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Geomicrobiology Journal
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Soil Fungi for Bioremediation of Pesticide
Toxicants: A Perspective
D. Mohapatra, S. K. Rath & P. K. Mohapatra
To cite this article: D. Mohapatra, S. K. Rath & P. K. Mohapatra (2021): Soil Fungi for
Bioremediation of Pesticide Toxicants: A Perspective, Geomicrobiology Journal, DOI:
10.1080/01490451.2021.2019855
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Soil Fungi for Bioremediation of Pesticide Toxicants: A Perspective
D. Mohapatra
a
, S. K. Rath
b
, and P. K. Mohapatra
a
a
Department of Botany, School of Life Sciences, Ravenshaw University, Cuttack, India;
b
Department of Life Sciences, Ramadevi Women’s
University, Bhubaneswar, India
ABSTRACT
Soil is the basis of all agroecosystem and its health is of utmost importance for the better prod-
uctivity and sustainability of agriculture but soil health is constantly deteriorating due to the add-
ition of xenobiotic compounds by various agronomicals and industrial applications. Pesticides are
widely used throughout the world for controlling the spread of various pests in agroecosystem
but the persistent nature and non-targeted toxicity of the compounds have also become the
major concern for agroecosystem and is directly hampering the yield of agricultural produce.
Hence, removal of these substances is of utmost importance and a variety of approaches are in
progress. Currently, the use of biological resources for the removal and degradation of these sub-
stances has emerged as a powerful tool. Numerous bacterial and fungal species with degradation
ability have been tried and established but very few attempts are made to make a comprehensive
evaluation of the potential of these agents. This review thrusts on the bioremediation efficiency of
soil fungi with an aim to make a comparative analysis and to critically evaluate their potential
application in the field. The species of Aspergillus,Allescheriella,Alternaria,Microsporum,Penicillium,
Phlebia,Paecilomyces,Trichoderma, etc. are known for their bioremediation potential. But
Aspergillus species are the most widely used for degrading almost all types of pesticides.
ARTICLE HISTORY
Received 30 July 2021
Accepted 14 December 2021
KEYWORDS
Soil fungi; bioremediation;
pesticides; agroecosystem;
soil health
Introduction
One of the biggest trademarks of the modern era is the syn-
thesis and application of hazardous organic compounds into
the natural habitat. These compounds include polychlorina-
tedbiphenyls, pesticides, chlorophenols, fuels, polycyclic aro-
matic hydrocarbons, and dyes (Diez 2010). Some of these
synthetic chemicals are very much resilient toward biodeg-
radation by natural flora as compared to the naturally syn-
thesized organic compounds (Chikere 2013; Mohapatra et al.
2018). Thus, the use of these chemicals and the residues
from their application have now become one of the major
problems throughout the world (Bhatt et al. 2021;
Mahmood et al. 2016; Mohapatra et al. 2018,2021; Sabarwal
et al. 2018). The amount of major pesticides (Table 1) used
in Indian agriculture gives an idea about the scenario of the
intensity of use of pesticides worldwide. The most seriously
affected sector by such activity is the soil of the agroecosys-
tem. Soil constitutes a complex interface between biotic and
abiotic elements that include minerals, water, gasses, organ-
isms, etc. For many centuries, healthy soils nurtured plants,
animals, microbes as well as humans for their growth and
survival. However, widespread development coupled with
excessive anthropogenic interference has gradually made this
soil unhealthy, and unfit for adequate plant nutrition in the
last few decades. Intensive use of fertilizers and agrochemi-
cals adds excessive heavy metals and toxic pollutants, which
are the main barriers in building up and maintaining the
life-support system in the soil. Heavy metals, pesticides, and
chemical contaminants in soil and water are the main con-
cern as they cannot be easily degraded into non-toxic forms
and have long-term environmental effects, which eventually
become part of the food chain. The unpleasant effects of
chronic exposure to these toxic and more or less persistent
chemicals are a series of life-threatening diseases like cancer,
renal failure, mental disorder, and paralysis. To mitigate
such environmental and non-target toxicity, several proc-
esses are being attempted to determine the ways to prevent
continuous soil health degradation by pesticides and to
improve soil health for maintaining sustainable productivity.
While many physio-chemical methods like ion exchange,
oxidation-reduction, filtration, chemical precipitation, elec-
trochemical treatments, and dialysis have been applied for
the removal of pesticides from the soil, none of these
approaches is hitherto considered fully effective in giving a
desirable outcome. Furthermore, higher costs, limited appli-
cation and low effectiveness to improve inherent soil health
make them almost abandoned or limited to a pilot-scale
exercise. Amid these uncertainties, bioremediation has
emerged as one of the most capable alternatives, which are
not only cost-effective but also an eco-friendly way of
removing these pollutants from soil. This is nothing but the
fine-tuning of the processes that nature is applying for mil-
lions of years to get rid of harmful chemicals from the
environment. The process of bioremediation involves the
use of available natural bio-resources, which include mainly
CONTACT P. K. Mohapatra pradiptamoha@yahoo.com Department of Botany, School of Life Sciences, Ravenshaw University, Cuttack 753003, India
ß2021 Informa UK Limited, trading as Taylor & Francis Group
GEOMICROBIOLOGY JOURNAL
https://doi.org/10.1080/01490451.2021.2019855
microorganisms, plants, bioamendments, etc., to reduce con-
taminants to undetectable, non-toxic, or at least to very low
and acceptable levels where they are no more threats to the
ecosystem. The environmental problems related to insecti-
cides are being approached by the use of native as well as
engineered organisms for safe, convenient, and economically
feasible technology (Bhatt et al. 2021; Kaur and
Balomajumder 2020a,2020b; Kumar et al. 2011; Mohapatra
et al. 2018). Consequently, quite a lot of biodegradation
methods for removal of organic compounds have been
developed and tried in recent years with an aim to remove a
specific or a broad range of chemicals from soil (Arbeli and
Fuentes 2007; Kumar et al. 2011; Mishra et al. 2020;
Mohapatra et al. 2003, Mohapatra et al. 2018; Prasad 2017;
Purohit et al. 2018).
Fungi are ubiquitous eukaryotic microorganisms, repre-
senting a diverse group of the organism from different envi-
ronments. These organisms are well-suited to almost all
habitats, from agriculture to the forest ecosystem, from mar-
ine to arctic. Fungal natural behaviors to work in the com-
munity with other organisms in the ecosystems helps them
to sustain in a wide range of ecosystems with other organ-
isms. They always prefer to syntrophically work with other
living beings and play an important role in most ecosystems
and are capable of regulating the flow of nutrients and
energy. Fungal metabolism breaks down advanced structures
Table 1. The major pesticides used in the Indian agricultural sector (compiled from http://ppqs.gov.in/statistical-database).
Sl. No Pesticide Category Consumption (2019–20) (metric tons) Consumption (2020–21) (metric tons)
1 Acephate I 405.91 356.90
2 Acetamiprid I 114.42 100.85
3 Buprofezin I 126.00 79.00
4 Carbaryl I 181.00
5 Carbofuran I 206.76 214.75
6 Cartap Hydrochloride I 358.19 374.41
7 Chlorantraniliprole I 105.00 135.34
8 Chlorpyriphos I 1430.62 1036.69
9 Cypermethrin I 674.65 343.91
10 Dimethoate I 367.51 209.59
11 Emamection Benzoate I 97.00 124.73
12 Fenvalerate I 667.33 149.73
13 Fipronil I 444.05 256.83
14 Imidachloprid I 371.99 317.17
15 Indoxacarb I 72.00 112.00
16 Lamda-cyhalothrin I 163.06 161.09
17 Malathion I 647.14 305.41
18 Methyl Parathion I 115.40
19 Monocrotophos I 551.02 351.91
20 Profenophos I 425.00 433.40
21 Quinalphos I 564.61 412.60
22 Thiamethoxam I 234.62 214.54
23 Triazophos I 156.00 48.00
24 Captan F 123.93 434.51
25 Carbendazim F 687.08 541.37
26 Carbendazim þMancozeb F 10.97 296.69
27 Copper Oxychloride F 298.84 316.32
28 Dodine F 56.26 166.79
29 Hexaconazole F 153.68 112.41
30 Mancozeb F 2181.46 1877.43
31 Propiconazole F 131.29 112.46
32 Propineb F 197.49 507.75
33 Sulphur F 3878.29 4245.13
34 Thiophanate-Methyl F 177.57 201.19
35 Thiram F 101.76 111.10
36 Zineb F 135.32 124.57
37 Ziram F 117.60 475.54
38 2,4-D Amine Salt H 1066.81 926.80
39 Atrazine H 346.26 287.82
40 Butachlor H 354.10 209.17
41 Glyphosate H 571.49 505.19
42 Isoproturon H 291.90 14.00
43 Metribuzin H 110.70 68.20
44 Pendimethalin H 198.94 149.00
45 Pretilachlor H 621.37 666.83
46 Aluminium Phosphide R 91.85 103.57
47 Zinc Phosphide R 123.27 89.25
48 Azadirachin BP 134.51 117.10
49 Bacillus thuringiensis BP 82.99 108.50
50 Beauveria bassiana BP 180.75 174.26
51 Neem based insecticides BP 184.91 565.90
52 NPV (H) BP 351.90 144.10
53 Pseudomonas fluorescens BP 401.22 592.68
54 Tricoderma spp. BP 215.28 969.00
55 Tricoderma viride BP 582.69 660.99
I: insecticide; F: fungicide; H: herbicide; R: rodenticide; BP: biopesticide.
2 D. MOHAPATRA ET AL.
of plant cells and complex cellular molecules like cellulose,
hemicelluloses, and lignin as well as very complex carbon-
based synthetic compounds regulating the C:N ratio and
contributing to soil nutrient recycling (Kumar et al. 2011;
Mishra et al. 2020; Prasad 2017; Purohit et al. 2018; Sharma
et al. 2012; Vidhya Lakshmi et al. 2008). Notwithstanding
certain species are adversely affected by the initial toxicity of
pesticides, pollutants, and heavy metals (Pointing 2001), but
in the process prolonged exposure, develop tolerance
(Jacobsen and Hjelmsø 2014; Rangasamy et al. 2017). Such
species may thus form an integral part of the remedial pro-
cess thereby becoming the most important bioremediation
process of soils (Adelowo et al. 2014; Ahmad 2020; Gadd
et al. 2014). The carbon, nitrogen, or energy from pesticide
molecules and other toxics substances, generated by their
degradation, are utilized by fungi for their growth and pro-
liferation. The mycelial networks act as a continuum in the
agroecosystems and accelerate the soil rehabilitation process.
They are also considered natural ecosystem engineers and
can be used to create environment friendly soils by the
mycoremediation process to reduce long-chain molecules,
disassemble toxins into simpler, less toxic substances
(Prasad 2017; Purohit et al. 2018). Different pesticides have
different modes of action (Table 2) and chemical compos-
ition when we consider their non-target toxicity and degrad-
ation potential. Quite a good number of reviews have been
published discussing the role of fungi in the degradation of
pesticides of diverse chemical structures. However, there is
no comprehensive report on the degradation potential of
soil fungi and their possible utilization in managing the soil
pesticide residues. This review presents the potential of the
soil fungal consortia to metabolically degrade pesticides that
are continuously being accumulated in the agroecosystems
due to commercial agropractices.
Bioremediation and environmental cleanup
Bioremediation is a natural process for the decontamination
of soil and groundwater using living organisms (Akhtar and
Mannan 2020; Raffa and Chiampo 2021; Robichaud et al.
2019; Wu et al. 2016). It involves waste management techni-
ques to effectively and sustainably remove pollutants from
the environment (Akhtar and Mannan 2020; Bhatt et al.
2021; Kumar et al. 2011). Different living organisms, cyano-
bacteria, algae, land plants, fungi, and bacteria also known
as bioremediators and can remove or detoxify these hazard-
ous wastes (Purohit et al. 2018). Bioremediation is aimed
primarily at reducing or cleaning soil from contamination
and improving soil health through environmental-friendly
approaches. Paul Stamets has coined ‘mycoremediation’and
refers to the use of one or more species of fungi to detoxify
the contaminated site (Purohit et al. 2018). It is a process by
using fungi to sequester/metabolically degrade toxic contam-
inants from soil and water in situ and ex situ (Akhtar and
Mannan 2020; Raffa and Chiampo 2021; Strong and Burgess
2008). Mycoremediation can be an economical, environmen-
tally friendly, and effective strategy to combat the growing
soil and water pollution problems. Fungi are the perfect
candidate to remove various pollutants in soil because of
their ability, such as robust fungal growth, vast hyphal net-
work, the production of versatile extracellular ligninolytic
enzymes, high surface area to volume ratio, resistance to
heavy metals, adaptability to withstand fluctuation of pH
and temperature and presence of metal-binding proteins
(Akhtar and Mannan 2020). The use of native and/or alien
microflora enhances soil health, even though there is a
potential danger of invasion by alien species.
The decomposition of organic molecules is influenced by
saprophytic fungi. These fungi secrete various extracellular
enzymes and acids that break down natural polymers, such
as chitin, keratin, pectin, lignin, etc. (Lamar and White
2001; Purohit et al. 2018). In addition, certain fungi also
produce little biocidal compounds or biostatic compounds
(e.g., trimethylamines produced by Geotrichum candidum)
involved with the bioremediation of complex organic mole-
cules, such as hydrocarbons and heavy metal (Hung et al.
2015; Khan et al. 2019). Others have meanwhile taken a hol-
istic approach to the use of plant, fungi, and bacterial con-
sortia. Many of these applications have shown encouraging
results in various toxins, pollutants, and dangerous chemi-
cals degradation (David et al. 2018; Singh 2006). Overall, the
combined inclusion of these approaches can be complemen-
tary and can maximize bioremediation prospects and effi-
ciency. Bioremediation partnering fungi opens the door to
speed up the process of soil cure. They can mediate greatly
by taking part in early degradation events and supporting
the remedial work of other organisms. Many fungi supply
energy by enzymatic contaminant degradation to stimulate
aggressive toxin degradation by other soil microbes (Prasad
2017; Purohit et al. 2018).
Soil fungi
Soil fungi encompass heterogeneous groups of organisms,
which mainly belong to Chytridiomycota,Zygomycota, and
Ascomycota. These phyla include mostly different species
belonging to genera Aspergillus,Allescheriella,Alternaria,
Acremonium,Beauveria,Cunninghamella,Cladosporium,
Engyodontium,Fusarium,Geomyces,Mortierella,
Microsporum,Penicillium,Phlebia,Paecilomyces,Rhizopus,
Stachybotrys, and Trichoderma (Pinedo-Rivilla et al. 2009;
Purohit et al. 2018; Tigini et al. 2009). They play a key role
in the decomposition of organic matter and the cycling of
carbon and nitrogen in soils. They are primarily saprophytes
and possess excellent ability to decompose complex organic
residues like cellulose, hemicellulose, and lignified tissues.
The fungal enzymes and pathways are also efficient to
degrade a variety of xenobiotic compounds. Extracellular
enzymes, such as monooxygenase, secreted by non-lignino-
lytic fungi lead to hydroxylation of polycyclic hydrocarbons
to non-toxic metabolites. They also show their potential as
soil bioremediators (but in the later stages of decomposition,
partially degraded recalcitrant polymers are often prevalent
(Aylward et al. 2013). The ecological succession in the
decomposition sequence of these fungi is linked to their spe-
cialized capacity to degrade complex polymers, such as
GEOMICROBIOLOGY JOURNAL 3
Table 2. Comprehensive information and mode of action of the pesticides discussed in this review (WHO 2005).
Pesticides Hazardous Class Structure Mode of action
Chlorfenvinphos Highly hazardous Ib Acetylcholinesterase (AChE) inhibitor.
Chlorpyrifos Moderately hazardous II Acetylcholinesterase (AChE) inhibitor.
Dimethoate Moderately hazardous II Acetylcholinesterase (AChE) inhibitor.
Lancer Slightly hazardous III Acetylcholinesterase (AChE) inhibitor.
Malathion Slightly hazardous III Acetylcholinesterase (AChE) inhibitor.
Methyl parathion Extremely hazardous Ia Acetylcholinesterase (AChE) inhibitor.
Monocrotophos Highly hazardous Ib Inhibition of ChE activities
Pirimiphos-methyl Slightly hazardous III Acetylcholinesterase (AChE) inhibitor.
Profenofos Moderately hazardous II Acetylcholinesterase (AChE) inhibitor.
Pyrazophos Moderately hazardous II Phospholipid biosynthesis inhibitor.
Dichlorovos Highly hazardous Ib Acetylcholinesterase (AChE) inhibitor.
Formothion Obsolete as pesticide O Acetylcholinesterase (AChE) inhibitor.
Fenitrothion Moderately hazardous II Acetylcholinesterase (AChE) inhibitor.
(continued)
4 D. MOHAPATRA ET AL.
Table 2. Continued.
Pesticides Hazardous Class Structure Mode of action
Endosulfan Moderately hazardous II Affinity for GABA (gamma- aminobutyric
acid) receptors in the brain and acts
as a noncompetitive GABA antagonist.
DDT Moderately hazardous II Affects the nervous system by interfering
with normal nerve impulses
Lindane Moderately hazardous II Affects the nervous system by interfering
with normal nerve impulses
Aldrin Obsolete as pesticide O Block gamma-aminobutyric acid
(GABA) activity.
Dieldrin Obsolete as pesticide O Act upon neurons in a global manner.
Chlordane Moderately hazardous II Affinity for GABA (gamma- aminobutyric
acid) receptors in the brain and acts
as a noncompetitive GABA
Heptachlor Obsolete as pesticide O Affinity for GABA (gamma- aminobutyric
acid) receptors in the brain and acts
as a noncompetitive GABA
a-HCH Moderately hazardous II Central nervous system stimulants
causing violent epileptiform
convulsions.
Pentachlorophenol Highly hazardous Ib Binds to mitochondrial protein and
inhibits mitochondrial ATP-ase activity.
b-cypermethrin Highly hazardous Ib Delay closure of the sodium channel,
resulting in a sodium tail
b-cyfluthrin Moderately hazardous II Interference with nerve signaling by
inhibition of the membrane sodium
channel systems in the
target organism.
3-phenoxybenzoic acid Interference with nerve signaling by
inhibition of the membrane sodium
channel systems in the
target organism.
Cyhalothrin Moderately hazardous II Interference with nerve signaling by
inhibition of the membrane sodium
channel systems in the
target organism.
(continued)
GEOMICROBIOLOGY JOURNAL 5
Table 2. Continued.
Pesticides Hazardous Class Structure Mode of action
Fenvalerate Moderately hazardous II Interference with nerve signaling by
inhibition of the membrane sodium
channel systems in the
target organism.
Allethrin Slightly hazardous III Interference with nerve signaling by
inhibition of the membrane sodium
channel systems in the
target organism.
Metolachlor Slightly hazardous III Inhibition of cell division and elongation
in plants due to interference with a
number of enzymes.
Alachlor Slightly hazardous III Elongase inhibition, and inhibition of
geranylgeranyl pyrophosphate (GGPP)
cyclization enzymes, part of the
gibberellin pathway.
Atrazine Unlikely to present acute hazard in normal use U Blocks electron transport on the
reducing side of PS II from QA to QB
which causes an increase in variable
chlorophyll fluorescence
Pyrazosulfuron ethyl Unlikely to present acute hazard in normal use U Inhibition of acetolactate synthase (ALS),
which is a key enzyme in the
biosynthesis of the branched-chain
amino acids, such as valine, leucine,
and isoleucine.
Sulfosulphuron Slightly hazardous III Inhibition of acetolactate synthase (ALS),
which is a key enzyme in the
biosynthesis of the branched-chain
amino acids, such as valine, leucine,
and isoleucine.
Chlorimuron-ethyl Unlikely to present acute hazard in normal use U Inhibiting biosynthesis of the essential
amino acids valine and isoleucine
Chlorsulfuron Unlikely to present acute hazard in normal use U Inhibition of cell division in the growing
tips of roots and shoots
Diuron Unlikely to present acute hazard in normal use U Inhibits photosynthesis by preventing
oxygen production and blocking the
electron transfer of photosystem II of
photosynthetic organisms.
Isoproturon Slightly hazardous III Inhibits photosynthesis by preventing
oxygen production and blocking the
electron transfer of photosystem II of
photosynthetic organisms.
Metsulfuron methyl Unlikely to present acute hazard in normal use U Inhibiting cell division in the shoots and
roots of the plant,
Glyphosate Unlikely to present acute hazard in normal use U Inhibition of the enzyme 5-
enolpyruvylshikimate-3-phosphate
(EPSP) synthase. The resulting
deficiency in EPSP production leads to
reductions in aromatic amino acids
that are vital for protein synthesis and
plant growth.
6 D. MOHAPATRA ET AL.
lignin and keratin, which most other fungi cannot use. As
such, their consortia with different species provide increased
effectiveness in the soil bioremediation process, often con-
sidered as very good xenobiotic degrading fungi.
Fungal diversity
The fungi are a huge diversity of organisms, ranging from
microscopic unicells to large macro fungi, as illustrated by
the famous mushrooms and toadstools, and by the large
fruit body, the giant puffball. Recent estimates show that
fungi include 72,065 species spread over 11 phyla under
7745 genera (Hawksworth et al. 1995). Out of this huge
group of organisms the member belonging to the family of
filamentous fungi like Aspergillus,Curvularia,Acrimonium,
etc., white-rot fungi like Pleurotus ostreatus,Trametes versi-
color, etc., marine fungus, extremophilic fungi, symbiotic
plant fungi like A. nidulans,T. viride, etc. are all capable of
degrading a wide range of pesticides of diverse chemical
structures and can be potential bioremediating agents
(Deshmukh et al. 2016). The dominating group in the mixed
population during remediation are Aspergillus,Penicillium,
Trichoderma (Rodrigues et al. 2020). These fungi can
degrade a wide range of chemicals, so are almost found in
all types of remediation processes. Moreover, this group is
present in both soil and marine biome making them avail-
able to all types of chemicals, though the degradation rate
depends on the availability of chemicals and exposure to the
particular chemical.
Fungi are considered to be the least explored group of
organisms to date thus exhibit their actual diversity in
nature (Blackwell 2011). The species known to us are only
those which have been characterized and isolated so far.
Therefore, it cannot be told with certainty the exact size of
the fungal biodiversity. It is estimated that more than 1mil-
lion species are as yet to be discovered, and more than
500,000 species may be only associated with insects (Driver
and Milner 1998). It is observed that the majority of the
fungi are most likely to spend some part of their life cycle
in the soil ecosystem and are directly or indirectly associated
with the soil biome thereby playing a complex role in the
soil and acting as fundamentals for soil biomes
(Hawksworth et al. 1995). The number and range of species
present in soil can, however, be hardly assessed in absolute
terms as the techniques available for soil isolation and fun-
gal detection are hitherto limited. Nevertheless, it is evident
that relatively few species have been isolated or reported
from the soil. There appears to have been no critical evalu-
ation of the number of species so far isolated from the soil,
although only those that grew after soil dilution and
excluded terrestrial macrofungi and plant pathogens. It is
not easy to detect exactly which fungi are in a soil sample
since the overwhelming nature of the vast majority of spe-
cies is one of the biggest problems (Blackwell 2011). This is
a popular phenomenon and estimates show that only 17%
of the known fungi can be cultivated easily in culture
(Blackwell 2011).
The best global predictors for soil fungal richness and
community composition are controlled by climatic factors,
followed by edaphic and spatial patterning. The richness of
all fungal and functional groups, apart from ectomycorrhizal
symbionts, are causally not related to plant diversity; this
suggests that feedback on the plant and soil has no impact
on global soil fungi diversity. Fungi follow similar patterns
of biogeography as plants and animals except for several
large taxonomic or functional groups, which run against
overall patterns.
The functional attributes of soil fungi
There can be considerable mycelium in a soil sample associ-
ated with a single fungus. In mycorrhizal fungi, a single net-
work can spread over many plants in the metarhizospheric
zone (Binkley and Fisher 2019). The traditional ‘fairy ring’is
a ring of fruit bodies that mark the periphery of a single
mycelial organism in the case of free-living soil fungi. The
current study suggests that these fungi can even be spread
over 100 hectares in soil and one such example is 880 hector
of soil fungi mycelia network found in Malheur National
Forest in Oregon (Barnard 2000). This wide distribution of
these organisms in soil is both due to spore formation and
mycelial growth, which varies from species to species.
Christensen (1989) described somewhat twenty functions of
fungi in soil but the main function is that they act as pri-
mary degraders of complex organic matters in soil. Many
soil fungi have other roles and interactions; one of them
being the mycorrhizae, which have been most widely
studied. The contribution of mycorrhizal association varies
and includes direct plant feeding, plant germination support,
and the prevention of pathogens by niche exclusion from
invading pathogenic microbes (Binkley and Fisher 2019).
Mycorrhizal associations range from free-living plant-fungal
associations to apparently intracellular habits of endomycor-
rhizal fungi that exist only within the host plant tissue.
There is also great variation in the specificity of mycorrhizal
associations, which have spatial and temporal variations as
well as that with the host plants. Although many mycor-
rhizal fungi can be associated with a large number of differ-
ent host plants showing a broad host range, others, like
Russula species, are very much specific to their host.
Moreover, one host can provide support in one rhizosphere
for several different mycorrhizal fungi too (Pinton et al.
2007). In many other interactions and associations, soil
fungi interact with plants, arthropods, nematodes or may
also act as fungal pathogens. In contrast, arthropods and
other invertebrates may be used as a food source for some
fungal species. However, fungi should be considered as
active organisms or as sleeping propaguli in the soil. The
latter may be associated with non-soil environments with
primordial life cycles, the soil being an intermediate reser-
voir. For example, the entomophthoralean fungi are patho-
gens of different arthropods, in particular, Diptera and
Homoptera (Keller 1991). On arthropods entomophthoral
spores germinate, and mycelial growth helps to invade the
host. Infected arthropods tend to move to higher plant
GEOMICROBIOLOGY JOURNAL 7
surfaces where the fungus binds the dead arthropod to the
plant through a haustoria process. Spores are then released,
most of them falling down to the ground and remain asleep
until a new host is available. Therefore, soil fungi can be
summarized as a vast range of organisms that can grow
actively freely, be closely associated or dormant with
other organisms.
Soil fungi as regulatory determinants
The biodiversity and interaction between soil organisms and
plants are also essential in agriculture. However, the quality
and extent of the soil activities in soil are determined by
their density, diversity, and interaction. Soil fungal diversity
and the range of metabolic activities function as the regula-
tory determinant of soil productivity and nutrient status.
Agrochemicals subsidies and use interfere with these proc-
esses of the ground and frequently determine the microbial
density, diversity, and extent of interactions. Many of the
soil microbes are not yet explored, and their role in the soil
agroecosystem is yet to be properly understood. The activity
of fungi in the decomposition of the wood is greater and
tends to decrease when bacteria become more dominant in
later decomposition phases. Fungi use a variety of extracellu-
lar enzymes (EEA), including protease, cellulase, b-glucosi-
dase, and chitinase to obtain food from both simple and
complex molecules (Cullings et al. 2008;
Zif
c
akov
a et al.
2011). These EEAs are responsible for degrading cellulose,
plant protein, starch, hemicellulose, and animal compounds
like chitin. Phosphatases, extracellularly synthesized by
plants and microorganisms, play a key role in the phos-
phorus cycle in which plants and microorganisms can
develop inorganic phosphorus, the sole form of phosphate
(Caldwell 2005; Wang et al. 2014). The ligninolytic enzymes
involved in lignin degradation, along with various xenobiotic
aromatic compounds, Laccase and polyphenol oxidase are
vital for mobilizing nutrients from recalcitrants (Burke and
Cairney 2002; Chan-Cheng et al. 2020; Pizzul et al. 2009;
Zhang et al. 2020). Different EEAs may be used to indicate
biological balance, fertility, quality, and the condition of the
soil (Srinivasulu and Rangaswamy 2013) and are involved in
the transformation of nutrients of the soil, energy metabol-
ism, and various compounds (Ekenler and Tabatabai 2003).
Inter-species differences in EEA activities related to C, N,
and P-metabolism could be caused by soil fungal differences
affecting C, N, and P-cycling (Diagne et al. 2020; El-
Sherbeny et al. in press; Li et al. 2013a; Muneer et al. 2021).
The interaction between soil fungal enzymes may usually
accelerate the decomposition of organic matter in the soil,
and nutrients like N and P may be released. This soil
organic matter breakdown in soil colloids can have a direct
effect on organic matter absorption (Li et al. 2013b). So, soil
fungus and the enzymes produced by them are not only
have the capability to degrade pesticides in soil but also are
a deciding factor for the regulation of soil heath.
Factors influencing bioremediation of soil
The soil with its biotic and abiotic processes is the most
important component of the agroecosystem. Soil is made up
of an endless range of components, but these are divided
into four main components: organic, mineral, water, and air.
The ideal ground for most plants is made up of 45% min-
eral, 25% water, 25% air, and 5% organic matter. In fact, the
proportions of these four components vary greatly in natural
conditions. But farming soil, compared with natural ecosys-
tems soil, has many anomalies and alterations and is rich in
additional nutrients. So, several factors govern the fate of
soil and the degradation of xenobiotics in agroecosystem
soil. Some of these factors which are of utmost importance
for the bioremediation of soil by soil fungi are dis-
cussed below.
Bioavailability
Pesticide bioavailability is a major limitation in the bio-
remediation of contaminated soil. In this context, the
amount of pesticides, which can be easily absorbed and
metabolized through microbes can be defined as bioavailable
pesticides. At low pesticide concentration, microbes fail to
generate energy that causes biodegradation due to which the
catabolic gene systems never get activated. At low pesticide
concentration the distance between the molecule and the
microorganisms increases, minimizing the probability of
contact (Coche et al. 2018). In the case of the co-metabolic
situation, this process is determined by the type and content
of the alternate metabolites to support microbial growth
(Bhardwaj et al. 2020; Odukkathil and Vasudevan 2016;
Raimondo et al. 2020). Microbial cells may degrade pollu-
tants at low pollutant concentrations in a less nutrient envir-
onment, while low environmental nutrients reduce their
growth rate, which eventually leads to reduced pesticide
intake. Sometimes in a growing stage of microbes, the bio-
availability of a higher amount of pesticides facilitates deg-
radation, resulting in a higher biodegradation rate. In
addition, enzyme-catalyzed reactions form the basis of bio-
degradation in most cases, due to which in a system with
low pesticide concentrations and microbial growth, the rates
of the enzyme reaction are therefore lower, typically follow-
ing Michaelis-Menten kinetics (Bhardwaj et al. 2020;
Odukkathil and Vasudevan 2013,2016; Raffa and
Chiampo 2021).
Physiochemical factors
Soil physicochemical factors, such as pH, temperature, mois-
ture, soil inorganic matter, and soil texture have a substan-
tial impact on the degradation of pesticides. The
biodegradation of most of the pesticides follows an enzyme-
driven process, due to which the biodegradation process can
be improved by optimizing the abiotic factors. The growth
of most microbes also depends on the environmental condi-
tions, that influence the rate of production of biosurfactants
and consequently determine the bioavailability of pesticides.
8 D. MOHAPATRA ET AL.
Physicochemical factors, such as pH, temperature, and C/N
ratio, also affect microbial biosurfactant production and thus
regulate pesticide bioavailability (Odukkathil and Vasudevan
2013,2016; Raffa and Chiampo 2021). Thus the abiotic fac-
tors are to be optimized if pesticide-contaminated soil is to
succeed in its bioremediation.
Temperature and pH
Microbial growth is always determined by the temperature
and pH of the environment in which they grow.
Optimization of these factors can help in augmenting the
process of bioremediation. The variation of pH in a stable
system, like soil, is comparatively less than in water, but it is
difficult to optimize the pH and temperature of the soil.
Because most microbial species can survive in a certain pH
range only and pH affects nutrient availability, optimization
of pH is a key step in bioremediation of contaminated soil.
Soil temperature also controls microbial activity and survival
as well as the rate of decomposition of organic matters
including pesticides (Ajiboye et al. 2020; Erguven 2018).
Natural organic matters are actively involved in sequestering
pesticides and their metabolites in the soil thereby strongly
affecting the distribution (Ajiboye et al. 2020; Erguven 2018;
Shareef and Shaw 2008). Ding and Wu (1995) reported that
dissolved organic matters (DOM) improved pesticide solu-
bility and reduced the content of sequestered pesticides on
the surface of the soil; therefore, it is likely that the binding
of pesticides to DOM can have significant environmental
effects. Pesticide-DOM binding mechanisms include hydro-
gen bond, electrostatic interactions, donor-acceptor mecha-
nisms, and charge transfer, hydrophobic interaction, ligand
exchange, and van der Waal’s force. Many of these mecha-
nisms can function simultaneously, depending on the
molecular nature of the pesticide and the chemical compos-
ition, functionality, and structure of the organic matter in
the soil (Erguven 2018; Senesi 1992). The largest fraction of
DOM comprises humic substances (HS) whose binding cap-
acity has a prominent role to play in the destiny of pesti-
cides in water systems (Ajiboye et al. 2020). Therefore, these
substances can be used as model systems to elucidate the
interaction mechanism between pesticides and DOM
(Alfonso et al. 2017; Nguyen et al. 2018; Prosen et al. 2007).
The soil sorption of the ionic pesticides (e.g., imazaquin),
has been found to rest on the electric potential of the soil’s
surface (Rocha et al. 2002). Bipyridinium cations, like para-
quat, are the most eminent members of the ionized pesticide
group, with an ionic interaction mechanism that is expected
to influence their electrostatic adsorption in the soil. The
adsorption of weakly acidic pesticide like phenoxyalkanoic
acid depend on the changes in the percentage of the dissoci-
ated acid and is one of the major mechanics of interaction
with the soil (Celis et al. 1999). Changes in pH and ionic
strength have a common effect on organic chemicals adsorp-
tion, as expected, in determining binding mechanisms where
the electrostatic effect is considered important.
The rate of degradation of organic compounds in soil
increases with the soil temperature increases (Alfonso et al.
2017; Robichaud et al. 2019). Thibault and Elliott (1979)
reported that for every 10 C rise in temperature there is an
increase in the growth of microbes due to organic com-
pound desorption, which increases the bioavailability of the
compounds for microbes to degrade. The optimum pH and
temperature, required for the degradation of a compound,
can vary with the chemical structure QSAR and microbes
thereof. In this area, studies also support the need to opti-
mize these factors so that pesticides contaminated soils can
be successfully bioremediated. The best conditions for the
biodegradation of pesticides in soils vary with soil com-
pounds and organisms, but degradation rates in acidic pH
were found to be slow in comparison to alkaline and neutral
pH due to increased stability of chemical groups in acidic
pH (Raffa and Chiampo 2021; Reid et al. 2000). In addition,
temperature and pH variations also affect the production
and emulsification activities of biosurfactants. The tempera-
ture and pH of the soil should therefore be optimized to
enhance microbial activity and the availability of pesticides
to microbes and thereby successfully bioremediate pesticide
pollutants in the agroecosystem.
Soil moisture
The soil moisture content has a significant impact on the
bioavailability and degradation of soil retained pesticides.
Soil saturation with an excess of water creates an anoxygenic
environment resulting in anaerobic respiration that is less
energy-efficient and slows down the biodegradation process.
On the other hand, soil humidity content is also critical for
chemical availability and microorganism enhancement and
proliferation (Gopal et al. 2007; Meena et al. 2020).
Volatilization and adsorption rates of pesticides in soil are
influenced by soil conditions (texture, moisture, organic
matter content, and pH) (Arias-Est
evez et al. 2008). Well
aerated humid soil facilitates a faster rate of degradation and
removal of environmental contaminants. Therefore, the
maintenance of the moisture content of the soil is important
for increasing the bioavailability of the pesticide in soil, and
only those pesticide groups, which are available in water-sol-
uble form may be degraded. The mass transfer of soil con-
taminants into the solution phase drives the bioavailability
of the xenobiotics (Kumar et al. 2006). Consequently,
increased capacity for microbial conversion does not neces-
sarily lead to a higher rate of biotransformation if mass
transfer from the soil phase to the solution phase is the lim-
iting factor (Gopal et al. 2007). The rate of degradation of
pesticides with a decrease in moisture content has been sig-
nificantly reduced in the majority of the bioremediation
studies and vice versa because of reduced bioavailability
(Arias-Est
evez et al. 2008).
The redox potential of soil determines the metabolic type,
nitrogen fixation, biological activity, and growth of the
microorganism in soil. It also governs the enzymatic activity
of the organism in soil (Husson 2013). Research also shows
that fungus develops better in conditions >250 mV but bac-
teria develop better under condition <0 mV, where the
growth of the fungus is significantly reduced (Seo and
DeLaune 2010). So, redox potential is a very important fac-
tor that decides the fate of xenobiotic compounds in the
GEOMICROBIOLOGY JOURNAL 9
natural environment and hence is a major factor affecting
the bioremediation of these compounds.
Soil composition
While determining the best bioremediation approach to field
conditions, the soil composition has an important consider-
ation. It affects the distribution of the contaminated soil’s
nutrients, oxygen, microbes, water, and pollutants. These
compounds in soil, organic and inorganic, may bind and
reduce the bioavailability of pesticides (Alfonso et al. 2017;
Nguyen et al. 2018; Robichaud et al. 2019). A range of
environmental factors, including redox potential and moisture
content, can affect the rate of biodegradation of pesticides in
soil. Adsorbed pesticides’bioavailability is dependent on the
degree of adsorption and the desorption rate. Sorption and
adsorbent properties differ among different soil types and
pesticides, which significantly determines the degradation
and removal of these compounds in the agroecosystem.
Surfactants
Surfactants are surface-active agents with a wide range of
characteristics including reduced surface and liquid inter-
facial tensions. They are amphiphilic molecules that are
aggregated into micelles. The lipophilic part of the surfac-
tants does not form hydrogen bonding in an aqueous phase,
thus increasing the free energy of the system. Therefore, to
stabilize the reduction in free energy, water gets adsorbed to
the pesticide by removing the hydrocarbon tail due to which
the structure of a micelle is formed where the hydrocarbon
part of the surfactant positions toward the center.
Consequently, the hydrophilic part comes in contact with
water (Odukkathil and Vasudevan 2013,2016). The forma-
tion of micelle reduces the interfacial tension among immis-
cible fluids so that they can be miscible by creating
additional surfaces. Thus a single interface consisting of a
miscible and immiscible component becomes a smaller
interface between the two components. The formation of
micelle allows the division into the central hydrophobic
pseudo-phase core of hydrophobic structures to permit solu-
bility. This can cause a compound to disperse beyond its
water solubility limit in a solution thus enhancing its sus-
ceptibility to degradation (Odukkathil and Vasudevan 2013,
2016; Raffa and Chiampo 2021).
The surfactants can improve solubilization of soil pollu-
tants in surfactants mediated bioremediation, which in turn
improves bioavailability (Li and Chen 2009). The bioavail-
ability of organic contaminants is improved through both
synthetic surfactants and biosurfactants. The ability of syn-
thetic surfactants, such as Afonic 1412-7, Tween80, Brij 30,
and Triton X-100 to improve the bioavailability of hydro-
phobe organically contaminated materials has experimented
(Odukkathil and Vasudevan 2013). In the bioremediation of
contaminated soil and surface environments, biosurfactants
are more ecologically acceptable than synthetic surfactants.
Biosurfactants, many of whose potential benefits are
improved biodegradation and bioremediation, are involved
as key elements in physical, chemical, and behavioral activity
of microbes. Many pesticides have little solubility, which is
determined by their partition coefficients (log K
OW
), and are
often presented as a second organic phase, a major limiting
factor in biodegradation, in the soil environment. These
hydrophobic pesticides are still linked to soil, and because
the interface tension between aqueous and organic phases is
unavailable for biodegradation. Under these limited condi-
tions, many indigenous microbes in the soil produce biosur-
factants to weaken the soil-toxicant interaction and act on
them to degrade. The presence of biosurfactants induces
micelles or pesticide pseudo-solubilization for the hydro-
philic microbes to interact with a hydrophilic micelle, which
contains a pesticide. This pseudo solubilized pesticide con-
tains micelle unites membranes and directly transfers a con-
taminant molecule to the outside of a microbial cell
membrane (da Silva et al. 2021; Miller and Bartha 1989;
Otzen 2017).
While bacterial species have been well explored in the
production of biosurfactants, relatively less fungi are known
to produce biosurfactants. Some of the fungi, which show
the ability to produce biosurfactants, are Candida lipolytica
(Rufino et al. 2007, Sarubbo et al. 2007), Candida ishiwadae
(Thanomsub et al. 2004), Candida bombicola (Bhardwaj
et al. 2013; Cavalero and Cooper 2003; Felse et al. 2007),
Candida batistae (Konishi et al. 2008), Ustilago maydis
(Alejandro et al. 2011), Aspergillus ustus (Kiran et al. 2009),
Trichosporon asahii (Chandran and Das 2010), and
Penicillium 8CC2 (Sena et al. 2018), which have been
studied so far. Many of these products are known to pro-
duce low-cost biosurfactants. Sophorolipids, which is a type
of glycolipids, is the main type of biological agents produced
by these strains. Fungal biosurfactants are found very effect-
ive in accelerating the degradation of contaminants by facili-
tating their desorption (da Silva et al. 2021). However, in
spite of the fact that biosurfactants can be found in the bio-
remediation of contaminated soils and subsurface environ-
ments in a more environmentally acceptable way than in
synthetic surfactants, their high production costs limit their
wide application.
Among fungal biosurfactants, sophorolipids have taken
the lead due to their high rate of production and both
hydrophilic and hydrophobic properties (Alejandro et al.
2011). In addition, other fungal biosurfactants like mannory-
lerythriol lipids, xylolipids, polyol lipids, and cellobiose lip-
ids are also additives for the formulation of colloids with
various xenobiotics (da Silva et al. 2021). These surfactants
are used to improve the bioavailability of insecticides to tar-
get plants for efficient pest management (Kumar et al.
2021). Lipopeptide biosurfctants produce by Penicillium
charysogenum SNP5 have high efficiency to reclaim contami-
nated soil (Camargo-De-Morais et al. 2003; da Silva et al.
2021). The emulsifiers synthesized impact the bioavailability
of pesticides in the soil environment (da Silva et al. 2021;
Eldin et al. 2019).
Organic amendments
The use of organic waste in farm land specifically has been
practiced for centuries because of its fertilizer characteristics
10 D. MOHAPATRA ET AL.
and contributions to the physico-chemical and biological
stabilities of soil (Robichaud et al. 2019; Said-Pullicino et al.
2004). However, in recent years the focus has been on
assessing the impacts of exogenous organic carbon resources
on soil pesticide mobility, behavior, and metabolism
(Blackshaw et al. 2005; Nguyen et al. 2018; Song et al. 2008).
In terms of agronomics, adding organic waste has an envir-
onmental aspect, as these amendments play a significant
role in determining the fate of xenobiotic compounds, such
as heavy metals, aromatic carbohydrates, and pesticides, and
also increase biological activity and fertility of the soil
(Bohme et al. 2005;B
€
uy€
uks€
onmez et al. 1999; Wanner et al.
2005). Microbial activity increases with organic amendments
because of the availability of simple organic nutrients like
sugar and amino acid availability. Additional organic modi-
fications increase the organic matter dissolved in soil, which
facilitates the release and movement of pesticides (Ajiboye
et al. 2020).
Organic modifications may accelerate biodegradation due
to structural changes in the porosity of soil and stimulation
of microbial activities (Worrall et al. 2001). The amendment
of organic waste to the soil contributes to the enhancement
of humic acid, fulvic acid, and other active humic com-
pounds (Plaza et al. 2003). The buffering capacity of the
soils and the promotion of soil structure are improved by
organic amendment, which creates favorable conditions for
aeration and retention of moisture (Marschner et al. 2003).
Additional organic amendments also encourage soil fungi to
positively modify their growth and metabolic potential. The
quantitative augmentation of fungal biomass and enzymes
acts non-selectively on pesticides and helps to initiate pesti-
cide degradation. Thus, these pesticides are partially
degraded by fungal activities and are more likely to lead to
bacterial mineralization. Increased bacterial soil population
also increases with soil organic contents (Chiu et al. 1998,
Purnomo et al. 2010). Organic amendments often change
the rate and path to the degradation of the pesticides in soils
by the nature of the organic modification and its effect on
the microbial community (Purnomo et al. 2010; Raffa and
Chiampo 2021; Robichaud et al. 2019). The majority of
organic lignocellulose materials are susceptible to attacks by
fungi, which release extracellular enzymes that function as a
catalyst for pesticide degradation by bacteria and enhance
the degradation process. The addition of degradable organic
components also improves biotransformation and pesticide
cometabolism by increasing microbial activity (Robichaud
et al. 2019).
The organically dissolved matter has similar characteris-
tics to the active surface agents, such as surfactants that
reduce surface tension and increase compounds solubility
making them available to the soil microbes (Li et al. 2005).
Soil modifications, such as spent wheat bran, mushroom
compost, biogas litters, poultry litter, charcoal, coir pith
compost and vermicompost, leaf compost, and farm manure
amendments during the biological restoration process,
invariably stimulate microbial degradation. However, some-
times organic amendments may lead to a reduction of the
degradation power of the organism, especially when too
many organic matters create an anaerobic situation in the
soil ecosystem. This, in turn, has an adverse impact on cer-
tain microbial populations, which are key to contaminant
degradation (Moorman et al. 2001). The addition of auxil-
iary carbon source/organic amendments also sometimes
lower soil degradation since the auxiliary carbon source
becomes the preferred substrates, available in plenty, than
the toxic xenobiotic compounds in soil (Ajiboye et al. 2020;
Ma et al. 2014; Supriya and Dileep 2009).
Degradation of insecticides by fungi
Organophosphates
A variety of organophosphates, with a diversity of structural
complexity, have been found to be efficiently degraded by
many fungal species. Soil-derived fungus Aspergillus oryzae,
isolated from the soil using 100 mg/l of monocrotophos,
showed tolerance up to 900 mg/l of monocrotophos in liquid
medium (Bhalerao and Puranik 2007; Fauriah et al. 2021).
Aspergillus sp. degraded chlorpyrifos very efficiently in a
continuous flow bioreactor as compared to degradation by
bacteria (Abraham et al. 2016; Yadav et al. 2015). The gen-
eral degeneration of the insecticides up to 1000 mg/l was
caused by two phosphatase-producing isolates A. flavus and
A. sydowii. By enhanced phosphatase activity, soluble phos-
phorus increased clearly under the action of these species in
the presence of chlorpyrifos. In soil amended by organic
matter, the mineralization of insecticide residues were higher
than in un-amended soil (Ajiboye et al. 2020; Hasan 1999;
Robichaud et al. 2019). Pirimiphos-methyl and lancer were
utilized by A. fumigatus as sole sources of phosphorus,
which not only sufficed the fungal growth but also the
residual phosphorus increased in the medium. Similarly, A.
flavus and A. sydowii isolates utilized lancer and malathion
(1.65 g/l) as a source of phosphorus (Hasan 1999). A. sydo-
wii,A. niger, and A. flavus were the best degrading species
by producing more than 50% of the biomass in enrichment
culture. But in the case of A. terreus and A. fumigatus they
were able to produce <50% biomass but were able to
degrade the pesticides (Hasan 1999). A. niger degraded
dimethoate by overproduction of OP acid anhydrase
(Phosphotriesterase) (Liu et al. 2001). A tolerant Aspergillus
niger MRU01 was found capable to degrade malathion most
efficiently followed by dimethoate, chlorpyrifos, and para-
thion (Mohapatra et al. 2021). As seen in the case of bac-
teria, the OP degrading activity of the fungus was enhanced
significantly by the presence of Cu
2þ
(Liu et al. 2001;
Mulbry and Karns 1989). Liu et al. (2001) observed that a
tolerant strain of A. niger was unable to degrade parathion
and dichlorovos but was able to degrade formothion and
malathion with almost with equal efficiency thus indicating
a moderate chemical specificity of the enzyme.
Fungi are also known to degrade OP insecticides by
enhanced phosphatase activity (Hasan 1999; Mohapatra
et al. 2018,2021). Two fungi A. flavus and A. sydowii, iso-
lated from wheat straw, were the first reported fungi capable
of degrading organophosphate pesticides (pyrazophos, piri-
miphos-methyl, malathion, dimethoate, profenofos, lancer)
GEOMICROBIOLOGY JOURNAL 11
and utilizing these compounds as sole phosphorus (Hasan
1999). A. flavus was also reported to degrade monocroto-
phos at pH 8 and temperature 30 C (Jain et al. 2014;
Thirugnanam and Senthilkumar 2016). Two other species of
Aspergillus namely A. niger and A. flavus degraded mono-
crotophos in a phosphorus-free liquid medium and A. niger
showed better degradation ability than the other Aspergilus
spp. (Abraham et al. 2016; Jain and Garg 2013). Penicillium
smithii and A. niger metabolically degraded dichlorvos effi-
ciently in a nutrient-enriched medium through a co-meta-
bolic mechanism. There was increased growth performance
of the fungi in the culture with the insecticide on prolonged
exposure indicating the utilization of the insecticide as a
substrate by the fungus (Mohapatra 2006). Phosphatase
from A. sydowii was very effective against pyrazophos fol-
lowed by lancer and malathion (Hasan 1999, Mohapatra
Table 3. Fungal species tested for the degradation of various pesticides.
Pesticides Fungal spp. References
Chlorfenvinphos P. citrinum,A. fumigatus,A. terreus, and
T. harzianum
Oliveira et al. 2015
Chlorpyrifos Aspergillus sp., A. niger,A. terreus,A. niger MRU01,
T. viride,Alternaria alternata,Cladosporium
cladosporioides,Fusarium sp.,C. cladosporioides
Hu-01; Acremonium sp., Fusarium oxysporum,P.
funiculosum,Acremonium strictum
Bharathkumari and Sivakami 2018; Chen et al.
2012; Hussain et al. 2007; Jaiswal et al. 2017;
Kulshrestha and Kumari 2011; Mohapatra et al.
2021; Mukherjee and Gopal 1996; Silambarasan
and Abraham 2013a,2013b; Yadav et al. 2015
Dimethoate A. niger,A. flavus,A. sydowii,A. niger MRU01 Hasan 1999; Liu et al. 2001; Mohapatra et al. 2021
Lancer A. flavus,A. niger and A. sydowii,A. flavus,
A. tamarii
Hasan 1999
Malathion A. flavus,A. sydowii,A. niger MRU01 Hasan 1999; Mohapatra et al. 2021
Methyl parathion A. niger,A. niger MRU01 Mohapatra et al. 2021
Monocrotophos A. orysae,A. niger, and A. flavus Jain and Garg 2013; Jain et al. 2014
Pirimiphos-methyl A. fumigatus,A. flavus,A. niger, and A. sydowii Hasan 1999
Profenofos A. flavus,A. sydowii Hasan 1999
Pyrazophos A. flavus,A. niger, and A. sydowii Hasan 1999
Dichlorovos A. niger Liu et al. 2001
Formothion A. niger Liu et al. 2001
Fenitrothion Cunninghamella elegans ATCC36112 Zhu et al. 2010
Endosulfan A. sydoni,A. niger,A. tamarii JAS9, A. terreus,A.
flavus,P. chrysogenum,T. harzianum
Ahmad 2020; Bhalerao 2013; Bhalerao and Puranik
2007; Goswami et al. 2009; Hussaini et al. 2013;
Katayama and Matsumura 1993; Silambarasan
and Abraham 2013b
DDT T. harzianum,T. viride,A. flavus,A. parasitacus Katayama and Matsumura 1993; Mehrotra
et al. 2004
Lindane A. niger Hussaini et al. 2013
Aldrin A. flavus,P. notatum Mehrotra et al. 2004
Dieldrin A. flavus,P. notatum Mehrotra et al. 2004
Chlordane A. niger Singh and Dwivedi 2004
Heptachlor A. niger Singh and Dwivedi 2004
a-HCH Aspergillus sp. Javaid et al. 2016
Pentachlorophenol T. harzianum,T. piluliferum,T. aureoviride Sing et al. 2014
b-cypermethrin A. niger,A. oryzae M-4 Chen et al. 2011a; Kaur and Balomajumder 2020b;
Zhao et al. 2016
b-cyfluthrin T. viride 2211, A. niger,A. terricola,Phanerochaete
chrysoporium
Saikia and Gopal 2004
3-phenoxybenzoic acid A. oryzae M-4 Zhu et al. 2016
Cyhalothrin Aspergillus sp. (PYR-P2) Kaur and Balomajumder 2020b
Fenvalerate Cladosporium strain HU Chen et al. 2011c
Allethrin Fusarium proliferatum CF2 Bhatt et al. 2020
Metolachlor A. flavus,A. terricola,Trichoderma spp., P. oxalicum
MET-F-1
Chang et al. 2020; Nykiel-Szyma
nska et al. 2020;
Sanyal and Kulshrestha 2003,2004
Alachlor Trichoderma koningii,Trichoderma spp. Nykiel-Szyma
nska et al. 2018,2020
Atrazine A. fumigatus,A. ustus,A. flavipes,Rhizopus
stolonifer,Fusarium moniliforme, F. roseum,F.
oxysporum,Penicillium decumbens,P.
janthinellum, P. rugulosum, P. luteum, T. viride,A.
fumigatus and A. terreus
Bravim et al. 2021; Herrera-Gallardo et al. 2021;
Hock et al. 2020; Oliveira et al. 2015; Singh
et al. 2008
Pyrazosulfuron ethyl Penicillium chrysogenum,A. niger He et al. 2006; Sondhia et al. 2013
Sulfosulphuron Trichoderma sp. Yadav and Choudhury 2014
Chlorimuron-ethyl A. niger Sharma et al. 2012
Chlorsulfuron Penicillium sp., A. niger Boschin et al. 2003; Gul and Ahmad 2020
Diuron A. fumigatus,A. terreus,Mortierella sp. Ellegaard-Jensen et al. 2013; Oliveira et al. 2015
Isoproturon A. fumigates,A. terreus Oliveira et al. 2015
Metsulfuron methyl A. niger,Penicillium sp. (DS11F), Trichoderma sp. Boschin et al. 2003; He et al. 2006; Vazquez and
Bianchinotti 2013
Glyphosate A. terreus,A. niger,T. viridae,Fusarium oxysporum,
A. oryzae A-F02,Penicillium IIR,P. chrysogenum,
P. notatum, and T. harzianum
Adelowo et al. 2014; Bujacz et al. 1995; Eman
et al. 2013; Fu et al. 2017; Klimek et al. 2001;
Krzy
sko-Łupicka and Orlik 1997; Krzy
sko-Lupicka
et al. 1997
12 D. MOHAPATRA ET AL.
et al. 2021). Mukherjee and Gopal (1996) reported that A.
niger was able to degrade chlorpyrifos and the degradation
of the insecticide increased with the increase in the duration
of incubation. Total degradation of chlorpyrifos (applied at
300 mg/kg) was observed by the use of A. terreus in media
supplemented with nutrient and media with no nutrient but
only supplemented with insecticide within 24 and 48 h,
respectively (Silambarasan and Abraham 2013a). Diazinon
was found degraded effectively by A. oryzae and A. niger
(Hamad 2020) and Cunninghamella elegans (Zhao et al.
2020). Glyphosate was primarily used by A. niger through
the breakdown of carbon-phosphorus bond (C-P), resulting
in the release of sarcosin and a group of phosphates. As the
source of phosphorus for fungal growth, the phosphate
group was used. The sarcosine released was possibly
degraded further to other metabolizable products. A small
fraction of the glyphosate has been broken down by cleavage
of the carbon-nitrogen (C-N), the release of aminomethyl
phosphonic acid, and growth of fungus escalated in the
presence of pesticide which depicts the degrading ability of
the organism Trichoderma viridae,A. niger, and Fusarium
oxysporum (Adelowo et al. 2014). A Malathion-tolerant
Aspergillus (Aspergillus niger MRU01), as well as the wild
type, were able to degrade four organophosphates malathion,
dimethoate, chlorpyrifos, and parathion after 5 days of incu-
bation of the organism in liquid Czapek Dox medium but
the ability of the tolerant strain was significantly higher than
of the wild type (Mohapatra et al. 2021).
Two soil fungi T. viride and A. niger were able to degrade
chlorpyrifos (Hussain et al. 2007). A. terreus, and
Verticillium sp. DSP were also reported to have efficient
chlorpyrifos degrading ability (Jayashree and Vasudevan
2007). Chlorpyrifos was more efficiently degraded by mixed
populations of fungi which included Alternaria alternata,
Cephalosporium sp., Cladosporium cladosporioides,
Cladorrhinum brunnescens,Fusarium sp., Rhizoctonia solani,
and T. viride, in liquid culture, rather than a single strain
(Jaiswal et al. 2017). Chlorpyrifos was efficiently used as a
source of carbon and nitrogen by Acremonium sp. strain
GFRC-1 and 83.9% degradation of the chemical was
achieved in Czapek Dox medium with full nutrients. The
major biodegradation product was detected as desdiethyl
chlorpyrifos (Kulshrestha and Kumari 2011). Kaur and
Balomajumder (2020a) reported a newly isolated
Acremonium sp. (MK514615), which was able to decontam-
inate carbamate-contaminated soil. A new fungal strain
Cladosporium cladosporioides Hu-01 was able to degrade
chlorpyrifos and utilized 50 mg/l of chlorpyrifos as the sole
carbon source and showed tolerance to a high concentration
of chlorpyrifos (up to 500 mg/l) and the optimum condition
being 26.8 C and pH 6.5 within 5 days (Chen et al. 2012).
A. terreus showed the highest mineralization potential fol-
lowed by A. tamarii,A. niger, but T. harzianum and
Penicillium brevicompactum showed moderately mineralized
ability in the presence of cyolan, malathion, and dursban
(Omar 1998). A study on the degradation of chlorpyrifos
using six fungi showed the degradation efficiency in the
order of T. viride >P. funiculosum >A. flavus >Fusarium
oxysporum >Rhizopus arrhizus >Acremonium strictum
which showed T. viride was the most efficient organism to
degrade the insecticide (Bharathkumari and Sivakami 2018).
Dichlorvos was effectively degraded by Cunninghamella aele-
gans,Fusarium solani,Talaromyces atroroseus,A. oryzae,
and Penicillium sp. and Trichoderma atroviride (Sun et al.
2019; Zhang et al. 2021). Zhu et al (2010) observed that
Cunninghamella elegans ATCC36112 could degrade fenitro-
thion up to 81% within 5 days of treatment. The fungi P. cit-
rinum,A. fumigatus,A. terreus and T. harzianum not only
showed resistance against chlorfenvinphos but were also able
to degrade it (Oliveira et al. 2015). A list of the organophos-
phates and the degrading organism is given in Table 3.
Organochlorines
With respect to various OC insecticides, most of the work
has been done on the fungal degradation of endosulfan. A.
sydoni strain was found to use endosulfan as a source of car-
bon in broth medium as well as in soil microcosm by effi-
ciently degrading it (Goswami et al. 2009). It has been
reported that the fungus degraded both a-and b-endosulfan
with almost equal efficiency (95 and 97%, respectively)
through oxidative and hydrolytic pathways in 18 days of
incubation (Goswami et al. 2009). A. niger completely
removed technical grade endosulfan within 12 days at
400 mg/ml concentration under laboratory condition
(Bhalerao and Puranik 2007). It was also seen that the fun-
gal degradation of the insecticide produced various less toxic
products like endosulfan sulfate, endosulfan diol, and an
unidentified metabolite, but the increase in CO
2
evolution
proved metabolic utilization of the insecticide and its deg-
radation products (Bhalerao and Puranik 2007). An indigen-
ous A. niger (ARIFCC 1053) strain was able to tolerate and
degrade higher concentrations up to 1000 mg/l of endosulfan
(Bhalerao 2013). Degradation of endosulfan was also
observed in A. terreus,A. niger,A. flavus, and P. chrysoge-
num, which was not very remarkable in the beginning three
days but increased significantly with an increase in incuba-
tion duration (Ahmad 2020; Mukherjee and Mittal 2005). A.
tamarii JAS9, which was isolated from soil contaminated
with endosulfan, grew well in 1000 mg/l concentration and
was able to tolerate up to 1300 mg/l (Silambarasan and
Abraham 2013b). The most important feature of this strain
was that it was able to degrade the more persistent endosul-
fan sulfate (Silambarasan and Abraham 2013a). Ahmad
(2020) isolated microbes from cultured soil which included
Bacillus subtilis,A. niger,A. flavus, and P. chrysogenum
which were able to degrade 10 mg/l of endosulfan. A. niger
degraded endosulfan (59%) more efficiently than lindane
(29%) (Hussaini et al. 2013).
Microbial degradation of DDT has been recorded since
1960 but reports on fungal degradation are limited.
Katayama and Matsumura (1993) observed that T. harzia-
num was able to degrade dieldrin, DDT, pentachloronitro-
benzene, pentachlorophenol, and endosulfan but was not
able to degrade hexachlorocyclohexane. Singh and Dwivedi
(2004) reported that many strains of T. viride could
GEOMICROBIOLOGY JOURNAL 13
metabolize DDT by producing DDNS, DDE, and DDA. A
work on A. flavus and A. parasiticus showed that these
organisms were able to convert DDT to DDE in nutrient-
enriched medium but were not as efficient as T. viride
(Mehrotra et al. 2004). The presence of DDT promoted a
higher rate of formation of reactive oxygen species in fungal
cells than the controls resulting in bioremediation of DDT-
contaminated soil (Russo et al. 2019). The authors also
reported DDT tolerance in Trichoderma hamatum FBL 587
and Rhizopus arrhizus FBL 578 in the presence of different
carbon sources. However, degradation of DDT was not done
very efficiently by different strains of the soil fungi, which
showed the recalcitrant behavior of the chemical in the
environment. Microbial consortia were helpful for the
removal of the residual DDT from the environment.
Synergistic interaction of a consortium of the brown-rot
fungus Fomitopsis pinicola and the bacterium Ralstonia pick-
ettii was found effective for DDT biodegradation, as com-
pared to the individual species (Purnomo et al. 2020). A
consortium of Mucor sp., Fusarium sp., and Trichophyton
sp. could grow effectively in DDT contaminated soil con-
taining >2mg/kg residue (Nasution and Bakti 2018).
Two fungal strain A. flavus and P. notatum were able to
degrade aldrin and dieldrin metabolically, which increased
with inoculum density and incubation time (Mehrotra et al.
2004). A syntrophic relationship was found between bacteria
and Aspergillus to efficiently degrade these chemicals as
compared to individual strains. Similar syntrophy between
Pseudomonas urticae and A. niger also caused degradation of
chlordane and heptachlor (Singh and Dwivedi 2004).
Efficient degradation of a-HCH was seen in Aspergillus spe-
cies in nutrient-enriched media but metabolic use of the
chemical is yet to be reported (Javaid et al. 2016). A white-
rot fungus Pleurotus ostreatus could degrade aldrin and diel-
drin through the epoxidation and hydroxylation reactions.
In some studies esterification, deoxygenation, dehydrogen-
ation, dechlorination, demethylation reactions have been
reported during the fungi-mediated transformation of these
pesticides (Akhtar and Mannan 2020; Purnomo et al. 2017).
Thirty-three fungal strains isolated from saw dust were able
to degrade pentachlorophenol out of which those which
showed the highest tolerance were identified to be T. harzia-
num,T. piluliferum, and T. aureoviride, which showed up to
99% sequence similarity and Cunninghamella bainieri
showed 98% sequence similarity when identified using
molecular identification technique (Sing et al. 2014). Xiao
et al. (2011) reported that Phlebia tremellosa,P. brevispora
and P. acanthocystis removed about 71, 74, and 90% of
heptachlor, respectively when incubated for 14 days. A list of
the organochlorines and the degrading organism is given in
Table 3.
Pyrethroids
There are not many reports on the fungal degradation of
pyrethroids but there are reports that cyanobacteria and
algae can degrade the compounds (Chandrakala 2016;
Samantarai 2006). Some literature is, however, available to
show the efficiency of Aspergillus to degrade pyrethroids.
Five fungal strains Trichoderma viride 5-2, T. viride 2211, A.
niger,A. terricola, and Phanerochaete chrysoporium were
tested for the ability to degrade b-cyfluthrin and all the
strains were able to degrade the chemical (at 5 mg/ml). The
highest degradation of the compound was seen in T. viride
5-2, followed by T. viride 2211. The degradation of the com-
pound followed the first-order kinetics with a fast degrad-
ation rate during the first 7 days and was proportional to
the growth of the fungi. The strain T. viride 5-2 produced
five degradation products after 20 days of growth, out of
which three were identified as 3(2,2-dichlorovinyl)-2,2-
dimethyl cyclopropanoic acid, a-cyano-4-fluorobenzyl-3-
(2,2-dichlorovinyl)-2,2-dimethyl cyclopropane carboxylate,
and a-cyano-4-fluoro-3-phenoxy benzyl alcohol (Saikia and
Gopal 2004). A. niger ZD11 isolated from heavy pesticides
contaminated soil showed production of pyrethroid hydro-
lase and was able to degrade a diverse type of pyrothroids,
thereby using the products as a sole carbon source (Chen
et al. 2011a,2011b; Liang et al. 2005; Zhu et al. 2016). Chen
et al. (2011a) reported that A. niger could completely
degrade 100 mg/l of 3-phenoxybensoic acid within 22 h and
was able to degrade 54.83% of b-cypermethrin (50 mg/l) in
7 days, which was quite higher than the rate of degradation
of these chemicals reported in some bacteria, such as
Stenotrophomonas strain ZS-S-01 and Ochrobactrum lupini
DG-S-01 (Chen et al. 2011b). A study on the novel filament-
ous fungus A. oryzae M-4 strain isolated from soy sauce koji
was able to degrade 80.62% of 100 mg/l of 3-phenoxybenzoic
acid within 5 days of incubation following first-order kinet-
ics (Zhu et al. 2016).
A co-culture study on Bacillus licheniformis B-1 and A.
oryzae M-4 was able to degrade 78.85% of 100 mg/l of
b-cypermethrin and the concentration of 3-phenoxybenzoic
acid was reduced to 0.05 mg/l after 72 h of incubation. This
co-culture of strains degraded b-cypermethrin efficiently and
3-phenoxybenzoic acid completely (Zhao et al. 2016). The
degradation of 3-phenoxybenzoic acid, which is a toxic
intermediate from co-metabolic degradation of b-cypermeth-
rin, plays an important role in limiting the degradation of
the parent compound as it inhibits b-cypermethrin degrad-
ation by retarding the growth of the bacterial strain.
Another intermediate is gallic acid, which is obtained, from
cometabolic degradation of 3-phenoxybenzoic acid, inhibits
the degradation of 3-phenoxybenzoic acid, and hinders the
growth of the fungus (Zhao et al. 2016). But the co-culture
of both these organisms was able to degrade both these
intermediate inhibitors successfully, thereby facilitating the
faster rate of removal of the parent compound (Zhao et al.
2016). A newly isolated Aspergillus sp. (PYR-P2) from conta-
minated soil was able to degrade 500 mg/l of pyrethroid
mixture (cypermethrin, cyfluthrin, and 14 cyhalothrin). The
strain was able to completely degrade all three insecticides
within 15 days in a liquid medium and was able to signifi-
cantly reduce the amount of pyrethroids, when augmented
in the soil system (Kaur and Balomajumder 2020b). A
Cladosporium strain HU was capable of degrading several
pyrethroids and also 3-phenoxy benzaldehyde and was able
14 D. MOHAPATRA ET AL.
to tolerate high concentrations up to 1200 mg/l. The degrad-
ation ability of fenvalerate of this strain showed activity over
a wide range of pH and temperature without showing lag
phase (Chen et al. 2011c). A study on the degradation kinet-
ics of the fungal strain, Fusarium proliferatum CF2, showed
that this strain was able to utilize 50 mg/l of allethrin as the
sole carbon source in minimal salt medium and tolerated
high concentrations of allethrin (up to 1000 mg/l). The opti-
mum condition for the degradation was at pH 6 and tem-
perature 26 C, causing completely degradation of allethrin
within 144 h (Bhatt et al. 2020). This showed that enzymatic
activities of the fungal strain gave them the ability to effect-
ively degrade b-cypermethrin and its metabolites except for
permethric acid, which makes it an important biodegradable
organism (Deng et al. 2015). The fungal degradations of sev-
eral other pyrethroids like phenanthrin (Wu et al. 2016),
allethrin (Bhatt et al. 2020), and cyphenothrin (Huang et al.
2020) have also been reported. The list of the pyrethroids
and the degrading organism is given in Table 3.
Degradation of herbicides
Chloroacetanilides
Many diverse types of herbicides are metabolically and co-
metabolically degraded by soil fungi. Metolachlor microbial
degradation was considered solely due to the mixed fungal
culture of Aspergillus terricola and Aspergillus flavus in soil
with an initial concentration of 20 mg/g against uninoculated
and inoculated sterile soils and the degradation was found
to be 49.21% (Sanyal and Kulshrestha 2003). The combined
culture of fungus was able to degrade almost 100% of the
herbicide, though there was a decrease in net degradation
when the fungi were applied separately (Sanyal and
Kulshrestha 2003). Penicillium oxalicum MET-F-1 isolated
from activated sludge was able to degrade 50 mg/l of metola-
chlor when coupled with 0.1% glucose and 0.1% yeast
extract, up to 88.6%, within 384 h under optimal conditions
(Chang et al. 2020). The evidence from the metabolites
formed during the degradation of metolachlor showed that
the parent chemical was hydrolyzed by dechlorination,
hydroxylation, and dealkylation and that aniline was in
minor fractions of the metabolites formed (Sanyal and
Kulshrestha 2004). A filamentous fungus Trichoderma
koningii was able to efficiently degrade alachlor up to 90%
after 72 h when applied separately and about 80–60% in the
presence of 1–5 mM of copper. More than 99% degradation
of the compound after 168 h was caused due to mitigation
of reactive oxygen species. The enzyme systems mainly
involved in the degradation were laccase and Cytochrome
P450 enzymes (Nykiel-Szyma
nska et al. 2018).
Eight fungal strains of Trichoderma spp. were investigated
for their ability to degrade of alachlor and metolachlor and
the test fungi were able to degrade 80–99% of alachlor and
40–79% of metolachlor after 7 days of incubation. The trans-
formation of herbicides was performed mainly by hydroxyl-
ation and dechlorination reactions. After 7 days of
application of the spores of T. koningii IM 0956, T. citrino-
viride IM 6325, T. harzianum KKP 534, T. viride KKP 792,
and T. virens DSM 1963 there was significant increase in the
roots and shoots of rapeseed seedlings, after being treated
with alachlor or metolachlor. All the strains taken in the
study also showed plant growth-promoting traits (Nykiel-
Szyma
nska et al. 2020).Paecilomyces marquandii was
reported to effectively degrade alachlor, in the presence of
zinc (Słaba et al. 2009). Cyanide hydratase was the key
enzyme involved in alachlor biodegradation by P. marquan-
dii, which was conformed from proteomic analysis
(Szewczyk et al. 2015). Słaba et al. (2015) also reported
accelerated degradation of alachlor by P. marquandii under
optimized and controlled conditions in liquid batches.
Sulfonylureas
Total biodegradation of two sulfonylureas-chlorsulfuron and
metsulfuron-methyl were achieved using Aspergillus niger in
a nutrient-rich medium where degradation was achieved due
to hydroxylation of benzene ring and cleavage of sulfony-
lurea bridge (Boschin et al. 2003). Even at high concentra-
tions of around 2 mg/ml chlorimuron-ethyl, A. niger,
isolated from soil samples, showed survival in the liquid
media. The degradation of the chemical by fungi was by
extracellular enzymes, which converted the chemical into a
more simple form (alcohols, fatty acids, aldehydes, and
ketones) to be used for growth and maintenance by the
microorganism. Fungal consortium with Aspergillus as the
major partner with species of Fusarium and Alternaria was
more effective in removing chlorimuron-ethyl from soil and
water (Sharma et al. 2012). A. niger and a Penicillium sp.
were able to degrade 97–99% of the parent compound
within 78–96 h of incubation in potato dextrose broth sup-
plemented with chlorsulfuron, but the same amount of deg-
radation was not achieved when the organism was
inoculated in perfused soil cores which indicated the organ-
isms could not directly metabolize the chemical (Gul and
Ahmad 2020). In chlorsulfuron contaminated soil, the soil
fungal population was dominated by the members of
Ascomycetes, and Basidiomycetes were represented by very
few species (Medo et al. 2020). Several species of
Aspergillus, Penicilium, and Fusarium have been utilized for
removal of sulfonylurea herbicides more efficiently as a con-
sortium than as solo bioremediatory (Javaid et al. 2016).
Two pyrazosulfuron ethyl degrading fungi Penicillium
chrysogenum and A. niger were isolated from the soil of rice
field were able to degrade the chemical in soil under labora-
tory conditions following a degradation pathway through
hydrolytic cleavage and sulfonylurea bridge (Sondhia et al.
2013). Four bacteria, nine filamentous fungi, and twenty
actinomycetes isolates had the capability of degrading met-
sulfuron-methyl, and an unknown strain of Penicillium sp.
(DS11F) showed the best performance among the test
organisms (He et al. 2006). Strains of Trichoderma,
Penicillium, and Mucor isolated from Argentina agricultural
soils were capable of using metsulfuron-methyl as a sole car-
bon and energy source. Trichoderma sp. showing the best
performance (Vazquez and Bianchinotti 2013). Yadav and
Choudhury (2014) isolated a Trichoderma species from
GEOMICROBIOLOGY JOURNAL 15
wheat rhizosphere soil, which was able to survive up to 2 g/l
of herbicide concentration and could survive in a minimal
broth rich in sulfosulfuron.
Fungal species like A. fumigatus,A. ustus,A. flavipes,
Rhizopus stolonifer,Fusarium moniliforme, F. roseum,F. oxy-
sporum,Penicillium decumbens,P. janthinellum,P. rugulo-
sum,P. luteum, and T. viride were able to degrade atrazine
in a basal salts medium supplemented with sucrose using N-
dealkylation of either alkyl amino group (Singh et al. 2008)
.Atrazine, diuron and isoproturon could not be degraded
only by fungi like A. fumigatus and A. terreus but a fungi-
bacteria consortium was effective to remove these herbicides
from the medium (Oliveira et al. 2015). This indicated that
the initial degradation was performed by bacteria and the
subsequent metabolism by the fungi. The fungal isolates viz
Aspergillus fumigatus,A. niger,A. nidulans,Trichoderma eri-
naceum,Fusarium verticillioides, and Penicillium citrinum
showed not only the resistance to atrazine but also high
mycelial growth in the presence of the herbicide (100 mg/
ml) indicating the metabolic degradation (Bravim et al.
2021; Herrera-Gallardo et al. 2021; Hock et al. 2020). In
most cases, it has been observed that from among the isolate
from atrazine contaminated habitat Fusarium spp. showed
an enhanced potential for detoxification of atrazine (Bravim
et al. 2021; Esparza-Naranjo et al. 2021). Similarly, the basi-
diomycetian strains like Pluteus cubensis SXS320,
Gloelophyllum striatum MCA7, Agaricales MCA17, Datronia
caperata MCA5, Pycnoporus sanguineus MCA16, and
Polyporus tenuiculus MCA11 could cause degradation of
atrazine, which was driven under nitrogen deficiency (Henn
et al. 2020).
Mortierella sp. isolates from agricultural soil were able to
degrade 19–54% diuron after 7 days when incubated at
24 C, but one strain among the species Mortierella sp.
LEJ701 was able to degrade almost 50% diuron within
10 days of incubation. But the species Mortierella sp. LEJ701
could not use diuron as carbon or nitrogen hence required
an additional carbon and nitrogen source for degrading
diuron (Ellegaard-Jensen et al. 2013). White rot fungi have
shown high efficiency to use diuron as a sole source of car-
bon and energy. They cause degradation of diuron due to
high levels of ligno cellulogic, antioxidant, and ctrochrome
P-450 monooxygenase enzyme activity (Hu et al. 2020). Li
et al (2021) have noted that Aepergillus,Pycnoporous,
Pleutus,Trametes,Neurospora,Cunninghamella, and
Mortierella as the reported fungal genera with high effi-
ciency to degrade diuron.
OP herbicides
Many fungal species belonging to the genera Aspergillus,
Fusarium,Mucor,Penicillium,Scopulariopsis, and
Trichoderma have been reported to tolerate and/or degrade
glyphosate (Carranza et al. 2017; Fu et al. 2017; Kunanbayev
et al. 2019; Zhan 2018). Sailaja and Satyaprasad (2006)
reported that application of glyphosate had no effect on spe-
cies of fungus-like Aspergillus,Fusarium, and Penicillium but
showed enhancement in the growth of Trichoderma which
was due to the herbicide induced alteration of pH causing
the acidity of the soil. Eman et al. (2013) reported local fun-
gal strains like A. niger FGP1, A. terreus PDP1, A. terreus
BGCS3, A. tamarii PDCS1, and A. flavus WDCS2 have a
high tolerance against glyphosate. After 16 days of incuba-
tion degradation of glyphosate in the liquid media was 90.6,
96.7, and 99.6% by A. flavus WDCz2, A. tamarii PDCS1, A.
flavus WDCZ2, respectively indicating a high degradation
efficiency of A. flavus WDCZ2 (Eman et al. 2013). Adelowo
et al. (2014) isolated three fungi T. viride,A. niger and
Fusarium oxysporum from soil and found these species
showed enhancement in growth in the presence of glypho-
sate, T. viride showed the highest growth enhancement while
A. niger showed the lowest. The enhancement in growth was
due to the degradation of glyphosate mainly through the
cleavage of Carbon-Phosphorus (C-P) bond, which resulted
in the release of a phosphate group and sarcosine (Carranza
et al. 2017; Spinelli et al. 2021)(Table 3).
Conclusion
In recent decades increased use of pesticide in agriculture
have augmented the crop yield but consequently, the con-
tamination of agroecosystem has emerged as an urgent issue.
Many different mechanisms have been proposed and experi-
mented with to get rid of such contaminants, and bio-
remediation stands as a viable and sustainable mechanism.
Different bacterial and fungal species have been isolated
from the contaminated habitat and have been evaluated for
their potential for use in bioremediation. Experiments in the
laboratory and field have shown that a large number of pes-
ticides are metabolically degraded by soil fungi but some are
degraded co-metabolically on supplementation of the cul-
tures/soil with alternate carbon sources. The rate of co-
metabolism is determined by the type of the added carbon
source. Among soil fungi, the members of Ascomycetes have
shown an enhanced resistance against the pesticides and
Aspergillus sp. have often been observed as a common mem-
ber of the fungal groups in pesticide-contaminated habitats.
In this review, it is observed that there is a great degree
of variation among various species and strains of fungi with
regards to their tolerance and biodegradation potential. The
fungal-fungal and fungal-bacterial synergism are more effi-
cient in bioremediation of contaminants than the perform-
ance of a single species/strain. Several studies have shown
the efficiency of bioremediation in vitro. However, field
studies in-situ are required to be extensively made to inte-
grate the passible factors that would influence the efficiency
of bioremediation. The fungal enzymes and biosurfactants,
so far known to degrade pesticides can be tried in a more
detailed way to optimize the mycoremediation process.
Acknowledgements
The authors thank the STD, and OHEPEE, Government of Odisha for
their support and DST, Government of India for infrastructure grant
under DST-FIST programme to the Department.
16 D. MOHAPATRA ET AL.
Disclosure statement
No potential conflict of interest was reported by the authors.
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