem and are plasma membrane con-

tained proteins [101]. The P1B-type ATPases (also known as HMAs) have a major

role in translocation of metal ions against their electrochemical gradient by using

ATP as energy. All living organisms including humans, yeast, and plants contain

HMAs [100]. In Arabidopsis, HMA4 is found to be more expressed in vascular tis-

sues of root, stem, and leaves. It has been characterized and its role in Cd detoxica-

tion has been conrmed in Arabidopsis [100]. HMA4 plays a role in xylem loading

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of Zn and Cd, and hence in the control of translocation to shoots in Arabidopsis

halleri and Thlaspi caerulescens. A. halleri plants (from a Cd-hyper-tolerant acces-

sion) with a lowered expression of HMA4 translocated less Zn from the root to the

shoot and were more sensitive to Cd and Zn treatments [103].

Work of Courbot et al. [100] leads to conclusion that elevated expression of

HMA4 P1B-type ATPase is an efcient mechanism for improving Cd/Zn tolerance in

plants under conditions of Cd/Zn excess by maintaining low cellular Cd2+ and Zn2+

concentrations in the cytoplasm. HMA2 and HMA4 are the only P1B ATPases iden-

tied so far, which are predicted to have a long C-terminal domain. It is interesting

that the CC dipeptides and the His-rich domains are found in the prolonged C ter-

mini of HMA2 and HMA4 and not in the N-terminal domain, where HMA domains

are always found. The N-terminal end of HMA1 also harbors a poly-His domain

[104] (Table 13.1).

13.4.5 Detoxication by Storage intheVacuoles

In the hyper-accumulation of Zn, Ni, and Cd, an enhanced capacity of metal stor-

age in leaf vacuoles seems to play an important role [98]. ABC transporter is

involved in many physiological processes. Several members of ABC transporter

are involved in vacuolar sequestration of metals. Hmt1 is found to be involved in

transport of PC-Cd complexes in the vacuoles of S. pombe [89]. Some members of

the MATE family were shown to function as cation antiporters that remove toxic

compounds from the cytosol by exporting them out of the cell or sequestering them

to vacuole [105].

Members of CDF family involved in the cytoplasmic efux of metal cations from

cytoplasm to organelles like Zn2+, Cd2+, and have been named MTP (metal tolerance

protein). CDFs are highly expressed in A. halleri and T. caerulescens: MTP1, MTP,

and MTP11. AtMTP1 suggested being involved in Zn tolerance and basal Zn accu-

mulation in leaves. MTP11 and especially MTP8 are close homologues of ShMTP8

that give Mn tolerance when expressed in yeast and when ectopically overexpressed

in A. thaliana [93] (Fig. 13.4).

Mechanism of metal transport in plant cell. Heavy metals enter into cytosol

through metal transporter from cytosol into vacuole via metal transporters.

HM—High Metal

LM—Low Metal

ZIP—Zinc-regulated transporter, iron-regulated transporter protein

NRAMP—Natural resistance-associated macrophage protein

CAX—Cation exchanger

ABC—ATP-binding cassette

MT—Metallothioneins

PC—Phytochelatins

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13.4.6 Metal-Binding Genes

Plants have been authentically proved to minimize the harmful effects of metal tox-

icity, by pathways or methods relating to the binding of heavy metals to cell wall

and its transporation [106, 107]. Generally, the synthesis of metal-binding peptides

based on the production of metallothioneins and phytochelatins [108]. Usually, che-

lation is the most common intracellular system for the maintenance of low concen-

trations and detoxication of free metals in plant cytoplasm that can be achieved by

thiol compounds (which contain sulfhydryl/thiol groups; such as a tripeptide gluta-

thione, GSH, γ-Glu-Cys-Gly; phytochelatins, PCs; metallothioneins, MTs), and

Table 13.1 Gene expression of various abiotic stress conditions in plants

Type of stress Genes involved Plant species References

Heat AtCaM3 A. thaliana Xuan etal. [145]

Heat AtCaM7 A. thaliana Lu etal. [146]

Heat OsCAM1-1 O. sativa Wu etal. [147]

Salt GmCaM4/5 Glycine max Park etal. [148]

Heat TaCaM1-2 Triticum aestivum Liu etal. [149]

Salt AtCML8 A. thaliana Park etal. [150]

ABA, droughts, salt AtCML9 A. thaliana Magnan etal. [151]

Salt AtCML18/CaM15 A. thaliana Yamaguchi etal.

[152]

Heat, cold, ABA AtCML24/TCH2 A. thaliana Delk etal. [153]

ABA, salt AtCML37/38/39 A. thaliana Vanderbeld and

Snedden [154]

ABA, drought AtCML42 A. thaliana Vadassery etal. [155]

Cold, heat, drought,

Salt.ABA

OsMSR2 O. sativa Xu etal. [156]

Heat AtPP7 A. thaliana Liu etal. [157]

Heat AtCBK3/CRK1 A. thaliana Liu etal. [158]

Cold, heat, Salt.ABA,

H2O2

AtCRCK1 A. thaliana Yang etal. [159]

Cold AtCRLK1 A. thaliana Yang etal. [160, 161]

ABA, H2O2, ROS,

dehydration

OsCCaMK/DMI3 O. sativa Shi etal. [162]

ABA, ROS ZmCCaMK Z. mays Ma etal. [163]

ABA, salt TaCCaMK Pisumsativum Pandey etal. [164]

Salinity DHAR Oryza sativa Chen and Gallie [165]

Drought, ozone DHAR A. thaliana Ushimaru etal. [166]

Cu, Zn AtZIP2, AtZIP4 A. thaliana Puig etal. [94]

Cd OsIRT1, OsIRT2 O. sativa Clemens etal. [95]

Mn ShMTP A. thaliana Delhaize etal. [167]

Cu/Ag HMA5–8 A. halleri Courbot etal. [100]

Zn/Co/Cd/Pb HMA1–4 A. halleri Courbot etal. [100]

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also by non-thiol compounds (such as organic acids and amino acids) [109–118].

Studies revealed that peptides that have either histidines (GHHPHG) 2 (HP) or

cysteines (GCGCPCGCG) (CP) be engineered to Lam B and expressed on the sur-

face of E. coli. Surface demonstrated that CP and HP enhanced the bioaccumulation

fourfold and twofold [119].

13.4.7 Phytochelatins

Phytochelatins are a family of cysteine-rich, thiol-reactive peptides that attach many

toxic metals and metalloids, producing good messengers for genetically better

phytoremediation system [120]. The general structure of PCs is (g-Glu-Cys) n-Gly,

where n differs from 2 to 11 [121]. Practical descriptions of an unusual phytochela-

tin synthase, LjPCS3, of Lotus japonicus, have been acknowledged in an extensive

variety of plant species and some microorganisms [122].

Phytochelatins plays an important role in biosynthesis and detoxication of

heavy metals [122]. After production, PCs combine with heavy metal ions and make

possible their transportation as complex into the vacuole, where they nally pro-

duce complexes of high molecular mass, which is the key method that utilizes to

bind heavy metal ions in both plants and yeasts [122]. Genes concerned in the

production of PCs are phytochelatin synthases, such as g-glutamyl cysteine trans

peptidase [121].

Fig. 13.4 Mechanism of metal transport in plant cell. Heavy metals enter into cytosol through

metal transporter from cytosol into vacuole via metal transporters

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Overexpression of phytochelatin synthase in Arabidopsis thaliana seedlings

causes tolerance of arsenic, but hypersensitivity to cadmium and zinc [123]. The

same fact was observed in other transgenic plants with diverse PCS genes show

diverse phenotypes, including heavy metal tolerance (by or lacking accumulation)

and hypersensitivity to heavy metal ions [124]. Overexpression of pytochelatin syn-

thase in tobacco: distinctive effects of AtPCS1 and CePCS genes on plant response

to cadmium [125]. Phytochelatin synthase of Thlaspi caerulescens increases toler-

ance and accretion of heavy metals when expressed in yeast and tobacco. PCs are a

group of cysteine-rich, thiol-reactive peptides that combine many toxic metals and

metalloids, thus producing best messenger for genetically improved phytoremedia-

tion pathways [120].

13.4.8 Metallothioneins (MTs)

Apart from PCs that are the product of enzymatically formulated peptides, MTs are

formulated resultantly by mRNA translation [126]. While PCs in plants may pri-

marily deal with Cd detoxication, MTs appear to elaborate the attraction with a

larger series of metals such as Cu, Zn, Cd, and As [127]. MTs demonstrated unusual

properties and performance that depends on their presence in a type of plants and

are extremely mottled in terms of their molecular characteristics and structural qual-

ities [128]; they probably contain a number of various activities in plants than a few

other living creatures. In plants, these ligands are concerned to negate the toxicity of

HMs by cellular sequestration, homeostasis of intracellular metal ions, and metal

transport modications [129–131].

Additionally, MTs play an important role in HM detoxication, actively involved

in cellular-related events including ROS scavenger [132], maintaining of the redox

level [133], repair of plasma membrane [134], cell proliferation, and its growth and

repair of damaged DNA [135]. There are numerous endogenous and exogenous

agents other than HMs that are able to bring the synthesis and expression of MTs.

Of these, osmotic stress, drought, intense temperatures, nutrient deciency, release

of different hormones, natural and dark-induced tissue decay, injuries, and viral

infections can be mentioned [12, 127, 136].

Ectopically expressed MTs in transgenic plants are proved to increase their toler-

ance towards metal intoxication. Kumar etal. [137] showed that OSMT1e-p, a type

1 MT extracted from a salt-tolerant rice genotype (Oryza sativa L. cv. Pokkali),

participated in tolerance for copper and zinc toxicity when ectopically expressed in

transgenic tobacco. They evaluated that tobacco plants in which gene have been

inserted possessed to hold more quantity of Cu2+ and Zn2+ in their roots or lower

leaves, considerably decreasing the HMs ions transportation and quantity in leaves

and harvestable plant parts. Zhigang et al. [138] accomplished that the ectopic

expression of BjMT2, a metallothionein type 2 from Brassica juncea, in Arabidopsis

thaliana enhanced copper and cadmium tolerance at the seedling phase but intensely

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decreased root growth when there was no heavy metal treatment. This tendency may

propose that ectopic expression of MTs in transgenic plants may proceed in host

plant in a nonspecic method and in a different way effect the organ growth.

13.5 Conclusion andFuture Prospects

The present review outlines the impact of abiotic stresses on plants. Most of the

investigations done so far mainly described the genetic investigation of plants against

abiotic stress; this review involved in genetically based defense and detoxication

pathways mainly Ca and ROS signaling, transportation, chelation, and detoxication

has been discussed in detail. Under stress conditions, plants activate specic mole-

cules which enhance plants tolerance and the development of defense mechanisms in

it. It has been observed that the activation of defense genes cascade transmit various

signals in cell organelles under various biotic stress conditions [40].

As abiotic condition produces oxidation stresses that overexpressed a number of

stress-induced proteins, this review could provide fundamental information about

antioxidant and regulatory genes production. As ROS have regulatory function as

signaling molecules, this feature may open a gateway to physiological, molecular,

and evolutionary research perspectives. Due to the importance of ROS, it is central

to modern plant biology to obtain a comprehensive understanding of the processes

where ROS have regulatory roles. Studies revealed that ROS signaling with ozone

as a tool is signicant for the transmission of distinct from of ROS signals to chro-

matin reformation and transcriptional regulation [140].

Therefore, the elaboration of transcriptomics and proteomics analysis will be

more helpful in understanding the bioinformatics and mutant studies. It has been

observed that ROS signals play an important role in intercellular Ca2+ signals,

Ca2+ inux which can be regulated by various PAs and Spm4+ proteins [141].

There is need to highlight the mechanism and interrelation of Ca2+-efux systems

with ROS and AtMPK6-signaling under biotic and abiotic stresses along with the

overexpression of regulatory genes against stress. Xing etal. [142] observed various

proteins kinesis like MKK2-MPK4/MPK6 and MEKK1 activation in salt, cold,

drought, and wounding stress can phosphorylate MPK4 which is also signicant in

abiotic stress signaling. Studies also revealed that ABA is a key hormone in induc-

ing abiotic stress responses in plants like barley showed pronounced effect of ROS

and PAs in salt-sensitive variety then salt-tolerant plants [143]. Likewise, ABA is an

important hormone under biotic and abiotic stress in plants [144]. But the links

between ABA and MAPKs under biotic and abiotic stresses has not yet been prop-

erly studied at cellular and subcellular levels in plants. Similarly, little work has

been done on K+/Na+ -signaling pathways under biotic and abiotic stresses like

Ca2+ homeostasis. Therefore, plants responses towards multiple abiotic stresses

would be another interesting area of future transgenic hyper-accumulator plants

production.

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In future, it is needed to identify molecular interaction of chelators with metal

transporters. Further investigation is needed to determine more about functional

signicance and biological role of transporter genes, especially, overexpression in

plants with greater biomass to increase their potential use in phytoremediation pro-

cesses. Furthermore, these genes can provide better understanding in the analysis of

gene regulation in metal-rich environment as well as metal-decient environment.

Similarly, silencing of transporter genes in edible crops may decrease metal bioac-

cumulation in food chain. Likewise, there is a greater area of exploration in terms of

3D structures of protein and functional analysis of the candidate genes. In future,

gene cloning and plant transformation can be done to determine efciency of metal

transporter genes in transgenic plants. Thus, the application of powerful genetic and

molecular techniques may surely be helpful in designing of hyper-accumulator

transgenic plants for bioremediation.

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369© Springer International Publishing AG 2017

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_14

Chapter 14

Engineered Nanomaterials

forPhytoremediation ofMetal/

Metalloid- Contaminated Soils:

Implications forPlant Physiology

DomingoMartínez-Fernández, MartinaVítková, ZuzanaMichálková,

andMichaelKomárek

Abstract Nanomaterials, including engineered nano-sized iron oxides, manganese

oxides, cerium oxides, titanium oxides, or zinc oxides, provide specic afnity for

metal/metalloids adsorption and their application is being rapidly extended for envi-

ronmental management. Their signicant surface area, high number of active sur-

face sites, and high adsorption capacities make them very promising as cost-effective

amendments for the remediation of contaminated soils. The alleviation of the toxici-

ties of metal/metalloids by their immobilization in the soil stimulates the growth

and development of plants during phytoremediation, but there is a body of evidence

indicating that nanomaterials themselves can yield both benecial and harmful

effects in plant systems at the physiological, biochemical, nutritional, and genetic

levels. Nanoecotoxicological studies are providing a good understanding of their

interactions with plants, and an increasing number of publications have attempted

to clarify and quantify their potential risks and consequences for plants. However,

many results are contradictory and the safety of engineered nanomaterials still

represents a barrier to their wide, innovative use in phytoremediation. Both their

positive and negative effects on plants will have to be taken into account to evaluate

their applicability, and the scientic community faces a challenge to understand

deeply the factors which can determine their relevance in environmental science and

technology.

Keywords Nanoparticles • Oxides • Stabilization • Immobilization • Nanoremediation

• Toxicity

D. Martínez-Fernández (*) • M. Vítková • Z. Michálková • M. Komárek

Faculty of Environmental Sciences, Department of Environmental Geosciences,

Czech University of Life Sciences Prague,

Kamýcká 129, Prague 6- Suchdol, Prague 165 21, Czech Republic

e-mail: domingo.marfer@gmail.com; martinez-fernandez@fzp.czu.cz; vitkovam@fzp.czu.cz;

michalkovaz@fzp.czu.cz; komarek@fzp.czu.cz

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14.1 Introduction

Soil contamination by metals (e.g., Cd, Cu, Mn, Ni, Pb, Zn) and metalloids (e.g.,

As, Sb) is a current global phenomenon endangering safe agricultural production

and groundwater quality. Therefore, society, economics, and science are involved

together in the common need for novel and environmentally friendly techniques for

soil remediation. Recently, nanotechnology has offered a new generation of envi-

ronmental remediation technologies that can provide cost-effective solutions to

some of the most challenging environmental clean-up problems [1]. While various

industrial sectors produce a large number of products containing nanomaterials,

nanotechnology is also used in environmental management. Nanoparticles (NPs,

materials with at least two dimensions between 1 and 100nm) and nanomaterials

(NMs, materials with at least one dimension smaller than 100nm) [2–4] have the

potential to revolutionize agricultural systems, environmental engineering, safety

and security, water resources, and numerous other life sciences [5]. The smaller

particle sizes, the higher specic surface area, and thus the higher number of reac-

tion sites for metal adsorption represent the main advantages of NMs [6]. Nano-

sized metal oxides, including nano-sized iron oxides, manganese oxides, aluminum

oxides, zinc oxides, titanium oxides, and cerium oxides, have specic afnities for

metal/metalloids adsorption and their application has been rapidly extended for

environmental tasks [7–9]. Although they can exist naturally in the environment,

they can also be produced/engineered intentionally [10], through methods that are

becoming simpler, more effective, and cheaper. Concerning the economic aspect, it

is clear that the doses of these compounds required for adsorption are lower when

applied as engineered NMs because of their huge reactivity, while the contact times

needed for the metal/metalloids adsorption are shorter in comparison with conven-

tional adsorbents. In this sense, their application seems to be protable [11].

14.2 Nanoparticles forEnvironmental Remediation

For sites contaminated by metal/metalloids, successful remediation is complicated

by the fact that these pollutants do not degrade spontaneously, and it is not usually

possible to excavate all the contaminated soil. Therefore, chemical stabilization of

the metal/metalloids in these soils, through adsorption, surface precipitation, struc-

tural incorporation, or ion exchange, is a viable option for such sites as this technol-

ogy immobilizes the contaminants in the soils and thus reduces their mobility,

bioavailability, and bioaccessibility. In fact, this rst step—the reduction of their

toxicity—is crucial for the establishment of a vegetation cover on the contaminated

sites during phytoremediation. The use of engineered NMs for remediation pur-

poses can then enhance the natural attenuation processes, a key mechanism for the

re-establishment of sustainable environmental systems. It has been claimed that

nanotechnology has great potential as an environmentally cleaner technology,

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including alleviation of the toxicities of various metal/metalloids [12]. As a result,

several studies have appeared in various journals dealing with the metal-

NMs- mediated diminution of metal toxicity [13–15]. In general, this remediation

technology involves (a) NMs transport along with the solution to the contaminated

zone; (b) attachment to soils in the contaminated zone; and (c) reaction with the

target contaminants to form less toxic or less mobile products [6]. The principal

removal mechanisms for the most common inorganic contaminants can be divided

into ve categories: (1) adsorption (Cr, As, U, Pb, Ni, Se, Co, Cd, Zn, Ba); (2)

reduction (Cr, As, Cu, U, Pb, Ni, Se, Co, Pd, Pt, Hg, Ag); (3) oxidation (As, U, Se,

Pb); (4) precipitation (Cu, Pb, Cd, Co, Zn); and (5) co-precipitation (Cr, As, Ni, Se).

The overall reaction processes are strongly inuenced by a number of factors, in

particular the NMs chemical properties and structure, the presence of more than one

contaminant species, the pollutant characteristics, and the hydrogeochemistry of the

aqueous environment (pH, redox conditions, natural dissolved species, etc.) [6, 16].

Since the amendments applied for the stabilization process need to be cost-

efcient and suitable for different soil types and should not pose a risk to environ-

mental compartments, application of engineered NMs in remediation technologies

provides a very interesting alternative to soil excavation and dumping, ex situ soil

washing, etc., because these are generally disruptive and costly. Nanoparticles have

been studied as adsorbents of metals and their characteristics (i.e., large surface

area, high number of active surface sites, low intra-particle diffusion rates, and high

adsorption capacities) make them very promising for the cost-effective treatment of

polluted soils [17, 18]. Then, nanoremediation, dened as the use of nanoparticles

for environmental remediation, has the potential not only to reduce the overall costs

of cleaning up large-scale contaminated sites, but can also reduce clean-up time,

eliminate the need for treatment and disposal of the contaminated soil, and reduce

the availability of some contaminants [19–21]. This is reected in the increasing

number of publications on this subject and the rising level of funding for remedia-

tion projects [6, 22–26].

To date, researchers have mainly focused their attention on the removal of metals

from aqueous solutions rather than soil-bound metals, which may be absorbed by

plants and subsequently spread into the human food chain. The nanoremediation of

contaminated soils is a topic that has been researchedless, compared to the removal

of pollutants from water or wastewater [21]. Based on this idea, soil columns can be

set up for ex situ remediation, and a liquid suspension of NMs can be added to

extract or to immobilize the contaminants (typically metals); the species adsorbed

onto the NMs can be removed by applying mild gravitational (centrifugal) or mag-

netic (in the case of magnetic NMs, such as magnetite) gradients [21]. As a result of

these previous leaching experiments with aqueous solutions, and thanks to the dem-

onstrated potential of engineered NMs, they have been gradually incorporated into

new in situ strategies for phytoremediation. Most of the information about their

behavior in aqueous systems can be extrapolated to the soil solution, which is

important since the physical/chemical properties of NMs are one of the most impor-

tant factors that control their behavior in the environment although obviously it

must be modied to take into account the new conditions in the soil. In this context,

14 Engineered Nanomaterials forPhytoremediation…

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different hydrochemical parameters—such as pH, Eh, ionic strength, and aqueous

chemistry—can change the aggregation kinetics and transformation of engineered

NMs and their subsequent behavior. Similarly, natural organic matter alters their

stability through electrostatic and steric interactions. The transformation process for

NMs is also altered by a conuence of factors, depending on the characteristics of

the NMs and of the environmental receptors.

Although the use of plants for phytoremediation and their capacity to accumulate

and tolerate high concentrations of metals have been explored, and a signicant

amount of literature is available, the same is not true regarding NMs, and understand-

ing the response of plants to NMs would be a key element in identifying mechanisms

involved in stress tolerance and NMs toxicity [27]. According to Juganson etal. [28],

936 was the total number of publications (as sum of all the materials) found in the

Thomson Reuters Web of Science™ for “environmental remediation” using NMs

(for a search done on March 19th, 2015), with 303 publications for nTiO2, 219 for

nFeOx, 110 for nAg, 74 for nZnO, 36 for nCuO, 16 for nCeO2, and the rest for other

NMs like fullerenes and carbon nanotubes (See chart in Fig. 14.2).

14.2.1 Main Types ofNanomaterials fortheAdsorption

ofMetals andMetalloids

Metal NMs display size-dependent properties, such as magnetism (magnetic NPs),

uorescence (QDs), or photocatalytic degradation (metal oxide NPs), that have bio-

technological applications in sensor development, agrochemical degradation, and

soil remediation [29]. Nanoparticles and nanomaterials are mainly classied accord-

ing to their dimensionality, morphology, or uniformity [30], but the classication

according to their chemical properties is the most accepted and useful.

14.2.1.1 Iron Nanooxides: Nanogoethite, Nanomaghemite,

Nanomagnetite

Iron oxides represent natural components of soils and exist in many forms, includ-

ing mainly goethite (α-FeOOH; prevailing in temperate climatic areas), hematite

(α-Fe2O3; prevailing in warm-dry climate zones), maghemite (γ-Fe2O3), and magne-

tite (Fe3O4). These iron oxides play a crucial role in soil systems due to their ability

to adsorb potentially toxic elements such as metals and metalloids [31–33]. Synthetic

iron-based NMs are thus interesting candidates for the removal of metals and metal-

loids from contaminated waters and soils, or their stabilization therein, due to their

increased specic surface area and modied surface structure, which strongly affect

their reactivity and chemistry [34–37]. During the remediation process, they can be

applied directly as nano iron oxides or in the form of their precursors (i.e., nZVI,

nano zero-valent iron) ([38] and references therein). The use of iron-based NMs for

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the in situ immobilization of trace elements limits the potential leachability of

metals/metalloids and thus prevents their transport into deeper soil layers and

groundwater [39].

Due to its abundance and the presence of surface hydroxyl groups, goethite

is signicantly involved in the transport and transformation of nutrients and

contaminants—including inorganic/organic anions, cations, and some gases [40].

Synthetic nanogoethite (nFeOOH) has been successfully used for the removal of Cu

from aqueous solutions, showing photocatalytic activity and a high adsorption

capacity for Cu [41].

Maghemite is a common weathering product in soils of temperate, tropical, and

subtropical climatic regions, usually formed during the oxidation of magnetite.

Synthetic maghemite (nFe2O3) is a promising material for the removal of inorganic

contaminants as it is readily available, inexpensive, and can be easily separated and

recovered because it is magnetic [42]. Nanomaghemite has been deeply investigated

due to its efcient removal of the most-toxic form of As (arsenite, As (III)) [43–47].

Moreover, it has been demonstrated that it is an important scavenger of Cr(VI), Pb(II),

Cd(II), Cu(II), and Zn(II) from aqueous solutions and thus could be a useful sorbent

for water and soil remediation [4, 42, 48, 49]. The use of nFe2O3 has been reported to

be useful for promoting the growth of plants in a contaminated soil [50], mainly due to

the immobilization of Zn from the soil pore water (available to plants) with the conse-

quent reduction of its toxicity to the roots and aerial parts. Adsorption of Pb(II) by

nFe2O3 occurs mainly through the formation of inner-sphere complexes, while Cd(II)

is likely adsorbed as a mixture of inner- and outer-sphere complexes [42]. The effec-

tiveness of the adsorption of these metal/metalloids is affected by the modication of

the atomic structure on the particles surface with decreasing size of nanomaghemite

[43]. Also, the presence of other components in the soil solution—such as citrate com-

plexes and organic acids [51], or other nutrients [52]—inuences the sorption process.

For example, PO43− has been described as a competitor for arsenite and arsenate immo-

bilization by nanomaghemite [53] due to their similar outer electronic structures.

Magnetite is a mixed-valence magnetic iron oxide, containing Fe2+ and Fe3+, and

it can be formed in the soil through (bacteria assisted) weathering of ferrihydrite

[54]. Immobilization of As in soils using nanomagnetite (nFe3O4) was performed by

Zhang etal. [32], who reported higher stabilization efciency of nFe3O4 compared

to iron sulde or nZVI.In another study, nFe3O4 proved to be an efcient amend-

ment for the removal of Pb from aqueous solutions, yielding fast adsorption with a

maximum capacity of 36mgPbg−1 [17]—which was much higher compared to, for

example, goethite [55]. Moreover, the behavior of Pb was not affected by the pres-

ence of other ions such as Ca, Ni, Co, or Cd. Additionally, desorption and regenera-

tion tests showed that nFe3O4 can be used repeatedly without loss of their adsorption

capacity [17]. Shen etal. [56] investigated the inuence of pH, temperature, and

particle size on the adsorption of metals from aqueous solution by nFe3O4. Under

room temperature at pH 4 with an average particle size of 8nm, ≥85% of Cu2+,

Cd2+, Ni2+, and Cr6+ were removed, yielding the maximum at pH>7 for divalent

metals and at pH 2 for hexavalent Cr. In contrast, coarse particles showed values of

maximum adsorption capacity about seven-times lower [56].

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Recently, the synthesis and utilization of iron NMs (nFeOx) with novel properties

and functions have been widely studied, both for their nano size and for their magnetic

characteristics [11, 36, 57–59]. Typical iron NMs syntheses involve routes including

chemical precipitation [60], sol-gel, hydrothermal, dry vapor deposition, surfactant

mediation, microemulsion, electro-deposition, and sonochemical methods [36]. Iron

oxide composites such as clay–iron oxide magnetic composites and magnetic zeolites

can be synthesized and used for the removal of metallic contaminants from water [61].

If the size of the magnetic NPs is reduced to below a few nanometers, they become

superparamagnetic. Using an external magnetic eld, these particles change their

direction. Therefore, the remediation efciency may be enhanced by combining metal

binding and selective adsorption properties with separation of magnetic nano-sorbents

from the system, since their magnetic behavior (either ferromagnetic or superpara-

magnetic) depends on the particle size [36]. Gómez-Pastora etal. [11] illustrated that

engineered nFeOx have very high adsorption capacities for metals/metalloids in

polluted waters; moreover, their magnetic properties facilitate their collection from

the solution—allowing their further reuse. The recovery of magnetic NMs by the use

of magnetic gradients [62, 63] represents a promising alternative for sorbent

applications.

14.2.1.2 Nano Zero-Valent Iron (nZVI)

A vast number of studies have demonstrated the applicability of nano zero-valent

iron (nZVI) as an amendment for remediation of metal/metalloid-polluted water

systems [6, 11, 64–66] or trace elements immobilization in contaminated soil [16, 39,

67–69]. The possible mechanisms by which nZVI stabilizes metal/metalloids

include adsorption and/or surface precipitation, redox reduction, and co- precipitation

in the form of metal iron oxides or oxyhydroxides [6] ([66] and references therein).

The particles of nZVI have a core-shell structure, which gives them characteristics

typical of both iron oxides (sorption) and elemental Fe0 (reduction) [6]. The iron

core (up to 98% Fe) is covered by a shell composed of iron oxides and hydroxides

(FeO, Fe2O3, FeOOH). Furthermore, the surface of nZVI particles has a signicant

inuence on their stability and mobility in the environment and it prevents their

rapid oxidation. Their increased specic surface area results in much higher reactiv-

ity but, on the other hand, the reaction of particles smaller than 20nm is so fast that

their reaction capacity may be depleted before they get to the contaminant. Thus,

attaining the optimal balance between the reactivity and lifetime of nZVI needs to

be guaranteed for in situ applications [66, 70, 71]. Detailed overviews of nZVI

reactivity have been provided by O’Carroll etal. [6] and Yan etal. [72], while the

reaction mechanisms have been recently reported by Filip etal. [73].

When nZVI is exposed to air or water, it is oxidized, forming a layer of iron

oxides or hydroxides on the surface that is responsible for the subsequent adsorption

process [6, 64, 74]. The reaction process is strongly dependent on pH.Under alkaline

conditions, a negatively charged surface is favorable to metallic cations adsorption,

while the high pH values limit the adsorption of metallic anions. Therefore, nZVI

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can interact with arsenate and chromate oxyanions at low pH, mainly through

electrostatic interaction with the positively charged groups on the surface [45, 56, 75].

Different forms of nZVI are available, including powder, mineral oil suspension,

or aqueous suspension. The synthesis methods include formation of the nanomate-

rial from atoms or molecules through physical/chemical methods (nucleation, vapor

condensation, precipitation, agglomeration) and physical/chemical methods to

breakdown a bulk material to the nanoscale size (thermal decomposition, thermal

reduction of oxide compounds, or pulsed laser ablation) [66, 71]. The synthesis

method inuences the size, shape, and composition of iron NPs and thus their actual

reactivity. Particles of nZVI prepared by the reduction of goethite or hematite are

generally bigger (up to 100 nm) and of irregular shape, while formation using

NaBH4 provides smaller, regular-shaped particles up to few tens of nm in size [66].

The borohydride reduction of ferrous salts is the most common method of nZVI

synthesis for laboratory-scale experiments [71, 76]. However, the industrial appli-

cation is precluded since highly reactive particles with a signicant tendency to

agglomerate are produced by this procedure and the overall expenses are high [70, 71].

The easy and cost-effective synthesis of highly reactive NMs, such as nZVI, is a

priority of the academic community and nZVI producers. In this context, a method

using leaf extracts from different trees was performed to obtain “low-cost” nZVI

particles [77].

In order to prevent particle aggregation and to improve nZVI reactivity, various

innovative surface modications/coatings have been developed [6, 66]. Among the

traditional agents for nZVI stabilization are poly(acrylic acid), poly(methyl meth-

acrylate), poly(ethylene glycol), polyaspartate, and others. Several stabilizers of

natural origin have been tested also, including xanthan gum, guar gum, potato

starch, alginate, and chitosan. Bimetallic NPs represent nZVI particles coated with

noble metals [6]([78] and references therein). Although the surface properties of

nZVI may change, the modiers generally ensure the transport of stabilized NPs

and the reaction with the target contaminant only in the polluted zone. The specic

surface of Fe0 can be 4–15m2g−1, while up to >40m2g−1 can be reached for surface-

stabilized particles [6, 71]. The use of several composites such as bentonite-nZVI

has been reported also [79].

Due to its abundance, easy accessibility, high reactivity, and efciency for risk

element stabilization, nZVI has become a widespread remediation amendment both

on a laboratory scale and for in situ applications. However, as nZVI is a redox-active

material, this being important for treatment of redox-sensitive elements (e.g., As,

Cr) [14, 80], the impact on soil microbial communities needs to be investigated.

According to Li etal. [81], rapid and complete oxidation of Fe0 eliminates its effects

on bacteria due to passivation within a few hours. Signicant impacts on microbial

diversity were reported for nZVI-treated soil contaminated with Pb, whereas there

were no effects on microbial activity in Zn-contaminated soil [16].

Efcient treatments of contaminated soils with nZVI have been reported, result-

ing in Pb and Zn immobilization [16, 69] as well as decreased availability of

As [39]. However, nZVI was a less efcient amendment for in situ soil remediation

of Cr, compared to other sorbents studied by Chrysochoou etal. [82], while success-

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ful remediation of Cr from ore processing residue [83] and wastewater [14, 80] was

achieved upon the application of nZVI.Further studies are needed in order to assess

the long-term impacts of nZVI on the environment in terms of contaminant stabili-

zation, while minimizing the effects on soil characteristics.

14.2.1.3 Manganese-Based Materials

Manganese oxides (including hydroxides and oxyhydroxides) together with Fe

oxides occur naturally as erosion products in almost all soil types, mainly as coat-

ings on soil particles and pores or in the form of concretions and nodules with

poorly crystalline (even amorphous) structure [84]. Compared to Fe oxides, Mn

oxides are generally less abundant in soils but appear more efcient in the immobi-

lization of some metals [85, 86]. This is mainly due to their large specic surface

and (usually) low value of pH at point of zero charge (pHpzc)—the reason for their

negative surface charge in usual soil conditions [87]. Their specic structure, formed

by sheets (layers) or tunnels in most cases, allows the accommodation of water

molecules or various cations in interlayer or tunnel regions [84]. Manganese oxides

possess strong oxidative properties and thus take part in many oxidation- reduction

and cation exchange reactions. For this reason, Mn oxides are not suitable amend-

ments for soils contaminated with Cr as they are able to readily oxidize Cr(III) to the

more toxic and mobile Cr(VI) [88, 89]. On the other hand, this oxidizing nature can

be benecial in the case of contamination with As; Mn oxides have proved efcient

in the oxidization of the more mobile and toxic As(III) to As(V) [90–93].

Due to their promising properties, many studies focused on the synthesis and test-

ing of engineered Mn nanooxides, which are potential agents for environmental

clean-up. Manganese oxide NPs can be prepared both by classical chemical routes

[94–97] and by biotechnological means, using the activity of microorganisms like

bacteria or fungi [98–100]. In fact, biogenic oxidation of Mn(II) represents also the

prevailing route for the formation of Mn oxides in soil. Although thermodynamically

favored, Mn(II) oxidation in the environment solely by chemical means is very slow.

On the other hand, when the process is mediated microbially, the reaction rate can be

increased by several orders of magnitude [101, 102]. As in the case of other nanoad-

sorbents, the rst studies dealing with these materials focused mainly on their syn-

thesis, characterization, or adsorption properties with respect to targeted compounds;

in this case, metals/metalloids. Based on these data, possible applications— including

remediation—can be proposed. In this context, application of Mn nanooxides for

soil remediation appears relatively safe as nanoscale biogenic Mn oxides are natural

and ubiquitous soil components. Although numerous studies have been published

dealing with the adsorption performance of Mn-based NPs [100, 103–108], their

application for the direct remediation of contaminated soil, together with assisted

phytoremediation, is still rather scarce. Della Puppa et al. [109], together with

Michálková etal. [110] and Ettler etal. [111], studied the adsorption properties and

stabilizing potential of partially nanoscale amorphous Mn oxide (AMO) with regard

to Cd, Cu, Pb, Zn, and As in contaminated soils. In these studies, after application to

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contaminated soils, AMO was able to decrease signicantly the amount of targeted

metal/metalloids in the soil solution, even showing a higher sorption capacity for Cd,

Cu, and Pb than engineered nanomaghemite [51, 110]. On the other hand, higher

dissolution of this agent in acidic conditions—connected with unwanted oxidation

and dissolution of soil organic matter—was recorded. For this reason, AMO appears

a suitable amendment for neutral and slightly alkaline soils.

14.2.1.4 Other NMs

In addition to Fe- and Mn-based NPs, there exists a wide variety of novel, engi-

neered NMs potentially usable in the remediation of soil and water contaminated

with metals/metalloids (see chart in Fig. 14.2, according to Juganson etal. [28]). To

date, one of the most studied NMs is nTiO2, also the most studied photocatalyst

worldwide. Besides its applications targeting the decomposition of various organic

compounds, dyes, etc., the process of photocatalytic reduction can be used to

remove various toxic metal ions as well. Numerous studies have examined the

potential of nTiO2 for the reduction of highly mobile and toxic Cr(VI) to Cr(III)

[112–114] and the immobilization of toxic As(III) species [115]. Other materials

(Zn/Al-based nanocomposites) were shown to be very promising as they not only

behaved as adsorbents but also had photocatalytic properties, being able to adsorb

the highly toxic Cr(VI)—that was subsequently reduced photocatalytically to

Cr(III) [116]. Nanoparticles of hydrous Ce oxide were reported as another material

suitable for adsorption of Cr(VI) from aqueous solution [117]. Nanoparticles of

nMgO, nTiO2, and nZnO were found to be efcient adsorbents for Cr in soil con-

taminated by leather factory waste, decreasing signicantly the exchangeable Cr

fraction while increasing the residual fraction [118]. Carbon nanotubes represent

another promising type of engineered material, being efcient in the adsorption of

various divalent metals from aqueous solution [119]. In the study of Jośko etal. [120],

application of multiwalled carbon nanotubes reduced the phytotoxicity of sediment

contaminated with various organic and inorganic contaminants.

14.2.2 Nanomaterials intheEnvironment

The global production of engineered NMs was estimated to be 260,000–309,000

metric tons in the year 2010; of which about 8–28, 0.4–7, and 0.1–1.5% were esti-

mated to have ended up in soils, water bodies, and the atmosphere, respectively

[121]. The use of NMs in environmental remediation will inevitably lead to the

release of NMs into the environment and subsequent ecosystems. Once in the envi-

ronment, NMs may persist for a long time or be taken up by organisms and trans-

ferred between organisms of different trophic levels, thus acting as an ecotoxicological

hazard, and undergo biodegradation or bioaccumulation in the food chain [121–123].

Plants are considered to represent both the rst sink for the accumulation of NMs

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from the surrounding environment and the point of entry for their bioaccumulation

in the food chain [124]. For this reason, emerging studies have focused on the gen-

eral consequences of NMs uptake by plants, as their effects on the biomass produc-

tion and plant response are very relevant to phytoremediation. Although

nanotechnology has the potential to solve problems that cannot be solved by the

full- scale products, one important aspect in nanoremediation, the safety of NMs,

still represents a barrier to their wide innovative use and is hindering their full appli-

cation; hence, intensive studies must be done before their use. As concluded recently

by Schaumann etal. [125], further assessment of environmental impacts on the fate

and effects of NPs is needed.

14.3 Consequences ofNanomaterials forPlants

Nanoecotoxicology is a branch within toxicology which focuses on measuring the

toxicity of NMs that enter into contact with organisms like plants, bacteria, sh, and

invertebrates [126]. A good understanding of the interactions of NMs with the plant

system is of paramount importance for assessing their toxicity and trophic transport

[127, 128]. To understand and quantify the potential risks for plants, the mobility,

bioavailability, toxicity, and persistence of manufactured NMs need to be studied.

An increasing number of published studies have attempted to understand the inter-

actions between NMs and plants, and several reviews have already examined the

implications of NMs in food crops [8, 129–131]. As can be seen in Fig. 14.1, which

shows the general trends of the effects of NMs on plants according to the published

literature to date, there is sufcient evidence that NMs can yield both benecial and

harmful effects in plant systems at the physiological, biochemical, nutritional, and

genetic levels. The interactions between plants and NMs can shed light on the envi-

ronmental consequences of nanotechnology, but, in contrast to the huge amount of

research done on the bulk chemicals as environmental hazards, the research on NMs

toxicity is markedly scarce [128] and it needs to be improved.

There are many factors which must be taken into account during nanotoxicologi-

cal studies, and this makes it very complicated to understand the real consequences

for plants since even small differences in the design of the experiments can produce

different results. For example, for most NMs, relatively high concentrations are

needed to cause observable toxicity in plants and the toxicity threshold is species

dependent [132, 133]. Owing to their insolubility in water, NMs in general have a

limitation for toxicity experiments [128]. Moreover, most plants showed visible

signs of recuperation from NMs toxicity—indicating that the toxicity was tempo-

rary [4]. Auffan etal. [134] pointed out that chemical stability under physiological

redox conditions appears to be a condition for the non-toxicity of metallic NMs.

Nevertheless, metallic NMs with strong oxidative or reductive properties can be

cytotoxic and genotoxic. Consequently, the nanotoxicological research on the

uptake and accumulation of NMs by plants, and their subsequent response, has

sometimes generated controversial data [27, 135, 136]. However, when taken

together, the apparent differences in the toxicity of NMs to plants may arise from

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their chemical (reactivity) and physical properties (size, form, aggregation) and the

dose (high or low concentrations), exposure time, plants (species, age, physiologi-

cal status), and experimental conditions (in hydroponics, soils, eld, glasshouse,

etc.) used. Many of the responses to NMs by plants have been evaluated to nd out

how NMs improve growth when used as soil amendments to reduce or mitigate the

toxicity of contaminants; but their effects must be studied in isolation, to ascertain

the effects of the NMs themselves on the plants. If the addition of NMs to a con-

taminated soil can potentially ameliorate metal-induced damaging effects on

growth, by the reduction of metal availability and toxicity, this stimulation of growth

may mask the potential negative effects caused by NMs. The effects on key physi-

ological processes in plants of engineered NMs with potential use for phytoreme-

diation, reported to date, are described below.

14.3.1 Germination

Seed germination tests represent one of the simple and rapid tools for assessing the

phytotoxicity of NMs. The recorded effects of NMs on seed germination fall into all

possible classes—being negative, nil, or positive, depending on the kind of NM, the

Fig. 14.1 Matrix of the described effects of NMs on plant physiology during environmental reme-

diation. The matrix has been created using the general effects of each NM from the literature

compiled in this book chapter. In some cases, contradictory responses have been detected (two

colors in the same box), denoting that the toxicity of NMs to plants is not completely understood

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species, and the concentration used [10, 27, 129]. For example, strong to total

inhibition of germination after seed exposure ton Fe3O4 has been reported for lettuce

(Lactuca sativa), radish (Raphanus sativus), cucumber (Cucumis sativus), and spin-

ach (Spinacia oleracea) [137, 138]. However, for the same NMs, Barrena etal. [139]

reported low to nil toxicity against cucumber and lettuce germination. Nano-CuO

did not affect germination, but inhibited growth of Zea mays seedlings [140] and

enhanced the seed germination and shoot-to-root ratio of lettuce [141]. Exposure of

pea (Pisum sativum) seeds to nZnO had no impact on germination [142].The effects

of nZVI on germination have been reported to be concentration dependent [143, 144].

Although widely used, germination tests have to be interpreted carefully. When

evaluating the inuence of NMs on seed germination in relation to the phytoreme-

diation of soil, attention should be paid to the experimental approach as different

experimental designs may give different results. Classical ecotoxicological studies

dealing with NMs and their effects on seed germination are usually performed with

the seeds directly exposed to NMs in the suspension. But, in the context of assisted

phytoremediation, this experimental approach appears to be not very suitable. In

this case, NMs are applied to soil—the aim being to immobilize contaminating met-

als, decrease their solubility and toxicity, and promote thus the plant growth. The

soil solution from contaminated soil amended with NMs thus represents a system

completely different to that of a pure NMs suspension. For this reason, NMs des-

tined for use in assisted phytoremediation should be tested not just directly in sus-

pension; their inuence on the soil solution composition and, subsequently, the

inuence of the soil solution obtained or the amended soil itself on seed germination

should be examined too, as the results of these tests could vary signicantly.

14.3.2 Uptake ofNMs by theRoots

Most of the available studies on phytotoxicity of NMs have focused mainly on tox-

icity symptoms of plants, and relatively few have examined the mechanisms of NMs

phytotoxicity, uptake, translocation, and bioaccumulation [136]. The roots are the

rst organ which can suffer from NMs interference in the soil, and for that reason

there is an urgent need to evaluate the impacts on plant physiology of NMs, together

with their potential ecotoxicity and interactions with the key processes in the rhizo-

sphere [145–147]. Possible interactions of NMs with plant roots include adsorption

onto the root surface, incorporation into the cell wall, and uptake by the cell

[148, 149]. For NMs to enter the root stele, they have to either cross the cell wall and

plasma membrane of an endodermal or exodermal cell or cross a root cell wall of a

cell exterior to the endodermis/exodermis and move into the stele symplastically

[135]. They may be transported from one cell to another through plasmodesmata.

However, the exact reasons why only some plant species readily take up several

NMs are still unknown and remain to be explored [131]. To cross an intact cell wall,

it has been hypothesized that NMs have to move passively through a cell wall pore,

suggesting that plant uptake is highly size selective. Even so, NMs could be

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incorporated passively into the apoplast of the endodermis; they would be then

subjected to the highly size-selective permeability of the membranes before reach-

ing the central cylinder [135]. It is generally assumed that it is difcult for NMs

bigger than 20nm to penetrate through the cell wall since the cell wall pore sizes

vary from 2 to 20nm [4, 8, 150–152]. According to Burello and Worth [153], NMs

with a diameter larger than 20–30nm act often as bulk materials; thus, the “true

nanoeffects” are attributable to NMs with smaller size. Roy and Bhattacharya [4]

suggested that NMs can enter plant cells by binding to carrier proteins, through

aquaporins and ion channels, by creating new pores, or by binding to organic chemi-

cals in the environmental media. Endocytosis may be also an important pathway by

which NMs enter plants [154] since NMs could theoretically activate membrane

receptors and induce endocytosis.

The high reactive capacity of NMs—due to their high specic surface area—can

stimulate their adhesion to the epithelial root cell wall. Nanomaterials of all compo-

sitions also have the potential to aggregate, due to Van der Waals forces or other

interactions [155]. As a result, NMs may aggregate along the roots, blocking their

proper water uptake and disturbing thus the whole plant physiology and ultimately

affecting their growth and development [3]. Aggregation of NMs appears to change

the color of the roots surface by covering the epithelial cells [152, 156]. It can affect

also the interactions of the plant with the external medium through mechanical dis-

ruption of membranes and cell walls, blocking the pores and diminishing the root

hydraulic conductivity [156]. Despite their adherence to the surface, due to binding

and electrostatic attraction by a limited number of cell surface cation exchange and

binding sites on the negatively charged root surface [157], some NMs do not seem

to move through the surface of the roots. In this way, inhibition of plant growth may

not derive directly from chemical phytotoxicity of NMs. Instead, toxicity may result

from the physical interactions between the NMs and plant cell transport pathways

[129, 156]. Anyway, even though potential aggregation might dramatically increase

the size of the NMs and reduce their mobility [149, 158], Whitley et al. [159]

showed that NMs may remain unaggregated in soil pore water for an extended

period of time, suggesting that NMs are likely to be bioavailable to plants.

Although some NMs can be found in plant cells and tissues [8], no uptake or

toxicity has been reported specically for nFeOx, which may be due to the adher-

ence of these materials to soil particles. Zhu etal. [124] reported no measurable

uptake of nFeOx by pumpkin (Cucurbita maxima) grown in either soil or sand, or

of nFe3O4 by lima bean plants. Similarly, Wang etal. [160] reported no uptake of

nFe3O4 (25nm in diameter) by pumpkin plants. For poplar, it appeared that some of

the nZVI penetrated through the membrane and was internalized in the root cells [161].

Zhou etal. [162] reported the adsorption of nCuO (55nm in size) onto the Triticum

aestivum root surface. As defended by Lü etal. [163], metallic NMs can affect the

epithelial root cells but they seem not to have an important effect in the xylem of the

plants. In spite of the adsorption of nCeO2 aggregates on the root surface observed

by Majumdar et al. [164], Ce accumulation increased linearly with increasing

exposure concentrations, corroborating previous studies in other edible plants

like tomato [165], rice [166], soybean [167], corn [168], and cucumber [169].

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The concentration-dependent linear increase in Ce accumulation in roots suggests

uptake through simple diffusion. According to Pradhan etal. [170], nano-sized Mn

oxides (nMnOx) were readily taken up from soil by the roots of Vigna radiata and

transported to the leaf, where nMnOx acted as a cofactor in a series of enzymatic

reactions during the assimilation of nitrate into organic nitrogen compounds. In a

nano-ZnO translocation study by Hernandez-Viezcas etal. [171], the use of μXRF

images led the authors to conclude that Zn was obtained from nZnO by the roots,

but it was not adsorbed on the root surface. This interesting research indicated an

important concept because the authors were careful to separate the effects produced

by nZnO from those related to and produced by Zn released from the NMs.

The type of roots and their architecture, their age, and the species are also crucial

factors in the response to and uptake of NMs by roots. For example, lignin can act

as a barrier to reduce the permeability of foreign materials in cells. This could be a

reason why, in the study by Ma etal. [161], nZVI was able to enter the root cells of

poplar plants, while the relatively high lignin content in the cell wall of Typha lati-

folia prevented nZVI from passing through it.

14.3.3 Translocation andAccumulation

The available literature indicates vaguely that NMs are found in plant cells and tis-

sues, and even though some studies report NMs internalization in roots, no translo-

cation to the shoots was found [4, 130]. However, NMs could potentially be taken

up by plant roots and transported to shoots through vascular systems, depending

upon the composition, shape, size, and plant anatomy [129]. There are many

physico-chemical differences between plant species—such as variations in hydrau-

lic conductivity, cell wall pore size, and root exudate chemistry—that could inu-

ence NMs bioaccumulation. For example, Zhu etal. [124] observed accumulation

of nFe3O4 (min. size 20nm) in Vigna radiata (mung bean) and pumpkin grown in

an aqueous medium, but did not observe bioaccumulation when conducting this

same experiment using Phaseolus lunatus (lima bean) or Phaseolus limensis, in

either soil or sand. Corredor etal. [172] investigated xylem transport of NMs by

injecting graphite-coated iron NMs into the pith cavity of the leaf petiole of pump-

kin plants, and a very homogenous population of approximately 46-nm NMs was

found in the xylem at a distance from the injection site, suggesting that NMs larger

than 46nm were not transported. Recently, lack of uptake and translocation for

nFe2O3 were demonstrated by Martínez-Fernández etal. [156], without the presence

of these NMs in the sap of Helianthus annuus, possibly as the result of aggregation

of NMs on the root surface. Another study examined the importance of size on

uptake by exposing wheat plants to nTiO2 ranging from 14 to 655nm [173], con-

cluding that NMs greater than 140nm were not taken up and NMs greater than

36nm were not translocated into the aerial portions of the plants. The accumulation

and translocation of nCeO2 were dose dependent [174], but they accumulated

mainly in the root tissue [166, 175]. Birbaium etal. [176] also reported no uptake

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after exposing 3–5-week-old corn plants to 37-nm-diameter nCeO2 for 14 days. The

work by Schwabe etal. [177] showed that nCeO2 with a size range of 17–100nm is

at least partially available for uptake by pumpkin. Translocation of Ce has, however,

been documented in previous studies of cucumber [169], corn [178], and beans [164].

Ma etal. [161] reported that nZVI was able to move into the root cells of poplar

plants while such internalization was absent in the case of Typha latifolia, maybe

because the aggregate of nZVI was too large for the xylem tissues to transport. In

both cases, upward transport to the shoots was insignicant. Although Lin and Xing

[179] reported accumulation of the nZnO in the protoplast of endodermal cells, they

found no evidence that the particles were translocated into the shoots and leaves,

possibly as the result of NMs aggregation in the exposure media. Wang etal. [180]

observed xylem- and phloem-based transport and biotransformation of nCuO

(20–40nm) as well as nCuO transport from roots to shoots via the xylem and trans-

location back to roots via the phloem.

In contrast, recent studies show that Au, Ag, CuO, and ZnO NMs are readily

taken up and translocated by plants, either as NMs or in their ionic form [181].

Nanomaterials may accumulate and/or increase the concentrations of the compo-

nent metal in the fruits/grains of agricultural crops, have detrimental or benecial

effects on the agronomic traits, yield, and productivity of plants, induce modica-

tions in the nutritional value of food crops, and transfer within trophic levels. So, it

is important to establish whether a more predominating trend of NMs accumulation

exists and whether the metals involved follow the same trend as the chemical form

available in the soil or in the water [27].

14.3.4 Water Balance

Because of their relevance to the proper growth, nutrients uptake, stress, and

biomass production of plants, more studies of NMs are needed at the root–soil inter-

face, including measurements of plant water relations [182]. As described above,

many researchers consider that the observed toxicity exerted by NMs in plants is

based on physical plant–NMs interactions. The presence of NMs on the root surface

could alter the surface chemistry of the root such that it affects how the roots interact

with their environment [183]. For example, it is known that metals/metalloids can

reduce the root hydraulic conductivity, with consequent decreases in plant water

content, turgor potential, and growth [184, 185], but how metallic NMs inuence the

transport of water through the roots is not known. Martínez-Fernández etal. [50, 156]

found a reduction of the root hydraulic conductivity in plants of H. annuus treated

with nFe2O3 in hydroponic culture, but no changes in the internal water status of

plants grown in a contaminated soil treated with the same NMs. The work by

Trujillo-Reyes etal. [186] suggests that the reduction in dry biomass production in

plants exposed to nFeOx is most likely due to the particle aggregation on the surface

of the root, which affected water entrance, resulting in growth reduction. Asli and

Neuman [187] also found that exposure to nTiO2 (30nm) caused a reduction of the

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water transport capacity in sections of primary roots, their effect being proportional

to the exposure time in hydroponic culture. The nTiO2 also inhibited transpiration,

reduced root hydraulic conductivity in Zea mays root apices, caused cell wall pores

to constrict, and resulted in minor inhibition of shoot and root growth. Other work

suggests that NMs increase the expression of aquaporins in roots [188, 189], maybe

as a response by the plant to compensate a reduction in root hydraulic conductivity.

Trujillo-Reyes etal. [190] showed that, although Fe ions/NMs did not affect the

water content of lettuce plants, Cu ions/NMs reduced their water content, root length,

and dry biomass. Nano-CuO (size: <50nm) was reported to reduce the transpiration

volume in plants [191], also related with an up-regulation of proline- biosynthesis

genes under nCuO exposure in Arabidopsis thaliana [192]. Nano-CuO stress induced

high accumulation of proline (a water-stress indicator in plants), and the degree of

accumulation was associated closely with the nCuO concentration [193, 194]. The

accumulative transpiration rate in plants indicated that transpiration was highest for

the controls and gradually decreased as the concentration of nZVI increased in pop-

lar, Typha latifolia [161], and peaplants after nZnO exposure [142, 179].

14.3.5 Nutrients Uptake

A damaged water transport system implies a lower capacity to pass water to the

shoot, affecting the transport of all the dissolved elements and causing a deciency

of them in the shoot, according to the plant requirements. Generally speaking, expo-

sure to NMs involves changes in the nutritional status of the plants and development

is negatively affected, but positive effects have been documented as well. Studies

revealed that the lower uptake of nutrients is related to the fact that NMs clog the

root openings and inhibit both hydraulic conductivity and nutrient uptake in roots

[156, 187], although NMs with high specic surface areas may also help to seques-

ter nutrients on their surface.

The effects of NMs on plant nutrition have been reported in few studies [195,

196]. Martínez-Fernández etal. [156] detected a signicant reduction of trace ele-

ments concentrations in shoots and roots of H. annuus exposed to nFe2O3, without

changes in their concentrations in the sap. However, the concentration of Mo in the

roots increased with the dose of nFe2O3, maybe due to the close relationship between

the Fe and Mo uptake systems and because the uptake of Mo can be facilitated

under higher Fe concentrations in the external medium [197]. Iron oxide NMs have

been reported as facilitators of iron and photosynthates transfer to the leaves of

peanut [40]. In bean (Phaseolus vulgaris), nCuO decreased the shoot Fe, Zn, and Ca

levels, but not that of Mg, while K showed little change and Na increased [195].

Aluminum, Ca, and Zn concentrations in roots and leaves were higher in plants

exposed to Cu NMs, compared with the control treatment [190]. Silicon NMs used

to mitigate the Cu toxicity increased the contents of Mg, Ca, K, and P in the root and

shoot of pea plants [15].

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There are different effects related to the nitrogen uptake and metabolism in plants

during the interaction with NMs. Manganese NMs affected the assimilatory process

by enhancing the net ux of nitrogen assimilation in mung bean plants [170]. Nano-

sized TiO2 can also have a positive effect on plants through promotion of the uptake

of nitrate, which accelerated the transformation of inorganic nitrogen into organic

nitrogen [27], due to increased nitrate reductase activity, and could also protect

chloroplasts from aging in soybean and A. thaliana [198, 199]. Exposure to nTiO2

also increased biomass, photosynthesis rate, and enzyme activity in spinach [200–202],

related to enhanced N2 xation from nitrogen photoreduction and the stimulation of

RuBisCo activity. Cai etal. [203] reported stimulation of the removal of Cu and

nitrate from aqueous solution, after the application of bimetallic Fe/Ni NMs. On the

other hand, the presence of nZnO in the environment is potentially hazardous to the

Rhizobium–legume symbiosis system [142], this interaction being an important fac-

tor for plant growth and crop productivity as it provides bioavailable nitrogen to the

plants. The presence of nZnO in the rhizosphere affected the early interactions

between rhizobia and the host plant as well as nodule development, and subse-

quently delayed the onset of nitrogen xation [142].

14.3.6 Oxidative Stress

Nanomaterials can mediate signicant elevations in reactive oxygen species (ROS)

generation and its subsequent consequences (such as membrane damage), as well as

the modulation of antioxidant defense system components and cellular redox

homeostasis in plants. Iron nanooxides can signicantly increase the antioxidant

enzyme activities, but their effects seem to be related more to the changes in the

mineral composition in the plant than to the presence of nanoscale forms of Fe

[190]. Iron NMs toxicity studies have primarily focused on Fe(II) and its oxides,

and little is known about the toxicity specic to others NMs such as nZVI.However,

nZVI produces Fe(II) and iron oxides through oxidation, and nZVI can produce free

radicals which are highly reactive and cause oxidative stress [204]. This could be

one of the mechanisms behind the toxic effects of nZVI on plants. Further studies

on A. thaliana evidenced that nZVI triggered high plasma membrane H+-ATPase

activity, resulting in stomatal opening that was vefold higher than in unexposed

plants [205]. Nano-MnOx has been reported to increase the activity of the electron

transport chain by binding with the CP43 protein chain of photosystem II [206]. The

nMnOx enhanced the oxygen evolution process, being a part of the water splitting

complex in the light reaction of photosynthesis, hence improving the photophos-

phorylation capacity [206]. Nano-CuO stress also induced modulation of anti-

oxidant enzymes activity, and nCuO treatment caused oxidative damage to rice

seedlings, as evident from high ROS-scavenging antioxidant enzymes activity and

enhanced malondialdehyde levels [193], and maximally disrupted the plant-defense

system by oxidative stress [207]. The accumulation of nTiO2 in plants does not

appear to induce oxidative stress in the leaves [173]. Biochemical assays with nZnO

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indicated increases in the specic activity of CAT (in the root, stem, and leaves of

Prosopis juliora-velutina), but no evidence of chlorosis, necrosis, stunting, or wilt-

ing, even after 30 days of treatment [171]. In ROS formation and release, the con-

version of fatty acids to toxic lipid peroxides occurs, leading to the disruption of

biological membranes [208] and consequently the entrance of and damage by NMs

and metals, causing TBARS (thiobarbituric acid reactive species) formation—

which damages the membrane permeability. This specic report showed that, by

increasing the concentrations of the nZnO, higher values for the TBARS were

observed. Silicon NMs protect pea seedlings against Cr(VI) phytotoxicity, by reduc-

ing Cr accumulation and oxidative stress and up-regulating the antioxidant defense

system and uptake of nutrient elements [15].

14.3.7 Chlorophylls

Nano-sized materials may interact with the proteins associated with photosystems,

the starch-synthesizing machinery, and/or carbohydrate translocation [209]. Since

chlorophyll content is considered as an index of the total light harvesting complex

and the electron transport components, present in chloroplast membranes [210], it

is used as a stress indicator in plants. Studies of the bioavailability of nFe2O3 in A.

thaliana, performed by Marusenko etal. [211], suggested that the Fe-NMs were not

used for chlorophyll production. Iron NMs reduced the accumulation of chloro-

phylls in the leaves of Lactuca sativa [190] and Helianthus annuus [156], this effect

being related to the reduction of the root hydraulic conductivity and the transport of

dissolved nutrients from the solution, especially for Mg since this nutrient is associ-

ated with the synthesis of chlorophylls. In an experiment with nCeO2, Zhang etal.

[131] related a reduction in the chlorophylls content with the physical adsorption of

the NMs on the root surface, and the consequent blockage of Mg uptake by the

roots. Nano-CuO was reported to decrease chlorophyll content signicantly in

wheat [212], soybean [213], and A. thaliana [192], and in Vigna radiata in an

invitro experiment [194]. On the other hand, A. thaliana plants treated with bulk

ZnSO4 had a smaller amount of chlorophyll and were shorter compared with the

plants treated with nZnO [211]. In an in vitro experiment carried out with

Petroselinum crispum by Dehkourdi and Mosavi [214], nTiO2 caused a signicant

increase in the chlorophyll content of seedlings. Higher chlorophyll contents were

also recorded in leaves of Brassica juncea treated with Ag NMs [215].

14.3.8 Genotoxicity

Plants have been used as indicator organisms in studies of genotoxicology, facilitat-

ing data interpretation for a complete understanding of the effect of NMs [216–218].

Genotoxicity may be produced by direct interaction of NMs with the genetic

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material, by indirect damage from NM-induced ROS, or by toxic ions released from

soluble NMs [174, 219, 220]. Nanoparticles that cross intracellular membranes

(diameter between 8 and 10nm) may be able to reach the nucleus, through diffusion

across the nuclear membrane or transportation through the nuclear pore complexes,

and interact directly with DNA [220] and inuence DNA replication and transcrip-

tion of DNA into RNA.However, as expected, this effect is very conditioned by

their size. The NMs aggregates could also mechanically damage the chromosomes.

Nano-CuO is able to enter the nucleus of plant cells and mediate direct oxidative

damage to DNA [221]. Nano TiO2 (~100nm in size) was found to be genotoxic as

well as cytotoxic in plant systems [222]. Kumari etal. [223] showed a direct rela-

tionship between the increase in the number of aberrations and the increase in the

concentrations of the NMs, by analysis of changes in the chromosome morphology

caused by nZnO in root cells of Allium cepa, and explained these results based on

ROS activation. According to López-Moreno etal. [174], the toxicity may rise

either due to the interaction of the DNA with the Zn ions leached out from the nZnO

or its direct interaction with the nZnO.But the absence of nZnO in plant tissues, as

shown by the XANES results, failed to conrm the main reason behind the geno-

toxic response in soybean. It is not clear from these studies whether the genotoxicity

in plants is caused by the NMs themselves or their biotransformation within the

plants. Ma etal. [129] pointed out that one of the most urgent needs in plant–NMs

interaction studies is to determine the genetic response of the plants and the genes

that are up-regulated/down-regulated in plants exposed to NMs, but this knowledge

is in its infancy still. More research needs to be focused on the differences in toxic-

ity of NMs in relation to their respective bulk counterparts, and on the effects of the

ions produced inside or outside the organism exposed to the NMs.

14.3.9 Growth andBiomass Production

During recent years, the number of peer-reviewed papers related to nanoecotoxicol-

ogy has increased exponentially. On the one hand, there are abundant references for

positive effects of NMs on growth and biomass production: nTiO2 in spinach

[200–202]; nCeO2 increased root and stem elongation in cucumber [167]; nCuO

enhanced lateral root formation in A. thaliana [192]; nCuO enhanced the lignica-

tion of root cells in Glycine max [213]; nFeOx increased root elongation in pumpkin

[160]; etc. On the other hand, negative effects also appear very often in the related

literature: Al2O3 NMs in tobacco plants [224]; nCuO reduced root length in

Landoltia punctate [225]; nCuO in L. sativa and M. sativa [196]; nCeO2 decreased

stem elongation in corn and inhibited root elongation in alfalfa and tomato [167].

Many researchers have considered the biomass production as a response to the

stress, concluding that some external factors (treatments) affect it positively or neg-

atively according to the stimulation or inhibition of the growth, development, and

productivity of plants in comparison to untreated controls. However, the biomass

production is only the mere consequence of the huge combination of the positive

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and negative effects (Fig. 14.1) at different plant physiological levels (Fig. 14.2).

In contaminated soils, the application of engineered NMs can decrease the contami-

nant availability in the soil solution and thus enhance the growth of plants.

Nevertheless, as described above, there are many steps previous to the complete

understanding of the real consequences for plants exposed to NMs. In fact, and

generally speaking, long-term exposure to metallic NMs affects growth negatively

[226–228]. Only when the individual effects of engineered NMs on plants and their

applicability are properly evaluated, can concise conclusions be obtained to decide

about their use during phytoremediation tasks.

Sunower plants treated with nFe2O3 in a Zn-contaminated soil showed a 25%

increase in shoot biomass, related to the Zn-adsorption capacity of the NM [50].

These results highlight the applicability of this NM as an amendment during phy-

toremediation due to its immobilization of metals in the soil, stimulating the growth

of plants by making the contaminants less available. However, an additional experi-

ment with the same species in hydroponic culture showed that treatments with the

Fig. 14.2 Once in the soil, the engineered NMs can interact with the metal/metalloids in the soil

solution, aggregate, or interact with the roots of the plants. According to the balance among all the

positive and negative effects at different plant physiological levels, the plant will show the overall

effect on biomass production; but, even when apparent changes in growth are not manifested,

intrinsic and important effects can happen in the plants. In accordance with Juganson etal. [28],

the chart shows the percentages of each type of NM within the total number of publications in the

Thomson Reuters Web of Science™ for “environmental remediation” by NMs

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same nFe2O3 reduced the functionality of the roots, changed the nutrient status of

the plants, and led to reductions in the shoot macronutrient concentrations and chlo-

rophyll content [156].

Aspects like stem and root elongation, root gravitropism, architecture of the root

system, and number of lateral roots can help to describe the direct effect that NMs

have on the growth of plants. Trujillo-Reyes etal. [190] found that iron NMs (Fe/

Fe3O4) reduced root size and changed root architecture, as well as affecting the root

water content and the chlorophylls accumulation in the leaves of Lactuca sativa.

Exposure of peas to nZnO had an impact on root length [142], decreasing the num-

ber of rst- and second-order lateral roots. When the concentration of nZnO was

increased in the medium, the shoot and root lengths declined. Extended treatments

with nZnO also resulted in shorter root length than in controls without the NMs in

radish, rape, ryegrass, lettuce, corn, and cucumber during seed incubation [132].

Controversial results were also found for nZnO treatments, which increased the

lengths of the shoots and roots compared with the control for peanut [229] and

Brassica juncea [230]. Ma etal. [231] observed that, although nCeO2 inhibited the

root elongation of lettuce, six other plant species were unaffected. Also, the dose is

crucial: Ma etal. [232] noted that nCeO2 at concentrations less than 250mgL−1

signicantly increased A. thaliana biomass, but, above 500mg L−1, biomass and

chlorophyll production were reduced and lipid peroxidation was evident.

According to the Thomson Reuters WoS, 770 peer-reviewed papers on nano-

ecotoxicology that corresponded to the keywords “nano* AND ecotoxic*” were

published between 2006 and March 2015 [28]. The rapidly increasing number of

scientic publications on ecotoxicity of NMs over the past decade has inspired sev-

eral review articles summarizing the existing data in the eld. However, each review

has focused on specic aspects and parameters of NMs testing; therefore, it is dif-

cult to get an overview of all the factors (and their values) that might inuence the

toxicity of NMs. All these results will be an important factor to take into account

with regard to the applicability of NMs for long-term use in phytoremediation tasks,

but they will be especially useful when the causes become clear.

14.4 Limitations andDrawbacks oftheUse ofNMs

The use of NMs is not exempt from limitations. For example, the proneness to rapid

passivation of some NMs [233], susceptibility to geochemical conditions [234], and

possible environmental and human health threats of various NMs [235]. The high

reactivity and heterogeneous size distribution of NMs may have adverse impacts on

the sorption efciency, which negatively affects the long-term performance and

overall applicability. The mobility and sorption capacity of such particles are limited

by three principal mechanisms: (1) nanoparticle aggregation followed by gelation,

caused by poor colloidal stability, (2) nanoparticle oxidation/corrosion followed by

the formation of corrosion precipitates, and (3) nanoparticle trapping from solution

by interaction with other components (i.e., mineral surfaces and organic matter) or

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via microbial removal [71]. In order to prevent them, the particles can be coated

with certain organic or inorganic materials. The development of various surface

stabilizers and modiers—such as biopolymers and alginate—has made the NMs

much more versatile [11, 59, 236].

In spite of their nano size, NMs may be able to pervade very small spaces in the

subsurface and remain suspended in groundwater, allowing the particles to travel

farther than larger, macro-sized particles; in practice, the NMs used currently for

remediation do not move very far from their injection point [237]. In fact, the intro-

duction of NMs into soil is one of the most difcult aspects to consider in in situ soil

bioremediation [21]. Since NMs have the potential to aggregate [155], either during

manufacture or during wastewater treatment, this may dramatically reduce their

bioavailability, mobility, and toxicity [158, 238], and consequently limit their effec-

tiveness [233, 239]. The aggregation of some NMs supports the need for polymer or

other coatings to modify their surface, in order to improve mobility [240].

The regulatory framework generally assumes that NMs possess toxicity and

risk equivalent to those materials with larger particles, but the smaller size of NMs

results in entirely different physico-chemical properties. Since knowledge about

NMs regarding their interaction with biota and their toxicity is scarce, their full-

scale application and usage for soil remediation is still problematic. For example,

the report by The Scientic Committee on Emerging and Newly Identied Health

Risks does not even mention or dene engineered NMs.

14.5 Prospective Work Plan inPhytoremediation withNMs

As described in this book chapter, NMs appear to offer faster and cheaper remedia-

tion solutions, and their use at sites around the world is beginning to be explored.

Comprehensive utilization of nanotechnology at the present time and unprecedented

application of NMs in products will certainly create signicant amounts of new-

generation waste in the near future [30]. NMs are a reality, but future research

efforts need to be directed towards nding new methods for nanoremediation, rec-

ognition of the biological effects of NMs in the environment, and creation of the

bases of nanobiomonitoring. Recently, a database working group was established in

the framework of the European Union Nano Safety Cluster [241]—which high-

lights that research efforts are necessary to promote science-based regulations for

nanotechnology. Detailed research on the biogeochemical behavior of NMs in soil

systems, and on the potential advantages and drawbacks of their use in chemical

stabilization combined with phytoremediation, is being undertaken by the scientic

community, but a much broader view is still needed about their use. Each step in this

research will have the potential to provide better knowledge for the use of engi-

neered NMs during remediation tasks and thus to provide a very signicant benet

to society, by evaluating the impacts and safety of NMs application to soils contami-

nated with metal/metalloids.

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Most experimental studies with NMs have been conducted in batch systems

using model compounds and model media, instead of natural systems. Understanding

the fate of NMs and their effects in natural environments also requires more realistic

experimental setups [125]. Mesocosm experiments complement laboratory experi-

ments and, as a best practice, can be combined with laboratory experiments to fur-

ther develop the process of understanding the aging and functioning of NMs [125]

as well as their transformations through their life cycle [242]. So, more realistic and

holistic studies are needed in future investigations, including long-term experiments

in more complex environmental media. Critical knowledge gaps and incomplete

studies mean that the mechanisms for the removal of metals by NMs from contami-

nated soil proposed by different researchers are often contradictory [14]. These

discrepancies in the literature can be primarily related to methodological and

experimental shortcomings, such as inadequate NMs characterization, lack of

consideration of NM aggregation or dissolution, lack of proper controls, or the use

of environmentally irrelevant NM concentrations and/or exposure conditions.

However, it is now evident that, under certain circumstances, NMs are bioavailable

and toxic to several key terrestrial ecoreceptors [135].

New perspectives regarding the combined use of engineered NMs have been

proposed and open a huge eld for new research. For example, magnetic Fe-NMs

have been demonstrated to increase the reactivity and cation exchange capacity of

biochar [243, 244], increasing the uptake of nutrients by the plants since the mag-

netic Fe-NMs can also increase nutrient availability and decomposition of soil

organic matter. Also, due to their antimicrobial properties, NMs may increase the

resistance of plants to stress and produce indirect plant growth stimulation. Detailed

knowledge of NMs ecotoxicity to bacteria and other soil microorganisms is also

lacking. It is essential to understand the diversity of the aspects involving engi-

neered NMs and plants if major advances in new elds are to be made.

Acknowledgments Domingo Martínez-Fernández is grateful for nancial support from the post-

doctoral grant (19835/PD/15) nanced by the “Consejería de Educación y Universidades de la

CARM”, through the “Fundación Séneca-Agencia de Ciencia y Tecnología de la Región de

Murcia”. Michael Komárek is thankful for the support from the Czech Science Foundation (project

15-07117S). The English revision by Dr. David J.Walker is also acknowledged.

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A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_15

Chapter 15

Phytoremediation Application: Plants

asBiosorbent forMetal Removal inSoil

andWater

RashaH.Mahmoud andAmalHassaneinMohammedHamza

Abstract Phytoremediation for metal-contaminated soils was started about 40

years ago, and the phytoremediation for organic pollutants is more recent.

Phytoremediation has gained extensive attention and much progress in remediation

of inorganic and organic contaminants and as the means for enhanced phytoreme-

diation. Phytoremediation of various inorganic pollutants such as Cd, Cr, Pb, Cu,

Zn, Co, Ni, Se, Cs, and As has been extensively studied. This is mainly based on the

use of natural hyperaccumulator plants with exceptional metal-accumulating capac-

ity, which can take up metals to concentrations at least an order of magnitude greater

than the normal plants growing in the same environment. These plants have several

benecial characteristics such as the ability to accumulate metals in their shoots and

an exceptionally high tolerance to heavy metals.

Keywords Phytoremediation • Heavy metal contamination • Hyperaccumulator

plants • Phytoextraction • Phytostabilization • Phytovolatilization • Rhizoltration

R.H. Mahmoud (*) • A.H.M. Hamza

Biochemistry Department, Faculty of Science, King Abulaziz University,

Jeddah, Saudi Arabia

Biochemistry and Nutrition Department, Faculty of Women, Ain Shams University,

Cairo, Egypt

e-mail: dr.rasha.asu@gmail.com; amal_hamza@hotmail.com

guarino@unisannio.it

406

15.1 Introduction

Phytoremediation for metal-contaminated soils was started about 40 years ago, and

the phytoremediation for organic pollutants is more recent. Phytoremediation has

gained extensive attention and much progress in remediation of inorganic and

organic contaminants and as the means for enhanced phytoremediation.

Phytoremediation of various inorganic pollutants such as Cd, Cr, Pb, Cu, Zn, Co,

Ni, Se, Cs, and As has been extensively studied. This is mainly based on the use of

natural hyperaccumulator plants with exceptional metal-accumulating capacity,

which can take up metals to concentrations at least an order of magnitude greater

than the normal plants growing in the same environment. These plants have several

benecial characteristics such as the ability to accumulate metals in their shoots and

an exceptionally high tolerance to heavy metals.

At present, there are totally more than 400 species of hyperaccumulator plants

for As, Cd, Mn, Ni, Zn, etc. Phytoremediation is a general term including several

processes, in function of the plant-soil-atmosphere interactions. For heavy metal-

contaminated soil, four processes of phytoremediation are recognized: phytoextrac-

tion, phytostabilization, phytovolatilization, and rhizoltration. The rst two

mechanisms are the most reliable. The different forms of phytoremediation require

different general plant characteristics for optimum effectiveness [1].

15.2 Denition andConcept

Phytoremediation can be dened as the process, which uses green plants for the

relief, transfer, stabilization, or degradation of pollutants from soil, sediments, sur-

face waters, and groundwater. Some plant roots can absorb and immobilize metal

pollutants, while other plant species have the ability of metabolizing or accumulating

organic and nutrient contaminants [2]. Multifarious relationships and interactions

between plants, microbes, soils, and contaminants make these numerous phytoreme-

diation processes possible. The term phytoremediation, from the Greek phyto, means

“plant”, and the Latin sufx remedium, “able to cure” or “restore”. It can be used for

a wide range of organic and inorganic contaminants [2]. Phytoremediation processes

are most effective where contaminants are present at low to medium levels, as high

contaminant levels can inhibit plant and microbial growth and activity [3].

Mechanisms involved in the uptake, translocation, and storage of micronutrients are

the same involved to translocate and storage heavy metals [1].

Phytoremediation is considered an economical and environmentally friendly

method of exploiting plants to extract contaminants from soil [4]. This process is

relatively cost-effective compared with other remediation techniques. However, a

thorough economic analysis for this process is unavailable. Most phytoremediation

studies are directed at the biological, biochemical, and agronomic processes [5]. An

economic outlook, instead of simple estimates of the cost advantages of phytoreme-

diation over other techniques, has not been reported.

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15.3 Advantages andLimitations ofPhytoremediation

Mechanisms

Phytoremediation, like other remediation technologies, has a range of both advan-

tages and disadvantages. The most positive aspect of using phytoremediation is as

follow: (1) more cost-effective; (2) more environmentally friendly; (3) applicable to

a wide range of toxic metals, and (4) more aesthetically pleasing method. On the

other hand, phytoremediation presents some limitations. It is a lengthy process, thus

it may take several years or longer to clean up a site and it is only applicable to

surface soils [6].

Prior to phytoremediation eld trials, extensive research was performed in labo-

ratories and greenhouses. Some of this work explored the effects of plants on

removal of contaminants from spiked soil and soil excavated from contaminated

sites. Many of these experiments provided valuable insights into the types and spe-

cic mechanisms of phytoremediation of organic contaminants [7]. Some organic

compounds can be transported across plant membranes. Of these, the low molecular

weight compounds can often be removed from the soil and released through leaves

via evapotranspiration processes (phytovolatilization). Some of the non-volatile

compounds can be degraded or rendered non-toxic via enzymatic modication and

sequestration in plants (phytodegradation, phytoextraction). Other compounds are

stable in the plants and can be removed along with the biomass for sequestration or

incineration.

15.4 Basics ofPhytoremediation Process

The discovery of metal-accumulating properties in certain plants leads to the devel-

opment of phytoremediation technology. Research in the eld of phytoremediation

is aiming to develop innovative, economical, and environmentally compatible

approaches to remove heavy metals from the environment. Even apart from the

metal hyperaccumulating property of the plants, the presence of ground cover with

plants helps to shield people from direct contact with the soil and prevents the blow-

ing of contaminated dust around the neighbourhood [8].

15.5 Types ofPhytoremediation Technologies

Depending upon the process by which plants are removing or reducing the toxic

effect of contaminants from the soil, phytoremediation technology can be broadly

classied as follows [9].

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15.5.1 Phytoextraction

This is the process of using pollutant-accumulating plants to remove metals or

organics from soil by concentrating them in harvestable plant parts.

15.5.2 Phytotransformation

This is the partial or total degradation of complex organic molecules by their incor-

poration into plant tissues.

15.5.3 Phytostimulation

In this process, the release of plant exudates or enzymes into the root zone stimu-

lates the microbial and fungal degradation of organic pollutants.

15.5.4 Phytostabilization

This is a method that uses plants to reduce mobility of contaminants (both organic

and metallic contaminants) by preventing erosion, leaching, or runoff and to reduce

bioavailability of pollutants in the environment, thereby preventing their migration

to groundwater or their entry into the food chain [10].

15.5.5 Phytovolatilization

This is the technique of using plants to volatilize pollutants or metabolites.

This technology can be used for volatile organic carbons (VOCs) and for the few

inorganics that can exist in volatile forms such as selenium and mercury [10].

15.5.6 Rhizo-Filtration

This is the use of plant roots to absorb or adsorb pollutants, mainly metals, but also

organic pollutants, from water and aqueous waste streams.

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15.5.7 Pump andTree

This method is the use of trees to evaporate water and simultaneously to extract pol-

lutants from the soil [11].

15.5.8 Hydraulic Control

It is the controlling of water table and soil eld capacity by plant canopies [12].

15.6 Plant Selection Considerations

Plant species for phytoremediation are selected based on their root depth, the nature

of the contaminants and the soil, and regional climate. The root depth directly

impacts the depth of soil that can be remediated. It varies greatly among different

types of plants and can also vary signicantly for one species depending on local

conditions such as soil structure, depth of a hard pan, soil fertility, cropping pres-

sure, contaminant concentration, or other conditions [13].

The cleaning depths are approximately phytoremediation. It has been reported

that for phytoremediation, grasses are the most commonly evaluated plants [14].

They have been more preferable in use for phytoremediation because compared to

trees and shrubs, herbaceous plants, especially grasses, have characteristics of rapid

growth, large amount of biomass, strong resistance, effective stabilization to soils,

and ability to remediate different types of soils [2]. They are pioneers and usually are

adapted to adverse conditions such as low soil nutrient content, stress environment,

and shallow soils [15]. The large surface area of their brous roots and their inten-

sive penetration of soil reduce leaching, runoff, and erosion via stabilization of soil

and offer advantages for phytoremediation. Wild plants such as grasses can produce

closures above ground quickly and reduce dispersion of the dust of tailings [16].

Shrubs and trees produce extensive canopy cover and produce deep roots to pre-

vent erosion in the long term. In addition, shrubs or trees provide high nutrient to the

grass while lowering water stress and improve soil physical properties [17]. Many

trees can grow on land of marginal quality, have massive root systems, and their

above-ground biomass can be harvested with subsequent resprouting without dis-

turbance of the site. However, the cost for planting trees is high and the growth rate

is low [18].

To achieve a stable persistent cover, it is important to use a mixed culture and

combine grasses, shrubs, and trees in revegetation programs of mining soils because

they represent two functional types of plants with different roles in the improvement

of mine soils. For a longer duration, as considered for most phytoremediation

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processes, it cannot be expected to clean up the soil only by one plant species used

exclusively in monoculture. Grasses, with their highly developed root system, can

stabilize the soils and reduce erosion, while legumes can add nitrogen to the soil,

preparing the establishment of other plant species typical of later stages of succes-

sion [19].

Perennial grasses develop a large plant biomass in a relatively short time and are

recognized as heavy metal-tolerant biosystems, accumulating high levels of these

elements. However, the shorter growing period of the seasonal owering plants is a

better option in phytoremediation over perennial plants, as it can be harvested yearly

or seasonally, and the area can be replanted with subsequent seasonal owering

plants [20]

For phytoremediation, it is better to use plant species adapted to the climatic and

soil conditions of the area to be de-polluted [18]. Use of indigenous plant species is

generally favored because they show tolerance to imposed stress conditions, require

less maintenance, and present fewer environmental and human risks than non-native

or genetically altered species [17]. However, particular non-native plant may work

best remediation of specic contaminant and can be safely used under circum-

stances where the possibility of invasive behavior has been eliminated [21].

15.7 Heavy Metal Removal by Phytoremediation

15.7.1 Heavy Metals in Soil

Heavy metals are the major environmental contaminants and pose a severe threat to

human and animal health by their long-term persistence in the environment. The

remediation of soils contaminated by heavy metals is a cost-intensive and techni-

cally complex procedure. Conventional remediation technologies are based on bio-

logical, physical, and chemical methods, which may be used in conjunction with

one another to reduce the contamination to a safe and acceptable level. In spite of

being efcient, these methods are expensive, time-consuming, and environmentally

destructive [22].

15.7.2 Sources ofMetal Pollution

Geological and anthropogenic activities are sources of heavy metal contamination.

Sources of anthropogenic metal contamination include industrial efuents, fuel pro-

duction, mining, smelting processes, military operations, utilization of agricultural

chemicals, small-scale industries (including battery production, metal products,

metal smelting, and cable coating industries), brick kilns, and coal combustion [23].

One of the prominent sources contributing to increased load of soil contamination

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is disposal of municipal wastage. These wastes are either dumped on roadsides or

used as landlls, while sewage is used for irrigation. These wastes, although useful

as a source of nutrients, are also sources of carcinogens and toxic metals. Other

sources can include unsafe or excess application of (sometimes banned) pesticides,

fungicides, and fertilizers [23]. Additional potential sources of heavy metals include

irrigation water contaminated by sewage and industrial efuent leading to contami-

nated soils and vegetables [24].

15.7.3 Metal Toxicity

All plants have the ability to accumulate “essential” metals (Ca, Co, Cu, Fe, K, Mg,

Mn, Mo, Na, Ni, Se, V, and Zn) from the soil solution. Plants need different concen-

trations for growth and development. This ability also allows plants to accumulate

other “non-essential” metals (Al, As, Au, Cd, Cr, Hg, Pb, Pd, Pt, Sb, Te, Tl, and U),

which have no known biological function [25]. Moreover, metals cannot be broken

down, and when concentrations inside the plant cells accumulate above threshold or

optimal levels, it can cause direct toxicity by damaging cell structure (due to oxida-

tive stress caused by reactive oxygen species) and inhibit a number of cytoplasmic

enzymes. In addition, it can cause indirect toxic effects by replacing essential nutri-

ents at cation exchange sites in plants [26].

15.7.4 Soil Metal Groups

Metals are natural components in soil. Based on their role on physiological activi-

ties, they can be divided in two groups: (1) Essential heavy metals (Fe, Mn, Cu, Zn,

and Ni) which are micronutrients necessary for vital physiological and biochemical

functions of plant growth. They are constituents of many enzymes and other pro-

teins and all plants have the ability to accumulate them from soil solution, (2) Non-

essential metals (Cd, Pb, As, Hg, and Cr) have unknown biological or physiological

function and consequently are non-essential for plant growth [27]. Both groups are

toxic to plants, animals, and humans above certain concentrations specic to each

element. High contents of both essential and non-essential heavy metals in the soil

may inhibit plant growth and can lead to toxicity symptoms in most plants [28].

However, some plant species have the ability to grow and develop in metallifer-

ous soils such as near to mining sites. Such plants can be used to clean up heavy

metal-contaminated sites. Willow (Salix viminalis L.), maize (Zea mays L.), Indian

mustard (Brassica juncea L.), and sunower (Helianthus annuus L.) have been

found to be highly tolerant to heavy metals. Vetiver grass (Vetiveria zizanioides)

showed tolerance to Pb and Zn and it can be used for revegetating Pb/Zn mine tail-

ings. Populus species are examples of plants widely used to remediate heavy metal-

contaminated soils [29].

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15.7.5 Heavy Metals

Heavy metals are natural constituents of the earth’s crust. Their principal character-

istics are an atomic density greater than 5 g cm−3 and an atomic number >20. The

most common heavy metal contaminants are Cd, Cr, Cu, Hg, Pb, and Zn. From the

geochemical point of view, trace elements are metals whose percentage in rock

composition does not exceed 0.1%. The occurrence of heavy metals in soils can be

the result of two main sources:

Natural source: Heavy metals occur naturally in the soil environment from the

pedogenetic processes of weathering of parent materials at levels that are regarded

as trace (<1000mg kg−1) and rarely toxic [30].

Anthropogenic sources: Human activities, such as mining, smelting, electroplat-

ing, energy and fuel production, power transmission, intensive agriculture, sludge

dumping, and melting operations, are the main contributor to heavy metal contami-

nation. Heavy metals in the soil from anthropogenic sources tend to be more mobile,

hence bioavailable than pedogenic, or lithogenic ones. The industry of mining and

processing metals is a major source of farmland heavy metal contamination [31].

15.7.6 Heavy Metal Phytoavailability

Bioavailability and phytoavailability are terms used to describe the degree to which

contaminants are available for absorption or uptake by living organisms that are

exposed to them. Plants respond only to the fraction that is “phytoavailable” to them

[32]. For heavy metal phytoremediation (and phytoextraction in particular), bio-

availability of metals in contaminated soils is a crucial factor regulating heavy metal

uptake by plant roots. However, metal phytoavailability is a complex phenomenon

that is dependent on a cascade of related factors [33].

15.7.6.1 Soil pH

Soil pH directly inuences the phytoavailability of metals as soil acidity determines

the metal solubility and its ability to move in the soil solution. Metal cations are the

most mobile under acidic conditions, while anions tend to be absorbed to oxide

minerals in this pH range [18].

15.7.6.2 Soil Texture

Texture reects the particle size distribution of the soil and thus the content of ne

particles like oxides and clay [34]. Particle size distribution can inuence the level

of metal contamination in a soil. Fine particles (<100 μm) are more reactive and

have a higher surface area than coarser material.

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15.7.6.3 Soil Organic Matter

Soil organic matter is frequently reported to have a dominant role in controlling the

behavior of trace metals in the soil. The organic matter is one of the factors that may

reduce the ability of metals to be phytotoxic in the soil due to metal-organic com-

plexation [35].

15.7.6.4 Redox Potential

The redox potential is one of the most soil properties that affect changes in metal

speciation. Redox potential in soil is established by oxidation-reduction reactions

resulting from microbial activity [36].

15.7.6.5 Root Zone

Plant root can inuence heavy metal phytoavailability by modifying the soil proper-

ties in the rhizosphere. The plant enzymes exuded from the roots should play a key

role in the transformation and chemical speciation of heavy metals in soils, which

facilitate their uptake by plant [37].

15.8 Phytoremediation Technologies inRemoving Soil Metals

15.8.1 Phytoextraction

This technology involves the extraction of metals by plant roots and the translocation

thereof to shoots. The roots and shoots are subsequently harvested to remove the con-

taminants from the soil. Salt etal. [38] reported that the costs involved in phytoextrac-

tion would be more than ten times less per hectare compared to conventional soil

remediation techniques. Phytoextraction also has environmental benets because it is

considered a low impact technology. Furthermore, during the phytoextraction proce-

dure, plants cover the soil and erosion and leaching will thus be reduced. With succes-

sive cropping and harvesting, the levels of contaminants in the soil can be reduced [39].

15.8.2 Phytostabilization

Also referred to as in-place inactivation, it is primarily used for the remediation of

soil, sediment, and sludges. It is the use of plant roots to limit contaminant mobility

and bioavailability in the soil. The plants’ primary purposes are to (1) decrease the

amount of water percolating through the soil matrix, which may result in the

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formation of a hazardous leachate, (2) act as a barrier to prevent direct contact with

the contaminated soil, and (3) prevent soil erosion and the distribution of the toxic

metal to other areas [8].

Phytostabilization can occur through the sorption, precipitation, complexation,

or metal valence reduction. It is useful for the treatment of lead (Pb) as well as arse-

nic (As), cadmium (Cd), chromium (Cr), copper (Cu), and zinc (Zn). Some of the

advantages associated with this technology are that the disposal of hazardous mate-

rial/biomass is not required and it is very effective when rapid immobilization is

needed to preserve ground and surface waters. The presence of plants also reduces

soil erosion and decreases the amount of water available in the system [21].

Phytostabilization has been used to treat contaminated land areas affected by

mining activities and Superfund sites. The experiment on phytostabilization by

Jadia and Fulekar [40] was conducted in a greenhouse, using sorghum (brous root

grass) to remediate soil contaminated by heavy metals and the developed vermi-

compost was amended in contaminated soil as a natural fertilizer. They reported that

growth was adversely affected by heavy metals at the higher concentration of 40 and

50 ppm, while lower concentrations (5–20 ppm) stimulated shoot growth and

increased plant biomass. Further, heavy metals were efciently taken up mainly by

roots of sorghum plant at all the evaluated concentrations of 5, 10, 20, 40, and 50

ppm. The order of uptake of heavy metals was: Zn > Cu > Cd > Ni > Pb. The large

surface area of brous roots of sorghum and intensive penetration of roots into the

soil reduces leaching via stabilization of soil and is capable of immobilizing and

concentrating heavy metals in the roots.

15.8.3 Rhizoltration

This technique is primarily used to remediate extracted groundwater, surface water,

and wastewater with low contaminant concentrations [41]. It is dened as the use of

plants, both terrestrial and aquatic, to absorb, concentrate, and precipitate contami-

nants from polluted aqueous sources in their roots. Rhizoltration can be used for

Pb, Cd, Cu, Ni, Zn, and Cr, which are primarily retained within the roots [21].

Sunower, Indian mustard, tobacco, rye, spinach, and corn have been studied for

their ability to remove lead from water, with sunower having the greatest ability.

Indian mustard has a bioaccumulation coefcient of 563 for lead and has also

proven to be effective in removing a wide concentration range of lead (4–500mg

L−1) [8]. The advantages associated with rhizoltration are the ability to use both

terrestrial and aquatic plants for either in situ or ex situ applications. Another advan-

tage is that contaminants do not have to be translocated to the shoots.

An experiment on rhizolteration by Karkhanis etal. [42] was conducted in a

greenhouse with duckweed and water hyacinth (Eichornia crassipes) to remediate

aquatic environment contaminated by coal ash containing heavy metals.

Rhizolteration of coal ash started from 0, 5, 10, 20, 30, 40%. Simultaneously, the

physicochemical parameters of leachate have been analyzed and studied to under-

stand the leachability. The results showed that pistia has high potential capacity of

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uptake of the heavy metals (Zn, Cr, and Cu) and duckweed also showed good poten-

tial for uptake of these metals next to pistia. Rhizoltration of Zn and Cu in case of

water hyacinth was lower as compared to pistia and duckweed. This research shows

that pistia/duckweed/water hyacinth can be good accumulators of heavy metals in

aquatic environment [43].

15.8.4 Phytovolatilization

This technique involves the use of plants to take up contaminants from the soil,

transforming them into volatile forms, and transpiring them into the atmosphere

[21]. Mercuric mercury is the primary metal contaminant that this process has been

used for. The advantage of this method is that the contaminant, mercuric ion, may

be transformed into a less toxic substance (that is, elemental Hg). The disadvantage

to this is that the mercury released into the atmosphere is likely to be recycled by

precipitation and then redeposited back into lakes and oceans, repeating the produc-

tion of methyl-mercury by anaerobic bacteria.

15.9 Metal Uptake by Plants

This depends on the concentration of soluble and bioavailable fraction of metals in

the soil solution. The bioavailable fraction of metal in the soil can be determined by

the Potential Bioavailable Sequential Extraction (PBASE) procedure [18]. Even

though chemical extraction won’t extract metal from the soil in a manner identical

to that of a plant root system, it can be used as a reliable method for assessing the

bioavailability of metals bound to soil particles [44].

Plants extract and accumulate metals from soil solution. Before the metal can

move from the soil solution into the plant, it must pass the surface of the root. This

can either be a passive process, with metal ions moving through the porous cell wall

of the root cells, or an active process by which metal ions move symplastically

through the cells of the root. This latter process requires that the metal ions traverse

the plasmalemma, a selectively permeable barrier that surrounds cells [10].

In a polluted soil, the concentration of bioavailable pollutants tends to reduce

over time due to physical, chemical, and biological processes. Because of this rea-

son, aged soils are more difcult to phytoremediate [10]. It is known that to enhance

metal solubility, plants either excrete organic ligands or lower the soil pH in the

rhizosphere. To improve metal solubility in the soil solution, synthetic chelates such

as ethylenediaminetetraacetic acid (EDTA), nitrilotriacetic acid (NTA), pyridine- 2-

6-dicarboxylic acid (PDA), citric acid, nitric acid, hydrochloric acid, and uorosi-

licic acid can be used in phytoremediation studies [45]. The addition of excess

chelating agents may increase the chances of leaching the metals from the soil to

groundwater. If the metal concentration in the soil is near to the phytotoxic levels,

addition of lime or organic matter reduces the metal solubility [10].

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15.9.1 Phytoremediation ofAs, Cd, Pb, andZn

Arsenic pollution is one of the major concerns in the world due to its chronic effects

on the health of human beings. Recently, it was proposed that phytoremediation

could be an effective tool for arsenic clean up [46]. Research in this eld has mainly

concentrated on arsenic contamination in the aquatic environment. Studies have

been done to remove arsenic from contaminated soil and revealed that Chinese

brake fern (Pteris vittata) is an efcient As accumulator. This plant is not suitable

for a region like Oklahoma, where the climate is too dry, even though it can be used

with higher metal 20 concentrations. Also, the concentration of Zn affects the

growth of P. vittata. A study has shown that a concentration of 1242mg Zn kg−1 in

soil causes phytotoxicity to the ferns [46]. Cadmium is present in most of the zinc-

contaminated sites. Different plants such as indian mustard (Brassica juncea), wil-

low clones (Salix), alpine penny-cress (Thlaspi caerulescens), sunower (Helianthus

annus), and corn (Zea mays) are able to accumulate Cd. Brassica juncea was able

to accumulate cadmium from a soil with a concentration of 200mg Cd kg−1 in soil.

Experiments showed that Thlaspi caerulescens can be a good phytoremediator in a

soil with 390mg Cd kg−1. Helianthus annus and Zea mays were also found as good

accumulators in soil with a cadmium concentration of 90mg kg−1 [47].

There are many plants that can accumulate lead in a very high concentration in

its different parts. Brassica juncea can be effectively used as a phytoremediator for

soils with lead contamination up to 500mg Pb kg−1 of soil. Helianthus annus and

Zea mays have been grown in a soil with a concentration of 16,000mg Pb kg−1 [48].

Research using Piptatherum miliaceum (Smilo grass) has shown that this species

can be used for remediating the metal contamination in a soil with 300–1500mg Pb

kg−1 concentration [49]. Thlaspi praecox is able to accumulate a considerable

amount of Pb from soil with a concentration of 67,940mg Pb kg−1 [50]. Hemidesmus

indicus has been shown to remove 65% of the lead effectively from a soil having

10,000 ppm of lead concentration [51]. Most of the superfund sites in US are con-

taminated with zinc. Studies showed that Piptatherum miliaceum (Smilo grass) can

be used for 21 phytoremediation in a soil with 100–600mg Zn kg−1 concentration

[49]. Helianthus annus and Zea mays have been grown in soil with a concentration

of 75,000mg Zn kg−1 and found to accumulate zinc in their harvestable parts [48].

15.9.2 Plants asBiosorbents forHeavy Metals Removal

inWaste Water

Wastewater is a mixture of pure water with large number of chemicals (including

organic and inorganic) and heavy metals, which can be produced from domestic,

industrial and commercial activities, in addition to storm water, surface water, and

ground water [52]. Due to the danger of the entry of chemicals into wastewater, it

must be treated before the nal disposal. Many physical, chemical, and biological

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methods have been developed for the treatment of wastewater. It is reported that

biological methods are more interesting for wastewater treatment and one of the

branches of biological method for wastewater treatment is phytoremediation [53].

The concept of this method is based on the using of plants and microorganisms in

the same process as to remove the pollutants from environment [54].

Among phytoremediation techniques, articial wetlands (AW) is known to be as

the most effective technology to treat wastewater. The AWs can promote biodiver-

sity via preparation of alarge habitat for a wide number of wildlife such as the rep-

tiles, rodents, shes, and birds. It should be noted that the selection of suitable

species of plants is important for the implementation of phytoremediation [53].

The selected species must contain the following features: (1) high ability to

uptake both organic and inorganic pollutants; (2) high ability to grow faster in

wastewater; and (3) should be easy to control. It should be also noted that the ability

of pollutant removal varies from species to species, plant to plant within a genus

[55]. The rate of photosynthetic activity and plant growth have a key role during the

implementation of phytoremediation technology for the removal of low to moderate

amount of pollutants [56]. In addition to water hyacinth, plants like Water Lettuce

(Pistia stratiotes), Duckweed (Water lemna), Bulrush (Typha), Vetiver Grass

(Chrysopogon zizanioides), and Common Reed (Phragmites australis) have been

successfully implemented for the treatment of wastewater containing different types

of pollutant [57]. Nowadays, human health is being threatened with the release of

polluted wastewater in presence of heavy metals into the environment.

Lasat [58] has shown that plants are successful in removing the heavy metals.

The use of plants as biosorbents for the removal of heavy metals is considered to be

inexpensive, effective, and eco-friendly technology. Phytoremediation can be con-

sidered advantageous if the plant is considered to be as solar-driven pump which can

concentrate and extract particular type of elements present in the polluted wastewa-

ter. The root of the plant helps to absorb the pollutants existing in the wastewater,

particularly the heavy metals and will help in improving the quality of water [59].

Water hyacinth has been widely studied in the laboratory at pilot and large scale

for the removal of organic matter present in the waste water in comparison to other

aquatic plants. Although water hyacinth is known to be a persistent plant all over the

world, it is being widely used as a main resource for waste management and agri-

cultural process [60]. Both the eld and laboratory studies have shown that water

hyacinth is capable of removing large number of pollutants present in the swine

wastewater [61]. Duckweed and water hyacinth are being considered for the treat-

ment of dairy and pig manure-based wastewater [59]. The treated wastewater in the

presence of water hyacinth for the duration of 25 days resulted in the reduction of

solids, calcium, magnesium, and total hardness. Wastewater from duck farm was

treated by water hyacinth and resulted in 64, 23, and 21% removal of COD, TP, and

TN, respectively [62]. In combination of water hyacinth and duckweed for treating

dairy wastewater, it could remove 79% of total nitrogen and 69% of total phospho-

rus [57].

Chen et al. [63] demonstrated that 36% of nitrogen and phosphorus could be

removed from swine wastewater using water hyacinth. Also reported among the

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different forms of nitrogen, ammonical nitrogen was found to be removed to a

greater extent when compared to other forms of nitrogen.

Ismail etal. [64] showed the efciency of water hyacinth and water lettuce for

the uptake of nitrate, ortho-phosphate, nitrite and ammoniacal nitrogen. It was

found that water hyacinth exhibited better performance for reducing nitrate in com-

parison to orthophosphate. Valipour etal. [65] in their latest study showed that the

roots of water hyacinth are primarily involved in the transportation, where the

shoots resulted in the accumulation of considerable amount of nutrients (N and P)

in comparison to the root area.

Liao and Chang [66] ranked the heavy metal removal rate based on the ability of

water hyacinth to remove (Cu > Zn > Ni > Pb > Cd) and showed that higher and

lower removal efciency belonged to Cu and Cd, respectively. Xiaomei etal. [67]

used water hyacinth for the removal of Zn and Cd from wastewater and also mea-

sured the concentration of Cd and Zn absorbed in different parts of water hyacinth

(stem, leaves, roots, owers). It was observed for the presence of 2040mg kg−1 of

Cd and 9650mg kg−1 of Zn accumulated in the roots of water hyacinth. According

to Shaban etal. [68], to treat 1L of wastewater contaminated with 1500mg L−1

arsenic requires 30 g of dried water hyacinth root for a period of 24 h estimated

chromium(III) removal from the aqueous solution and found the removal rate to be

87.52% with 10mg Cr/1 solution. Gupta and Balomajumder [69] found that water

hyacinth can uptake more than 99% of phenol in a single and twofold solution of Cr

and Phenol (at 10 mg L−1) in 14 and 11 days, respectively. Padmapriya and

Murugesan [70], during their study for the removal of heavy metals in aqueous solu-

tion using water hyacinth, found Langmuir and Freundlich models tted well for the

biosorption of all the metal ions.

15.10 Fate ofAbsorbed Metals inPlant

The metals absorbed in a plant can accumulate in various parts of the plant. For an

effective phytoremediation process, the metals should be accumulated in a harvest-

able part of the plant. Brake fern, one of the major plants for arsenic phytoremedia-

tion, accumulated almost 95% of arsenic taken up into the aboveground biomass.

The arsenic concentration in the brake fern root was the least when compared to the

other parts. The highest concentration was reported in old fronds followed by young

fronds, ddle heads, and rhizomes [71]. Arsenate usually enters the plant root

through the phosphate uptake system, and to limit the toxicity, the plant chemically

reduce As(V) to As(III) in the roots. In the case of Indian mustard, a large portion of

absorbed As remains in the root itself and a small amount of arsenic is transported

to the shoots; however, the addition of water-soluble As-chelators can increase this

fraction [72]. In most plants, the major portion of absorbed Cd remains in the root

of the plant and only some is translocated to the shoots [72].

Sunower accumulates zinc mostly in the stem (437.81mg Zn kg−1 dry weight)

and lead in roots (54.53mg Pb kg−1 dry weight). In the case of corn, lead and zinc

R.H. Mahmoud and A.H.M. Hamza

guarino@unisannio.it

419

were accumulated more in leaves (84.52mg Pb kg−1 dry weight) (1967mg Zn kg−1

dry weight) [48]. Hemidesmus indicus 22 accumulates lead in the shoots [51] and

Smilo grass accumulates lead in roots and zinc in shoots [49]. Experiments on

Thlaspi praecox revealed that Zn and Cd accumulate in the shoots and their concen-

tration in the shoots is linearly correlated with total soil Zn and Cd concentrations,

thus conrming that the plant can be used for the phytoremediation of soil contami-

nated with Zn and Cd. At the same time, 80% of the accumulated lead is immobi-

lized in the roots [50].

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A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_16

Chapter 16

Nutrient Management Strategies forCoping

withClimate Change inIrrigated Smallholder

Cropping Systems inSouthern Africa

DavieM.Kadyampakeni, IsaacR.Fandika, andLawrentL.M.Pungulani

Abstract Sound management of soil nutrients is critical for optimizing crop

vegetative and reproductive development and realizing high yields in irrigated crop-

ping systems. This paper discusses the work done in Africa and presents lessons

from other parts of the world for improved nutrient management under irrigation.

Considering the rising temperatures and erratic rainfall as a consequence of climatic

change and depleted soil nutrients as a result of continuous cropping, this review

offers remedial options for managing soil fertility while optimizing water use and

crop yields. The paper intends to inform agricultural policy makers and help farmers

and organizations in Africa to manage soil nutrient and water resources efciently

and achieve high yields. Importantly, this discussion should stimulate further

research in nutrient and water management under varying ecological scenarios

of southern Africa to provide a cogent basis for climate change adaptation

interventions.

Keywords Irrigation management • Nutrient depletion • Nutrient use efciency

• Water use efciency

D.M. Kadyampakeni (*)

Soil and Water Sciences Department, University of Florida, Citrus Research and Education

Center, 700 Experiment Station Rd, Lake Alfred, FL 33850, USA

International Water Management Institute, PMB CT 112, Cantonment, Accra, Ghana

e-mail: dakadyampakeni@yahoo.com

I.R. Fandika

Department of Agricultural Research Services, Kasinthula Agricultural Research Station,

P.O.Box 28, Chikwawa, Malawi

e-mail: fandika68@gmail.com

L.L.M. Pungulani

Department of Agricultural Research Services, Chitedze Agricultural Research Station,

P.O.Box 158, Lilongwe, Malawi

e-mail: lawrentp@yahoo.co.uk

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16.1 Introduction

Agricultural water management in southern Africa uses about 6% of cultivated area

(Table 16.1). This is far below the irrigation potential pegged at 37% of cultivated

land [1]. Average annual precipitation estimated at 9.07×1012m3 and annual agri-

cultural water withdrawals of approximately 20.41 × 109m

3 suggest increasing

potential for water extraction and irrigated agricultural production [1]. One con-

straint for intensifying irrigated production is lack of site-specic nutrient guide-

lines for managing crops under irrigation. Farmers have been managing irrigated

crop nutrients by trial and error, for example, using recommendations for crops

grown during the rainy seasons due to lack of guidelines for irrigated nutrient man-

agement. Besides this, the recommendations that have been developed in the region

are limited to a few crops such as sugar cane and rice. Considering the gross nutrient

mining in sub-Saharan Africa, there is need for integrated sustainable agricultural

systems of nutrient management by (1) applying modest amounts of mineral fer-

tilizer according to site-specic recommendations, (2) improving water storage and

nutrient retention through efcient use of manure and household waste, and

(3) properly timed or split application of mineral fertilizers and appropriate tillage

and conservation measures [2–5]. In view of lack of these guidelines, this paper

documents the merits and demerits of nutrient management options that could be

tested and adapted to specic agro-ecologies of southern Africa under irrigated con-

ditions. This approach, together with judicious use of nutrient- and water-efcient

and high yielding crop varieties, should lead to increased expansion of cultivated

area and improved crop yields, water use, and food security, while helping farmers

mitigate the effects of climate change. The paper discusses several nutrient manage-

ment options for increasing crop yield, mitigating adverse climate change scenarios

and conserving water quality.

16.2 Nutrient Management Strategies

16.2.1 Fertilization withSoluble Inorganic Fertilizer Sources

ViaIrrigation

The practice of applying water via irrigation is commonly called fertigation. It is a

practice of applying macronutrients particularly N, K, and S in irrigation water [6].

This method is used extensively in commercial agriculture and horticulture (vegeta-

bles, fruits trees, and other high value crops) to supply additional nutrients or cor-

rect nutrient deciencies detected in plant tissue analysis. Fertilizer injection during

middle one-third or the middle one-half of the irrigation is recommended for ferti-

gation using microirrigation to prevent the nutrients from accumulating near the soil

surface or leaching beyond the root zone and terminating fertilizer application

before irrigation completion.

D.M. Kadyampakeni et al.

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425

Table 16.1 Water use patterns for southern Africa: agriculture water withdrawals, irrigation potential, irrigated area, and precipitation (FAO [1])

Country

Potential

irrigation (ha)

Average annual

precipitation

(106m3)

Agriculture water

withdrawals (106m3)

Total cultivated

area (ha)

Total water- managed

area (including wetlands

and valley bottoms) (ha)

Angola 3,700,000 1,258,790 211 3,300,000 400,000

Botswana 13,000 241,825 80 380,000 7939

Burundi 215,000 35,460 222 1,351,000 104,430

Congo 7,000,000 3,618,120 112 7,800,000 13,500

Lesotho 12,500 23,928 0.6 334,000 2637

Malawi 161,900 139,960 810 2,440,000 118,290

Mozambique 3,072,000 827,161 550 4,435,000 118,120

Namibia 47,300 235,253 213 820,000 9573

Rwanda 165,000 31,932 102 1,385,000 102,500

South Africa 1,500,000 603,926 7836 15,712,000 1,498,000

Swaziland 93,220 13,678 1006 190,000 49,843

Tanzania 2,132,221 1,012,191 4632 5,100,000 184,330

Zambia 523,000 767,700 1320 5,289,000 255,922

Zimbabwe 365,624 256,729 3318 3,350,000 193,513

Total 19,000,765 9,066,653 20,413 51,886,000 3,058,597

16 Nutrient Management Strategies forCoping withClimate Change inIrrigated…

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16.2.2 Merits andDemerits

Benets of fertigation over traditional broadcast or drop-fertilizing methods include

(1) timely correction of in-season nutrient deciencies, (2) application nutrients in

synchrony with crop demand, (3) use a low volume of water, at low pressure, result-

ing in low energy costs, (4) improved nutrient and water use-efciency due to

increased root density and reduced nutrient leaching, (5) reduced incidence of pest

and weed invasion, and other plant diseases, due to a reduced wetted area and drier

soil surface [6–10].

Challenges associated with use of fertigation are ascribed to its requirement of

properly designed irrigation system and skilled irrigation management, currently

lacking in most parts of southern Africa, with the exception of South Africa. Thus,

hands-on training and demonstrations of fertigated agriculture would improve

Southern African farmer’s skills in fertigation. In addition, fertigation may not work

well with ood or furrow-irrigated systems because a lot of the nutrients may be

deposited near the inlet. Also, application of fertilizers with high amounts of anhy-

drous NH3 and Ca2+, Mg2+, and HCO3

− may precipitate CaCO3 and MgCO3, causing

emitter clogging. The invention of functional and cost-effective sand and screen

ltration devices has helped to overcome the clogging problems and has ultimately

resulted in an expanded use of micro-irrigation in areas with low-quality water [10].

16.3 Use ofOrganic Manures

Typical compositions of organic animal and plant manures are presented in

Table 16.2 [11–14]. The composition of animal manure and crop residues and such

typical composition ranges could be used as guidelines and provide the basis for

calculating manure application rates inlocal smallholder cropping systems.

16.3.1 Merits andDemerits

Organic manures (1) increase plant-available N, P, and other micronutrients organic

matter complexation, (2) improve soil organic matter content, (3) increase soil mois-

ture retention, improve soil structure, and increase inltration rate, and (4) reduce

Al3+ toxicity in acid soils by complexation with organic matter [6].

The low nutrient content and bulkiness restricts greatly the distance manure can

be transported, often no more than 10 km. Chemical composition of manure is

highly variable; it is sometimes difcult to apply a specic amount of nutrients

when manure is spread. Mineralization of manure is dependent on many factors and

not well-controlled by the producer, thus there is potential for nitrate leaching [15].

D.M. Kadyampakeni et al.

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427

16.4 Integrated Inorganic andOrganic Fertilizer Sources

PlusIrrigation

Several studies showed the importance of good irrigation (supplemental during the

rainy season or full irrigation during the dry season) and nutrient management. For

example, Kim etal. [16] demonstrated that maize plants responded to N and water

simultaneously resulting in ~ 61 to 68% greater water use efciency and 48–44%

greater fertilizer use efciency due to improved crop growth, evapotranspiration

(ET), and N transport to the root zone. Other studies showed similar benets in

irrigated maize, rice, and vegetables, using either inorganic fertilizer sources or a

combination of organic and inorganic fertilizers [14, 17–21]. Several cropping sys-

tems that have been tested and could be adapted to local conditions in southern

Africa are presented in Table 16.3.

Table 16.2 Typical composition for a number of organic wastes and manures

Nutrient source N (%) P (%) K (%) Mg (%) C (%)

Schumann [13]

Animal manure 1–4 0.22–0.87 0.58–1.66 naana

Agricultural wastes 0.5–2.3 0.04–0.26 0.96–1.31 na na

Agricultural industrial wastes 0.5–1.2 0.06–1.75 0.33–1.66 na na

Miller and Donahoue [12]

Cattle 2–8 0.2–1.0 1–3 1–1.5 na

Poultry 5–8 1.0–2.0 1–2 2–3 na

Swine 3–5 0.5–1 1–2 0.1 na

Sheep 3–5 0.4–0.8 2–3 0.2 na

Cattle 2–8 0.2–1.0 1–3 1–1.5 na

Poultry 5–8 1.0–2.0 1–2 2–3 na

Mafongoya etal. [11]

Acacia karro 2.0 0.25 na na 4.90

Acacia nilotica 1.1 0.46 na na 3.80

Colososper mopane 1.0 0.24 na na 3.76

Gliricidia sepium 0.9 0.43 na na 4.15

Hay 2.2 0.17 na na 4.32

Acacia karro 2.0 0.25 na na 4.90

Manure (poultry) 2.3 1.5 2.7 na 15.2

Manure (Ruminant) 1.0 0.3 1.7 na 9.0

Solid waste 0.28 0.15 0.58 na 2.54

aNot applicable

16 Nutrient Management Strategies forCoping withClimate Change inIrrigated…

guarino@unisannio.it

Table 16.3 Examples of fertilizer management options that could be tested and adapted for agro-ecologies in sub-Saharan Africa

Country

Rate (kgha−1)

Irrigation rate Cropping system Irrigation system Soil texture SourceN P K

Malawi 120 na 92 40mm very 3–4

days, 40mm every

week, 40% ASWD

Maize (Zea mayz L.) Furrow Sandy loam Fandika etal. [17]

Malawi 120 na 92 40mm very 3–4

days, 40mm every

week, 40% ASWD

Maize (Zea mayz L.) Furrow Sandy loam Fandika etal. [17]

Iowa, USA 30, 140, 250 30 30 510mm, 640mm

per season

Maize (Zea mayz L.) Center-pivot Sandy loam Al-Kaisi and Yi [36]

UK 200 30 134 20mm every day Maize (Zea mayz L.) naaSandy loam Ogola etal. [37]

India 60, 120, 180 30 30 874mm, 1274mm

of irrigation plus

rainfall per season

Rice (Oryza sativa) Flood Loamy sand Mahajan etal. [38]

Philippines 200 25 40 na Rice-rice (Oryza

sativa)

Flood Clay and silty

clay

Dobermann etal. [39]

Indonesia 120 17 33 na Rice-rice (Oryza

sativa)

Flood Clay and clay

loam

Dobermann etal. [39]

Vietnam 80 17 25 na Rice-rice (Oryza

sativa)

Flood Clay Dobermann etal. [39]

China 120 17 33 na Rice-rice (Oryza

sativa)

Flood Clay loam Dobermann etal. [39]

China 90 19 62 na Rice-rice (Oryza

sativa)

Flood Silty clay

loam

Dobermann etal. [39]

China 120 17 33 na Rice-rice (Oryza

sativa)

Flood Silty clay

loam

Dobermann etal. [39]

India 120 17 33 na Rice (Oryza sativa)-

wheat (Triticum

aestivum)

Flood Silty loam Dobermann etal. [39]

guarino@unisannio.it

India 100 21 41 na Rice (Oryza sativa)-

mungbean (Vigna

radiata)-sesame

(Sesamum indicum L.)

Flood Clay Dobermann etal. [39]

Colorado,

USA

134, 179 112 Na 780mm per season Onion (Allium cepa) Drip Clay loam Halvorson etal. [40]

Colorado,

USA

134, 179 112 Na 2230mm per

season

Onion (Allium cepa) Furrow Clay loam Halvorson etal. [40]

Canada 270 30, 60, 90 200 0.47Lh−1, 1–2 h

daily

Tomato (Solanum

lycopersicum)

Drip Sandy Liu etal. [41]

Jordan 113, 170 55 183 381mm per season Potato (Solanum

tuberosum)

Drip (including

rainfall)

Clay loam Mohammad etal. [42]

Jordan 79, 157, 236 39, 58 Na 343 per season Squash (Cucurbita

spp.)

Drip Silt loam Mohammad [43]

Colorado,

USA

168, 224, 280 na Na 735, 1064mm per

season

Maize (Zea mayz L.) Furrow Silty clay Halvorson and Bartolo

[44]

Nebraska,

USA

197, 243 22, 80 Na 64–178mm per

season

Maize (Zea mayz L.) Center-pivot Silt loam Djaman etal. [45]

North Dakota,

USA

90, 135, 180,

225

57, 67 Na 106–136mm per

season

Maize (Zea mayz L.) Center-pivot Silt loam,

silty clay

Derby etal. [46]

New Jersey,

USA

560 336 504 152mm per season

(plus 620mm rain)

Maize (Zea mayz L.) Trickle Sandy loam Karlen etal. [47]

Oklahoma,

USA

118, 236, 354 na Na na Maize (Zea mayz L.) Center-pivot Loam Freeman etal. [48]

North Dakota,

USA

100, 200 34 140 na Maize (Zea mayz L.) na Loam, sandy

loam

Wienhold etal. [49]

(continued)

guarino@unisannio.it

Portugal 80, 160, 240 145 224 17, 98, 114, 256,

390, 527mm of

irrigation plus rain

Potato (Solanum

tuberosum)

Sprinkler Silt, ne sand Ferreira and Carr [50]

China 240 53 100 258, 443mm

irrigation, plus

579mm rain and

298, 414mm

irrigation plus

665mm rain per

season

Rice-rice (Oryza

sativa)

Border Silty clay Ye etal. [51]

Italy 53, 105 26 Na 185–346mm

irrigation plus

169mm rain,

190–373mm plus

120mm rain per

season

Maize (Zea mayz L.) Drip Silty clay

loam

Di Paolo and

Rinaldi [52]

Burkina Faso 260, 353 49, 99 50, 60 320mm irrigation

plus 573mm rain,

317mm irrigation

plus 180mm rain

per season

Tomato (Solanum

lycopersicum)

na Sandy loam Sangare etal. [14]

Burkina Faso 405, 835 87, 156 237, 255 298mm irrigation

plus 558mm rain,

155mm irrigation

plus 294mm rain

per season

Cabbage (Brassica

oleracea)

na Sandy loam Sangare etal. [14]

Table 16.3 (continued)

Country

Rate (kgha−1)

Irrigation rate Cropping system Irrigation system Soil texture SourceN P K

guarino@unisannio.it

Burkina Faso 128.5 39, 53 117, 220 406, 482mm of

irrigation per

season

Carrotn (Daucus

carota)

na Sandy loam Kangare etal. [53]

Burkina Faso 415, 468 79, 114 97, 440 302mm irrigation

plus 27mm rain,

175mm irrigation

plus 402mm rain

per season

Lettuce (Lactuca

sativa)

na Sandy loam Sangare etal. [14]

Burkina Faso 313, 453 63, 107 357, 438 412, 598mm of

irrigation per

season

Lettuce (Lactuca

sativa)

na Sandy clay Sangare etal. [14]

Burkina Faso 247, 396,

439, 486,

504, 601

58, 68,

90, 95,

120

285, 331,

383, 413,

496, 678

175–359mm of

irrigation and

29–538mm of rain

per season

Lettuce (Lactuca

sativa)

na Sandy clay Sangare etal. [14]

Lebanon 240, 360, 480 na na 490mm by drip

and 850mm by

sprinkler irrigation

Potato (Solanum

tuberosum)

Drip, Sprinkler Clay Darwish etal. [54]

India 60, 120, 180 30 30 1279, 1679mm of

irrigation plus

rainfall per season

Rice (Oryza sativa) Flood Loamy sand Mahajan etal. [38]

China 240, 270 60, 75 75 412–425, 528–545,

529–547mm of

irrigation plus

538–802mm of

rainfall per season

Rice (Oryza sativa) Flood Sandy loam Liu etal. [21]

(continued)

guarino@unisannio.it

Australia 150 15 na 310, 380mm of

irrigation

Wheat (Triticum

aestivum)

na na Whiteld etal. [55]

Florida, USA 176, 220, 330 49 247 1.1, 2.6, 3.8,

4.6mm per day of

surface or

subsurface drip

irrigation

Tomato (Solanum

lycopersicum)

Surface or

subsurface drip

Sandy Zotarelli etal. [19, 20]

Florida, USA 73, 82, 145,

164

na na 0.9, 21.1mm per

day of surface or

subsurface drip

irrigation

Zucchini squash

(Cucurbita pepo L.)

Surface or

subsurface drip

Sandy Zotarelli etal. [18]

South Dakota,

USA

56, 112, 168 na na 540, 480, 480mm

of irrigation plus

rainfall per season

Maize (Zea mayz L.) na Silty clay

loam

Kim etal. [16]

aNot available

Table 16.3 (continued)

Country

Rate (kgha−1)

Irrigation rate Cropping system Irrigation system Soil texture SourceN P K

guarino@unisannio.it

433

16.5 Mixed andRotational Cropping Systems

The mixed cropping and crop rotation will be important in managing fragile envi-

ronments. Examples of mixed cropping systems would include cereal-legume or

legume-vegetable, or cereal-vegetable arrangements. Use of legumes, in particular,

would reduce the emission of greenhouse gases such as NO, CO2, and N2O due to

their ability to symbiotically x N compared with cereals [22]. Examples of legumes

adapted to Africa include cowpea (Vigna unguiculata L.), common bean (Phaseolus

vulgaris L.), soya bean (Glycine max L.), groundnut (Arachis hypogaea L.), pigeon

pea (Cajanus cajan), chickpea (Cicer arietinum), bambara groundnut (Vigna sub-

terranea), and lentil (Lens culinaris). They occupy well over 15 million hectares in

Africa with yields in the order of 200–1400kgha−1 [23].

16.5.1 Merits andDemerits

The advantages of mixed and rotation cropping include increased productivity from

the same piece of land through (1) better use of solar radiation, (2) increase in nutri-

ent and water use efciency, and (3) better control of weeds, pests, and diseases [24].

These cropping systems, particularly crop rotation, might be limited where land

sizes per farm family are fairly small (0.5ha or less).

16.6 Mitigating Greenhouse Gas Emissions inIrrigated

Crop Production

Many environmentalists have been concerned with greenhouse gas (GHG) emis-

sions in irrigated cropping systems due to N volatilization and denitrication losses

(NO, N2O) [25–27]. In southern Africa, Meixner etal. [28] found that NO uxes are

largely controlled by soil moisture. Thus, optimizing water management and fertil-

izer use would reduce the impact of N losses on GHG emissions. For example, in

the various agroecosystems studied, the greatest NO emissions (27ngN m−2s−1)

were found in the agricultural plots. In China, researchers found that improved man-

agement of irrigation, timing of fertilizer applications, and split fertilizations

increased maize yields and reduced N2O and NO emissions by 7 and 29%, respec-

tively, with 7 to 14% greater yield in irrigated maize [29, 30]. The remedial and

mitigation options for reducing N-related GHG emissions include reduction in N

fertilizer use through an increase in fertilizer use efciency, preferential use

of NH4NO3 instead of urea, improved timing of fertilizer application, the use of

nitrication and urease inhibitors, improving the fertilizer uptake efciency of

16 Nutrient Management Strategies forCoping withClimate Change inIrrigated…

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434

crops in tropical agriculture, and intercropping cereals with legumes [22, 31–35].

Quan tication of the effects of manures and fertilizers on GHG emissions in irri-

gated systems for development of climate change mitigation strategies is lacking in

southern Africa. Thus, climate-change scenarios investigated in developed countries

should provide important insights for developing and modifying nutrient manage-

ment strategies for southern Africa’s irrigated cropping systems.

16.7 Water Quality Monitoring inIrrigated Cropping

Systems inSouthern Africa

Systems for monitoring water quality in irrigation schemes are nonexistent in sub-

Saharan Africa [14]. There is a need for developing guidelines for total nutrient

loads for macro- and micro-nutrients because excessive application rates may nega-

tively affect aquatic life and other important terrestrial organisms. While the pri-

mary goal of a nutrient management program is to increase crop yield, the corollary

objective should be to conserve water quality for other ecological uses. Periodic

monitoring of drainage ditches and in-eld water sampling in irrigation schemes

and 3–4 year experiments that compare various nutrient and water application rates

would provide the requisite benchmarks for developing thresholds for fertilizer

application rates for maximizing crop yield and water use while conserving envi-

ronmental quality.

16.8 Conclusions

The paper presented selected options for improved nutrient management in irriga-

tion systems for adaptation to southern Africa farming systems. Technologies that

conserve water and increase nutrient use efciency will be important in helping

farmers realize high yields and greater farm incomes. Long-term goals for good

nutrient management in southern Africa include (1) sustaining environmental qual-

ity by minimizing nutrient leaching to surcial and ground water sources, (2)

replenishing soil fertility in nutrient-depleted soils, and (3) developing site-specic

nutrient and water management recommendations for smallholder farming systems.

The rst necessary tasks will need on-farm adaptive-verication nationwide trials to

validate and modify the elite technologies in concert with selected crop genotypes.

Participatory technology selection by farmers, coupled with detailed nancial anal-

ysis, will help in ensuring that water and nutrient management technologies identi-

ed are acceptable and economically feasible.

D.M. Kadyampakeni et al.

guarino@unisannio.it

435

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A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_17

Chapter 17

Phytoremediation ofLandll Leachates

PrasannaKumarathilaka, HasinthaWijesekara, NanthiBolan,

AnithaKunhikrishnan, andMeththikaVithanage

Abstract Municipal landll leachate is a complex refractory wastewater which

consists of extensive level of organic compounds, ammonia, and heavy metals.

Contamination of water by landll leachate has become a serious environmental

concern worldwide due to its adverse impact on human health, aquatic organisms,

and agricultural crop production. In recent years, constructed wetland (CW) has

received promising attention in the treatment of landll leachate, because of its cost-

effective and eco-friendly nature and simplicity in operation, in addition to higher

treatment efciency. Hence, the present chapter is mainly focused on providing a

concise discussion of the CWs and its phytoremediation attributes for the remedia-

tion of landll leachate. Natural wetland plant species and short rotation coppice

(SRC) have been introduced to remove contaminants from landll leachate.

Different processes such as phytoextraction, phytodegradation, phytovolatilization,

rhizoltration, phytostabilization, rhizo-redox reactions, sedimentation, adsorption,

and complexation involve to remove nutrients (i.e., nitrogen and phosphate), heavy

metal(loid)s, biological oxygen demand (BOD), and chemical oxygen demand

(COD) to a great extent in CW systems. In addition, well-managed SRC systems

save millions of dollars by eliminating the leachate transportation and treatment

process which were earlier practiced. Further, there are a number of examples where

phytoremediation has failed due to excessive leachate application and lack of man-

agement practices. Therefore, it is obvious that successful transfer of phytoremedia-

tion technologies from the laboratory to the eld is a crucial step in terms of removal

efciency.

P. Kumarathilaka, B.Sc. • M. Vithanage, Ph.D. (*)

Environmental Chemodynamics Project, National Institute of Fundamental Studies,

Hantana Road, Kandy 20000, Sri Lanka

e-mail: prasannakumarathilaka@gmail.com; meththikavithanage@gmail.com

H. Wijesekara, M.Phil. • N. Bolan, Ph.D.

Global Centre for Environmental Remediation, Faculty of Science and Information

Technology, University of Newcastle, Callaghan, NSW, Australia

e-mail: wijesekara84@gmail.com; Nanthi.Bolan@newcastle.edu.au

A. Kunhikrishnan, Ph.D.

Chemical Safety Division, Department of Agro-Food Safety, National Academy of

Agricultural Science, Wanju-gun, Jeollabuk-do 55365, Republic of Korea

e-mail: a.kunhikrishnan@hanmail.net

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440

Keywords Wetlands • Evapotranspiration • Macrophytes • Nutrients • Heavy

metal(loid)s

17.1 Municipal Solid Waste Dumpsites

In many developing and developed countries, disposal of waste to open landll sites

is the most common method of waste management (Fig.17.1). For instance, it is

reported that approximately 90% of landlls in South and Southeast Asia are non-

engineered open disposal facilities [1]. From an economic point of view, these open

dump sites provide simple and cheap means of waste disposal. On the contrary,

numerous contaminants including liquids, gases, and dusts may pollute the

Waste disposal (%)

Country

Turkey

Mexico

Poland

Greece

Hungary

New Zealand

Slovakia

Iceland

Portugal

Gfinland

Italy

USA

Spain

South Korea

France

Norway

Luxembourg

Germany

Belgium

Austria

Denmark

Sweden

Japan

Netherlands

Switzerland

Ireland

Australia

Czech..

United..

100

Fig. 17.1 Percentage of waste disposal in landlls in numerous nations. Reproduced from Bolan

NS, Thangarajan R, Seshadri B, Jena U, Das KC, Wang H, Naidu R (2013) Landlls as a bioren-

ery to produce biomass and capture biogas. Bioresource Technol 135: 578–587 [3], with permis-

sion from Elsevier

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surrounding environment, if the open landll sites are not properly managed [2, 3].

In this regard, soil and groundwater contamination, atmospheric pollution by gas

emissions, and the consequent adverse health problems are particularly obvious in

the open dump sites.

Depending on the source of landll material, the contaminant type and the extent

of deleterious effects vary signicantly. Landlls could be divided into three main

categories based on the waste materials received, namely (1) industrial waste, (2)

municipal waste, and (3) a combination of industrial and municipal wastes [4]. As a

result, different landll sites play different role towards environmental contamina-

tion. For example, industrial waste is consisted of various kinds of hazardous mate-

rials including heavy metal(loid)s. These components can be easily leached from

the landll site to the groundwater and surface water bodies. On the contrary,

municipal waste is mainly composed of variety of organic materials. Subsequent

decomposition of these organic materials by microorganisms results in the release

of leachate and emission of volatile organic compounds (VOCs) and greenhouse

gases (GHGs) [5–7]. VOCs are generally more toxic at trace level than many other

inorganic compounds and are implicated in carcinogenic and mutagenic effects in

humans and animals. In addition, geomembranes are not effective in preventing the

transport of VOCs, since they are able to diffuse readily through geomembrane

polymers even in engineered landlls [8].

Non-biodegradable contaminants including heavy metal(loid)s are of signicant

concern in managing landll sites, since these can be accumulated in soils and con-

taminated water bodies due to leaching and runoff. There have been a number of

reports demonstrating the contamination of soil around the landll site by various

heavy metal(loid)s, such as Copper (Cu), lead (Pb), nickel (Ni), zinc (Zn), and cad-

mium (Cd) [9, 10]. The microbial activity and soil quality can be seriously affected,

consequently producing unfavorable conditions for plant growth. Deleterious

effects of various heavy metal(loid)s depend upon not only the total metal concen-

tration, but also the bioavailable fraction of a particular heavy metal(loid). However,

owing to changes in soil properties, bioavailable fraction of heavy metal(loid)s can

be increased, leading to potential toxicity to living organisms [4, 11].

17.2 Environmental Issues ofLeachate

The main environmental aspect related to open dump sites is the discharge of leachate

into the environment. More precisely, leachate contamination is the result of mass

transfer process. Over its whole life-cycle process, the waste in landll undergoes

biological, chemical, and physical transformations. Primarily, three physical stages,

solid phase (waste), liquid phase (leachate), and gas phase, can be identied in the

landll [12]. Since leachate contains high concentrations of contaminants including

heavy metal(loid)s, xenobiotic organic compounds, dissolved and suspended

organic matter, high biological oxygen demand (BOD), and chemical oxygen

demand (COD), most detrimental concern associated with leachate discharge is that

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of soil, groundwater, and surface water contamination [4]. However, leachate char-

acteristics may vary due to the climatic variations, solid waste composition, and age

of the landll. In most climatic conditions, precipitation and snow accelerate the

saturation of waste in the landll with subsequent generation of leachate [13]. In

addition, aerosols and malodors released during landll handling and treatment is a

serious concern. It is reported that some persistent pollutants such as dissolved car-

bon originate from household chemicals, industries, and co- deposited hazardous

waste [14]. Leachate discharge into the soil and groundwater systems is likely to

undergo attenuation processes including dilution, biodegradation, and other

physico-chemical processes such as evaporation and adsorption [12]. Consequently,

the interactions between these processes and the leachate load lead to the formation

of leachate plume. The extent of leachate plume will determine the environmental

risks to the surface water ecology and to human health via water supplies from

groundwater and surface water [15]. Furthermore, terrestrial and aquatic plant spe-

cies could be seriously damaged due to the uptake of contaminants with high

concentrations.

It is obvious that the leachate impact on groundwater and surface water quality

correlated with improper landll management strategies in the past. Different fac-

tors such as landll sitting, design, operation, maintenance, and cost mainly govern

the discharge of leachate into the environment [16]. Similarly, large number of land-

lls is located on the ground or on a slope, hence accumulation of leachate could be

a negative factor due to geotechnical stability [2]. Since there is no lining or leachate

draining system in those open dump sites, the leachate problem persists for a quite

long time, possibly many decades even after the landll closure [17]. These cases

showed that the importance of proper management practices is a must to control the

effect of leachate. Figure17.2 shows leachate generation and ow into the environ-

ment from non-engineered landll sites or open dump sites.

Fig. 17.2 Non-engineered landll operation site (a) and subsequent production of landll leach-

ate (b)

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17.3 Pollutants inLandll Leachate

The landll leachate has been identied as an intense pollutant causing severe con-

tamination to the soil, water, and air ecosystems (Fig.17.3). Due to the complexity

of pollutants in landll leachate, different ways are used to explain these pollutants.

Basically, these pollutants are grouped into heavy metal(loid)s, xenobiotic organic

Recycling and reuse

Waste generation

Residential

Commercial and institutional

Industrial

Refuse derived fuel

(RDF) manufacture

Composting Dispose to open

landfill sites

Pyrolysis for

biochar

production

Biogas

production

Electricity

generation

Fertilizer

applications

Environmental

remediation

applications

Electricity

generation /

Transportation

Landfill leachate production

• High concentration of heavy metal(loid)s

• Inorganic macro components

• High BOD and COD

• Xenobiotic organic compounds

• Accumulation in soils

• Contamination of water bodies

• Uptake by plants

Detrimental effect to terrestrial and aquatic organisms

Aerobic and anaerobic decomposition

Runoff and infiltration

Greenhouse gas

emissions

Release of volatile and

semivolatile organic

compounds

Fig. 17.3 Different waste management strategies and adverse effects of landll leachate

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compounds (XOCs), organic compounds, and inorganic compounds [5]. However,

some of these groups are not mutually exclusive. On the contrary, most of the pol-

lutants do not exist in the leachate as their unique fundamental molecules or ions,

but present as complex matrices such as organic-metal(loid) derivatives. Microbial

species and their toxins can be also counted as pollutants, since some of these spe-

cies such as eubacteria and archaea are found to be dominated in landll leachate

[18–20].

The presence of alloys, paints (i.e., lead-based paints), automotive parts, lamp

laments, iron scraps, ceramics, and batteries (i.e., nickel–cadmium batteries) in a

landll contributes to the presence of heavy metal(loid)s in leachates [2, 21, 22].

Many researchers have reported a wide variation of heavy metal(loid) concentra-

tions in leachates, thereby expressing the potential risk to the environment [2, 23].

Besides the sample handling techniques and protocols, the colloidal matter have a

greater afnity to heavy metal(loid)s in leachate, therefore the heavy metal(loid)

concentration may depend on the colloid content in leachate [5, 24]. The precipita-

tion and sorption of heavy metal(loid)s should be uncounted to understand their fate

and transport in the environment and in treatment plants.

Landll leachate is one of the predominant wastewater types containing a wider

range of XOCs such as benzene, phenol, trichloroethene, and chlorinated aliphatics.

However, relatively low concentrations (i.e., <1mg L−1) of individual XOCs are

reported in many cases [5, 25]. Most of these XOCs are derived from re-retardants

(i.e., tri(2-chloroethyl) phosphate), insect repellents (i.e., N,N-diethyltoluamide),

and pharmaceuticals (i.e., uoxetine and ibuprofen) [19, 26–28]. Some of the XOCs

(e.g., toluene and phenol) are reported to cause serious carcinogenic and teratogenic

effects to animals through chronic exposure [29].

Due to the decomposition of different types of wastes originated from slaughter

houses, sh markets, and households, high concentrations of organics (e.g., recalci-

trant substances such as humic and fulvic acids) and inorganic compounds (e.g.,

nitrate (NO3−), nitrite (NO2−), ammonium (NH4+), sulfate (SO42−), chloride (Cl−),

uoride (F−)) can be found in leachates from open dumpsites, thereby causing seri-

ous environmental pollution [2, 30]. Therefore, comprehensive characterization of

landll leachate is essential to design their appropriate treatment methods. In any

case, temporal and spatial variation of the quantity and quality of landll leachate is

the most difcult issue in leachate management.

17.4 Conventional Treatment Methods forLandll

Leachates

Landll leachate treatment technologies can be divided into three basic types, leach-

ate recirculation, biological, and physicochemical. In most cases, combinations of

these treatment technologies or integrated systems are used to design sustainable

leachate treatment facilities. Additional environmental benets such as energy crops

and biogas production are also associated with some of these treatment methods [31].

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17.4.1 Leachate Recirculation

Recirculation of landll leachate into the landll tip or into the landll bioreactor is

one of the cost-effective treatment methods. Many advantages have been reported

with this technique such as reduction of leachate volume and time required for

waste stabilization, enhanced biogas production, and improved leachate quality

[32]. Combined treatment of landll leachate with the domestic sewage is also iden-

tied as another type of leachate treatment method [31, 33]. A lab-scale study for

treating a mixed substrate as landll leachate and domestic sewage has been suc-

cessful, concluding the importance of operative strategies such as pretreatment of

leachate [34]. However, the leachate transfer methods do not behave as ultimate

treatment methods, since they potentially pose disadvantages such as increment of

efuent concentrations and inhibition of microbial activities during leachate recir-

culation [31].

17.4.2 Physico-chemical Treatment

A wide range of physico-chemical treatment methods are used for the treatment of

landll leachate. These include air or ammonia-stripping, adsorption, membrane l-

tration, coagulation–occulation, chemical precipitation, chemical and electrochemi-

cal oxidation (i.e., ozonation), evaporation, reverse osmosis, photoelectrooxidation,

and sedimentation–otation [23, 31, 35, 36]. Basically, these physico-chemical treat-

ment methods lead to a reduction of suspended solids, colloidal particles, color,

pathogens, and toxic compounds including excess nutrients in the leachates. For

example, activated carbon adsorption is used as a common method for the removal of

dissolved organics and heavy metal(loid)s [37]. Further, waste materials such as

waste steel scrap and slag are used in this type of treatment methods [38]. In most

cases, a combination of these physico-chemical treatment methods has been used for

achieving the targets and generally reects better treatment abilities [31, 39]. However,

the expensive nature for establishment and handling of these techniques limits their

usage signicantly and raises greater challenges for their commissioning.

17.4.3 Biological Treatment

Biological treatment methods are associated with the microorganisms (i.e., biodeg-

radation through aerobic and anaerobic microorganisms) and plant species (i.e.,

phytoremediation). These techniques are mainly used for the treatment of organic

pollutants and nutrients in leachate. Aerated lagoons, sludge processing reactors,

biolm reactors, and biolters are examples of techniques that use aerobic biologi-

cal processes [31]. Diffusers or mechanical aerators are generally used to supply

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aerobic conditions to these systems. Examples of anaerobic biological processes are

anaerobic digesters, lters, and sequencing batch reactors [40]. Constructed wet-

land (CW) is a common type of biological treatment method [41]. Due to the inex-

pensive nature, reliability, and simplicity, these biological techniques are widely

used in landll leachate treatment. However, limited effectiveness of treatment pro-

cesses with aging and low efciency for treating old landll leachates are identied

as drawbacks associated with some of these biological treatment methods [31, 35].

17.5 Phytoremediation ofLandll Leachate

Advanced physico-chemical and biological leachate treatment technologies require

continuous budget and energy supply and sufcient technical capabilities for the

operation and maintenance of the equipments [16]. Therefore, high-tech solutions

are not sustainable for many landll sites, particularly in developing countries.

Hence, a sustainable leachate treatment strategy with economical and technical fea-

sibility and climatic compatibility is the only viable option in such regions. Over the

last few decades, CW system has been recognized as an appropriate and practical

alternative for landll leachate treatment, making it safe to discharge into the sur-

rounding environment [42–44]. Besides their small ecological footprint, CW sys-

tems possess similar aesthetic value as natural wetland systems.

Constructed wetland systems consist of different media types, and typically,

same species of emergent plants. Constructed wetland system is mainly classied

into free water surface system (FWS) and subsurface ow system (SSFS). In FWS

(Fig.17.4a), oxygen is prevalently introduced into the wetland via algal photosyn-

thesis and atmospheric diffusion. In SSFS, leachate ows underneath and through

the plant rooting media, and subsequently leachate level is maintained below the tip

of the substratum. In other words, SSFS may act as xed-lm bioreactors [45]. In

terms of fewer issues arising from odors, disease-related vectors, and public expo-

sure, SSFS is highly recommended for landll leachate treatment [46]. SSFS is of

two types, horizontal and vertical (Fig.17.4b, c). In the horizontal ow systems

(HFS), the leachate is fed into the inlet and continues its way under the surface of

the bed in a more or less horizontal path until it reaches the outlet zone. Conversely,

in the vertical ow systems (VFS), landll leachate is fed on the whole surface area

through distribution system and passes the lter in a more or less vertical path. In

VFS, greater oxygen transport is involved compared to the HFS [47, 48]. As a result,

VFS is more efcient for removing ammoniacal nitrogen (NH3-N) and organic mat-

ter from landll leachate. The efciency of landll leachate remediation achieved

by CW depends upon different factors including the type of media used (sand,

gravel, clay, or silt), availability of microorganisms, and selectivity of plants (mono-

culture or mixed beds) [49].

In general, CW systems are receiving untreated or partially treated leachate

[50]. For example, reed beds are considered unsuitable for primary treatment of

high strength landll leachate due to the toxicity of leachate to the reed. However,

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aerobic biological pre-treatment of landll leachate may lead to subsequent removal

of contaminants effectively by reeds in CW system [51]. In the case of SSFS, suf-

cient mechanical pretreatment is required to remove excessive amount of sus-

pended solids, since they may cause ltration-bed clogging and consequent surface

ow. Aerated lagoons have been recognized as the prevalent pretreatment method

for treating concentrated landll leachate from landll operations [52].

Outlet

Fine gravel

Sand and gravelCollection pipe with drain pipe laterals

Root-bed media (clay, gravel)

Water level

Macrophytes

(a)

(b)

(c)

Macrophytes

Gravel layer

Coarse media pre-filter

Macrophytes

Inlet pipe

Gravel-rooting media

Fig. 17.4 Layout of different types of CW systems (a) surface ow (b) subsurface-horizontal ow

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17.5.1 Physico-chemical Properties ofSoil

Soil physico-chemical parameters play a major role for the purication capacity as

well as longevity of CW systems. Physico-chemical parameters including pH,

moisture content, bulk density, total organic carbon, and total nitrogen content could

be effectively utilized to determine the status of soil quality [53]. More precisely,

hydraulic conductivity (K) is a key factor that is closely correlated with the ef-

ciency of landll leachate treatment. On comparing different types of soils, sandy

and gravelly soils possess high K values, whereas clay soils have low K values [51].

Therefore, the water movement through sandy and gravelly soils occurs rapidly, and

hence, soil–water contact is decreased. Similarly, in clay soils also, rapid water

movement and reduced soil–water contact take place. On the contrary, it is reported

that silt or loamy soils mixed with sand and gravel may enhance soil–water contact

at a great extent [51]. It has been experimentally observed that ammonia (NH3),

COD, total nitrogen, salinity, and conductivity were low when landll leachate was

percolated via clay and sandy soil column compared to the sandy soil alone. It can

be attributed to adsorption of dissolved components including humic and fulvic acid

onto soil particles, ion exchange, or precipitation [54].

Particle size and pore size distribution in a particular soil in CW systems has

gained a signicant attention among the soil physical parameters. Bruch etal. [55]

assessed ve different lava and one uviatile operating sand lters for their differ-

ences in pore size distribution, specic inner surface area, and cumulative pore vol-

ume. The results revealed that these soil physical parameters had an inuence on

purication capacity and hydraulic conductivity. The alteration of soil hydraulic

properties may inuence the hydrology of CW system, thereby affecting the removal

efciency. During the wetland construction process, occurrence of soil disturbance

including the mass grading of local landscape and redistribution of the upper soil

horizons is observed [56]. Soil disturbance involves compaction and the vertical

integration of soil horizons and regolith, which leads to new soil textures and change

in bulk density. Additionally, loss of macrostructure and increment of clay amount

may change the way water is held within the soil matrix. A study by Campbell etal.

[57] observed that soils in CW consisted less organic matter, greater bulk density,

and increased rock fragments. It is reported that compaction removes all macro

structure within the soil matrix in terms of long-term operation basis [56].

17.5.2 Importance ofRhizosphere Microbiology

Since the microorganisms are the rst organisms which deal with the pollutants in

landll leachate, they should be having an own mechanism to grow and overcome

extremely toxic conditions in CW systems [58]. The health of soil ecosystems basi-

cally depends on the biological processes including decomposition of organic mat-

ter and nutrient cycling. Soil microorganisms involve transformations of organic

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matter and mineralization of nitrogen (N), phosphorus (P), and sulfur (S) [59].

Additionally, microbes play a crucial role in the degradation of complex chemical

compounds into simpler components, which can be easily absorbed by plant spe-

cies. Wetland plant species are capable of growing in environments where their root

system is submerged. These plants transport air towards the root system by diffusive

and/or convective mechanisms via specialized gas channel tissues called aeren-

chyma [60]. Subsequent leakage of air from the root system may provide aerobic

conditions within the water-saturated soil instead of anaerobic conditions. Hence,

soil rhizosphere provides habitats for both aerobic and anaerobic microorganisms.

On the contrary, organic compounds such as sugars, alcohols, and acids which are

released by the plant species into the rhizosphere act as a carbon source for micro-

organisms [61]. In this way, soil itself facilitates appropriate habitats for different

types of microorganisms such as bacteria, fungi, and yeasts [49].

Assessing soil microbial parameters such as microbial biomass and basal respi-

ration and biochemical parameters such as enzyme activities can predict the

response of microbes to environmental changes including temperature and mois-

ture, as well as pollution [62]. It is obvious that changes in microbial biomass lead

to detrimental impacts in soil health. Therefore, different microbial testing tech-

niques can be used to investigate soil quality with reliable and accurate measure-

ments [51]. For instance, uorescence in situ hybridization (FISH) technique can be

utilized to recognize the number and relative distribution of bacterial species and

their strains. Thus, this technology facilitates insights into the diversity of microor-

ganisms in CW systems. A study by Sawaittayothin and Polprasert [63] revealed

that the predominance of bacteria such as heterotrophys and autotrophs is respon-

sible for BOD5 removal from the landll leachate, and FISH technique was used for

evaluating phylogenetic identity, morphology, and number of microorganisms.

17.5.3 Selection ofPlant Species forLeachate Remediation

Typical natural wetland plant species such as cattail (Typha latifolia L.), reed

(Phragmites australis), rush (Juncus effuse L.), yellow ag (Iris pseudacorus L.),

and mannagrass (Glyceria maxima) are used in CW [64]. Table17.1 summarizes

various plant species used in CW systems in different countries around the world

and Fig.17.5 shows successful utilization of differentplants for the treatment of

reverse osmosis rejected concentrate at Medawachchiya, Sri Lanka. The inherent

features of wetland plant species such as extremely high transpiration rates, frost

resistance, disease resistance, and tolerance to high heavy metal(loid) concentra-

tions make them successful for remediation purposes. Additional characteristics

including ease of rooting, fast establishment, quick growth, extensive rates of pho-

tosynthesis, and elevated usage of water make them successful in CW systems [42].

Further, the clear advantage of using vegetation species in a CW is provision of

supporting media for biological activities. Moreover, ecological advantages

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Table 17.1 Use of different plant species for the removal contaminants in landll leachate

Landll

leachate

location CW conguration Plant species Operational parameters of CW Reference

Wola

Pawlowska,

Poland

1m high lysimeter of 0.6m in

diameter and 0.28m3 in volume

Willow (Salix amygdalina L.) Hydraulic loading rate (HLR)—1, 3

and 5mm day−1

Białowiec

etal. [69]

Reed (Phragmites australis)

Rhinelander,

USA

Trees were planted at a spacing of

3m within rows and 4m between

rows

Hybrid poplar (Populus nigra L. × P.

maximowiczii A.Henry ‘nM6’)

Thermal dissipation probes were

used to measure sapow

Zalesny etal.

[42]

Vrhnika

municipal

centre,

Slovenia

Length of 25cm long cuttings were

planted in 12L pots

Poplar (Populus deltoids Bartr. cl. I-69/55

(Lux),

Compost-soil mixture in the range

of 1:2 on a volume:volume basis

was utilized

Justin etal.

[44]

Two willow species (Salix viminalis L. and

Salix purpurea L.)

Da-Liao MSW

site, Taiwan

Each system consisted of two types

of basins in series, free water surface,

and subsurface ow microcosm

Reed (P. australis) HLR was 0.014m3 m−2 day−1Yang and Tsai

[101]

Cattail (Typha orientalis)

Virens (Dracaena sanderiana)

Oneida country

landll, USA

Cuttings were planted in a split plot

design with eight blocks at a spacing

of 1.2 × 2.4m

Poplar (Populus maximowiczii) Drip irrigation was used to apply

treatments and application rate was

22.7L tree−1

Zalesny etal.

[102]

Lamby way,

Poland

Three glass tanks (80L × 5W × 55 H

(cm)) were used

Reed (P. australis) Tanks were subjected to continuous

recirculation as in a horizontal ow

Bialowiec

etal. [64]

Willow (Salix viminalis × burjatica)

Sanitary

landll site,

Slovenia

Constructed as a vertical ow–

horizontal ow system. Vertical beds

covered 41m2, 22m2 (5m × 4.4.m),

and 19m2 (5.3m × 3.6 m) with a

depth of 0.8m. The horizontal bed

covered 270m2 (13.5m × 20 m) with

a depth of 0.4m

Reed (P. australis) HLR—0.5cm day−1Bulc [103]

Cattail (T. latifolia)

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Landll

leachate

location CW conguration Plant species Operational parameters of CW Reference

Solid waste

transfer station,

Thailand

Constructed horizontal subsurface

ow system. Four ponds of 1m wide,

3m long and 1m depth

Cattail (T. augustifolia) Different HLR—0.01, 0.028 and

0.056m3 m−2 day−1

Chiemchaisri

etal. [104]

Dragonja

landll site,

Slovenia

Consisted of two interconnected beds

with a horizontal subsurface ow.

The rst bed was 21m long, 10m

wide and 0.9m deep. The second bed

was 23m long, 10m wide and 0.8m

deep.

Reed (P. australis) HLR—3cm day−1Bulc etal.

[105]

Solid waste

dump site,

Slovenia

Consisted of six interconnected beds

with a horizontal and vertical

subsurface water ow

Reed (P. australis) After the pretreatment process,

pumps all the water in 3 h on the

landll cover where willows and

reeds are planted

Justin and

Zupancic [74]

Willow (S. purpurea)

Saginaw

Township,

USA

System consisted of aeration, settling,

intermittent vertical sand ltration

and surface ow wetland treatment

Cattail (T. latifolia) The total system detention time was

180 days

Kadlec and

Zmarthie [106]

Wilmington,

USA

FWS [14.3m × 4.1m × 31cm (L ×

W × H)] was used

Sweet ag (Acorus calamus) Flow volume for the wetland was

11m3 and detention time was 10–12

days

Kozub and

Liehr [83]

Soft rush (Juncus effuses)

Arrow arum (Peltandra virginica)

Pickerel weed (Pontederia cordata)

Lizard’s tail (Saururus cernuus)

Softstem bulrush (Scirpus validus)

Burreed (Sparganium androcladum)

(continued)

17 Phytoremediation ofLandll Leachates

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Table 17.1 (continued)

Landll

leachate

location CW conguration Plant species Operational parameters of CW Reference

North-western

region,

Bulgaria

Consisted of VFS, 123mm in

diameter and 900mm in height

Reed (P. australis) Used three different ow rates, 40,

60, and 82 mL min−1, respectively

Lavrova and

Koumanova

[45]

Iowa, USA Subsurface ow dimension was 15.5

× 6 m; L × W

Stiff goldenrod (Solidago rigida) Other local plant species including

curly dock (Rumex crispus), bull

thistle (Cirsium vulgare), stinging

nettle (Urtica dioica), willow (Salix

spp.) and cannabis (Cannabis

sativa) were dominated with time

Nivala etal.

[79]

Nonthaburi

province,

Thailand

Horizontal subsurface ow and free

water surface ow systems were

examined

Cattail (T. latifolia) Bed was lled with mixed sand and

clay

Ogata etal.

[43]

Upstate

NewYork,

USA

Subsurface ow dimension was 33 ×

3 × 0.6 m; L × W × H

Reed (P. australis) Leachate application rate was

1.4m3 day−1 and residence time of

15 days

Sanford etal.

[107]

Bangkok,

Thailand

Subsurface ow dimension was 4 ×

0.5 × 0.5 m; L × W × H

Cattail (T. angustifolia L.) Different hydraulic retention (HRT),

1, 3, 5 and 8 days, were studied

Sawaittayothin

and Polprasert

[63]

Northeastern

part of Ankara,

Turkey

Operated three subsurface wetland

systems. Two of them were vertical

ow mode and one being horizontal

ow mode. The dimensions were 1L

× 0.5W × 0.4 H (m)

Cattail (T. latifolia) Leachate application rate was 10L

day−1

Yalcuk and

Ugurlu [47]

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453

including carbon sequestration, erosion control, pollution prevention, and enhanc-

ing landscape appearance are some of the prevailing benets of wetland plant

species [44].

In the last couple of decades, short rotation coppice (SRC) has been introduced

to remove contaminants from landll leachate [51]. In this management strategy,

plant species which possesses multiple shoot growth are cut down particularly on a

3-year rotation period. Consequently, harvested biomass could be used as a sustain-

able energy source as well as a CO2 neutral fuel. It has been identied that willow

(Salix sp.), poplar (Populus sp.), and eucalyptus (Eucalyptus sp.) are the most preva-

lent species in SRC systems. For instance, in Sweden, 14,500ha of willow coppice

are grown commercially [65]. Both willow and poplar possess a rapid growth rate,

1–3m year−1, in addition to high plant densities and high biomass yields [51]. Some

studies indicated that planting densities, between 12,000 and 25,000 trees ha−1, have

been successfully established [66]. Additionally, a study by Mitchell et al. [67]

revealed that biomass yield from willow SRC laid between 2.2 and 13.5 oven dried

tonnes (odt) ha−1 year−1. It appears that SRC can be successfully utilized for treating

not only landll leachate, but also different sources of pollutants such as municipal

wastewater, sewage sludge, and agricultural efuents, since SRC possesses exten-

sive transpiration rates. For example, Salix cinerea possesses high evapotranspira-

tion rates ranging from 16.4 to 27.4L m−2 [51].

Fig. 17.5 Reed (Phragmites sp.) plants in constructed wetland systems

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17.5.4 Application ofVegetation Filter

Vegetation lter is a term commonly applied to explain the soil–plant treatment

system in CW.More precisely, different phytoremediation processes are involved to

remove detrimental substances from contaminated water and soil with the help of

wetland plant species [11, 48]. Different types of phytoremediation processes and

their signicance in the removal of contaminants are summarized in Table17.2.

In terms of hot and dry climatic conditions, evapotranspiration can be utilized in

soil–plant systems for treating landll leachate. It seems obvious that greater amount

of leachate volume decreases due to evapotranspiration of soil–plant systems [68].

It has been experimentally observed that USEPA recommended plant species such

as willows, poplars, and reeds have been successfully utilized in recent years for

landll leachate evapotranspiration [69]. It is well-known that higher transpiration

rate sometimes exceeds the annual rainfall of willow stands leading to reduced

groundwater level [70]. For instance, Agopsowicz [71] observed that introduction

of willows for landll leachate removal enhanced the evapotranspiration efciency

by 1.28–5.12 folds more than the evaporation efciency from the plantless soil sur-

face. This study also found that evapotranspiration of 3-months old willow sprouts

was 1.6–1.8 fold greater than an average rainfall rate in Poland. Similarly, Białowiec

etal. [72] noticed that transpiration of 3-months old sprouts of Salix amygdalina L.

leads to evapotranspiration between 80 and 90%. Nevertheless, landll leachate

evapotranspiration efciency is basically dependent on the physico-chemical char-

acteristics of landll leachate, as well as the plant species used [73, 74]. For instance,

various dissolved compounds in landll leachate may induce negative impacts on

plant species in two ways. Firstly, biomass growth rate is inhibited and results in the

reduction of evapotranspiration [69]. Transpiration efciency coefcient (β−1—ratio

of biomass transpiration; unit—g d.m. mm−1) may provide a clear view related to

Table 17.2 Phytoremediation processes in wetland plant species

Process Importance

Phytoextraction Contaminants are taken up from rhizosphere, transported and

translocated to above ground shoots.

Phytodegradation Plants take up, store, and metabolize or convert toxic contaminants

to non-toxic by-products.

Phytovolatilization Plants extract volatile metal(loid)s and organic compounds and

release them into the atmosphere.

Rhizoltration Plant roots grown in aerated water precipitate and concentrate toxic

components and pollutants are broken down by soil

microorganisms.

Phytostabilization Plants stabilize the pollutants rendering them harmless and control

soil erosion and water inltration, as well as humidication and

lignication of organic compounds.

Rhizo-redox reactions Plants alter the speciation of heavy metal(loid)s, thereby affecting

their bioavailability and mobility.

P. Kumarathilaka et al.

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455

landll leachate inuence on plants [75]. If harmful compounds are available to

plants, transpiration coefcient value reduces signicantly. For example, although

β−1 values typically lies between 1.9 and 4.9 g d.m. mm−1 for willow, Białowiec

etal. [69] have reported β−1 values in the range 0.12 and 0.45 g d.m. mm−1. Secondly,

it is known that dissolved substances are responsible for reducing the difference

between soil-water tension and soil-water tension at air entry. As a result, evapora-

tion of landll leachate could be remarkably decreased [70].

It is obvious that when designing a soil–plant system for landll leachate treat-

ment, linear relationship between biomass growth and transpiration is a particular

concern. Further, proper management strategies and fertilization may stimulate the

plants’ growth and increase its biomass [69]. For example, the dark color of the sand

at the top of waste heap increases soil temperature, which enhances evaporation.

Additionally, inherent characteristics of organic soil types such as higher water

retention and capillary suction may increase evaporation in landll site. Additionally,

since some wetland plant species belongs to non-food crop category (i.e., willow),

transformation of disease-related pathogens via food chain does not occur.

Despite leachate treatment difculties, its characteristic features including high

concentrations of plant macro- and micro-nutrients (N, K, Mg, Ca, Zn, and B) facil-

itate the potential of landll leachate reuse as a valuable fertilizer for growing

energy crops such as short rotation willow coppice [73]. There are several examples

in literature regarding the use of landll leachate irrigation on tree growth, exhibiting

its fertilizing capability. For instance, Justin etal. [44] detected a positive relation-

ship in the biomass production of willow and poplar plantations treated with landll

leachate possibly due to the fertilization properties of landll leachate. The usage of

leachate showed up to 155% increment in the aboveground biomass compared to

the control which was treated with water. Further, poplar was observed as the most

efcient plant in biomass production owing to its high leaf production [44].

Tree sap ow has been recognized as a good surrogate indicator with respect to

water usage of a particular plant Smith and Allen [76]. In general, heat is utilized as

an indicator to measure sap ow in trees, and heat pulse, heat balance, and thermal

dissipation methods are commonly used for measuring sapow [76]. A study by

Zalesny etal. [42] demonstrated that sap ow of hybrid poplars (Populus nigra L. ×

P. maximowiczii A.Henry ‘NM6’) showed a negative relationship with temperature,

wind speed, rainfall, and vapor pressure decit. In addition, they observed an

increase in sap ow as sapwood area increased from 43.8 to 122.3cm2. Further,

exploitations to the stand were reported as 2.8 and 11.3mm d−1 for two consecutive

years, respectively, exhibiting great capacity for reducing landll leachate amount.

It is known that the tips of major roots as well as lateral roots of plant species

may release oxygen (O2) into the rhizosphere [77]. As a result, a layer of O2 around

the roots ranging from 1 to 3mm thickness is formed, whereas the thickness of the

layer is determined by the actual redox status of the rhizosphere [60]. Accordingly,

redox conditions in the subsurface may determine the aerobic processes such as

nitrication and anaerobic processes such as methanogenesis and denitrication.

Membrane inlet mass spectrometry (MIMS) can be utilized to monitor dissolved

gases such as O2, CO2, and CH4 within the treatment bed of a vegetation lter treat-

ing leachate at a particular landll site [78]. Williams etal. [60] observed a positive

17 Phytoremediation ofLandll Leachates

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456

correlation between the dissolved oxygen prole throughout the CW bed and the

distribution of willow roots in the soil. Therefore, it seems obvious that depth distri-

bution of roots in a vegetation lter bed provides important information with respect

to oxygenation potential.

17.6 Pollutant Diminution inLandll Leachates

byPhytoremediation

17.6.1 BOD andCOD Removal

In a CW, the biological reactions are controlled by microorganisms via biolm for-

mation on the bed material [49]. Photosynthetically produced O2, which helps in the

growth of microbes, may transport to the root zones in CW systems. It seems obvi-

ous that relatively low COD removal efciency at the start-up time in CW system

could be attributed to the formation of active microorganisms. Additionally, leach-

ate treatment performance enhances when the CW system matures [47]. The BOD5/

COD ratio is an important parameter which explains whether organic compound of

landll leachate is biodegradable or not. For example, if the BOD5/COD ratio in

landll leachate is quite low, it means a majority of the organic compounds are

non- biodegradable and vice versa [15]. Moreover, a low organic loading rate may

lead to low COD removal efciency in CW systems. In case of later stage of leach-

ate, a sizeable fraction of COD ow along the CW systems leads to intermittent and

poor COD removal efciency [79]. In terms of removal efciency, temperature does

not affect signicantly. Nevertheless, hydraulic retention time (HRT) is apparently

responsible for greater COD removal rates. In this context, HFS gains higher COD

removal than VFS, possibly due to the higher HRT values [47]. Moreover, aeration

may increase BOD5 and COD removal efciencies in CW systems. Several demon-

strations of BOD and COD removal efciency from landll leachate in CW systems

are presented in Table17.3. A study by Nivala etal. [79] found higher BOD5

efciencies, up to 97%, with supplemental aeration. However, in the absence of

aeration, the BOD5 removal rate was between 75 and 81%.

17.6.2 Nitrogen andPhosphate Removal

It is known that different forms of N such as nitric oxide (NO), NO3−, and NH3 are

highly soluble in water and can enter the water bodies along with landll leachate

discharges. Thus, the removal of excess N from landll leachate is particularly

important in terms of water quality. It is well-established that numerous removal

processes including nitrication, denitrication, uptake by plants and microorgan-

isms, NH3 volatilization, and adsorption onto cation exchange are involved in the

removal of N in CW [46]. Table17.3 summarizes the N and phosphate (PO43−)

removal efciency from landll leachate.

P. Kumarathilaka et al.

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Table 17.3 Raw leachate characteristics and treatment efciency of CW systems

Plant species Metal(loid)

Raw leachate concentration (mg L−1) Leachate treatment efciency (%)

ReferenceNH3-N Total P COD BOD Metal(loid) NH3-N Total P COD BOD Metal(loid)

Reed (Phragmites

australis)

Cr 198 5.5 2800 204 0.27 100 100 96 92 70 Lavrova and

Koumanova [45]

Zn 0.25 100

Al 0.33 77

Stiff goldenrod

(Solidago rigida)

NA 253 NA 873 177 NA 98 NA 60 97 NA Nivala etal. [79]

Cattail (Typha

latifolia)

NA 29 NA 1650 NA NA 38 NA 7NA NA Ogata etal. [43]

Cattail (T.

angustifolia L.)

Cd 25 3.5 385 130 1 96 93 81 91 99 Sawaittayothin and

Polprasert [63]

Cattail (T. latifolia)Pb 122 NA 212 NA 8.45 67 83 42 NA >90 Yalcuk and

Ugurlu [47]

Ni 1.71 30

Reed (P. australis)Fe 496 2.3 485 76 3.90 51 53 50 59 84 Bulc [103]

Reed (P. australis)Fe 88 NA 1264 60 10 81 NA 68 46 80 Bulc etal. [105]

Cattail (T.

augustifolia)

NA NA NA 500 17 NA NA NA 63 44 NA Chiemchaisri etal.

[104]

Reed (P. australis)Cr 327 2.37 1508 193 0.66 42 38 41 65 33 Justin and Zupancic

[74]

Willow (Salix

purpurea)

Mn 1.05 12

Cu 0.15 18

Zn 0.34 33

As 0.04 34

Cattail (T. latifolia)Zn 382 2.8 NA NA 19 99 89 NA NA 16 Kadlec and Zmarthie

[106]

As 7 29

Cr 24 67

Cattail (T. latifolia)Cd 94 NA 72 12 0.003 NA NA 2 53 33 Martin etal. [108]

Reed (P. australis)NA 230 NA 2000 300 NA NA NA 88 91 NA Mæhlum [109]

Cattail (T. latifolia)

NA not available

guarino@unisannio.it

458

Temperature plays a major role in CW systems and has an effect directly or

indirectly on the nutrient uptake by plants and microbial activity [80]. From the

previous studies, it is demonstrated that nitrication efciency in CW systems

becomes inhibited at a water temperature of 10 °C and further decline rapidly at

about 6 °C [80]. Contrarily, at relatively higher ambient temperatures, signicantly

higher total Kjeldahl nitrogen (TKN) and NH3 removal were achieved. A study by

Akratos and Tsihrintzis [81] has shown that TKN and NH3 removal were higher at

temperatures above 15 °C.Similarly, Yalcuk and Ugurlu [47] demonstrated that the

highest ammonia-N (NH3-N) removal of 62.3% was achieved at higher ambient

temperatures.

Furthermore, organic matter oxidation and transformation of N by microorgan-

isms depend upon the ambient temperature [81]. It is known that oxygen concen-

tration is also temperature-dependent. In general, the solubility of oxygen increases

with decreasing temperature and vice versa [47]. However, it is important to main-

tain aerobic conditions in CW systems to avoid denitrication. During the day

time, photosynthesis results in oxygen generation which facilitates oxygen require-

ment for stabilization of organics and nitrication. Consequently, plant rhizosphere

aeration enhances aerobic decomposition processes including nitrication and

gaseous losses of N through denitrication [82]. In general, autotrophic nitrication

involves two successive reactions. Initially, ammonium is converted to nitrite by

ammonium oxidizing bacteria. Following that, NO3− oxidizing bacteria converts

nitrite to NO3−. The inuence of different ow rates and recirculation ratios for

removal of NH4-N has been successfully investigated by Lavrova and Koumanova

[45]. They revealed that lower ow rates achieved higher NO3-N concentrations,

whereas the effect of recirculation ratios was opposite. It is well-established that

1mg L−1 of dissolved oxygen (DO) is sufcient for the oxidation of ammonium.

According to the above study, the DO values during nitrication experiment ranged

between 5.2 and 8mg L−1.

In CW systems, denitrication mainly depends on the availability of NO3-N and

organic C.Additionally, denitrication also depends on environmental conditions

including pH, ambient temperature, the amount of DO, and the availability of sub-

strates for microbial attachment [83]. More precisely, mean N loss in CW systems

can be assigned to denitrication. It is obvious that lack of organic substrate may

inhibit denitrication process possibly due to the minimization of synthesis and

activity of denitrifying enzymes [47]. Nevertheless, CW systems are unable to

remove C completely, since plant litter and plant/root may contribute increment of

C to the system [84]. Therefore, low amount of C in CW is not a huge problem,

since it is vital for anaerobic respiration and denitrication process.

The possible processes for PO43− loss in CW system include sedimentation, adsorp-

tion, and biological transformations [85]. Low concentration of P in efuent can be

attributed to the uptake by plants and microorganisms [47]. Moreover, insufcient P

level in leachate may adversely affect biomass growth, and consequently, the treat-

ment capability in CW system. For this reason, denitrication rate, in particular, is

reduced to a great extent. Typically, landll leachate consists of low levels of minerals

with reactive iron (Fe) or aluminium (Al) or calcium (Ca) hydroxide, which are capa-

ble of stimulating PO43− precipitation to a great extent [86]. A study by Sakadevan and

P. Kumarathilaka et al.

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459

Bavor [87] explained that while P removal in a long-term basis is mainly dependent on

the substratum, litter, and Al/Fe component, plant uptake is less involved. It has been

experimentally observed that phosphorous removal in a CW system is a seasonal

dependence and particularly correlated to plant growth and consequent PO43− uptake.

Nevertheless, it appears that PO43− removal is less linked with temperature because it

is apparently governed by adsorption process [47]. As previously mentioned, ow

rates and recirculation ratios are responsible for PO43− removal efciency. A study by

Lavrova and Koumanova [45] demonstrated that higher retention time and higher

recirculation ratio eliminated a greater concentration of total P from the leachate.

17.6.3 Heavy Metal(loid) Removal

From the previous studies, it is well-understood that heavy metal(loid) removal in

CW systems is mainly governed by various biological and physico-chemical factors

including microbial activity, uptake by plant species, sedimentation, occulation,

precipitation, adsorption, complexation, oxidation, and reduction, and cation and

anion exchange [88, 89]. Typically, it is impossible to remove heavy metal(loid)s;

however, their physico-chemical characteristics can be modied by (im)mobiliza-

tion and subsequently managed [47]. Higher amounts of heavy metal(loid)s are

removed due to the binding processes in CW systems. Typically, heavy metal(loid)

ions possess positive charge; hence, they are rapidly adsorbed, complexed, and

bound with suspended particles. In terms of long-term removal, heavy metal(loid)s

are precipitated as their insoluble salts such as suldes, hydroxides, carbonates, and

bicarbonates and subsequently deposited within the wetland substrate [41].

Additionally, algae and microorganisms are able to take up heavy metal(loid)s avail-

able in the dissolved form in CW system [88].

The symplastic and apoplastic pathways provide a route towards absorption of

heavy metal(loid)s into the roots of plant species. Symplastic pathway, an energy-

dependent process, is mediated by specic or generic metal(loid) ion carriers or

channels. But in apoplastic pathway, it is the opposite; the metal(loid) ions of

metal(loid)-chelate complex penetrate the root via intercellular spaces [58]. More

precisely, plant roots are able to solubilize soil-bound heavy metal(loid)s by acidify-

ing the soil environment. Some enzymes (i.e., reductases) bound to the plasma

membrane can also reduce the soil-bound heavy metal(loid) ions. Additionally,

mycorrhizal fungi and root-colonizing bacteria increase the bioavailability of heavy

metal(loid)s in CW systems [90]. Hence, rhizospheric microorganisms enhance the

plant uptake of heavy metal(loid) ions.

In hyperaccumulator plants, heavy metal(loid)s absorbed by roots efciently

transport to the shoots via xylem system and the xylem loading process is governed

by membrane transport proteins. Therefore, in metal(loid) accumulators, xylem

loading process and translocation to shoot are stimulated by the complexation of

metal(loid)s with low molecular weight chelators such as organic acids, and the

metal(loid)s are converted into a less toxic form at any point of the transport path-

17 Phytoremediation ofLandll Leachates

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460

way [58]. There are several examples in the literature of the use of wetland plant

species for removing heavy metal(loid)s from landll leachate and are summarized

in Table17.3.

17.7 Advantages andDisadvantages ofPhytoremediation

ofLandll Leachate

17.7.1 Economic Benets

There is little information on detailed economic analysis related to feasibility of

landll leachate treatment systems. Nevertheless, from the limited studies, it is

well-understood that SRC treatment systems provide economically feasible dis-

posal options for landll leachates, if managed properly. In other words, on-site

utilization of landll leachate treatment in CW systems saves millions of dollars at

each site where it is implemented [73]. In Sweden, short-rotation willow coppice is

grown over an area of 14,500ha as a commercial crop. In the last few decades,

about 30 landll leachate treatment systems on restored caps or adjacent to the

landll are operated using willow plantations in Sweden [91]. It is well-known that,

due to high evaporation, willow plantations are able to reduce greater amounts of

leachate formation, and therefore, recycling of landll leachate back to establish

willow plantation leads to near zero net discharge of leachate [73]. It is reported that

short-rotation willow coppice can be harvested every 3–4 years and approximately

6–10 t dry matter ha−1 year−1 will be produced [65]. Subsequently, harvested dry

matter could be used to generate electricity and produce heat.

A number of case studies in the USA indicated that on-site phytoremediation

strategies save transportation and disposal costs signicantly. For example, around

2100 hybrid poplar trees were planted over an area of 5.5 acres in Jeffco landll,

name of the state, and subsequently, 14 million gallons of leachate have been pro-

cessed, thus saving a total amount of $810,000 [92]. Similarly, a republic landll in

Chicago consisting of more than 4000 hybrid poplar trees over a 7.5 acre processed

greater than 2.7 million gallons of landll leachate, thus saving $350,000in leach-

ate disposal costs in 2 years [92]. Furthermore, 3 acres of vetiver (Chrysopogon sp.)

were installed in a landll and processed three million gallons of leachate per year.

As a result, the disposal cost dropped from $0.13 to $0.015 per gallon [92].

Apart from energy production, the usage of SRC for landll leachate treatment

enhances the economic competitiveness of renewable energy systems. In other

words, landll leachate facilitates a viable option for producing SRC biomass, since

approximately 20–30% of the SRC production cost required for irrigation and fer-

tilization could be saved [51]. Hence, the leachate treatment cost could be offset by

selling the biomass. For instance, SRC treatment system in Sweden possesses US$

13–18 compared to the conventional leachate treatment plant, which costs about

US$ 10–27 [93]. So, there is an increasing trend to utilize energy crops for treating

landll leachate in the world, since this strategy addresses not only an environmen-

P. Kumarathilaka et al.

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461

tal issue but also energy production. More precisely, energy conversion technolo-

gies such as spark ignition gas engines can be applied to convert SRC biomass to

generate electricity [94].

Phytoremediation strategy can also be integrated with other novel treatment

technologies, thereby stabilizing or enhancing their existing functions simultane-

ously and avoiding individual drawbacks. Integrating CWs with other emerging

technologies such as membrane bio-reactor, electrochemical oxidation, and micro-

bial fuel cells has been studied over the last few years and has been proven to be

efcient for treating pollutants and for sustainable energy recovery [95, 96].

17.7.2 Phytotoxicity ofLandll Leachate

It is especially important to study the responses of plant species to landll leachate

stress, since phytoremediation strategies for landll leachate is gaining an increas-

ing focus in recent years [97]. As previously mentioned, landll leachate possesses

toxic components including organics and heavy metal(loid)s and may outweigh the

benecial effects of nutrients. Additionally, the high concentration of chloride and

sodium leads to high ionic strength in landll leachates [73]. The toxicity of landll

leachate may damage plant species when it is utilized for irrigation. However, the

plant toxicity depends on different factors such as plant species, soil type, irrigation

rates, and climatic conditions. Some of the deleterious impacts of landll leachate

are premature leaf senescence, leaf damage, less biomass production, and poor sur-

vival rates [73, 98].

There are several examples in the literature on the effect of landll leachate on

plant species. A study by Sang etal. [99] monitored several physiological changes

of maize (Zea mays L.) such as growth chlorophyll content, lipid peroxidation,

protein oxidation, and activities of antioxidant enzymes. The results revealed that

landll leachate affected the growth and chlorophyll level of maize seedlings.

In addition, lipid peroxidation and protein oxidation in leaf tissues were increased

to a great extent indicating plant stress landll leachate [99]. Phytotoxicological

tests play a major role in designing phytoremediation technologies for treating land-

ll leachate [100]. Such tests are able to indicate the maximum dose of treated

leachate, which does not affect the plant species negatively. Dimitriou etal. [73]

conducted a pot experiment to quantify the growth responses of ve different wil-

low clones with different leachate mixtures. The results showed that plant growth

rates were reduced with leachate irrigation. Additionally, they have suggested that

leaf length could be a useful stress diagnostic tool for use in situ showing a high

correlation to growth [73].

In spite of detrimental effects of landll leachate on plant species, many wetland

plant species and SRC are tolerant of high level of heavy metal(loid)s; however,

there is a concern about the risks of bioaccumulation of toxic heavy metal(loid)s via

food chain. Therefore, proper disposal and harvesting practices are necessary for

plant biomass which accumulates heavy metal(loid)s [48]. Similarly, accumulation

17 Phytoremediation ofLandll Leachates

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of heavy metal(loid)s in wetland substrates could have long-term implications. Due

to the oxidation and disturbance of wetland substrate, sedimented heavy metal(loid)

s can be released into the system [41].

17.8 Summary andConclusions

In the last couple of decades, an on-site treatment of landll leachate with the help

of CW is widely practiced in numerous nations in the world. It is obvious that CW

could be the ideal technology for landll remediation due to its cost-effective and

eco-friendly nature. Additionally, establishing vegetation in landll sites will facili-

tate erosion and hydraulic control by reducing inltration of rainfall. It appears that

the degree of success in terms of contaminant removal efciency by CW systems

varies depending upon the plant species selected, availability of microbial commu-

nity, climatic conditions, physico-chemical properties of soil, and CW congura-

tion. It is well-established that nutrients (i.e., N and PO43−), heavy metal(loid)s,

BOD, and COD can be successfully removed to a great extent by CW systems.

Additionally, well-managed SRC systems save millions of dollars by eliminating

the transportation and treatment process which were earlier practiced. Nevertheless,

some of the deleterious impacts of landll leachate may adversely inuence the

treatment efciency of wetland plant species.

Phytoremediation of landll leachate is still new and has to be developed. There

are a number of examples where phytoremediation has failed. Basically, this failure

can be attributed to excessive leachate application and lack of management prac-

tices due to poor understanding of the plant–soil system. The current knowledge and

understanding on the limitations may enhance future investigations in respect to

phytoremediation of landll leachate. For this reason, experience, investigations,

and eld trials are vital to forecast and certify that treated leachate as well as har-

vested wetland plants has almost detoxied attaining minimum risk to human

beings and the environment. Additionally, residual management is quite necessary

to overcome problems arising from the public. Fundamental investigations based on

phytoremediation of landll leachate is not sufcient to solve the problem; hence,

enough attention is to be given for developing large scale investigations with new

strategies and approaches together with integrated technologies. There have been

several investigations regarding landll leachate remediation by using SRC, though

results found to date have been promising. Even though landll leachate is utilized

as a means of fertilization for SRC to enhance yields, further investigations are

necessary to establish the full potential of this strategy. Also, assessing the microbial

interactions and their symbiotic activities on landll leachate treatment is also an

urgent necessity. Genetic engineering technology can be applied to enhance existing

traits or confer novel capabilities of plant species which are used for treating landll

leachate. Overall, it is obvious that successful transfer of phytoremediation tech-

nologies from the laboratory to the eld is a crucial step for the future of landll

leachate phytoremediation to be more efcient.

P. Kumarathilaka et al.

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A.A. Ansari et al. (eds.), Phytoremediation, DOI 10.1007/978-3-319-52381-1_18

Chapter 18

Phytomining ofRare andValuable Metals

LuísA.B.Novo, PaulaM.L.Castro, PaulaAlvarenga,

andEduardoFerreirada Silva

Abstract The exponential growth of low-grade mining ores and metal-polluted

soils around the world during the last decades is expected to continue at a higher

rate in the foreseeable future. Yet, the strategic and commercial importance of some

elements found in those sub-economic ores and soils, their elevated market prices,

and the corresponding environmental concerns have opened a window of opportu-

nity for phytomining. This phytoextraction-based technology uses the ability of cer-

tain plants to uptake valuable metals, producing a bio-ore from the harvested

biomass that allows metal recovery through smelting. Once applied at large scale,

phytomining may either function as a standalone operation to retrieve the desired

element or jointly with phytoremediation, nancing the costs of the latter. This

chapter reviews the advances of phytomining since its inception in the 1990s, focus-

ing on the results obtained to date, with gold, nickel, thallium, and rhenium.

Keywords Phytoextraction • Phytoremediation • Hyperaccumulation • Gold •

Nickel • Thallium • Rhenium

18.1 Introduction

The uptake of valuable metals by plants has fascinated scientists for nearly

three centuries. Ever since Beccher and Kunckel ascertained the presence of gold

in plants during the 1700s [1], they were promptly followed by illustrious

L.A.B. Novo, Ph.D. (*) • E.F. da Silva, Ph.D.

Department of Geosciences, GeoBioTec Research Center, University of Aveiro,

Campus de Santiago, 3810-193 Aveiro, Portugal

e-mail: novo@ua.pt; eafsilva@ua.pt

P.M.L. Castro, Ph.D.

Faculty of Biotechnology, Centre of Biotechnology and Fine Chemistry,

Catholic University of Portugal, 4200-072 Porto, Portugal

e-mail: plcastro@porto.ucp.pt

P. Alvarenga, Ph.D.

Department of Technologies and Applied Sciences, School of Agriculture,

Polytechnic Institute of Beja, Beja 7800-295, Portugal

e-mail: paula.alvarenga@ipbeja.pt

guarino@unisannio.it

470

contemporary chemists including Berthollet, Sage, Rouelle, Darcet, and Deyeux,

who also claimed to have found small amounts of the precious metal in vegetables

[2]. Then, after Lungwitz suggested the possibility of using plant tissue analysis for

gold bioindication in 1900 [3], several researchers have reported the occurrence of

this much sought-after metal in many plant species from different locations through-

out the twentieth century [4, 5]. These studies marked the beginning of biogeo-

chemical exploration, highlighting the utility of plants for prospecting of gold,

silver, or even uranium, along with other less valuable elements [6]. In the late

1970s, Jaffré etal. coined the term hyperaccumulation to characterize plants that

uptake nickel to concentrations surpassing 1000mg kg−1 [7], while in 1983, Chaney

suggested the use of hyperaccumulator plants for the reclamation of metal-polluted

sites, a process known as phytoremediation [8]. Phytoremediation is a low-cost,

solar-driven and environment-friendly alternative to conventional solutions that are

often impractical due to their prohibitive costs, unworkability, and detrimental side

effects [9–11]. Two phytoremediation categories are especially relevant for the miti-

gation of metal pollution in soil: phytostabilization and phytoextraction.

Phytostabilization is a management strategy to restrain metals in the rhizosphere of

metal-tolerant plants (known as metallophytes), averting their migration through the

soil and into aquifers [12, 13]. On the other hand, phytoextraction involves the

uptake of a considerable amount of a given metal by the root and its translocation

into the shoot. This method reduces the concentrations of metals in the soil, allow-

ing their safe disposal after harvest [9, 12]. Phytoextraction requires metallophytes

presenting fast growth- rate, high biomass yield, hyperaccumulation (or at least ele-

vated metal levels in the shoots combined with high biomass production), and bio-

concentration and translocation factors greater than 1 [12, 14].

Still, it was not until the 1990s that phytomining, the use of phytoextraction to

recover valuable metals from waste substrates, was proposed [15–18]. In recent

years, the ever-rising price of rare and valuable elements, the incapacity of conven-

tional mining to extract the totality of metals from mineral ores, and the build-up of

billions of tons of mine waste around the world have strengthened the importance of

phytomining, prompting its development [5, 19, 20].

18.2 Hyperaccumulation

The uptake and accumulation of metals in plants generally depends on the bioavail-

ability of these elements in their growth substratum. Hence, metal availability is

highly inuenced by the pH, oxygen content, nutrient balance, and coexistent inor-

ganic and organic compounds [21, 22]. For some metals, low solubility and strong

interactions with the organic or silicate matrix result in partial or complete unavail-

ability. Metals are also inuenced by a series of root mechanisms that encompass

cation exchange, exudation of low-molecular weight organic acids, chelating com-

pounds and enzymes, and the acidication of the rhizosphere through H+ secretion.

These processes, as well as symbiotic associations between plants and mycorrhizal

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fungi and bacteria [22, 23], have the potential to increment metal bioavailability and

promote their entry into root cells over passive or active absorption [21, 24, 25].

When in the root, metals can be stored or translocated into the aboveground parts,

usually via the xylem in a process mediated by membrane transport proteins, for

detoxication and sequestration in the vacuoles [25, 26].

Certain plants have the uncommon capacity of hyperaccumulating metals, both

essential and/or non-essential for their growth and development, to levels that can

surpass 2% of their dry biomass weight [26]. Hyperaccumulators are classically

described as plants that accumulate metals to concentrations10 to 100-fold higher

than those normally measured in the shoots of other plants growing in the same

environment [9, 27]. Thus, the threshold concentrations in aboveground plant tissue

of hyperaccumulators should be at least 100mg kg−1 for Cd, 1000mg kg−1 for Co,

Cu, Ni, As, and Se, and 10,000mg kg−1 for Zn and Mn, to mention a few metals [28,

29]. Recently, the denition of metal hyperaccumulation has been revamped by Van

der Ent etal. [30], who propose that the onset concentrations should be at least one

order of magnitude and 2–3 orders of magnitude greater than the typical levels

found in plants growing on metalliferous soils and normal soils (not metal-enriched),

respectively. Hence, the thresholds for Cu, Co, and Cr should be lessened to 300mg

kg−1 dry shoot weight, and the criterion for hyperaccumulation of Zn should be

lowered to 3000mg kg−1 dry shoot weight [30, 31].

Furthermore, hyperaccumulators must also exhibit a bioconcentration factor

(BF) and translocation factor (TF) greater than 1. The BF denotes the competence

of a plant to extract metals from the growing media and accumulate them. The TF

expresses the plant’s aptness to translocate metals from the root to the shoot. A TF

higher than 1 indicates that the plant is capable of transporting metals from roots to

its aboveground parts. The BF and TF are calculated according to Eqs. (18.1) and

(18.2), respectively [9, 32].

BF C

=Plant

Soil

(18.1)

=Shoot

Root

(18.2)

where CPlant is the metal concentration in the plant (mg kg−1), CSoil is the metal

concentration in the soil (mg kg−1), CRoot is the metal concentration in the root (mg

kg−1), and CShoot is the concentration of metal in the shoot (mg kg−1).

Plants can also be divided into obligate or facultative hyperaccumulators,

depending on whether or not they are restricted to metalliferous environments [31].

Hyperaccumulators have been identied in numerous studies and may occur in

nearly 500 species of vascular plants from 45 families of angiosperms, including

species pertaining to Asteraceae, Brassicaceae, Caryophyllaceae, Cyperaceae,

Cunoniaceae, Fabaceae, Flacourtiaceae, Lamiaceae, Poaceae, Violaceae, and

Euphorbiaceae [10, 33, 34]. From these species, 85–90% are obligate endemics to

metalliferous environments [31].

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18.2.1 Chelant-Assisted Phytoextraction

When metals are not sufciently available for plant uptake, chelating agents can be

used to increase their bioavailability in soil. Chelants promote the formation of

strong water-soluble complexes with metals desorbed from the soil solid phases by

decreasing the free-metal activity, which in turn causes the dissolution of previously

unavailable metals. When the chelant saturates, the solid phase becomes deprived of

metals, or the metal solubility equilibrium is reestablished, the chelation process

halts [35, 36]. The characteristics of the chelant and the soil matrix regulate the

quantity of bioavailable metals in the soil solution. The efcacy of a chelant in the

mobilization of metals is generally related to the stability constants of the corre-

sponding metal complexes [37, 38]. Stability constants can be utilized to classify

chelating agents according to their general effectiveness, but not regarding the value

of a specic chelant on different metals, due to the inuence of metal speciation in

a given soil [36]. Moreover, the biodegradation of metal complexes is highly depen-

dent of the metal type and is not correlated to the stability constant of the chelate

complexes [38]. Upon metal complexation by the chelating agent, plants may

uptake the metals through a number of mechanisms that include: (a) absorption of

free metals following their separation from the chelant (known as the split-uptake

mechanism); (b) absorption of the intact chelant-metal complexes; or (c) metal

exchange between plant metabolic ligands and the chelants [35, 36].

Natural and/or synthetic chelants have been broadly used in phytoextraction tri-

als, to increase metal bioavailability and, consequently, the extraction and transloca-

tion of metals to the shoot [10]. Chelating agents can be divided into two major

groups, inorganic and organic. The latter can be subdivided into synthetic aminop-

olycarboxylic acids (APCAs), natural APCAs, and low molecular weight organic

acids (LMWOAs) [39]. A large number of synthetic APCAs have been employed to

enhance phytoextraction: cyclohexylenedinitrilotetraacetic acid (CDTA), diethyl-

enetriaminepentaacetic acid (DTPA), ethylenediamine di-o-hydroxyphenylacetic

acid (EDDHA), ethylenediaminetetraacetic acid (EDTA), ethylene glycol tetraace-

tic acid (EGTA), hydroxyethylenediaminetriacetic acid (HEDTA), and hydroxye-

thyliminodiacetic acid (HEIDA) [10, 36, 39]. Among these, EDTA has gained

recognition as an efcient chelant to promote the uptake of a wide range of metals

by plants. Nevertheless, the reduced biodegradability, leaching risk, and toxicity of

synthetic APCAs led to the study of less hazardous chelating agents [35, 36]. In this

context, two natural APCAs, ethylenediaminedisuccinic acid (EDDS) and

nitrilotriacetic acid (NTA), have been suggested as viable alternatives to synthetic

APCAs, due to their elevated biodegradability, reduced toxicity, and chelating

potential [36]. Moreover, experiments with LMWOAs have also yielded promising

results for numerous metals. In comparison with APCAs, LWMOAs have the

advantage of avoiding excessive metal mobilization effects and leaching risks, given

their higher biodegradability and consequent lesser persistence in soils [35]. In

addition, considering that the roots naturally release many of these LMWOAs, they

pose very little phytotoxicity threat [35, 39]. Contrarily to their organic counter-

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parts—with the exception of thiourea (CH4N2S)—the use of inorganic chelants in

phytoextraction has been generally restricted to enhance the bioavailability of a

single element: gold [5, 19]. In its natural form, Au(0), it is not readily available for

plant uptake due to low solubility in soil. Hence, a number of inorganic chelants,

most of them cyanide- derived, are applied in order to chelate Au(0) and transform it

into Au(I) or Au(III). Gold is then ready for root absorption, although once in plant

tissue it is promptly reduced to Au(0) (over 90%), and only small quantities remain

as Au(I) and Au(III) [21]. However, the addition of inorganic chelants like ammo-

nium thiocyanate (NH4SCN), ammonium thiosulfate ([NH4]2S2O3), potassium bro-

mide (KBr), potassium cyanide (KCN), potassium iodide (KI), sodium cyanide

(NaCN), and sodium thiocyanate (NaSCN) has raised concerns due to their persis-

tence in soil, detrimental impact on soil microbiota, and potential to mobilize unde-

sirable amounts of elements such as As, Cu, Fe, Ni, or Zn into the groundwater [37,

40]. Table18.1 summarizes the list of most organic and inorganic chelating agents

used in phytoextraction studies.

18.3 Phytomining ofRare andValuable Metals

The procedure for a phytomining operation comprises a number of standard steps,

including: (1) locate a site (mine tailings, mineralized or polluted soils) with sub-

economic levels of the target metal; (2) plant a high biomass yield species with

aptitude to accumulate elevated amounts of the target metal (ideally a hyperaccu-

mulator) and tolerate other coexisting metals; (3) where necessary, apply a chelating

agent near plant maturity in order to increase the bioavailability of the target metal;

(4) harvest the plants when these reach maximum biomass production or exhibit

symptoms of decay; and (5) incinerate the harvested plant biomass to retrieve the

bio-ore, from which the target metal can be recovered through smelting. In addition

to these stages, several approaches and variables may inuence the outcome of a

phytomining trial of a given element.

18.3.1 Gold

In spite of numerous studies reporting the ability of plants to accumulate gold

throughout the twentieth century [4, 64, 65], the rst true gold phytomining experi-

ment was only conducted in the late 1990s [18]. In that pioneering greenhouse trial,

Brassica juncea plants accumulated gold in their aboveground tissues to a concen-

tration of 57 mg kg−1, from silica sand articially spiked with gold chloride to

achieve a gold concentration of 5mg kg−1. Prior to harvest, the substrate was treated

with NH4SCN to increase gold’s bioavailability and its uptake [60]. Another study

using the same auriferous substrate basis and a gold concentration of 3.8mg kg−1

tested the potential of ve root crops to extract gold upon treatment with ammonium

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thiocyanate and ammonium thiosulfate [58]. The results exhibited an average gold

concentration of 113mg kg−1 in the roots of Raphanussativus ‘oriental radish’ and

89 mg kg−1 in the roots of Daucus carota treated with NH4SCN (1 g kg−1) and

(NH4)2S2O3 (2 g kg−1), respectively. Assuming a biomass yield of 18 tons ha−1 (tops

+ roots), and the subsequent gold yield of 1450 g ha−1 for D. carota after treatment

with ammonium thiosulfate, a prot of US$ 7,550 would be expected.

Lamb and colleagues [61] employed an articial substrate prepared from silica

sand nely disseminated with gold to a concentration of 5mg kg−1, to study the

ability of B. juncea, Berkheyacoddii, and Cichorium intybus to uptake gold under

the effect of ammonium thiocyanate, ammonium thiosulfate, potassium bromide,

potassium cyanide, potassium iodide, and sodium thiocyanate. The data showed

Table 18.1 List of chelating agents that have been reported to increase metal bioavailability in soil

Type Chelating agent Element Reference

APCAs CDTAsPb [41]

DTPAsCd, Cu, Ni, Pb, Zn [42, 43]

EDDHAsPb [44]

EDDSnCd, Cr, Cu, Pb, Ra, U,

[38, 45–48]

EDTAsCd, Cr, Cu, Ni, Pb, Ra,

U, Zn

[42, 43, 46, 47, 49, 50]

EGTAsCd, Pb [41, 51]

HEDTAsCd, Cr, Ni, Pb [43, 50, 52]

HEIDAsCu [53]

NTAnAs, Cd, Cr, Cu, Pb, U,

[45, 48, 49, 51, 53]

LMWOAs Acetic acid Pb [41]

Citric acid Cd, Cr, Cu, Ni, Pb, Ra,

U, Zn

[42, 45–47, 54]

Fumaric acid Cd [55]

Gallic acid Cd, Cu, Ni, Pb, Zn [42]

Oxalic acid Cd, Cr, Cu, Ni, Pb, U,

[42, 45, 56]

Succinic acid U [56]

Vanillic acid Cd, Cu, Ni, Pb, Zn [42]

Gold-Specic Ammonium thiocyanateiAu [57–60]

Ammonium thiosulfateiAu [57, 58]

Potassium bromideiAu [61]

Potassium cyanideiAu [59, 61]

Potassium iodideiAu [61]

Sodium cyanideiAu [57, 59, 62]

Sodium thiocyanateiAu [61]

ThioureaoAu [57, 63]

s synthetic; n natural; iinorganic; o organic

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average gold levels of 97mg kg−1 and 326 mg kg−1 in B. coddii and B. juncea,

respectively, when hyperaccumulation was induced via KCN.

A 2003 greenhouse study evaluated the accumulation of gold in canola (Brassica

sp.), growing in soils presenting increasing concentrations of the metal: 1.25, 2.5,

and 5.0mg kg−1 [66]. The media was treated with potassium cyanide at a rate of 0.2

g kg−1 and induced the accumulation of approximately 60, 120, and 150mg kg−1 of

gold in the plants, according to each increment in the substrate gold levels. In addi-

tion, the authors made an economical assessment estimating a potential gross prot

of US$ 6,437 per hectare.

In 2005, a eld trial was carried out in Brazil to assess the feasibility of using B.

juncea and Zea mays to extract gold from oxidized ore containing 0.6mg kg−1 of the

precious metal [59]. One week before harvesting, the trial plots were treated with

ammonium thiocyanate (0.3 g kg−1), potassium cyanide (0.15 g kg−1), and sodium

cyanide (g kg−1), to induce gold hyperaccumulation. The highest concentrations, 30

and 39mg kg−1, were found in B. juncea after the application of KCN and NaCN,

respectively. Centered on these results, a later appraisal was made to determine the

prot obtained at the end of the process [67]. It was calculated that a prot of

approximately US$ 20,000ha−1 could be taken.

One year later, Rodriguez etal. [68] spiked soil with KAuCl4 to attain concentra-

tions of gold in the substrate of 5 and 10mg kg−1. Chilopsis linearis was allowed to

grow for 4 weeks after germination, until utilizing thiourea and ammonium thiocya-

nate as chelants (both as 0.76mg kg−1). Average gold levels in the stems of plants

developed in the soil holding 5mg kg−1 gold reached 296mg kg−1 with CH4N2S and

197mg kg−1 with NH4SCN.

In 2007, a new pot study screened some Australian native plants and exotic agri-

cultural species for their prospective use in cyanide-induced gold phytoextraction

[62]. The chosen plant species comprised Eucalyptus polybractea, Acacia decur-

rens, Sorghum bicolor, Trifolium repens, Bothriochloa macra, Austrodanthonia

caespitosa, and Microlaena stipoides. Crushed ore from the Davis stockpile at the

Stawell Gold Mine in Victoria (Australia), presenting mean gold levels of 1.75mg

kg−1, was picked as substrate. Following standard practice, 1 week before harvest-

ing, the ore was treated with sodium cyanide at the rates of 0.1 and 1 g kg−1. The

most signicant results were obtained with B. macra and T. repens, which presented

shoot gold concentrations of 24 and 27mg kg−1, respectively, under the effect of 1

g kg−1 NaCN.

A greenhouse experiment used mine tailings featuring 2.35mg of gold per kg of

ore, collected from an active mine (El Magistral) in the state of Sinaloa, Mexico

[69]. To evaluate the potential of Sorghum halepense for gold phytoextraction, the

plants were allowed to grow for 10 weeks after germination, and 2 weeks before

harvest, their respective pots were treated with different rates of thiourea (0.0076,

0.015 or 0.030 g kg−1); sodium cyanide (0.5, 1, or 2 g kg−1); ammonium thiosul-

phate (1, 2 or 4 g kg−1); and ammonium thiocyanate (0.32, 0.64 or 1.28 g kg−1). The

most efcient chelating agent was NaCN at a dose of 1 g kg−1, inducing the accu-

mulation of 23.9mg kg−1 of gold in the aboveground dry matter of S. halepense.

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Entering the current decade, Wilson-Corral etal. [70] reported the results of an

earlier gold phytoextraction study. The authors used gold-enriched silica sand

(3.8 mg kg−1), ammonium thiocyanate at a rate of 1 g kg−1 [19], and the plants

Amaranthus spp., S. halepense, Helianthus annuus, Sesamum indicum, Gossypium

hirsutum, Brassica campestris, and Amoreuxia palmatida. Although average con-

centrations of gold in plant tissue were not disclosed, levels above 304mg kg−1 were

found in B. campestris. Subsequently, the same research team carried out two new

eld and greenhouse experiments [57]. In the eld trial, average gold concentrations

of 19, 22, and 15mg kg−1 were attained in the leaves, stems, and roots of H. annuus,

grown in a 50m2 plot constructed over the tailings of the aforementioned Magistral

Mine in Mexico. These results were obtained inducing gold hyperaccumulation

with sodium cyanide at a rate of 1 g kg−1 of ore. Concerning the greenhouse study,

Kalanchoe serrata plants were cultivated in pots containing the same tailings, but 2

weeks before harvesting gold bioavailability was increased through the application

of sodium cyanide (1 g kg−1), ammonium thiocyanate (1.24 g kg−1), ammonium

thiosulphate (2 g kg−1), and thiourea (0.03 g kg−1). Average gold levels in the aerial

parts of the plant reached 9mg kg−1 and 10mg kg−1, when treated with NH4SCN

and (NH4)2S2O3, respectively. Based on the results of biomass yield and gold levels

in plant tissue from these studies, Wilson-Corral etal. suggested earnings of US$

15,098ha−1.

A 2014 pot experiment assessed the viability of three plant species (Lindernia

crustacea, Paspalum conjugatum, and Cyperus kyllingia), to extract gold from cya-

nidation tailings (1.68 mk kg−1 gold) located in the Sekotong District of West

Lombok Regency, Indonesia [71]. Ammonium thiosulfate (2 g kg−1) and sodium

cyanide (1 g kg−1) were applied to the corresponding pots to induce hyperaccumula-

tion. Though the addition of the chelants enhanced the uptake of gold to the shoots

by 106% ([NH4]2S2O3) and 30% (NaCN), the average concentrations have only

reached a maximum value of 0.6 mg kg−1 in P. conjugatum under the effect of

ammonium thiosulfate.

Table 18.2 summarizes some of the most relevant gold phytoextraction results

obtained to date.

18.3.2 Nickel

Conventional nickel mining usually requires ores with a cut-off grade above

30,000mg kg−1 to be economically feasible. However, few ore bodies present those

concentrations and the existing ones are becoming depleted [72]. On the other hand,

the weathering of ultramac rocks has generated large areas of serpentine soils that

are widely scattered around the globe. Nickel levels in serpentine soils are usually

within the range of 1000–7000mg kg−1 [24, 72], making them impractical for com-

mercial mining. The discovery of numerous nickel hyperaccumulators growing on

serpentine soils across the planet, has impelled the assessment of their potential to

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phytomine this metal [16, 17]. Table18.3 exhibits some plant species known to

hyperaccumulate nickel.

Nicks and Chambers carried out the rst phytomining experiment in 1995, using

the nickel hyperaccumulator Streptanthus polygaloides, an endemic serpentino-

phyte from California [17]. Results obtained in a serpentine soil presenting nickel

levels of 3,340mg kg−1 showed an average nickel concentration of 5,300mg kg−1 in

the shoots and a biomass yield of 4.8 tons ha−1. The authors suggested that upon

plant selection, an optimized strain could generate a biomass of 10 tons ha−1, con-

taining 10,000mg kg−1 of nickel. Considering the price of nickel at the time, US$

7.65kg−1, and a 50% return to the grower [84], the crop would be worth US$ 382,

plus US$ 131 relative to energy generation from biomass incineration.

In 1997, Robinson and colleagues conducted two separate studies to determine

the potential of Alyssum bertolonii and B. coddii for nickel phytomining [85, 86]. To

evaluate A. bertolonii, in situ experimental plots were prepared over serpentine soils

in Murlo (Tuscany, Italy), presenting nickel concentrations of 1,600mg kg−1. These

plots were fertilized with different N + P + K regimes during a 2-year period. The

best fertilizer treatment generated biomass and nickel yields of 9 tons ha−1 and

72kg ha−1, respectively. Those results translated to an economic return of US$ 539

according to the nickel price of the period, in addition to US$ 219 from energy pro-

duction [84, 85]. The eld trials with South African nickel hyperaccumulator B.

coddii resulted in a biomass yield of approximately 22 tons ha−1. Assuming a maxi-

mum concentration of 7,880mg kg−1 (shoots), 168kg of nickel would be extracted

per hectare and a return of US$ 1,260 plus US$ 288 (from energy generation) would

be obtained. Nevertheless, given that nickel concentrations in plants grown in

Table 18.2 Concentration of Au in plants from different phytomining trials

Media Plant species Chelant

Au (mg kg−1)

ReferencePlant Media

Tailings Bothriochloa macra NaCN 24 1.75 [62]

Brassica juncea KCN 30 0.64 [59]

Brassica juncea NaCN 39 0.64 [59]

Helianthus annuus NaCN 19 2.35 [57]

Kalanchoe serrata (NH4)2S2O310 2.35 [57]

Sorghum halepense NaCN 24 2.35 [69]

Trifolium repens NaCN 27 1.75 [62]

Zea mays NaCN 20 0.64 [59]

Silica sand Berkheya coddii KCN 97 5 [61]

Brassica campestris NH4SCN 304 3.8 [70]

Brassica juncea NH4SCN 57 5 [60]

Brassica juncea KCN 326 5 [61]

Daucus carota (NH4)2S2O389 3.8 [58]

Rapahanus sativus NH4SCN 113 3.8 [58]

Soil Chilopsis linearis CH4N2S 296 5 [68]

Chilopsis linearis NH4SCN 197 5 [68]

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experimental studies are lower than those found in wild plants, the authors sug-

gested a more conservative estimate. Hence, considering average shoot levels of

5,000mg kg−1, the production of nickel would be 110kg ha−1, worth US$ 670 plus

US$ 288 from energy revenue [84, 86].

Subsequently, a 2003 study with Alyssum murale Waldst. & Kit. and Alyssum

corsicum Duby announced maximum shoot nickel concentrations of 22,000 mg

kg−1 and biomass yield up to 20 tons ha−1 [72]. Based on these gures, the authors

predicted nickel extraction of 400kg ha−1, worth US$ 1,749 (after deducting pro-

duction and land rental costs, and 25% of the nickel value to support the expenses

of metal recovery and license and royalty fees).

Recently, Bani etal. [87] undertook eld trials at two ultramac vertisol sites

from Albania, to assess the inuence of plant density on nickel phytomining with A.

murale. In the rst site, two different plots presenting soil nickel levels of 3,100 and

2,060mg kg−1 were planted with A. murale at a density of 1 and 6 plants per square

meter, respectively. In the second site (3,300mg kg−1 nickel), A. murale was planted

at a rate of 4 seedlings per square meter. The results showed that plant densities of

1, 6, and 4 plants per square meter generated biomass yields of 10, 5, and 10 tons

ha−1. Accordingly, nickel production was of 77, 41, and 112kg ha−1. Assuming the

current price of nickel, these results would provide an economic revenue of US$

728, 388, and 1,059, respectively, suggesting that a plant density of 4 plants per

square meter would be the most protable strategy.

Still in 2015, another study evaluated the biomass yield of Alyssum serpyllifo-

lium Desf. susbp. lusitanicum T.R.Dudley & P.Silva (also known as Alyssum pin-

todasilvae), growing under natural conditions in the Portuguese massifs of Morais

Table 18.3 Some nickel hyperaccumulators from different locations across the planet

Plant species

Concentration (mg

kg−1) Location Reference

Alyssum bertolonii 13,400 Italy [73]

Stackhousia tryonii 41,260 Australia [74]

Sebertia acuminata 11,700 New Caledonia [7]

Rinorea niccolifera 18,000 Philippines [75]

Psychotria costivenia 38,530 Cuba [76]

Phyllanthus insulae-japen 38,720 Indonesia [74]

Jatropha sp. 13,500 Brazil [77]

Alyssum murale 13,160 Serbia [78]

Bornmuellera baldaccii 12,115 Albania [79]

Berkheya coddii 11,600 South Africa [80]

Alyssum heldreichii 11,800 Greece [81]

Thlaspi apterum 21,500 Bulgaria [81]

Phyllanthus nummularioides 26,560 Dominican Republic [74]

Psychotria cf. gracilis 10,590 Malaysia [74]

Streptanthus polygaloides 14,800 California, U.S. [82]

Alyssum pintodasilvae 9,000 Portugal [83]

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and Bragança [88]. The results depicted average biomass production of 6.3 tons ha−1

in the Morais massif and 8.1 tons ha−1 in the Bragança massif. The correspondent

nickel yield of 27.7kg ha−1 (Morais massif) and 27.4kg ha−1 (Bragança massif)

would be worth US$ 512 and 506, respectively.

18.3.3 Thallium

In light of its scarcity [24], high bioavailability for plant uptake [89, 90], and ele-

vated price [91], thallium has also been the target of some phytomining-related

studies.

In 1999, an investigation conducted over the tailings of a lead/zinc mine at Les

Malines (Les Avinières, France) discovered thallium concentrations up to 3,070mg

kg−1 in the shoots of Iberis intermedia [92]. Although average thallium levels on

aboveground tissues surpassed 1,000mg kg−1 and the biomass yield pointed to 15

tons ha−1, the authors assumed 800mg kg−1 and 10 tons ha−1, respectively, to ensure

a conservative economic assessment. Thus, the resulting 8kg of thallium per hect-

are would be worth US$ 2,400 (considering the price of thallium at that time, US$

300kg−1). The same study has also analyzed the feasibility of thallium phytomining

with Biscutella laevigata. With a biomass production of 4 tons ha−1, B. laevigata

would need an average shoot thallium concentration of 425mg kg−1 to reach the

proposed $US 500ha−1 necessary to make phytomining viable. The results showed

that approximately 39% of the plants topped this concentration threshold [89, 92,

93].

Other studies have also presented additional thallium hyperaccumulators that

could be suitable for phytomining. Escarré etal. [94] found average shoot thallium

concentrations of 250mg kg−1 (maximum levels reached 1,500mg kg−1) on Silene

latifolia, growing on the aforementioned Les Avinières region, France. Jia and col-

leagues collected samples of Brassica oleracea var. capitata growing in long-term

thallium-contaminated sites from Lanmuchang, West Guizhou Province, China

[95]. Shoot thallium levels up to 1,503 and 818mg kg−1 were determined in plants

harvested in soils of mining sites and alluvial soils, respectively.

18.3.4 Rhenium

Rhenium is one of the scarcest (7 × 10−8%) and most broadly dispersed elements on

Earth’s upper crust. Because of its rarity and distinguishing physicochemical prop-

erties, rhenium is also one of the most costly metals [96, 97]. Rhenium is usually

found in soils as perrhenate (ReO−

4), its most stable form, presenting great mobility

and solubility [98, 99].

18 Phytomining ofRare andValuable Metals

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Following different accounts of rhenium accumulation, both on eld [100, 101]

and laboratory trials [102], two rhenium phytomining experiments were carried out

to date [14, 103]. The rst, published by Bozhkov etal. [103], is actually the corol-

lary of a series of earlier communications by the same authors [104–106], in which

besides reporting the concentrations of rhenium in plants growing around the Asarel

mine in Bulgaria, the results of two pot experiments are also presented. One of the

pot trials employed Atriplex hortensis, Polygonum fagopyrum, Medicago sativa,

and T. repens growing on soil spiked with an aqueous KReO4 solution 2-weeks after

seed germination, in order to attain a soil rhenium concentration of 128.72mg kg−1.

A week after, plants were harvested and rhenium levels up to 3,150; 9,130; 46,586;

and 35,090mg kg−1 were found in A. hortensis, P. fagopyrum, M. sativa, and T.

repens, respectively, denoting their capacity to hyperaccumulate this metal. The

other pot test [106] used soil collected at the Asarel mine (5mg kg−1 rhenium) and

the plant M. sativa. Plant tissue was analyzed 10 and 35 days after sowing, exhibit-

ing maximum rhenium levels of 2,780 and 4,870mg kg−1 ash. Based on rhenium’s

price at the time, the authors suggested a revenue of US$ 21,915 tons−1, which

would be roughly US$ 74,617 per hectare based on their biomass yield projections.

However, these results are extremely optimistic, for they are calculated on top of

maximum rhenium concentrations in plant tissue (not average values), and do not

consider any process expenses. Furthermore, their studies fail to provide crucial

methodological details and results (such as number of replicates used or biomass

production), experimental consistency (the pot experiment design is poorly

explained), and statistical analysis (inexistent), averting to duplicate their work.

In 2015, a pot trial developed under greenhouse conditions assessed the potential

of B. juncea and Equisetum hyemale growing on organic substrate spiked with

KReO4 to obtain rhenium concentrations of 5, 10, 20, 40, and 80mg kg−1 [14]. The

plants were harvested 45 and 75 days after sowing. According to increasing sub-

strate rhenium levels, B. juncea presented shoot rhenium concentrations that ranged

from 1,553 to 22,617mg kg−1 at 45 days, and 1,348 to 23,396mg kg−1 at 75 days,

whereas in E. hyemale the concentrations varied between 74 and 925mg kg−1at 45

days, and between 87 and 714mg kg−1 at 75 days. The authors estimated that, con-

sidering production and bio-ore extraction costs, a prot of US$ 3,906ha−1 could be

feasible.

18.4 Perspectives

In spite of the technological progresses of the mining sector, current solutions are

incapable of extracting the totality of metals from mineral ores. Thus, billions of

tons of toxic mine waste with residual metals are scattered throughout the 0.4 × 106

km2of land estimated to be affected by mining activities around the world [107].

Many of these metals are much in demand for their high market prices or strategic

signicance [108, 109], opening a clear window of opportunity for phytomining.

The auspicious results obtained during the last two decades reveal that it may be a

L.A.B. Novo et al.

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feasible and inexpensive alternative to exploit low-grade ores and mineralized or

polluted soils. Moreover, as a still infant technology, phytomining could benet

from a series of scientic advances and approaches to improve the process prot-

ability and sustainability, and stimulate its commercial application, including: (i)

the use of genetic manipulation to enhance metal hyperaccumulation [26, 27]; (ii)

the utilization of plant growth promoting bacteria to reduce metal phytotoxicity and

increase biomass yield [23, 110]; (iii) where necessary, restrict chelating agents to

highly biodegradable compounds in order to avoid collateral effects on the soil

microbiota and leaching of metals into groundwater [37]; (iv) the selection of native

or naturalized non-invasive plant species to prevent environmental damage; and(v)

the development of target-specic phytomining to meet precise and highly lucrative

market needs—such as invivo gold nanoparticles desired for industrial, chemical,

electronic, and medical applications [111, 112].

Phytomining is an exciting plant-based technology that after further improve-

ment may be successfully applied at commercial scale. The work carried out so far

suggests that phytomining is viable either as a standalone operation to recover valu-

able metals, or as a tool to nance the expenses of concomitant processes like

phytoremediation.

Acknowledgments The authors gratefully acknowledge nancial support from the Portuguese

Foundation for Science and Technology (FCT) under grant No. SFRH/BPD/103476/2014.

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A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1_19

Chapter 19

Air Phytoremediation

StanislawW.Gawronski andHelenaGawronska

Abstract Air pollution presently is a challenge fo-r many areas of the world. Plants

are higher organisms that can best deal with this problem despite the fact often in

the air is a mixture of pollutants of different origin and toxicity. The world of plants

is very diverse and well adopted to changes in the environment, including air. This

large biodiversity allowed to select species with a very high tolerance, which are the

base for the discipline known as phytoremediation. All plants during their presence

in the environment run the process of phytoremediation, but some species tolerate a

very high concentration of selected pollutants. Moreover, they are able to uptake/

accumulate and next to degrade/detoxify in order to make them less harmful.

Tolerant plant species can be found in very extreme conditions but for phytoreme-

diation are useful plant species which besides being cultivatable, produce a large

leaf area and biomass. Urban areas often contribute in creating high polluted sites as

street canyons, road crossing, bus stops, and surrounding of heavy trafc freeway.

In all these places, air pollution can be mitigated by the presence of selected plant

species. Additionally, agronomic practices allow to maintain them on a polluted site

and to form them in conguration for optimal deposition of pollutants. Air phytore-

mediation in urban areas, where at present men spend most of the time, is strongly

desired and hard to overestimate if environment and human health and well-being

are the prospect.

Keywords Air pollution • Anthropocene • Bioremediation • Effective air phytore-

mediants • Green infrastructure

S.W. Gawronski (*) • H. Gawronska

Laboratory of Basic Research in Horticulture, Faculty of Horticulture,

Biotechnology and Landscape Architecture, Warsaw University of Life Sciences,

Ul. Nowoursynowska 159, 02-776 Warsaw, Poland

e-mail: stanislaw_gawronski@sggw.pl

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19.1 Introduction

Air pollution nowadays is among the biggest challenges in urban ecology having

negative impact on human health, well-being, and the health of the environment as

a whole and remains a major issue in many parts of the world. Changes in the

economy and society started with the industrial revolution referred to as the begin-

ning of the Anthropocene [1] have led to signicant changes in the environment,

and we now know for sure in the atmosphere as well. These changes either cover

appearance or increase the level of various often very toxic pollutants, among which

as the most important are particulate matter (PM), volatile organic compounds

(VOCs) with most often listed in this group polycyclic aromatic hydrocarbons

(PAHs), ozone (O3), nitrogen oxides (NOX), in some cities sulfur dioxide (SO2), and

also heavy metals. When pollutants are inhaled, they act as potential carcinogens,

mutagens, allergens, and teratogens causing a number of diseases, mainly in the

respiratory and cardiovascular systems including lung cancer that consequently

increases morbidity and mortality. WHO reports that in 2012, around seven million

people died—one in eight of total global deaths—as a result of air pollution expo-

sure. The organization estimates and conrms that air pollution is now the world’s

number one environmental health risk [2].

Current knowledge and public awareness of the negative impact tends to

seek effective measures to reduce the negative impact and as far as possible

“repair” already degraded environment. A strategy for reducing emission and

remediating already emitted pollutants is urgently needed. Once pollutants are

already in the ambient air, with few exceptions, the only option to clean up the

air is to use a nature-based, environmental-friendly biotechnology called phy-

toremediation. In this technology, appropriately designed or selected higher

plant species together with their microbiome are employed to remove pollut-

ants from the air and to degrade or detoxify them in a sustainable, conscious,

and controlled manner.

It is important to keep in mind that raked leaves litter and mown, grass as con-

taining impurities, should be treated as pollutants.

19.2 Air Pollutants andTheir Origin

In most cases, these pollutants are of anthropogenic origin, such as fossil fuel com-

bustion, non-exhaust vehicular emissions, household, industrial production, demo-

lition of old buildings, burning of tropical forests and slash-and-burn agricultural

technique, and climate change. All of the above points lead to endanger the environ-

ment. Fossil fuel combustion pollutes air with a number of contaminants, among

which, as the most often and dangerous are listed: particulate matter (PM), volatile

organic compounds (VOCs) with the most often listed in this group polycyclic aro-

matic hydrocarbons (PAHs), heavy metals (HMs), carbon oxides as the toxic CO

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and major greenhouse gas CO2, oxides of sulfur (mainly SO2) and various oxides of

nitrogen (NOX), ozone (O3), and black carbon (BC) as not so dangerous. The SO2

and NOX are main gases responsible for acid rains.

19.3 Particulate Matter

The number one air pollutants on a global scale are particulate matter. They are

one of the most harmful air contaminants, especially for infants and elderly [3]. A

characteristic feature of the PM is the ability to stay in the air for quite a long time,

i.e., from hours to several weeks. They are able to move on long distances from

the source of their emission as, for example, from Asia to the West Coast of the

United States [4]. Chemically, PM is a complex, heterogeneous mixture of differ-

ent chemicals with some of them characterized as very toxic. Besides, they them-

selves are toxic; they also serve as the nuclei collecting on their surface other

contaminants. Based on their aerodynamic diameter, there are several classica-

tions of PM.Most often in literature, the classication with four fraction sizes is

cited: large (10–100μm), coarse (2.5–10μm), ne (0.01–2.5μm), and ultrane

(below 0.01μm). Particulate matter of smaller diameter than 2.5μm (PM2,5) and

10μm (PM10) serves as an indicator of air quality. The recently published study

showed, based on available information, that trafc (25%), combustion and agri-

culture (22%), domestic fuel burning (20%), natural dust and salt (18%), and

industrial activities (15%) are the main sources of PM which constitute to cities’

air pollution. However, there are signicant differences in these numbers, between

various regions of the world [5].

Usually their occurrence is promoted by an incomplete combustion. Globally,

the main sources of PM emissions are the burning of fossil fuel and biomass for

industrial purpose, food preparation, and homes’ heating. However, now in many

parts of the world, in the urban environment, a major source of pollution is undoubt-

edly considered transport. In the case of air pollution by motor vehicles and trafc-

related PM, pollution from exhaust and non-exhaust sources is estimated to

contribute almost equally. It is not always realized that about 50% of trafc-related

pollution comes from non-exhaust sources. It is generated by brake, clutch, tire, and

road surface wear and due to picked up materials by vehicle trafc from roadside.

Approximately 50% of the wear debris generated during braking will be in size of

PM10 and will stay in the air, while the rest as bigger and heavier becomes soon

deposited on road surface or nearby roadside [6]. There is probability that some part

of pollutants greater than PM deposited on the road under the vehicles of heavy

weight becomes crushed to PM size. Brake wear can contribute up to 55% by mass

to total non-exhaust PM and more than 20% to total trafc-related PM pollution.

It should be noted that once pollutants are emitted into the atmosphere, our oppor-

tunities to remove them are very limited. In cleaning up the air from pollutants

undoubtedly participate rain but, also plants especially those with large leaf area

play very important role [6].

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19.4 Organic Pollutants

Organic pollutants are formed during incomplete combustion as a result of burning

wood or peat that takes place at lower temperatures and subsequently are emitted to

the air. The list of organic pollutants opens polycyclic aromatic hydrocarbons

(PAHs) which can exist in over 100 different combinations. The US Environmental

Protection Agency has compiled a list of 16 priority PAHs, as the most common and

harmful to humans and environment. In other countries this list might slightly differ.

They are present in the atmosphere from primary sources and from reactions of the

parent PAHs that can be chlorinated and nitrated [7, 8]. PAHs are a problem because

they are very persistent in the environment, which increases with the number of

rings. If they enter into the soil (washed off by rain from air, plants, or other sur-

faces), they will be strongly retained by the soil sorption complex, and the more dry

the soil is, the PAHs will be bound more strongly. Washed off PAHs from other

sources are transported with the storm to the pond or river. Individual PAHs vary in

behavior. Those with three and four rings in the air very easily turn into a vapor

form. The PAHs with ve or six rings are solid and are very persistent in the envi-

ronment. Their degradation processes are performed by wood-rotting fungi and go

very slowly. N-PAHs are about ten times more carcinogenic compared to the

unmodied, parent PAHs. Both N- and Cl-PAHs are long lasting in environment.

Plants have great difculties in uptake of PAHs and their degradation. Gaseous

three- and four-ring PAHs, as all of them, are hydrophobic, but they can pass through

the stomata to the interior of the plants, while those with ve and six rings probably

partly penetrate in by process of diffusion through the lipophilic wax layer. Most of

the plants are not able to degrade PAHs for this purpose; they “employed” some

microorganisms of microbiome or accompanying phyllobacteria and endobacteria

living on the surface of leaves or inside the host cells, respectively [9].

19.5 Heavy Metals

Combustion of fossil fuels releases into the environment quite a few of heavy metal

list of which open very toxic arsenic (As), lead (Pb), mercury (Hg) and chromium

(Cr) VI, cadmium (Cd), recently joined antimony (Sb), and in high concentration

zinc (Zn) and copper (Cu). Toxicity of all of the above is well recognized and

described. However, in the last years, it has been demonstrated that the behavior of

metals emitted during the combustion can be more complicated, and the new struc-

tures are formed by them which are much more toxic. During the combustion of

fossil fuels in high temperature, metals will transfer to gaseous form and in part can

be easily oxidized. In this form metals became extremely reactive with organic com-

pounds forming products known as environmentally persistent free radicals (EFPRs)

[10]. Formation of EPFR occurs when oxidized metals as CuO2, ZnO2, and FeO2 are

reduced by the delivery of electron from organic compounds [10]. The new struc-

tures become free radicals, lasting from days for phenoxy radicals to months and

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even years for semiquinone radicals [11]. It is worth noting that present in signicant

quantities in the air, nontoxic form of iron oxide after the entry into reactions with

organic gaseous pollutants becomes very reactive pollutant as well as forming

EPFR.But the question is open: how much of the Fe is in the reactive toxic form of

EPFR. This as can be assumed depend on whether Fe is also accompanied by con-

tamination with organic compounds. From deposits on the leaf surface, substantial

amounts of Fe [12] plants probably use part by themselves [13]. Both metals and

organic compounds of high molecular weight slowly, on the base of physical diffu-

sion, penetrate through the wax into the rst living cells of the epidermis where they

are sequestered [14]. The penetration of heavy metals into the leaf via the epidermis

is conrmed by a presence of higher level Pb and Cr in the leaves, because these met-

als are uptaken from the soil only in small quantities [15].

Zinc is a very important element for plants, very easily taken up from the soil

even in high quantities if available, but in site heavily polluted by vehicles, it is also

deposited on leaves from polluted air. Its presence in the air is also deposited on

leaves due to abrasion of tires, brake linings with Zn, and corrosion of crash barri-

ers. Manganese is added to the steel (0.8%) as a coolant to lower the temperature. It

is difcult, however, to judge on its level in plant tissues whether its presence is an

effect of pollution and if so in what proportion, because this element can be taken

up in high quantities when plants are irrigated with water from deep wells.

Copper as an essential element for plants is very desirable; however, in higher

concentrations, it is toxic to them. In the soil Cu is a medium bioavailable, but in

polluted air, it also is coming from brake lining wear and alloy bearings. In brakes

manufactured Cu is used because it is an excellent conductor of heat; during braking

temperature, it can reach up to 1000°C [6]. The high number of cars currently in use

release to the environment large amount of Cu that some of the countries intend to

limit content of this element in brakes to be no higher than 0.5%. Elements as Pb,

Ni, Cr, V, and generally also Fe are poorly uptaken from soil by the plants and in

small quantities translocated to the aboveground part, so their presence in leaves

and twigs indicates on air pollution as a main source [16, 17].

Iron is emitted by transport vehicles in large quantities to the air [6], and proba-

bly substantial part of it is in oxidized form. The oxidized form of Fe probably

reacts with secreted organic contaminants by vehicles and forms highly reactive

EPFR. The phenomenon is a subject of study, but it is possible that we should

change our opinion that this element is not so toxic as is commonly believed. At

present it is assumed with high probability that Fe in form of EPFR is very toxic.

Lead is not well uptaken by plants, and its distribution to the upper part of the

plants is very limited. Pb presence depends more on the air-contaminated exhaust

fumes (even now in small amount from unleaded gasoline), from brake linings with

Pb, and with wind picked up particles of soil contaminated with this element in

previous years.

Nickel and vanadium are frequent additions to the alloy steel, and the level of

both in the air increases in sites polluted by transport. A good example is Norway

where the level of vanadium in the soil on clean sites was higher than in polluted

ones, but the leaves of the plant growing on polluted sites by transport were more

contaminated with V [17].

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19.6 Black Carbon

Black carbon (BC) consists of pure carbon as a product of incomplete combustion

of fossil fuel and biomass and is commonly referred to as soot. Toxicological study

indicates that the BC is not pollutant, which directly harms human health as a com-

ponent of the ne PM, but it plays a role of a carrier for a wide variety of toxic

compounds on its surface. BC is the most strongly light-absorbing components of

PM.BC suspended in the air absorbs sunlight and generates heat in the atmosphere,

when deposited on snow and ice, which are light in color; thus, it facilitates increased

melting and diminishes their otherwise substantial reective capacity [18].

Therefore, BC contributes to climate change [19] and thus has a strong direct inu-

ence on global warming. Control of BC might slowdown this process [20], and it is

supposed that it will be more effective than our efforts to reduce CO2 emissions. We

can assume that BC similarly affects also temperature of the plant surface in addi-

tion to reduction of the access of light to the photosynthetic apparatus. Inhibition of

photosynthesis recorded in plants growing in the vicinity of coal mine partially can

be explained by the negative impact of BC on plants [21]. In the cocktail of air pol-

lutants, sometimes it is difcult to determine exactly the negative action of BC itself

since it is also a carrier of many others, often more toxic compounds. Also during

forest res, BC is one of the predominant components of the aerosol with its nega-

tive impacts on photosynthesis [22]. Coniferous compared to deciduous plant spe-

cies better withstand the stress conditions as in the rst smaller reduction of

photosynthesis is noted [22].

19.7 Gaseous Pollutants

Dominant route in the penetration of gaseous contaminants to the plants is stomata,

cell structure that controls the plant gas exchange, but non-stomatal entering of

gases takes place too. This way gets into the plant: CO2, CO, NOx, SO2, O3, and

gaseous two- and three-ring PAHs.

The CO2 released into the atmosphere is a major greenhouse gas, undoubtedly

considered as the main causative factor of climate change, but for plants, it is one of

the nutrients and promotes their better growth. A huge global surface of the plants

makes them a very important player in the removal of the gaseous pollutants from

the atmosphere. Biologically active surfaces that are most exposed to contact with

penetrating contaminants from the air as plasmalemma of plants’ or lung epithelium

have protective mechanisms. The contact layers are reaching in four major low

molecular mass oxidants: ascorbate (vitamin C), uric acid, reduced glutathione, and

alpha-tocopherol.

There are many different reactive forms of nitrogen present in the air environ-

ment. In urban areas, the main source of emissions of oxidized form of nitrogen

(NOx) is transport, with NO2 present in high quantities. The further fate of NO2

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depending on environmental conditions can be very different. A high temperature in

the presence of ultraviolet promotes the formation of O3, reactive radicals oxidizing

most of biological molecules, and NO.This process is reversible, and when envi-

ronmental conditions change, it may proceed in the opposite direction forming

again NO2 and O2. In the presence of H2O, other transformations create a mixture of

nitric and nitrous acids (HNO3, HNO2), both are of acid rain components. The third

possible direction of NO2 transformation is process of nitration, which in recent

years has devoted more attention. It was experimentally conrmed that NO2 in pol-

luted urban areas can promote the nitration of protein and peptide molecules. The

process of posttranslational modication of proteins by nitration, in an environment

polluted with NO2, runs very smoothly as shown in a study by Franze etal. [23],

where in a period from several hours to several days, 20% of the analyzed samples

were modied. For a long time, results of observations and studies indicated that

asthma and allergic diseases are enhanced by trafc-related air pollution, and it is

already conrmed that nitration enhances allergic responses [24]. Most studies of

this phenomenon were conducted on the protein Bet v 1, birch pollen, and it was

conrmed, on molecular level, as a posttranslational modication of this protein by

NO2 [25]. In a polluted environment, higher allergenicity and changes in the protein

composition of the pollen of Platanus orientalis was noted, one of the most impor-

tant species in urban forestry of the warmer area of the world [26]. Also, pollen of

invasive weed Artemisia artemisiifolia is more potent in allergenicity at elevated

levels of NO2 [27]. Crown evidences are samples of birch pollen collected in the

natural conditions, which showed that pollen from urban areas had a higher aller-

genic potential than pollen from rural areas, despite the fact that content of allergen

in analyzed samples was the same [28].

Nitrogen is the second most important element after carbon for plants, primarily

for protein synthesis. Not all forms of this important element for life are available

for plants including the most abundant atmospheric N2. In this situation plants, if it

is possible, uptake nitrogen in any available form if they only can do that, even in

the form of ammonium, which when uptaken in larger quantities can be toxic. In

light of the above question arises whether plants are able to use nitrogen from NO2.

Answers to this question were a subject of study by a team led by H.Morikawa [29].

They studied 217 of herbaceous and woody plant species in terms of their ability to

use NO2 as nitrogen source. Authors reported that plants are using nitrogen from

NO2 and found out that between tested species, huge differences exist in the capac-

ity of using NO2. They highlighted the group of species best in the ability to uptake

15N-labeled NO2 and named them as NO2-philic. The most efcient in the uptake

and assimilation of NO2 within woody species proved to be Magnolia kobus,

Eucalyptus viminalis, Populus nigra, Robinia pseudoacacia, E. grandis, E. globu-

lus, Populus sp., Sophora japonica, and Prunus cerasoides; within herbaceous

cultivated plant, efcient was Nicotiana tabacum, and from herbaceous plants, the

naturally growing along roadsides was Erechtites hieracifolia. Further research in

this laboratory evaluated the uptake of NO2 by 70 woody species (present in the

surroundings of a road) in a concentration 0.1μll−1 and compared with concentra-

tion 4μll−1 which was already inhibiting assimilation of NO2. The most valuable

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result of this experiment was identication of a group of species that tolerate high

level of NO2 and lead efcient assimilation at both high and low concentration of

this pollutant, among which valuable in this context appear Robinia pseudoacacia,

Sophora japonica, Populus nigra, and Prunus lannesiana [30]. It is interesting that

these species are also listed as good phytoremediants of heavy metals from soil.

A more dangerous and toxic to hemoglobic organisms (including humans) is the

carbon monoxide (CO), which naturally is secreted in small amounts by all living

organisms, from bacteria to humans. It is also one of the products from the combus-

tion appearing in the large quantities when oxygen is limited. Although for man it is

extremely toxic but plants tolerate higher levels of CO, what is conrmed by their

presence noted nearby sites where combustion of fossil fuels and biomass takes

place? The amount of CO detained by plants from the downloaded to the air was

assessed already many years ago. In the 1980s of the last century, one of the most

extensive experiments was performed [31], in which ability of 35 plant species to

accumulate 14CO and its further fate was studied. Several fold interspecies differ-

ences in 14C content in the leaves of tested species and differences in metabolism

were found. Absorbed CO in the corn (C4 plant species) was oxidized to CO2, while

in the bean plants (C3 species), part of the CO was also oxidized to CO2, but the

other part was reduced and incorporated into the amino acid serine. The list contains

17 surveys of woody species, among which stand out the following species: Ficus

variegata, Acer saccharum, A. saccharinum, Gleditsia triacanthos, Pinus resinosa,

P. nigra, and Fraxinus pennsylvanica. Also two shrub species Syringa vulgaris and

Hydrangea sp. are characterized by their high ability to x CO [31].

19.8 Organic Carbon

Organic carbon (OC) creates structures of organic compounds which are emitted by

plants directly into the air but also can be formed from organic precursor gases emit-

ted from anthropogenic and natural sources. Particles containing OC may also pose

some risk to human health [32]. Some plant species release signicant amounts of

organic compounds consisting OC called biogenic volatile organic compounds

(BVOCs) to defend themselves against herbivores, pathogens, and other stress fac-

tors. To this group of compounds, what belong mainly are isoprene, monoterpenes,

sesquiterpenes, and other C10-C15 BVOCs [33]. These stressful conditions cause the

creation in the tissues of leaves ROS against which the plant defends itself leading

oxidation of isoprene [34]. The phenomenon is still being investigated, but there is

evidence that high level of BVOC contributed ozone and EPFR formation. Ozone

and EPFR forming potential of plants should be taken into account in increasing of

the green infrastructure. Taha [35] suggests limit from the cultivation quantity of the

tree species that emit more than 2μgg−1h−1 isoprene (isoprene of micrograms per

gram of dry-leaf mass per hour) and 1μgg−1h−1 of monoterpenes. According to

Curtis etal. [33], Aesculus glabra exceeds this value considerably and Corylus col-

urna and Tilia americana slightly. Similar studies previously were conducted by

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Benjamin and Winner [36] who assessed the ozone-forming potential (OFP) of 308

species of trees and shrubs dividing them into three groups: low-OFP contributing

less than 1g ozone d−1, medium between 1 and 10g ozone d−1, and greater than 10g

ozone d−1. The results of evaluation of so many species grown in urban areas are

valuable source of information till present. In the group of low-emitting BVOC,

there are popular genera such as Malus, C. camphora, orange, and pear and species

Ginkgo biloba and Juglans nigra. The group of high-emitting BVOC belong to the

following genera: Salix, Quercus, Populus, Pinus, and Liquidambar. Similar assess-

ment in Europe to carry out Karl with coauthors [37] also noted the high emissions

of BVOC by species belonging to the previously mentioned genera. According to

Baraldi et al. [38], BVOC emissions are conducive to both high radiation and

temperature.

19.9 Plant Species forAir Phytoremediation

As already mentioned, when pollutants are emitted into the outdoor air, our oppor-

tunities to remove them are limited, and besides rain, only plants are participating in

this. At this point, it should be remembered that in parts of the world with higher

precipitation, substantial part of PM deposited on leaves is washed off from them

into the soil or on sealed urban area with every rainfall. Plants as autotrophic organ-

isms during photosynthesis are carrying out gas exchange, uptake of CO2, and

release of water vapor process fundamental for plants’ life importance. However,

simultaneously, pollutants present in the ambient air, list of which is very long, also

enter into the plants’ tissues. For the efcient running of this crucial process for

plants, they developed a series of defense mechanisms against uptaken air pollut-

ants. In consequence, plants, by the way, act as an efcient biolter removing from

the air, at the same time, many pollutants with varying actions of toxicity. In that

context trees and shrubs deserve special attention, Leaf Area Index (LAI) of which

is relatively high. LAI indicates how many times 1m2 of the ground, occupied by

plant(s), is covered by them. In this role the best are trees, LAI of which reaches ten;

for shrub, it usually ranges from ve to seven but also herbaceous plants like grass

lawns with the index around two plays a positive role.

The presence, in environment, of plants accumulating signicant amount of PM

enables them to act as a biological lter [39, 40]. In several studies signicant dif-

ferences between plant species in the level of PM accumulation were reported [41–

43] with some of them perfectly acting as a biological lter. Penetration of

contaminants by a biolter and stopping them depends on the turbulence around

plants’ organs and increases with faster air movement. Also leaves folding and the

presence of various formations on their surface as, for example, trichomes or hairs

play positive role in PM accumulation [41, 42]. Besides, metabolic uptake by plants,

the stickiness of the leaf surface, and their aerodynamic properties are involved in

higher PM accumulation [43], and electrostatic forces between heavy metal ions

settling on PM and leaf blades cannot be ruled out [44]. In the deposition of PM on

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the leaf surface, very important is the presence of the epicuticular wax, which the

main role is to protect plants against negative external factors. In fullling this pro-

tective role, important are the thickness of the wax layer, its structure, and chemical

composition. Plant physiologist distinguished 23 major forms of wax [45].

Accumulated on the surface, PM in part remains on it, while the other parts of PM

penetrate deeper and stuck in the wax layer (hereinafter designated as surface PM,

sPM and in wax PM, wPM, respectively) [41]. sPM can be easily washed off by

rain, while wPM stays in the wax for much longer time [46, 47].

In the temperate zone falling in the autumn, leaves contain signicant amount of

both sPM and wPM weight of which per 1 cm−2 often is around or even exceeds

40μg. It is not worthy that 40μgm−3 of air is designated as a norm according to

European Union standard. Comparison of these values shows the potential of what

vegetation has in cleaning of the air. Rain washes off PM both from the air and from

the surface of the leaves and thus cleans up the air, but it should be noted that between

rainfalls, just plants accumulate PM on their surface and in wax, thus ensuring con-

tinuous removal of pollutants from the air. Although part of PM that stuck in wax is

removed together with wax as leaves are getting older, a new wax is produced and

layered on the leaves’ surface into which another part of PM from the air will pene-

trate. Therefore, it can be assumed that in older leaves, wax layer already developed

remains in similar amount through the vegetation time; in spite of that, wax is entrained

by wind and rain [48]. Torn off by the wind, wax together with wPM as heavier than

just PM much faster falls at a certain distance from the tree/shrub, while the wax with

sPM is washed off by the rain, which gets into the soil under the plants. In recent years,

several research centers evaluate plant species in order to recommend them for culti-

vation in areas with PM-polluted air. Position of a given species on the lists drawn up

is variable and depends on location and local pollution and the weather conditions as

wind and rain [49]. However, there are species which are located at the top part of the

list, despite growing in different locations. The following tree species are listed as

tolerant to the pollution and accumulating signicant amount of PM, so they can be

considered as potential phytoremediants: Pinus sylvestris, Betula pendula, Pyrus cal-

leryana, Sorbus intermedia, Populus sp., Alnus spaethii, Robinia pseudoacacia,

Elaeagnus angustifolia, Sophora japonica, Fraxinus pennsylvanica, F. excelsior,

Quercus ilex, and Tilia x europaea “Pallida” [41, 50–54].

In the cities shrubs play also an important role as phytoremediants. Due to their

smaller size, they can be planted closer to the edge of the road, and, therefore, they

would be the rst objects on which emitted from vehicle engines pollutant are

directed. Proposed for PM phytoremediation shrubs are Pinus mugo; Syringa mey-

eri; Spiraea sp. and Stephanandra incisa; Taxus x media, T. baccata, and Hydrangea

arborescens; Acer campestre and Physocarpus opulifolius; Sorbaria sorbifolia; and

Forsythia x intermedia [50, 51]. An interesting group of plants for phytoremediation

is climbers, which occupy little land but produce a large leaf area directed vertically.

They can, therefore, be planted in downtown of old cities where land is usually

extremely limited as well as on screens of the highways. List of recommended spe-

cies for temperate zone is short and comprises Hedera helix, Parthenocissus tricus-

pidata, P. quinquefolia, and Vitis riparia ([55] and authors not published data). One

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of the most tolerant climbers in polluted urban environment is Polygonum aubertii

[56]. Although it is a little less decorative species but as fast growing and very toler-

ant to pollution, the species is recommended for places where garbage-collecting

containers are located as excellent for camouage of this place.

In phytoremediation, natural herbaceous vegetation plays an important role in

colonizing almost any available soil surfaces in urban areas and providing therefore

a surface for PM deposition. LAI of these species is much smaller but still on some

species like Achillea millefolium, Berteroa incana, Polygonum aviculare and

Brickellia veronicifolia, Flaveri trinervia, and Aster gymnocephalus, much greater

amounts of PM are deposited than in other species [57]. In case of Aster gymno-

cephalus, additionally accumulation of signicant amounts of heavy metals was

recorded [58].

19.10 Green Infrastructure Directed toPhytoremediation

The presence of any plant promotes phytoremediation, but its signicant level is

important for the environment and can be achieved only in properly planned green

infrastructure. Each technology has its own niche. In the case of phytoremediation,

it is very effective in mitigation of pollution but not from very highly polluted air (as

far as plants as living organisms can endure it); at the same time, it is a cheap tech-

nology and can be used on large scale. At present, our task is to recognize how this

process is run, what kind of mechanisms are involved, and which species are carry-

ing out phytoremediation most efciently. Once this technology is introduced in

practice, questions will always be raised on how to make it more efcient and cost-

effective in comparison to other technologies.

We should of course remember that general principle in environmental protec-

tion strategies is elimination or reduction of both the source and level of pollutants

emission. These concerns are also underlined by WHO [2] in recommendations of

a series of measures for mitigation of the negative impact of polluted air on humans

and environmental health and men’s well-being. Many of these actions are already

in the hands of city authorities, who carry them through introduction of several

legislations as, for example, restriction of wood and low-quality coal burning. Also

recommendations proposed in California like improving trafc ow, enforcing

speed limit, reducing number of stops and sharp turns, imposing vehicles’ weight

limits, and rerouting tracks are quite easy to implement [59]. All these can be applied

via administrative decisions. These actions play crucial role in prevention/reducing

pollutants emission, but unfortunately they do not eliminate pollutants that are

already being emitted into the atmosphere.

Having in mind the current status of air pollution, in many parts of the world,

strategies/technologies of pollution mitigation are urgently needed. When pollut-

ants are already in the atmosphere, the only option where men are able to lower the

pollution is to use environmental-friendly biotechnology called phytoremediation

where plants are used as a tool for cleaning up the air.

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Plants as organisms leading sedentary type of live possess several defense mech-

anisms against stresses including negative impact of air pollution. Plants possess

large ability to change or even create the environment in order to make it optimal or

close to optimal for them. They would achieve this goal much faster if they exist in

greater amount of what is conrmed by the environmental conditions prevailing in

the forest or park. The presence of plants in an appropriate number ensures and

improves several environmental parameters as increased humidity level, lowered

temperature, inltrated storm water, reduced wind and noise, and aesthetic values.

The team under the leadership of Dr. Nowak estimated that green infrastructure

in the United States saves the death of 850 people and 670,000 from the problems

of respiratory system (per year) as a result of exposure to air pollution [60]. Other

detailed evidence on the positive greenery effects obtained in NewYork is reported

by Lovasi etal. [61], who showed that children living in the streets with trees get

sick much less on asthma in their early childhood as compared with those who were

living nearby the streets without trees. Focusing green infrastructure on air purica-

tion should simultaneously also build health-supporting role of the ecosystem, like

the improvement of city climate, and create space for physical activity and health

conditions of the residents [62]. The efcient removal of the pollutants from the

outdoor air via phytoremediation is challenging especially that this technology is in

the early stages of its development. However, the ability of plants to change a cli-

mate in the city is building a hope to use them for effective removal of air pollutants

as well. Therefore, one of the rst tasks is to increase the green infrastructure, espe-

cially with trees since they have a large surface and thus are able to perform clean-

ing the air on greater scale, and additionally below them, there is a space for shrubs

and herbaceous plants. A good example of such activities is NewYork, where cam-

paign of planting of one million new trees by 2017 has become a model for copying

by many cities in the world. The city authorities have taken a number of other initia-

tives to increase green infrastructure, of which the most spectacular was/is to build

a linear High Line Park on the viaduct defunct railways of the city. On the 2.5km

long park, more than 160 species of plants from different habitats were gathered,

making this place the biggest and unique attraction in Manhattan [63].

It is obvious that the most important role in greenery plays larger areas with trees

as parks and forests surrounding the city as well as green belts along roads with

extensive trafc. This raises the question of how wide should be the belt of vegeta-

tion needed to perform effective biolter function. During examination of plants’

efciency as a biolter, a signicant reduction in particulate matter on leaves of

Quercus ilex was noted at a distance from 0 to 20m from the edge of the road. This

reduction was stronger in the lower part of the crown, at a height of 1.5–2m, which

on average is at the height of a man’s face [53]. A similar pattern of PM accumula-

tion on the leaves (at the same height) of the Tilia cordata grown in the park was

recorded by Popek etal. [64]; the greatest accumulation of PM was on trees grown

15m from the edge of the road with heavy trafc. In further distances signicant

decrease (around 50%) in the amount of PM was recorded. Interesting is the fact

that in the rst measuring point (3m from the roadway), a little less of PM was

accumulated on the leaves, that was probably due to their entrainment by moving

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vehicles and PM pumping into the biolter, which pose some resistance to pollut-

ants’ penetration [64]. The process of pollutants movement through biolter con-

sists of two interdependent phenomena which take place (1) deposition and (2)

dispersion. The effectiveness of these processes is demonstrated by results of

Al-Dabbous and Kumar [65], which showed 37% reduction in PM on the sidewalk

separated from the roadway with coniferous plants. This evergreen plant as a barrier

for PM is demonstrated by the fact that on its part facing to the road, 11% higher

amount of PM was noted when compared to the area without vegetation. Interesting

results on the evaluation of various barriers, i.e., vegetation barrier and solid barrier,

and of combined applications of both were presented by Tong with colleagues [66].

The vegetation barrier consists of conifers plants that, as evergreen, performed phy-

toremediative role all year round. Authors evaluate the ability of the plants in the

barriers to mitigate the pollutant emitted by vehicles as referred to trees’ height and

distance on which PM is moving. Both the height of the barrier (6 and 9m) and its

distance from the road did not have signicant effect on PM movement on further

distances. On the other hand, however, when they compared the width of the vegeta-

tive barrier (6m, 12 m, and 18 m), the results conrmed and expected positive

impact of the greater width. Authors of this interesting and very useful planning of

roads and surrounding studies recommend two options. The rst consists of a wide

vegetation barrier with high canopy density with two rows having tall plants condu-

cive to falling particles between the two rows and the second row that would be a

barrier for deection of pollutants. The second option is a combination of solid bar-

rier (6m high) with vegetation barrier (9m high) located behind the solid ones. It

can be assumed that, in addition to accumulation of pollutants, the solid barrier

would better protect against noise.

Vegetation should constitute a sort of sieve that allows to penetrate pollutants

through, but it should not be too dense to prevent increasing of pollution concentra-

tion. Additionally, if possible, architecture of the vegetation should direct pollut-

ants toward the ground in order to avoid deection of polluted airow over the

biolter [67]. In many cities, buildings are close to streets that form a kind of

canyon, where the dynamics of the air movement depending on the wind direction

differently behaves and the presence of trees in canyons might increase pollutant

concentrations. This is conrmed by measurements carried out in such conditions,

and authors suggest to consider the possibilities of the negative impact of urban

forestry in canyon street in the city [68]. Although the positive impact of urban

forestry in reduction of air pollution is widely known, but in this situation, it is

necessary to develop an optimal status of greenery. The results of Jin etal. [69]

with Platanus x acerifolia and Cinnamomum camphora, very common tree species

on the streets in China, indicated the optimal range for these species as 50–60%

and 1.5–2.0 for canopy density (CD) and leaf area index (LAI), respectively. These

optimal parameters can be obtained through plant species selection, their planting

pattern, and pruning. Additional reduction of air pollution by urban forest in the

canyon street can be achieved by creation of green walls on the buildings, which

are able to reduce NO2 and PM10 by 40% and 60%, respectively, as well as, although

less effectively, green roofs [70].

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The evaluation of the reduction of pollution by urban forest in the Leicester City

Center (UK) demonstrates that trees reduce concentration of ambient vehicular pol-

lutants, at pedestrian height, on average by 7% [71]. To know the dispersion of air

pollutants, a group of researchers from the Aarhus University developed Operational

Street Pollution Model (OSPM) which elaborated the behavior of air masses in the

street canyons, taking into account the impact of wind resulting in higher concentra-

tions of pollutants in the leeward side than in windward ones [72]. The model in the

following years has been further developed for urban areas [73]. The program has

been proved as very useful, and now, it is used by many companies from the area of

planning and urban lands management on different continents.

To assess the phytoremediation potential of urban forestry, iTree and UFORE

models that rened and took into account general parameters like crown volume,

species-specic effects, plant health conditions, different designs, etc. were devel-

oped [74].

19.11 Agronomic Recommendation Supporting theAir

Phytoremediation

Air pollution is usually a mixture of a number of pollutants; therefore, it is neces-

sary to grow the species known as highly tolerant to adverse environmental condi-

tions. Plants developed a number of defense mechanisms including creating a

barrier on their surface to stop pollutants from entering inside of the plants and, in

case when pollutants would get in, to transform and maintain them in the nontoxic

form. In the polluted air dominated with contamination such as PM (vehicle and

household emission) and NO2 (heavy track diesel emission), it is recommended to

cultivate species tolerant to these pollutants or with some risk for the success, spe-

cies of the same genus. Trees growing in the canyon streets and in front of the bio-

lter surrounding roads should be some kind of sieve, through which the air ows

to encounter small obstacles and deposit the pollution on subsequent encountered

leaves, twigs, and branches. In the case of schools, kindergarten and playground

pollutants should well penetrate into the biolter (to avoid deection of airow) and

then must be stopped inside the biolters to ensure the greatest possible safety for

the children. Since cultivated plants possess characteristics for a given species/cul-

tivar architecture, crown size, shape, and canopy density, these parameters must be

taken into consideration in planning greenery in the specic sites.

In places with high pollutant emissions as crossroads, bus stops, and uphill

streets, it should be preferred that species are not only tolerant to air pollution but

also are good phytoremediants accumulating pollutants, that along with mown

grass, falling in autumn litter, can be raked and burned in an incinerator. Trees and

shrubs growing in urban areas are often formed by pruning for better fullling aes-

thetics goals, health, or rejuvenation purposes; however, in light of the obtained

results, we should form them also for optimal phytoremediation canopy density.

More resistant to pollution are deciduous trees, but on the autumn, they lose their

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foliage meaning that during winter they do not play a phytoremediative role. It can

be, to some extent achieved, by planting together with them with evergreen climbers

as, for example Hedera helix. It is worth noting also that young oak (usually up to

10 years) and hornbeam formed as a hedge do not shed their leaves until spring;

therefore, all this time, the phytoremediation functions are fullled.

Man uses plants as a building material, source of food, and energy and has

recently “employed” them as green liver for helping in cleaning up air in the

humanosphere.

Nowadays, men are trying to take advantage of this by “harvest” of plant bio-

mass saturated with pollutants via an operation called phytoremediation which con-

tributes in the purication of air of the humanosphere in order to make it more safe

both at present and for the incoming generations.

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S.W. Gawronski and H. Gawronska

guarino@unisannio.it

A.A. Ansari et al. (eds.), Phytoremediation, DOI10.1007/978-3-319-52381-1

Abscisic acid (ABA), 24

ACC deaminase (ACCD), 30

Acer platanoides, 117, 120–122, 124, 125

Acorus calamus, 302

Agricultural water management, 424

Agronomic phytofortication, 232

Agrostis tenuis, 263

Air phytoremediation

autotrophic organisms, 495

biolter, 495

environmental conditions, 500

humanosphere, 501

leaf surface, 496

natural herbaceous vegetation, 497

penetration, 495

PM deposition, 497

PM-polluted air, 496

pollutant emissions, 500

temperate zone, 496

tree species, 496

trees, 500

vehicle engines, 496

Air pollution

origin, 488–489

particulate matter, 489

public awareness, 488

toxic pollutants, 488

Allium cepa, 387

Alyssum bertolonii, 109

1-Aminocyclopropane-1-carboxylic acid

(ACC), 11, 27

Ammonium thiocyanate, 475

Amorphous Mn oxide (AMO), 376

Angiospermae, 106

Anthropocene, 488

Anthropogenic chemicals, 158

Antioxidant enzyme, 25

Antropogenic pollution

assessment, 114

ecosystems, 114

elemental determinations, 120

gamma activity, 120

gamma spectrometers, 120

grading soil contamination, 115

INAA, 119

industrial enterprises, 115

model region, 115

MPL average, 115

plant samples, 119

regional center, 115

sampling points, 119

short-lived isotopes, 120

Tula city, 115

Tula industry, 115

washed samples, 119

woody plants, 117, 119

Apocynum lancifolium, 31

Aquaculture, 308–320

Aquaculture in Southeast Asia, 310, 312

Aquaculture in Vietnam, 313

Aquatic macrophytes

Azolla caroliniana (Mosquito Fern), 268

Brassica juncea (Mustard Green), 269

classication, 267

Eichhornia crassipis (Water Hycianth), 268

Hydrilla verticillata (Hydrilla), 270

Lemnoideae (Duckweeds), 269

Pistia stratiotes (Water Lettuce), 269

Ricciocarpus natans, 271

Schoenoplectus californicus (Giant

Bulrush), 271

Index

guarino@unisannio.it

506

Aquatic macrophytes (cont.)

Spirodela intermedia (Duckweed), 270

Vallisneria spiralis, 271

Arabidopsis halleri, 209, 355

Arabidopsis thaliana, 265, 384

Arbuscular mycorrhizal fungi (AMF), 5

Arsenic, effects of, 158

Articial wetlands (AW), 295, 417

Atlantic, 188

Atriplex halimus, 31

AtZIP1, 354

AtZIP2 T-DNA, 354

Autochthonous bacteria, 55, 71

Autochthonous biodegrading bacteria, 59–62

Auxine, 60–61

Azolla caroliniana (Mosquito Fern), 268

Bac Kan Province, Northern Vietnam, 161

Bacopa monnieri, 265

Bacterial consortium, 67

BALANS model, 334, 335, 340

Baldur, 188

Bambara groundnut (Vigna subterranea), 433

Benzene biodegradation pathways, 78

Bioaccumulation

bioindication, 107–108

environment, 107–108

factor analysis, 143–144, 188

heavy metals, 124

iron, 132–133

Mn, 130

Ni, 134

Tula region, 117

in woody plants, 142–144

Zn, 133, 134

Bioaugmentation, 5, 54, 56, 63, 65–67, 70, 72

Bioconcentration factor (BF), 471

Bioelements

on Brassica, 185–187

on medicinal plants, 207–212

on woody plants, 225–228

Biolm formation, 8

Biogas, 302

Biogeochemical activity, 126

Biological ligands, 204

Biological oxygen demand (BOD), 441

Biomonitoring

aboriginal species, 125

anthropogenic load, 120

antropogenic pollution, 114–126

atmospheric deposition, 105, 106

bioaccumulation, 124

biogeochemical parameters, 120

bioindication, 104, 109

biomarker, 105

chromium, 121

concentrations, 120

copper, 122

environmental monitors, 104

indicator, 105

leaf biomass, 145

leaf chlorosis, 144

levels, 104

mechanisms, 104

metallurgical defense, 126

northwestern regions, 106

phytoremediation, 145

quantication, 105

quantitative characteristics, 104

requirements, 104

statistical analysis, 145

surface area, 106

toxic components, 105

toxic elements, 124

woody plants, 121, 145

Bioremediation, 54–57, 62, 63, 69, 72, 74, 75,

82, 83

metallurgical enterprises, 139

phytoextraction, 109–110

phytoremediation, 109

shrub leaves, 139

soils contaminaion, 137

Biosparging, 68–69

Biostimulation, 54–57, 63–65, 69–71

Black carbon (BC), 492

Blue revolution, 234

Brassica

cadmium and bioelements on, 185–187

general characteristic, 185

Brassica campestris, 200

Brassica chinensis, 197

Brassica juncea (Mustard Green), 160, 171,

173, 200, 216, 218, 262, 269, 386,

389, 411, 416

Cadmium

on Brassica, 185–187

on medicinal plants, 214–223

Cajanus cajan (Pigeon pea), 433

Calcium signaling, 348–349

Calendula ofcinalis, 214

Californium, 188

Canna indica, 302

Canopy density (CD), 499

Index

guarino@unisannio.it

507

Carbamazepine, 282, 283, 285

Cattail (Typha latifolia L.), 449

Chambers, 477

Chamomilla recutita, 211

Chelant-assisted phytoextraction

bioavailability in soil, 472

characteristics, 472

inorganic and organic agents, 472

inorganic chelants, 473

LMWOAs, 472

mechanisms, 472

natural/synthetic chelants, 472

organic and inorganic chelating agents,

473, 474

plant uptake, 472

solid phase, 472

stability constant, 472

synthetic APCAs, 472

Chemical oxygen demand (COD), 441

Chenopodium album, 31

Chickpea (Cicer arietinum), 433

Chilopsis linearis, 475

Chlamydomonas reinhardtii, 193, 216, 228

Chlorella vulgaris, 201–203

Colocasia tonoimo, 302

Common bean (Phaseolus vulgaris L.), 433

Composite tailings (CT), 22

CO2 neutral fuel, 453

Constructed Wetlands (CW), 280–283

coupling plant microbial fuel cells,

286–289

salt marsh plants, 281–282

substrates, 282–283

Contamination, 345

Converging technologies, 234

Core drill to collected soil, 58

Correlation analysis, 141

Cowpea (Vigna unguiculata L.), 433

Crinum americanum, 265

Crops, 184–187

Brassica

cadmium and bioelements on, 185–187

general characteristic, 185

cadmium and zinc compounds effects,

214–223

metal complexes on, 200–206

to toxic metal application, 188–191,

193–200

Cucumis sativus, 216, 380

Cucurbita maxima, 381

Cyclohexylenedinitrilotetraacetic acid

(CDTA), 472

Cyperus papyrus, 300

Cytokinins, 8

Daucus carota, 474

Denaturating gradient gel electrophoresis

(DGGE), 89

Detoxication strategy in plants

mechanisms, 352

metal transporters, 352–353

soil to root uptake, 353–354

vacuoles, 355

Diethylenetriaminepentaacetic acid

(DTPA), 472

Ecosystem services provision, 300–302

Effective air phytoremediants, 488, 497

Eichhornia crassipis (Water Hycianth), 193,

264, 268, 271, 414

Electrolyte leakage, 40, 41

Eleusine indica, 163, 166–168, 171, 172,

177, 179

Endocytosis, 381

Endophytic bacteria, 89

Environmental remediation, 372

Environmentally friendly technologies,

294–297

Environmentally persistent free radicals

(EFPRs), 490

Ethylene glycol tetraacetic acid (EGTA), 472

Ethylenediamine di-o-hydroxyphenylacetic

acid (EDDHA), 472

Ethylenediaminedisuccinic acid (EDDS), 472

Ethylenediaminetetraacetic acid (EDTA), 212,

415, 472

Eutrophic lakes, 294, 295, 297–300, 302

Eutrophication, 293, 294

Evapotranspiration, 453, 454

Extracellular defense strategy of plants,

345–346

Extracellular signaling, 347

Farmers, 424

Festuca arundinacea, 65, 70

Festuca rubra, 263

Field evaluation of heavy metal

Ha Thuong and Tan Long Mines, 173–175

Ha Thuong Field Experimental Site,

173–175

Tan Long Field Experimental Site, 175

Floating treatment bed (FTB), 297

Floating Treatment Wetlands (FTW), 295,

297, 302, 303, 318

Index

guarino@unisannio.it

508

Fluorescence in situ hybridization (FISH), 449

Fluoxetine, 444

Frank Laboratory of Neutron Physics of

JINR, 120

Free water surface system (FWS), 446

Gaseous pollutants, 492–494

Genes in calcium signaling, 349

Genetic phytofortication, 232

Genotoxicity, 386, 387

Gibberellins (GAs), 8

Gimnospermae, 106

Global aquaculture industry, 309

assessing water quality, 313–314

selecting plant species, 314–315

in Southeast Asia, 310–312

in Vietnam, 313

Glomalin, 5

Glycine max, 387

Glycophytes, 28

Gnaphalium suaveolens, 109

Gold, 473–477

Green biosynthesis, 184, 231

Green infrastructure

air purication, 498

biolter, 499

environmental protection, 497

High Line Park, 498

leadership, 498

legislations, 497

mechanisms, 497

parameters, 500

PM movement, 499

pollutants, 497

pollution, 500

Quercus ilex, 498

sedentary type, 498

vegetation, 499

Green revolution, 234

Greenhouse effect, 234

Greenhouse experiment, 475, 476

Eleusine indica, 172

Pityrogramma calomelanos, 171–173

Pteris vittata, 171–173

Vetiveria zizanioides, 172, 173

Greenhouse gases (GHGs), 441

Greenhouse study, 475

Groundnut (Arachis hypogaea L.), 433

Ha Thuong and Tan Long Mines, 173

Ha Thuong Field Experimental Site, 173

Halophytes, 28

Heat-shock proteins, 345

Heavy metal contamination

phytoavailability, 412–413

principal characteristics, 412

in soil, 410

soil metal groups, 411

sources, 410–411

toxicity, 411

in waste water, 416–418

Heavy metal(loid)s, 131, 441, 444, 449,

490–491

atmospheric deposition, 105

concentrations, 127

Heavy metal-contaminated soils, 158

potential plants for removal, 159–168

safety levels in soil, 158

Helianthus annus, 232, 382, 386, 416

Hill reaction activity (HRA), 185

Hordeum vulgare, 189, 232

Horizontal ow systems (HFS), 446

Horizontal subsurface ow (HSSF), 281

Humic and fulvic acids, 444

Hybrid oating treatment bed (HFTB), 297

Hydraulic control, 4, 409

Hydraulic retention time (HRT), 456

Hydrilla verticillata (Hydrilla), 270

Hydrocarbon-contaminated soil, 56–69, 82

Hydrocarbons in soil, 79–80

Hydrocarbons, chemical structure, 80

Hydrocotyle umbellata, 264

Hydrocotyle vulgaris, 302

Hydroxyethylenediaminetriacetic acid

(HEDTA), 472

Hydroxyethyliminodiacetic acid (HEIDA), 472

Hyperaccumulation

BF, 471

capacity, 471

characterization, 470

chelant-assisted phytoextraction, 472–473

denition, 471

metal availability, 470

metals, 470

obligate/facultative hyperaccumulators, 471

organic/silicate matrix, 470

plant tissue, 471

reclamation of metal-polluted sites, 470

TF, 471

and uptake, 470

Zn, 471

Hyperaccumulator, 406, 473

Hyper-accumulator plants, 406

essential nutrients, 344

gene expression, 356

nonbiodegradable chemical species, 344

Index

guarino@unisannio.it

509

root uptake system, 344

ROS species, 344

self-protection, 345

Hypericum perforatum, 211, 218

IAA, 8

IBR-2 reactor, 119

Ibuprofen, 444

Indigenous hyperaccumulator, 160, 179

Indole-3-acetic acid (IAA), 8

Indoleacetic acid (IAA), 34

Inorganic and organic fertilizer sources, 427–433

Instrumental neutron activation analysis

(INAA), 119

Integrated aquaculture–phytoremediation

systems (IAPS), 309, 312, 319, 320

Intercellular signaling, 347

Ipomoea aquatica, 288

Iron, 132, 133

Iron nanooxides, 372–374

Irrigation management, 426

Karwinskia humboldtiana, 229, 230

Karwinskia parvifolia, 229

Kosogorsky Metallurgical Works (KME),

116–118

Lactuca sativa, 380, 386, 389

Lago de Guadalupe, 294

Land farming, 54, 56, 63, 64, 69–71

Landoltia punctate, 387

Lead, 135–138

Leaf area index (LAI), 499

Lemnoideae (Duckweeds), 269

Lemnoideae minor, 270

Lentil (Lens culinaris), 433

Light expanded clay aggregate (LECA), 283, 285

Long-Range Transboundary Air Pollution

(LRTAP), 106

Low molecular weight organic acids

(LMWOAs), 472

Lycopersicon esculentum L., 350

Maghemite, 373

Magnetite, 373

Maize (Zea mays L.), 461

Maize production region (MPR), 186

Manganese, 130

Mannagrass (Glyceria maxima), 449

Matricaria recutita, 208, 211, 213

Maximum permissible level (MPL), 115, 130

Medicago sativa, 261

Medicinal plants, 207

cadmium and zinc compounds effects,

214–223

metal chelating agents on, 212–214

toxic metals and bioelements on, 207–212

woody trees as, 228–231

Membrane inlet mass spectrometry (MIMS), 455

Mercury, 197

Metal complexes on crops, 200–206

Metal pollution, 410–411

Metal toxicity, 411

Metal transporters, 352, 353

Metal uptake by plants, 415

Metal-binding genes, 356–357

Metallothioneins (MTs), 358–359

Metals absorbed, 418

Microbial inoculants-assisted

phytoremediation, 3

AMF, 5

atrazine, 12

bioaugmentation, 5

CEC, 12

characteristics, 12–13

degrading microbes, 11

heavy metals, 4

hydraulic control, 4

mechanisms, 4, 5, 8–11

metal cations, 12

natural/anthropogenic sources, 3

organic toxins, 4

painting plant leaves with microbial

suspension, 8

physical and chemical properties of

pollutants, 12

phytoextraction, 4

phytoltration, 4

phytoremediator, 8

phytostabilization, 4

phytovolatization, 4

and plant contributions, 7

plant root-colonizing microbes, 4

seedinoculation, 8

soaking plant roots, 8

soil management, 4

soil organisms, 3

soil pollution (see Soil pollution)

soil remediation, 4

soil washing, 4

symbiotic root colonizing microorganism,

4–5

Index

guarino@unisannio.it

510

Microorganisms isolated from hydrocarbon-

contaminated site, 60–61

Mine tailings, 475

Mining sector, 480

N, N-diethyltoluamide, 444

Nano Zero-Valent Iron (nZVI), 374, 375

Nanoagrochemicals, 233, 234

Nano-CuO stress, 384, 385

Nanomaghemite, 372–374

Nanomaterials, 379–389

consequences for plants

chlorophylls, 386

genotoxicity, 386–387

growth and biomass production,

387–389

nutrients uptake, 384–385

oxidative stress, 385–386

seed germination, 379, 380

translocation and accumulation,

382–383

uptake of NMs by roots, 380–382

water balance, 383–384

consequences for plants, 378–389

in environment, 377–378

Nanomaterials, adsorption of metals and

metalloids, 372–377

iron nanooxides, 372–374

manganese-based materials, 376–377

nZVI, 374–376

Nanoparticles for environmental remediation,

370–378

Nanoremediation, 371, 378, 390

Nano-sized metal oxides, 370

Nanotechnology, 370

Natural APCAs, 472

Natural Attenuation (NA), 62

NBIC technologies, 234

Nickel (Ni), 134–135, 476–479

Nicotiana tabacum, 265

Nitrilotriacetic acid (NTA), 415, 472

Non-hyper-accumulator, 344

Nutrient management strategies, 424

fertilization, 424

fertilizer management, 428–432

greenhouse gas (GHG) emissions, 433

merits and demerits, 426

mixed and rotational cropping systems, 433

organic wastes and manures, 427

water quality monitoring, 434

water use patterns, southern Africa, 425

Nutrient use efciency, 434

Ocean contamination, 308

Operational Street Pollution Model

(OSPM), 500

Oponent, 188

Organic carbon (OC), 494, 495

Organic manures, 426

Organic pollutants, 490

Organic toxins, 4

Osmotic stress, 38

Oxidative stress, 385

Particulate matter, 489

Pb/Zn mine, 411

Persistent organic pollutants (POPs), 106

Petroleum hydrocarbons (PHC), 21

pollutants, 6, 79, 80, 88

microbial degradation, 82

Petroleum spills into environment, 79

Petroleum-contaminated soil, 78

biological degradation, 82

bioremediation, 82–84

phytoremediation, 81–86, 88–91

rhizoremediation, 84–85

Petroselinum crispum, 386

PGPR-Enhanced Phytoremediation Systems

(PEPS)

development, proof and full-scale

application, 35

grass species, 34

green technology, 33

microbial communities, 34

microbial strains, 34

PGPR, 34

phytoremediation, 33 (see soil salinization)

Pharmaceutical and personal care products

(PPCPs), 278–280, 282, 283,

285, 289

Phaseolus limensis, 382

Phaseolus lunatus, 382

Phaseolus vulgaris, 191, 384

Photosynthetic electron transport (PET), 194

Phragmites australis, 65, 70, 267, 282, 288

Phytoavailability

bioavailability, 412

organic matter, 413

redox potential, 413

root zone, 413

soil pH, 412

texture, 412

Phytochelatins, 357, 358

Phytodegradation, 5, 88, 110

Index

guarino@unisannio.it

511

Phyto-DSS model, 330

Phytoextraction, 4, 25, 109–110, 159, 262,

408, 413, 470, 472–473

Phytoltration, 4

Phytofortication, 232, 233

Phytomining, 469, 470

genetic manipulation to enhance metal

hyperaccumulation, 481

gold, 473–477 (see Gold)

hyperaccumulation

(see Hyperaccumulation)

mine tailings, 473

mineralized or polluted soils, 473

mining sector, 480

nickel, 476–479

opportunity, 480

phytoextraction, 470

phytostabilization, 470

plant-based technology, 481

rhenium, 479–480

thallium, 479

Phytoremediation, 3, 65, 66, 81–82, 109,

163–168, 231, 232, 260, 268–271,

279–280, 295, 328–340

advantages, 29, 407

aquatic macrophytes, 267

Azolla caroliniana (Mosquito Fern), 268

Brassica juncea (Mustard Green), 269

Eichhornia crassipis

(Water Hycianth), 268

Hydrilla verticillata (Hydrilla), 270

Lemnoideae (Duckweeds), 269

Pistia stratiotes (Water Lettuce), 269

Ricciocarpus natans, 271

Schoenoplectus californicus (Giant

Bulrush), 271

Spirodela intermedia (Duckweed), 270

Vallisneria spiralis, 271

As, Cd, Pb, and Zn, 416

benets, 272–273

biological treatment, 445–446

biotic and abiotic plant stress factors, 32

BOD and COD removal, 456

cytoplasmic Na+, 30

denition, 406

description, 21

economic benets, 460–461

environmental issues, leachate, 441–443

excavation and soil removal, 32

future of, 273–274

halophytes, 29

heavy metal(loid) removal, 459–460

heavy metal-contaminated soil, 406

higher plants, 112–113

in situ treatments, 29

indigenous plants, 168–171

inorganic pollutants, 406

leachate recirculation, 445

limitations and challenges, 273, 407

living plants, 344

mechanism, 29, 260, 261

metal-accumulating properties, 407

metal-contaminated soils, 406

microbial inoculants-assisted

phytoremediation (see Microbial

inoculants-assisted

phytoremediation)

models, 328

application types, 338–340

detailed vs. robust, 328–329

plant-oriented models, 329–333

soil-oriented model, 333–336

uptake-process-oriented models,

336–338

municipal solid waste dumpsites, 440–441

nitrogen and phosphate removal, 456–459

overexpression, 360

PEPS, 43–44

petroleum-contaminated soil, 85–86,

88–91

PGPR, 30–31, 37

physico-chemical properties, 445, 448

phytoextraction, 262

phytostabilization, 263–264

phytotoxicity, 461–462

phytotransformation, 266

phytovolatilization, 265–266

plant roots, 21

plant salt stress and adaptation, 26–27

plant selection, 409–410

plant species, 449–453

plants used, 261

pollutants, landll leachate, 443–444

Reed (Phragmites sp.) plants, 453

revegetation, 32–33

rhizoltration, 264

rhizosphere microbiology, 448–449

soil horizons, 32

treating and improving soil, 308

types, 21, 407–409

vegetation lter, 454–456

waste management strategies and adverse

effects, 443

Phytostabilization, 4, 111, 159, 263, 264, 408,

413–414, 470

Phytostimulation, 408

Phytotransformation, 110, 266, 408

Phytovolatilization, 110, 265, 266, 415

Index

guarino@unisannio.it

512

Picea abies, 117, 121, 123

Pilot Field, 62–63

Piptatherum miliaceum (Smilo grass), 416

Pistia stratiotes (Water Lettuce), 269, 417

Pisum sativum (Pea), 200, 218, 380

Pityrogramma calomelanos, 168, 171–174, 179

Plant cell membranes, 39

Plant growth-promoting bacteria (PGPB), 78

Plant growth-promoting rhizobacteria (PGPR),

27, 30, 87–91, 94

Plant macro- and micro-nutrients, 455

Plant–metal interaction, 184

Plants, 184, 207–223

for heavy metal removal from soils,

159–168

medicinal, 207

cadmium and zinc compounds effects,

214–223

metal chelating agents on, 212–214

toxic metals and bioelements on,

207–212

woody, bioelements and toxic metals,

225–228

PLANTX model, 331

Poaceae, 34

Pollutant area, 56–59

Polyamines, 25

Polycyclic aromatic hydrocarbons (PAHs), 35,

488, 490

Polymetallic contamination, 126, 127, 130

Pontederia cordata, 302

Pontederia sagittata, 300

Populus nigra, 65, 70

Potamogeton pectinatus, 267

Potassium cyanide, 475

Potato production region (PPR), 186

Potential Bioavailable Sequential Extraction

(PBASE), 415

Proline-biosynthesis genes, 384

Pseudomonas, 35, 82, 92, 93

Pteris vitatta, 163, 166–168, 171–174, 177,

179, 262, 416

Pulse amplitude-modulated (PAM), 38

Pure plant oils (PPO), 186

Pyridine-2-6-dicarboxylic acid (PDA), 415

QUANT-2A, 119

Rapeseed production region (RPR), 186

Raphanus sativus, 380, 474

Reactive oxygen species (ROS), 24, 385

Reed (Phragmites australis), 449

Remediation. See Phytoremediation

Rhamnolipids, 8

Rhenium, 479

Rhizobium–legume symbiosis system, 385

Rhizodegradation, 110

Rhizoltration, 111, 114, 264, 408, 414–415

Rhizoremediation, 21, 84, 85

Rhizosphere bacteria (RB), 78, 88, 90,

94, 455

Rhizosphere microbiology, 448, 449

Ricciocarpus natans, 271

Root zone, 413

ROS signaling

antioxidative system, 350

APX genes, 351

Ca2+ signals, 348

Cu/Zn-SOD, 351

environmental stresses, 352

genes, 350

genetic control, 350–352

metal transport, 357

redox-sensitive proteins, 349

transduction pathways, 351

Rush (Juncus effuse L.), 449

Salix amygdalina L., 454

Salix cinerea, 453

Salix fragilis, 117, 121, 125

Salix viminalis, 411

Salt marsh plants, 281

Salt overly sensitive (SOS), 27

Salvia ofcinalis, 218

Sample preparation, 119–120

Sampling, 119–120

Scenedesmus quadricauda cells, 206

Schoenoplectus californicus

(Giant Bulrush), 271

Schoenoplectus tabernaemontani, 297

Scirpus validus, 282

Scrophulariaceae, 261

Sebetaria, 109

Seedinoculation, 8

Selenium, 214

Setaria italica, 189

Short rotation coppice (SRC), 453

Signaling networks, 346

Signaling strategy in plant

calcium signaling, 348–349

extracellular signaling, 347

genes incalcium signaling, 349

Index

guarino@unisannio.it

513

intercellular signaling, 347

signaling networks, 346

transduction, 346

Signals transduction in plants, 346

Soaking plant roots, 8

Sodium adsorption ratio (SAR), 23

Sodium cyanide, 475

Soil and air pollution, 121, 144, 145

Soil contamination by metals, 370

Soil excavation, 158

Soil management, 3

Soil metal groups, 411

Soil organisms

advantage of degrading ability, 4

degrading ability, 5

elimination, 3

plant establishment and survival, 3

Soil pollution

abiotic factors, 12

biological, 3

sources, 6–7

stress condition, 11

types, 8–10, 12

Soil remediation, 4

Soil salinization, 22, 36, 38–43

eld trials

NaCl, 41–43

salinity and PGPR, plant growth, 41

Lab/Greenhouse experiments

cell membrane integrity, 39–41

photosynthesis, salt inhibition, 38–39

salinity and PGPR, plant growth, 36

phytoremediation, 43–44

Soil salts

applications, 22

classication, 23

clay particles, 23

crop irrigation, 22

environmental impacts, 21

evapotranspiration, 22

global agricultural and economic

losses, 21

Na+, K+ and Cl−, 24

NaCl, 23

oil extraction, 22

plant nutrients and micronutrients, 23

salt stress and acclimation signaling

pathways, 25–28

salt stress and ROS damage, 24

Soil washing, 4

Sorbus aucuparia, 119

Soya bean (Glycine max L.), 433

Spartina maritima, 279, 283–285

Spinacia oleracea, 380

Spirodela intermedia (Duckweed), 270

Statistical analysis

Ca-Mg, 141

categories, 140

multivariate statistical analysis,

140, 143

Subsurface ow system (SSFS), 446

Sunower, 418

Symbiotic root colonizing microorganism, 5

Symphoricarpos albus, 128

Synthetic aminopolycarboxylic acids

(APCAs), 472

Tan Long Field Experimental Site, 175

Terminal restriction fragment length

polymorphism (T-RFLP), 89

Thalia dealbata, 302

Thallium, 479

Thiobarbituric acid reactive substances

(TBARS), 193

Thiourea (CH4N2S), 473

Thlaspi caerulescens, 262, 355, 416

Thlaspi praecox, 416, 419

Tilia cordata, 117, 120, 125, 498

Tolerant, 344, 345, 350

Tomato chloroplastic MDHAR, 351

Total Kjeldahl nitrogen (TKN), 458

Total petroleum hydrocarbon (TPH), 55, 78,

79, 82, 83, 85, 86, 89, 90

degradation, 84

microbial degradation, 83–84

Toxic metals

application, crops, 188–191,

193–200

and bioelements on medicinal plants,

207–212

on woody plants, 225–228

Transfer factor, 138, 140

Translocation factor, 188, 354–355, 471

Transporters, 352–353

Trees, 225–231

Triangle of U theory, 185

Trietaris europea, 109

Trifolium repens, 265

Triticum aestivum, 200, 216, 232,

265, 381

Tula region, 118

Tulachermet, 118

Typha angustifolia, 282

Typha latifolia, 267, 282, 288

Typha orientalis, 302

Typhalatifolia, 382, 383

Index

guarino@unisannio.it

514

USDA Natural Resources Conservation

Service, 23

Vacuoles, 355

Vallisneria spiralis, 271

Vanadium, 118

Vegetative lter strips (VFS), 316, 317

Verona, 188

Vertical ow systems, 446

Vertical subsurface ow (VSSF), 281

Vetiver (Chrysopogon sp.), 460

Vetiveria zizanioides, 172, 173, 175–178, 411

Vicia faba, 189

Vigna radiata, 382, 386

Vigna unguiculata, 189

Volatile organic carbons, 408

Volatile organic compounds, 441, 488

Wastewater treatment plants (WWTPs), 278,

280, 281, 286, 289

Water hyacinth, 417, 418

Water pollution, 294

Water quality in Eutrophic Lakes,

297–300

Water quality index (WQI), 300

Water use efciency, 427, 433

Wetlands, 446, 449, 453, 454, 460, 462

Willow (Salix sp.), 453

Woody plants, 126, 127, 130

bioelements and toxic metals,

225–228

Woody trees, medicinal plants, 228–231

Xenobiotic organic compounds (XOCs),

443–444

Yellow ag (Iris pseudacorus L.), 449

Zea mays, 416

seedlings, 380

Zinc (Zn), 133–134

Index

guarino@unisannio.it

ResearchGate has not been able to resolve any citations for this publication.

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