Towards sustainable physiochemical and biological techniques for the remediation of phenol from wastewater: A review on current applications and removal mechanisms

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Journal of Cleaner Production 417 (2023) 137810
Available online 24 June 2023
0959-6526/© 2023 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).
Review
Towards sustainable physiochemical and biological techniques for the
remediation of phenol from wastewater: A review on current applications
and removal mechanisms
Amina Bibi
a
, Shazia Bibi
b
, Mohammed Abu-Dieyeh
b
, Mohammad A. Al-Ghouti
a
,
*
a
Environmental Science Program, Department of Biological and Environmental Sciences, College of Arts and Sciences, Qatar University, Doha, P.O. Box: 2713, Qatar
b
Biological Sciences Program, Department of Biological and Environmental Sciences, College of Arts and Sciences, Qatar University, Doha, P.O. Box: 2713, Qatar
ARTICLE INFO
Handling Editor: Maria Teresa Moreira
Keywords:
Phenol contaminated wastewater
Low-cost adsorbents
Sustainable green technologies
Physiochemical and biological methods
Waste management
Sustainability and clean production
ABSTRACT
Phenol is a priority pollutant that presents a signicant threat to human health and natural systems when dis-
charged directly into the environment. Consequently, numerous technologies have been used and developed to
eliminate phenol from wastewater streams. These technologies can be categorized into physical, chemical, and
biological methods. While conventional treatment methods are highly efcient in phenol removal; some of these
techniques are not environmentally friendly and others are expensive. Therefore, sustainable, and green tech-
nologies are being employed and taken into consideration in the treatment of phenol wastewater due to their
effectiveness, affordability, and environmental compatibility. This review aims to highlight efcient green
physiochemical and biological methods of water treatment and demonstrate the mechanisms of phenol removal
in these technologies. Particular emphasis will be given to the use of low-cost adsorbents prepared from in-
dustrial and agricultural wastes for the efcient removal of phenol from wastewater as adsorption processes
show the highest cost-effectiveness among all the treatment technologies.
1. Introduction: phenolic compounds, reactivity, toxicity, and
fate
Water pollution is an increasingly pressing issue on a global scale,
primarily caused by factors such as rising water demand, population
growth, industrialization, urbanization, and agriculture activities. This
has resulted in the degradation and pollution of the environment,
adversely affecting water bodies, and ultimately affecting human health
and the environment. Several organic pollutants are responsible for the
reduction of water quality including phenolic compounds (Soto--
hernandez et al., 2017). Phenol is an aromatic organic compound with
the molecular formula C
6
H
5
OH, it comprises of an aromatic ring to
which a hydroxyl group is attached. These compounds are either formed
naturally by the action of different organisms or are produced by
numerous industries and released to the environment as wastewater
without appropriate treatment (Almasi et al., 2021; Soto-hernandez
et al., 2017). Fig. 1-A shows the sources of the various phenolic com-
pounds found in the environment. The introduced phenols are recalci-
trant contaminants that are resistant to degradation through physical,
chemical, and biological processes (Mohamed et al., 2020).
Consequently, the US EPA and the Canadian national pollutant release
inventory (NPRI) categorized phenol as one of the 129 specic priority
pollutants that must be remediated before discharge (US EPA, 2014). It
is estimated that over 10 million tons of phenolic compounds are dis-
charged into the environment (Alshabib and Onaizi, 2019) by the
petrochemical, pharmaceutical, leather, textile, and agrochemical in-
dustries. In addition, industrial processes such as paint, paper, pulp, and
pesticide production are also believed to be responsible for phenolic
compounds discharge into the environment (Alshabib and Onaizi, 2019;
Deng et al., 2011). Fig. 1-B shows the percent of phenolic compounds in
a variety of industrial efuents. The concentrations of phenolic com-
pounds in industrial efuents range from 1 mg/L and could reach up to
7000 mg/L (Mohd, 2020).
It is vital to handle phenolic compounds properly since these com-
pounds can adversely affect human health and biotic systems (Sar-
avanan et al., 2021). In terms of their impact on human health, phenolic
compounds are considered toxic, carcinogenic, and mutagenic. In
addition to that, phenols can lead to various health complications and
disorders including genotoxicity, muscle fatigue, dysfunction of the liver
and kidneys, metabolic and eating disorders, weight loss, diarrhea,
bronchoconstriction, irregular breathing, irritation of the ducts, coma,
* Corresponding author.
E-mail address: mohammad.alghouti@qu.edu.qa (M.A. Al-Ghouti).
Contents lists available at ScienceDirect
Journal of Cleaner Production
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https://doi.org/10.1016/j.jclepro.2023.137810
Received 19 November 2022; Received in revised form 1 June 2023; Accepted 14 June 2023

Journal of Cleaner Production 417 (2023) 137810
2
and central nervous system (Anku et al., 2017). The toxicity levels of
phenolic compounds range between 10 and 24 mg/L, and the lethal
blood concentration is ~1.5 mg/mL. Phenolic compounds can adversely
affect the environment by polluting soil and water bodies including
surface and groundwater (Alshabib and Onaizi, 2019; Anku et al., 2017;
Mohamad Said et al., 2021). Consecutively, these pollutants induce
changes in plant communities’ structure and bioaccumulate in birds and
sh, ultimately incorporating into food chains and adversely affecting
health (Garg et al., 2020; Mandeep et al., 2020). Fig. 1-C shows the
physicochemical characteristics of phenol and their maximum permis-
sible levels in different water bodies as recommended by the US EPA
(Alshabib and Onaizi, 2019). From these strict limits, it could be
deduced that it is necessary to sustain low concentrations of phenolic
compounds in water bodies to protect human and environmental health
(Raza et al., 2019).
Many derivatives of phenol are usually found in various water bodies
Abbreviations
AC Activated carbon
AGS Aerobic granular sludge
AOPs Advanced oxidation processes
CNTs Carbon nanotubes
CWAO Catalytic wet air oxidation
DESs Deep eutectic solvents
DNAPLs Dense Nonaqueous Phase Liquid
EPA Environmental protection agency
GO Graphene oxide
HRP Horseradish peroxidase
ILs Ionic liquids
IMO International maritime organization
LAC lignite-activated coke
LC
50
Lethal concentration 50
LLE Liquid-liquid extraction
LNAPLs Light Nonaqueous Phase Liquid
MBBR Moving bed biolm reactors
MNBs Micro-nano-bubbles
NAPLs Nonaqueous phase liquids
NF Nanoltration
NPRI National pollutant release inventory
RGO Reduced graphene oxide
RO Reverse osmosis
SLM Supported liquid membrane
TOC Total organic carbon
Fig. 1. A-Sources of phenolic compounds in water. B- Phenol concentration (%) in efuents of various industries (Raza et al., 2019). C- Physiochemical charac-
teristics of phenol and maximum permissible levels of phenol in water.
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Journal of Cleaner Production 417 (2023) 137810
3
including groundwater. These compounds come from nearby waste
dumping sites or spill sites. Phenol has a high solubility in water (83 g/
L), and has a short half-life in soils, thus it reaches groundwater easily.
Phenolic compounds enter water bodies and the aquatic environment
through direct discharge or runoff (Anku et al., 2017). When these
compounds reach water bodies, they transform into various in-
termediates. The transformation of phenolic compounds in the aquatic
environment is driven by various physical, chemical, and biological
interactions. In terms of physical interactions, phenol can interact with
nitric ions in the presence of UV radiation, leading to the formation of
intermediates such as 4-nitrophenol. Additionally, phenol photolysis
can result in the production of hydroquinone in the presence of charge
transfer complexes (Kinney and Ivanuski, 1969). On the other hand,
chemical interactions can also lead to the formation of intermediates, for
example, phenol can interact with hydroxyl radicals or nitric ions, which
leads to the formation of 2-nitrophenol (Moussavi, 1979). Additionally,
metal cations interact with various phenolic compounds, which leads to
their ionization and in turn enhances their solubility in water (Epa,
1979). Aerobic and anaerobic biological interactions also lead to alter-
ations in the structure of phenolic compounds, for instance, microor-
ganisms produce 4-chlorophenol upon encountering
4-chlorophenoxyacetic acid, and tetrachlorocatechol is generated as a
by-product of pentachlorophenol degradation. Additionally, chloro-
benzenes could also be degraded microbially, leading to the formation of
chlorocatechol (Anku et al., 2017).
A signicant portion of organic contaminants in nature exist as non-
aqueous phase liquids (NAPLs). NAPLs are organic contaminants that
are immiscible in water. NAPLs are grouped into two main categories
based on their density: Light Nonaqueous Phase Liquid (LNAPLs) and
Dense Nonaqueous Phase Liquid (DNAPLs). Due to their inherent
properties, these compounds persist in the pores of sediments as a
separate phase. Their fate in the marine environment is highly depen-
dent on their density. In wet porous media, mass transfer of NAPLs re-
quires the volatilization of these compounds into a gaseous phase,
following that, the gaseous NAPLs will dissolute into the aqueous phase,
and nally, the aqueous phase will sorb into the solid phase. In marine
sediments where water coats the soil particles, NAPL mass transfer into
the soil is not possible, and NAPLs need to be partitioned rst to the
aqueous phase, before sorbing into the sediments (Fitts, 2013; Lenhard
et al., 2005). Phenolic compounds are recalcitrant pollutants; therefore,
the accumulation of these contaminants and their intermediates in water
bodies and marine sediments is expected. Progressively, ocean oor
organisms such as bottom feeders that feed and forage near the ocean
bed will start to accumulate these compounds, and thus these pollutants
enter food webs and chains (Zhou et al., 2017). For example, chlor-
ophenols distribution patterns in marine sediments indicate that these
compounds could be accumulated in marine water by adsorbing the
suspended organic compounds due to their lipophilic nature (Xie et al.,
1986). Several freshwater and marine organisms showed high sensitivity
to phenol. For example, the highest sensitivity noted from marine or-
ganisms towards phenol was with Archaeomysis kokuboi where the LC
50
value was found to be 0.26 mg/L after 96 h (Noszczy´
nska and
Piotrowska-Seget, 2018). On the other hand, the freshwater organism
Cirrhinus Mmririgala recorded the maximum sensitivity with an LC
50
of
1.55 mg/L after 96 h (Duan et al., 2018). Due to the highest sensitivity,
the international maritime organization (IMO) has listed phenolic
compounds as the top 20 hazardous compounds that present high risks
to aquatic life (Panigrahy et al., 2022).
2. Phenol remediation technologies
The presence of phenol in industrial wastewater streams necessitates
its treatment to prevent adverse impacts on human health and the
environment. Moreover, effective treatment of these streams can
potentially yield a sustainable and renewable water source. A range of
physiochemical technologies, including distillation, nanoltration,
reverse osmosis, chemical oxidation, electrochemical oxidation, sol-
vents extraction, ozonation, advanced oxidation processes, photo-
catalytic oxidation, adsorption, and biodegradation, have been
employed for industrial wastewater treatment (Garg et al., 2020; Sar-
avanan et al., 2021a, 2012b; Villegas et al., 2016). Table 1 summarizes
the various technologies used for wastewater treatment and their ad-
vantages and disadvantages. Though these techniques have many ad-
vantages, they also have a few disadvantages mainly the high
operational and maintenance costs, high energy requirements, pre-
treatment requirements, low removal efciencies, scaling, fouling, lim-
itation due to selectivity, and pH level constraints, and use of hazardous
organic solvents (Gholami-Bonabi et al., 2020; Jim´
enez et al., 2018).
Additionally, the selection of the treatment process signicantly relies
on pollutant concentration, volume, and cost of the treatment process
(Guha Thakurta et al., 2018). Thus, in recent years, a shift towards the
use of environmentally compatible technologies in wastewater treat-
ment was noticed to reduce the operation costs and their impact on the
environment (Adeniyi and Ighalo, 2019).
This review aims to provide a comprehensive overview of the con-
ventional methods used for phenol wastewater treatment, explain the
mechanism involved in phenol removal, and emphasize green technol-
ogies that support sustainable practices. Additionally, this review pro-
vides inclusive insights into the various available options and serves as a
guide for selecting the most suitable technologies for treating phenol
wastewater considering factors such as initial pollutant concentrations,
costs, wastewater volumes, and environmental compatibility. Finally,
this review highlights the importance of adsorption technologies as a
cost-effective, sustainable, and environmentally friendly solution, which
could be adopted for the treatment of phenol-contaminated wastewater.
2.1. Physical treatment methods
2.1.1. Distillation
Many physical processes are used in wastewater treatment. These
processes do not require the use of chemicals. Steam distillation is one of
the physical recovery processes in which the steam volatile compounds
are volatilized, condensed, and collected in receivers. Steam distillation
separates immiscible liquids based on the volatility of the steam. In the
liquid phase, phenol has limited immiscibility with water, however, the
immiscibility disappears at a temperature of approximately 68 ◦C
(Saputera et al., 2021). Phenol can be distilled from polluted water by
the process, which begins when wastewater enters the distillation col-
umn, and when the temperature reaches around the boiling point of
water, and the pressure of 1 atm, the water evaporates and condenses
while the phenol is collected at the bottom of the column. This method
of treatment is suitable for highly contaminated wastewater owing to
the high operating costs and energy requirements (Gao et al., 2021).
A few studies investigated distillation processes in phenol removal as
noted in Table 2. Mohammadi & Kazemi investigated the use of a vac-
uum membrane distillation process for phenol wastewater treatment. It
was found that the optimum condition for phenol separation was a
temperature of 45 ◦C, a pH of 13, a concentration of 1000 mg/L, and a
pressure of 60 bar (Mohammadi and Kazemi, 2014). Additionally, the
distillation process could be combined with the extraction process to
recover phenolic compounds. These technologies use extraction solvents
to collect phenol out of the water as noted in Table 2, leading to
improved recovery rates. For example, in a study, phenol was recovered
using a steam distillation—extraction process using diethyl ether as the
extraction solvent. High recovery rates of about ~100% were achieved
at an initial phenol concentration of 30 mg/L and highly saline and
acidic conditions and similar trends were noted in other studies (Sapu-
tera et al., 2021).
It could be noted from these studies that distillation processes are
effective in separating various pollutants from wastewater and are
particularly benecial for pollutants with concentrations of ~1000 mg/
L. However, their application in wastewater treatment is not widely
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Journal of Cleaner Production 417 (2023) 137810
4
reported due to the high capital and operational costs associated with
them. Moreover, the use of these technologies results in the production
of concentrated wastes that require further treatment or disposal.
Therefore, the use of distillation processes is often not practical with
larger volumes of wastewater containing low levels of pollutants.
2.1.2. Membrane separation
In this process, a membrane is used to separate the components in a
solution by rejecting unwanted substances and allowing the others to
pass. These technologies can separate various pollutants such as salts,
dyes, and organic pollutants including phenol (Cevallos-Mendoza et al.,
2022). Membrane performance is highly dependent on membrane
properties, pollutant properties, and operating conditions as noted in
Fig. 2-A. This gure also shows the different types of membranes used in
water treatment and their applications. Membrane-based technologies
have advantages such as the absence of by-product generation, easy
installation, and low energy consumption. However, membrane-based
technologies also have a few barriers including stability and reus-
ability as these membranes have short lifetimes due to fouling (Obotey
Ezugbe and Rathilal, 2020).
Reverse osmosis (RO), nanoltration (NF), and ultraltration are all
different levels of membrane-based technologies that are being used for
phenolic wastewater treatment as reported in Table 2. Among these,
nanoltration is most used for the separation of phenol. In a study,
several nano-ltration membranes were tested for the removal of phenol
with an initial concentration of 1000 mg/L, it was noted that the highest
removal rate was achieved using DSS-HR98PP polymeric membrane
with 80% rejection at neutral pH levels (B´
odalo et al., 2009). Nano-
ltration could also be coupled with adsorption to improve the separa-
tion of phenol, especially with membranes that have larger pore sizes
where phenol can pass on easily. For example, a study showed that
phenol removal using adsorption/nanoltration was around 31% in the
presence of nanoparticles and only achieved 4% removal in the absence
of adsorption i.e., around 675% improvement compared to using
Table 1
Advantages and disadvantages of phenolic compounds remediation technologies.
Technology Mechanism Advantages Disadvantages
Physical
methods
Distillation Distillation is a technique for separating the
components from a miscible uid mixture by
selective evaporation and condensation
according to boiling points.
High recovery rates with high purity and
ability to reuse, fast separation, feasible for
highly concentrated solutions.
High operational costs, energy-intensive,
not appropriate for low concentration
solutions, and it might have impurities.
Membranes Membrane technologies separate
contaminants from water based on properties
such as size or charge.
Operates under various conditions and
concentrations, is selective, and allows the
use of hybrid systems.
High capital and operational costs, scaling,
and fouling of membranes.
Nanoltration A pressure-driven membrane process that
removes solutes with molecular weight in the
range of 200–1000 g/mol, typically from
aqueous mediums.
Effective in removing pollutants, simple to
operate, and maintain.
Cannot treat higher loads of pollutants,
membrane replacement is needed with
time.
Reverse osmosis RO membranes allow water to pass through
while rejecting solutes, such as low molecular
weight organic materials.
Highly effective at removing contaminants
and is energy efcient.
Cannot treat higher loads of pollutants,
high capital and operational costs, and
fouling.
Chemical
methods
Chemical oxidation Chemical oxidation requires the use of an
oxidizing agent in the treatment of
wastewater to oxidize the organic pollutants.
Oxidation of high and low concentrations
of pollutants could also be paired with
other technologies such as UV to improve
removal efciency.
The use of large quantities of hazardous
oxidizing chemicals, and partial oxidation
leads to the formation of by-products.
Electrochemical
oxidation
Electrochemical oxidation uses electric
current or a potential difference between two
electrodes (anode and cathode), with which
hydroxyl radicals or oxidizing species can be
generated and used for the oxidation of
pollutants.
Use fewer chemicals and harsh materials,
easy to operate and maintain.
This technology is dependent on electrical
energy, and it requires electrode
replacement frequently which has higher
operational costs.
Ozonation Ozonation produces hydroxyl (OH) radicals
through the decomposition of ozone (O
3
) that
are used in the oxidation of organic
pollutants.
Eliminate and reduce bioactivity, toxicity,
and biological effects of pollutants.
Expensive and not very effective for COD
and TOC reduction leads to the formation of
disinfection byproducts.
Photocatalysis Photocatalysis is a process in which light
energy is used to generate radicals that are
used for pollutant degradation
Oxidation could be done using solar
irradiation which is abundant and benign.
Occurs at ambient temperature and
pressure conditions with cheap, nontoxic,
and noncorrosive catalysts.
Most semiconductor materials are not
visible light active, require high band gap
energy and agglomeration of nanoparticles
can occur making it difcult to separate or
reuse catalyst from aqueous solution.
Extraction with
solvents
Extraction is a method that separates
compounds based on their relative solubilities
in two different immiscible liquids.
High concentrations of pollutants can be
extracted and thus collected as a product,
simple to operate.
Extraction solvents are volatile, ammable,
toxic, and expensive, and require the use of
large solvent volumes, multiple extraction
steps, and a long extraction period.
Adsorption Adsorption is a process where pollutants are
removed due to electrostatic interactions,
π
-
π
interactions, hydrophobic interactions, van
der Wal forces, and hydrogen bonding
between the adsorbent and adsorbate.
Cost-effective especially when waste
materials are used. Can be used for different
concentrations depending on the
adsorption capacity of the adsorbent.
Less effective at high concentrations, and
difcult to separate adsorbent from
adsorbate. Spent adsorbents should be
recycled or treated before discarding into
the environment.
Biological
methods
Phytoremediation Phytoremediation includes phytoextraction
and accumulation, and phytodegradation by
certain resistant plants.
Eco-friendly, cost-effective, and can
promote biodiversity.
Cannot operate at extremely high
concentrations due to toxicity and the
requirement of large spaces.
Enzymes Enzymes oxidize pollutants into simpler
organic compounds.
Operates on low and medium
concentrations, with no toxicity issues. The
degradation of pollutants is rapid and
selective.
Enzyme purication is expensive as the
process is dependent on reaction conditions
such as temperature and pH as well as
possibilities of enzyme deactivation.
Biodegradation Biodegradation requires the use of
microorganisms that consume pollutants as
carbon sources leading to the decomposition
of pollutants.
Eco-friendly, cost-effective, and works with
low and high-strength wastewater.
Sometimes, it cannot operate at extremely
high concentrations (toxicity) and requires
larger space, chemical treatment, and
aeration.
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Journal of Cleaner Production 417 (2023) 137810
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Table 2
Examples of green physical, chemical, and biological technologies for phenol remediation.
Type of
remediation
technique
Example Pollutant Initial
Pollutant
concentration
Removal
effectiveness
Remediation mechanism(s) Conditions Reference
Physical technologies
Filtration Double ltration with AC Phenol 0.3–0.1 mg/L 99% retention AC porosity, surface area,
and chemistry play
important role in the
adsorption of phenol.
– Fuentes et al. (2018)
RO-membrane Polyamide thin lm
composite RO
Phenol 1000 mg/L ~70% retention The electrostatic repulsion
between phenol and
membrane
pH 6.5 Mnif et al. (2015)
Chemical technologies
Extraction DESs-Choline chloride
(ChCl)-glycerol
Phenol – 98.3% H-bonds between the DES
and phenolic compounds
T 30 ◦C Yi et al. (2019)
Ozonation Ozone reactor Phenol 2000 ~35% Oxidation by ozone and
hydroxyl radicals
pH 9
T 25 ◦C
Haag and Hoigne
(1983)
Electrochemical
oxidation
Sn-doped Ti/PbO
2
Phenol 500 89% The electric current
generates hydroxyl radicals
and other oxidizing species
to degrade phenol
pH 5.5
T 30 ◦C
Li et al. (2013)
Electrochemical
oxidation
Ti/SnO2–Sb2O3–Nb2O5/
PbO2
Phenol 500 78% The electric current
generates hydroxyl radicals
and other oxidizing species
to degrade phenol
pH 7
T 20 ◦C
Yang et al. (2008).
Extraction Terpenoids and
hydrophobic eutectic
solvents
Phenol 500 mg/L >95% separation The presence of acceptor H-
bond regions in the solvent
would increase phenols
afnity.
T40 ◦C Rodríguez-Llorente
et al. (2020)
Adsorption AC Phenol 100 mg/L 434 mg/g Π-
π
interaction, electron-
donor–acceptor complex
formation, and H-bonding
pH 6, T 25 ◦C Mojoudi et al.
(2019)
Extraction by Ionic
liquid
1-ethyl-3-methyl
imidazolium
cyanoborohydride,
Phenol 100 mg/L Selectively
extracted 95% of
the phenol
Intermolecular interaction
between [BH
3
CN] anion
and phenol molecules.
T 30–80 ◦C Mathews et al.
(2019)
Ionic liquid 4-butyl-1-methyl
pyridinium bis
(triuoromethyl sulfonyl)
imide
Phenol 15,000 mg/L 96% extraction of
phenol
Hydrophobic interactions
between ILs and Phenol
pH 6, 25 ◦C Sas et al. (2020)
AOPs
Photocatalysis
Polymer, CNT, TiO
2
–NH
2,
and UV
Phenol 10 mg/L 99%
photodegradation
in 7 min
An increase in electron-hole
pairs on the catalyst surface
leads to higher
concentrations of reactive
hydroxyl radicals, which
lead to phenol degradation.
pH 5 Mohamed et al.
(2020)
AOPs
Photocatalysis
Acid-modied TiO
2
nanoparticles
Phenol 55 mg/L 99%
photodegradation
in 23 h
Phenol can be hydroxylated
by OH radicals and the
formation of Lewis acid Ti
3+
sites on the TiO
2
surface via
hydrogenation leading to
higher phenol degradation
T 20 ◦C Ling et al. (2015)
AOPs O
3
-calcium peroxide Phenol 5 mg/L 97% Calcium peroxide produces
hydrogen peroxide that
degrades phenol
pH 3
T 25 ◦C
Honarmandrad et al.
(2021)
AOPs - Fenton Fe (II)/H
2
O
2
Phenol 100 mg/L 100% degradation
in 9 min
Phenol oxidation is carried
out by hydroxyl radicals
generated from a reaction
between hydrogen peroxide
and iron (II) salts
pH~3, T25 ◦C Esplugas et al.
(2002)
Biological methods
Phytoremediation Hydrilla verticillata Phenol 100 mg/L 99% removal in 7
days
Transformation and
detoxication by
peroxidases and ROS
12-h
photoperiod
at 21/16 ◦C
Chang et al. (2020)
Enzymatic Laccase Phenol 376.44 mg/L 96% removal in 30
min
Catalytic oxidation of
phenol
pH 5, T50 ◦C Asadgol et al. (2014)
Enzymatic Peroxidase Phenol 100 mg/L 97.4% Catalytic oxidation of
phenol
pH (4.0–9.0),
T (20–60 ◦C).
Gonz´
alez et al.
(2006)
Biodegradation Pseudomonas
fredriksbergsis
Phenol 700 mg/L 90% removal in 96
h
Hydroxylase and oxygenase
enzymes are used to
biodegrade phenol and use
it as a carbon source
pH 7, T28 ◦C Aljbour et al. (2021)
Biodegradation Acinetobacter lwofi Phenol 500 mg/L 100% removal in
12 h
Enzymes such as
Hydroxylase and catechol
1,2-dioxygenase break
down phenol via the ortho-
cleavage pathway.
pH8, 33 ◦C Xu et al. (2021)
(continued on next page)
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Table 2 (continued )
Type of
remediation
technique
Example Pollutant Initial
Pollutant
concentration
Removal
effectiveness
Remediation mechanism(s) Conditions Reference
Hybrid technologies
Adsorption +
nanoltration
Silver nanoparticles-
Nanoltration
Phenol 180 mg/L ~100% First adsorption of phenol
by nanoparticles (increase
in the particle size) followed
by ltration with an NF
membrane,
pH7, - Naidu et al. (2016)
Adsorption +UV GO-UV Phenol 100 mg/L 95.95% hydrogen bonding,
π
-
π
interactions, electrostatic
interaction, H
2
O
2
oxidation.
pH6, T 35 ◦C Al-Ghouti et al.
(2022)
Distillation +
extraction
Steam distillation with
diethyl ether
Phenol – 91.8% Water evaporates leaving
phenol to be extracted
-, T 50 ◦C Bart´
ak et al. (2000)
Adsorption +
photocatalysis
AC/TiO
2
/CeO
2
Phenol 763 mg/L 50.91% Photocatalysis and
adsorption by the
negatively charged surface.
pH 8, T30 ◦C Dalanta and
Kusworo (2022)
Adsorption +UV AC-UV Phenol – 99% AC accelerated the
degradation of organic
compounds by catalyzing
O
3
to generate hydroxyl
radicals (⋅OH).
pH7, T25 ◦C Xiong et al. (2020)
Fig. 2. A- Overview of ltration technology and its role in the removal of pollutants. B- Overview of Phenol chemical oxidation process, and C- Extraction of phenol
using DESs.
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Journal of Cleaner Production 417 (2023) 137810
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nanoltration alone (Naidu et al., 2016).
Reverse osmosis is another emerging membrane-based technology
that removes organic impurities including phenolic compounds from
water by using pressure to force water molecules through a semi-
permeable membrane (Srinivasan et al., 2011). Since organic matter can
cause clogging in the reverse osmosis membranes, nanoltration tech-
nologies are used in advance. The coupling of reverse osmosis and
nanoltration is essential for the stabilization of pressure uctuations
that occur when using reverse osmosis systems alone, leading to higher
treatment efciencies. Several studies investigated the use of RO in
phenolic compound removal (Srinivasan et al., 2011). In a study,
phenol-contaminated wastewater with an initial concentration of 10
mg/L was subjected to RO. It was found that with RO alone, the phenol
concentrations in the efuent reached around 1.30 mg/L at pH 4, with a
removal efciency of 87%. To further improve the treatment process,
the wastewater stream was subjected to pretreatment using granular
activated carbon. It was noted that the overall efciency of the RO
system increased to 99% (Ipek, 2004). Li et al. conducted a study to
compare the effectiveness of nanoltration and reverse osmosis mem-
branes in removing phenol from synthetic wastewater. Different types of
nanoltration membranes (NF-90, NF-97, and NF-99) and reverse
osmosis membranes (RO-98pHt and RO-99) were tested on wastewater
containing phenol with levels below 1000 mg/L. The researchers found
that both nanoltration and reverse osmosis membranes had advantages
and disadvantages. Nanoltration showed a low rejection rate of phenol
(between 0.41 and 0.72) but had a high maximum ux rate (up to 180
L/m
2
/hr), while reverse osmosis had a high rejection rate (0.81) but a
low minimum ux rate (only 60 L/m
2
/hr). Generally, the study showed
that NF and RO could be considered effective methods for removing
phenol from wastewater (Li et al., 2010).
It could be implied from these studies that membrane technologies
could be used in the treatment of phenolic wastewater with concentra-
tions of approximately 1000 mg/L with retention levels of ~95–100%.
However, similarly to distillation technologies, membrane technologies
have high operational costs due to fouling and maintenance and high
rates of energy consumption, therefore their application could be
hindered.
2.2. Chemical treatment methods
2.2.1. Chemical oxidation
Phenol-contaminated wastewater could also be treated chemically.
Chemical oxidation is a destructive method that involves the use of
oxidizing agents in the treatment process to oxidize organic pollutants.
Many chemicals are used for the oxidation of organic pollutants
including hydrogen peroxide, chlorine, chlorine dioxide, permanganate,
and ferrate, these oxidants degrade organic compounds such as phenol
into simpler compounds or water and carbon dioxide as illustrated in
Fig. 2-B. These oxidants generate radicals that rapidly and non-
selectively react with organic compounds leading to their degradation
(Pal, 2017). Chemical oxidation processes are very advantageous owing
to their effectiveness, ability to operate under various conditions of pH
and temperature, and low operating costs, however, these processes
form recalcitrant pollutants during the process of oxidation, leading to
the generation of by-products, mostly when the oxidation process is
incomplete (Peings et al., 2015).
A comparative study investigated sulfatoferrate, potassium per-
manganate, and calcium hypochlorite’s ability to oxidize phenol. In this
study, the initial phenol concentration was 30 mg/L, and the experiment
was conducted at pH 9. After an hour, it was found that sulfatoferrate
was able to degrade 57% of phenol, potassium permanganate degraded
70% of phenol, and calcium hypochlorite had a removal efciency of
61%. It is worth mentioning that the use of permanganate could increase
manganese concentration in water whereas using hypochlorite could
lead to the formation of several chlorinated by-products. The authors
stated that using sulfatoferrate is safer due to the lack of by-product
generation (Peings et al., 2015). Matta et al. reported that synthetic
chloride green rust-H
2
O
2
was able to degrade 100% of phenol with an
initial phenol concentration of 50 mg/L in 1 min by hydroxylation/ox-
idation. Additionally, these compounds degraded 62% of total organic
carbon (TOC) in 24 h at neutral pH of 7 (Matta et al., 2008). Chamberlin
et al. used potassium permanganate to oxidize phenol with an initial
concentration of 125 mg/L at a high temperature of 95 ◦C. This oxida-
tion process achieved a maximum removal of 62%. Due to this partial
removal, another experiment was conduction for the degradation of
phenol using hypochlorite while maintaining alkaline conditions to
sustain the oxidation process. It was noted that the maximum removal
was achieved by adding 5000 mg/L of hypochlorite to a solution con-
taining 100 mg/L of phenol (Chamberlin et al., 1952). It could be
determined from the studies that it is vital to ensure that complete
oxidation is achieved especially in chlorination at lower pH levels since
partial chlorination could lead to the formation of chlorophenols, which
are more recalcitrant than phenol itself.
The Fenton oxidation process is another type of chemical oxidation
process in which the ferrous or ferric cation decomposes hydrogen
peroxide to generate strong oxidizing agents capable of degrading
organic and inorganic substances including phenolic compounds. Yavuz
et al. investigated the ability of the Fenton process to degrade phenol.
Around 98% removal was achieved at an initial phenol concentration of
500 mg/L, hydrogen peroxide (H
2
O
2
) concentration of 3000 mg/L,
ferrous sulfate (FeSO
4
) concentration of 1500 mg/L, and pH of 2.3. The
authors stated that under acidic conditions, the reaction between H
2
O
2
and Fe
2+
generated hydroxyl radicals that oxidized phenol molecules.
Although phenol removal was sufcient, the COD removal efciency
was limited (Yavuz et al., 2007).
Catalytic wet air oxidation (CWAO) is a process that involves
oxidizing a liquid waste stream with air or oxygen at elevated temper-
atures and pressures in the presence of a catalyst. One potential appli-
cation of CWAO is in conjunction with trickle bed reactors (TBRs),
which are often used for the treatment of wastewater. The combination
of CWAO and TBRs can enhance the efciency of wastewater treatment,
particularly when dealing with complex organic compounds and other
contaminants (Candan and Ayten, 2021). A trickle-bed reactor is a type
of xed-bed reactor that can be used to remove organic compounds from
wastewater (Makatsa et al., 2019). The reactor consists of a vessel lled
with a bed of catalyst particles over which the wastewater is passed. The
reactor operates in a continuous mode and the wastewater ows
downward over the catalyst bed while air or oxygen is introduced to
facilitate the oxidation of phenol and its derivatives. The effectiveness of
a trickle bed reactor for removing phenol and its derivatives depends on
several factors, including the catalyst type, catalyst loading, hydraulic
retention time, and initial concentration of the wastewater (Al-Huwaidi
et al., 2021). They offer several advantages including high contact ef-
ciency between the wastewater and catalyst, low catalyst cost, and low
energy requirements. Additionally, A trickle-bed reactor (TBR) can be
easily scaled up to meet the requirements of different wastewater
treatment applications (Mohammed et al., 2016). In a study, phenol was
oxidized in a TBR over Al–Zr- a pillared clay catalyst under various
experimental conditions. It was found that a complete conversion of
phenol with an initial concentration of 1000 mg/L was achieved in 180
min, with a temperature of 160, a gas velocity of 0.012 m/s, and a
pressure of 10 bar (Makatsa et al., 2019). Other studies also showed that
combining TBR technologies with other processes such as reverse
osmosis can signicantly improve the treatment process (Al-Obaidi
et al., 2018).
Chemical oxidation processes offer a solution for treating wastewater
contaminated with initial concentrations of up to 500 mg/L. These
processes have been shown to achieve oxidation efciencies ranging
from approximately 55%–100%. However, several challenges hinder
their application including the need for large quantities of oxidants and
the generation of by-products. In addition, these methods have safety
and environmental concerns due to the use of strong chemical agents.
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Journal of Cleaner Production 417 (2023) 137810
8
2.2.2. Electrochemical oxidation
Electrochemical oxidation is another method of wastewater treat-
ment that does not require the use of chemical reagents in the oxidation
process. Direct electrochemical oxidation oxidizes pollutants by their
adsorption to the anode surface by charge transfer reaction. On the other
hand, indirect oxidation uses an intermediate redox reagent present in
the solution that prevents the electron transfer between the electrode
and the pollutant preventing electrodes from fouling (Martín-Pozo et al.,
2022).
Many studies have investigated the use of electrochemical oxidation
of phenol as noted in Table 2. Saratale Rijuta et al. investigated phenol
oxidation using a Ti/PbO
2
electrode with the addition of Fe
2+
. Complete
removal of phenol with an initial concentration of 250 mg/L was
observed at 50 ◦C, pH 2, and a potential difference of 5 V (Saratale Rijuta
et al., 2016). Abou-Talab et al. recently investigated phenol removal
from petroleum wastewater using graphite electrodes as an anode and
stainless-steel electrodes as a cathode. Complete phenol removal from
an initial concentration of about 6.8 mg/L was achieved within 15 min
and under a current density of 3 mA/cm
2
(Abou-Taleb et al., 2021).
It is clear that electrochemical oxidation processes are a viable op-
tion for treating phenol-contaminated wastewater with initial concen-
trations of up to 500 mg/L. Such processes can achieve removal
efciencies ranging from 78% to 100%. However, limitations regarding
their application include the high equipment costs and energy re-
quirements, and the selection of appropriate anodic materials.
2.2.3. Ozonation
Ozonation is a common method of water treatment in which hy-
droxyl radicals (•OH) generated from ozone (O
3
) decomposition are
used to oxidize pollutants including phenolic compounds (Manas,
2021; Pavithra et al., 2017). In alkaline mediums, ozone can act as a
strong oxidizing agent, with a redox potential higher than hypochlorite,
and it has higher solubility in water when compared to oxygen. Ozone
can interact with pollutants and degrade them through two pathways.
Directly, where contaminants interact with ozone or indirectly through
hydroxyl radicals that are generated from the decomposition of ozone in
the aqueous medium as illustrated in Fig. 3-A. Wastewater treatment
with ozonate is preferred to low phenol concentrations to reduce costs,
therefore it is used as a nal disinfection step in wastewater treatment
facilities (Sorokhaibam and Ahmaruzzaman, 2014). Numerous variables
such as ozone dose, pH, and temperature impact the removal and
degradation of phenolic compounds in such treatments. Many studies
investigated the use of ozone in phenol removal. Turban & Uzman used
an ozone bubble column containing phenol with initial concentrations
ranging from 50 mg/L – 100 mg/L and stated that ozone was capable of
complete phenol oxidation within 40 min (Turhan and Uzman, 2008).
Wang et al. investigated the use of a self-design ozone generator for
phenol degradation. It was noted that the degradation rate relied on the
initial phenol concentration and reaction times. For example, 99%
removal was achieved in 30 min when the initial phenol concentration
was 100 mg/L, however, at an initial concentration of 3000 mg/L,
around 480 min were needed to completely degrade phenol (Wang et al.,
Fig. 3. A-Phenol ozonation process and reaction pathways. B- Classication of MNBs based on their size C-MNBs surface charges. D-Phenol degradation via free
radicals generated from the collapse of MNBs.
A. Bibi et al.

Journal of Cleaner Production 417 (2023) 137810
9
2016). As mentioned, it is important to maintain alkaline conditions
during the treatment process to ensure enough production of hydroxyl
radicals for the complete oxidation of phenol.
One advancement in the ozonation process is through the use of
micro-nano-bubbles (MNBs). Microbubbles are small-sized bubbles that
have a diameter ranging from 10 to 15
μ
m whereas nano-bubbles have a
diameter of fewer than 200 nm as illustrated in Fig. 3-B, collectively,
these bubbles are called micro-nano-bubbles (Wang et al., 2020). MNBs
have many important characteristics that could be benecial for envi-
ronmental applications. This includes their small size, large interfacial
area, high internal pressure, low rising velocity, and long residence time.
Additionally, MNBs demonstrate high stability, an exceptionally large
surface-to-volume ratio, a high rate of oxygen dissolution, and a great
ability to generate free radicals (Hu and Xia, 2018; Nirmalkar et al.,
2018). In general, MNBs are negatively charged at various pH conditions
as indicated by their zeta potential However, the surface charge of MNBs
can be modied by the surrounding environment’s composition,
providing the opportunity for ne-tuning their properties for optimal
performance in different applications as noted in Fig. 3-C. This feature of
MNBs is important for the interactions between these bubbles and other
compounds or pollutants in water as it determines the magnitude of
electrostatic attraction and repulsion in water treatment systems (Jia
et al., 2013; Takahashi, 2005). The bursting of the MNBs leads to the
generation of free radicals (Liu and Tang, 2019). During bubble collapse,
the Zeta potential usually increases and leads to the formation of free
radicals (Takahashi et al., 2007; Xiong et al., 2018). These generated
radicals have strong oxidizing power against a wide range of recalcitrant
organic pollutants such as phenol as shown in Fig. 3-D, leading to its
complete decomposition.
Many studies examined the use of MNBs in the elimination of re-
fractory pollutants such as phenols. Wu et al. explored the use of
microbubbles generated by cavitation for the degradation of phenol and
it was compared between two conventional bubbles over various pH
ranges. This study found that phenol degradation was more sufcient
using microbubble ozonation where it required only 50% of ozone as
compared to the conventional reactor. The half-time of phenol degra-
dation using a conventional reactor was 18 min whereas, at the micro-
bubble reactor, it took only 7 min at an initial phenol concentration of
9411 mg/L. Moreover, phenol degradation was found higher at high pH
levels since phenol dissociates at such pH levels, and the ozone mass
transfer rate was around 1.5 times higher compared to the conventional
bubbles reactor (Wu et al., 2019). Another important feature of MNBs
technology is that it could be used effectively in the in-situ treatment of
groundwater. Compared to MNBs, macro-bubbles have a short lifetime
and a small zone of inuence, thus their use in water treatment is less
efcient. Several studies have shown that MNBs can increase the dis-
solved O
2
levels in groundwater by increasing the mass transfer rate and
thus lead to the oxidation of dissolved organic pollutants (Hu and Xia,
2018; Liu and Tang, 2019).
To sum up, in alkaline conditions ozone can efciently treat many
organic pollutants, including phenol at concentrations up to 9000 mg/L
both in-situ and ex-situ. Many advancements have been reported in the
literature, yet further improvements are required to optimize the use of
MNBs with air instead of pure O
2
or O
3
to make the process more
economically feasible while maintaining adequate removal efciencies.
2.2.4. Advanced oxidation processes
Advanced oxidation processes (AOPs) are chemical technologies that
remove and oxidize soluble organic efuents from water based on the in-
situ generations of strong oxidants such as hydroxyl and sulfate radicals.
AOPs have been developed to overcome the limitations of ozone treat-
ment alone. These processes combine ozone with UV, catalysts, and
hydrogen peroxide to optimize the degradation of pollutants into water
and CO
2
(Ghime and Ghosh, 2020). In acidic and neutral environments,
ozone has low solubility and stability, thus the formation of hydroxyl
radicals becomes slow, and by using hydrogen peroxide or UV with
ozone, more hydroxyl radicals are produced to accomplish higher
oxidation levels. However, studies are required to understand the
different scenarios in terms of by-product generation, pollutant
decomposition pathways, and toxicity of end products.
An extensive study investigated several AOPs including O
3
, O
3
/
H
2
O
2
, UV, UV/H
2
O
2
, UV/O
3
, O
3
/UV/H
2
O
2
, Fe
2+
/H
2
O
2
and photo-
catalysis processes for the oxidation of phenol in aqueous medium with
an initial concentration of around 100 mg/L. Among all, the Fenton-
based process demonstrated the fastest removal rates for phenol in
wastewater with removal efciency ranging from 32 to 100% within 9
min. It was noted that the photocatalysis took the longest period of
degradation (150 min) where the removal efciency ranged between
42% and 77%. However, the authors stated that from an economical
point of view, ozonation was found to be more cost-effective (Esplugas
et al., 2002). In another study, phenol with an initial concentration of
1000 mg/L was photo-oxidated using UV-H
2
O
2
. At ambient temperature
and pressure and a pH of 3.5, it was found that 99% of phenol was
degraded within 90 min with an oxidant concentration of 34,000 mg/L
(Primo et al., 2007). It is evident that some AOPs could treat higher
concentrations of pollutants compared to ozonation alone; however, it
also requires large quantities of oxidants and has higher energy
requirements.
2.2.5. Photocatalysis
Photocatalysis is an advanced oxidation process in which light en-
ergy is used to drive chemical reactions. When the catalyst absorbs en-
ergy either from UV light or visible light, an excited electron (−)/hole
(+) pair is formed with photons of energy greater than or equal to the
band gap energy. Due to their activated state, the electron and hole
perform chemical reduction and oxidation of oxygen and water leading
to the production of reactive species as illustrated in Fig. 4. Holes can
oxidize adsorbed organic matter or water, producing hydroxyl radicals
HO•. Electrons, on the other hand, reduce O
2
to the superoxide radical
anion (O
2
•). These O
2
•and OH•radicals will in turn degrade organic
and inorganic pollutants including phenol into CO
2
and H
2
O. Commonly
used catalysts include zinc oxide and titanium dioxide. Photocatalysis is
a promising cost-effective technology that can effectively degrade a
wide range of toxic compounds including phenol. One of the main ad-
vantages of this technology is that it requires low energy to operate.
Since it uses sunlight as a source of energy, it signicantly reduces the
cost of operation compared to traditional treatment methods. Addi-
tionally, photocatalysis is a recyclable process, which means that the
catalyst can be recovered, and reused, thus reducing the need for
additional materials and minimizing waste generation. This results in
lower operating costs and makes the technology more sustainable in the
long run (Ansari et al., 2019; Mohamed et al., 2020; Pawar and Lee,
2015).
Many studies have investigated the use of photocatalysis in the
degradation of phenol as noted in Table 2. Belekbir et al. photocatalyzed
phenol using nanosized metal-impregnated TiO
2
under near-UV light. It
was found that complete phenol degradation under NUV-Vis light irra-
diation using TiO
2
–Cu was achieved with an initial phenol concentration
of 50 mg/L, neutral pH, and ambient temperatures. This process was
found to be slower when compared to the use of UV light, however,
economically, a longer time using near UV–Vis radiation counter-
weighted the expenses of utilizing UV irradiation sources (Belekbir
et al., 2020). Even though many studies have analyzed the effects of
photocatalysis pollutant degradation but faced an issue with how to
accelerate the interfacial charge transfer economically. Thus, Pardeshi &
Patil used a slurry batch reactor to degrade phenol using sunlight instead
of UV lamps (approximately 4% UV light and 43% visible light) as an
energy source to reduce operational costs. The experiments were per-
formed at ambient temperatures in the presence of ZnO (250 mg/100
mL), and it was found that phenol solution with an initial concentration
of 75 mg/L was entirely degraded to CO
2
and H
2
O within 8 h (Pardeshi
and Patil, 2008). From these studies, it is evident that photocatalysis
A. Bibi et al.

Journal of Cleaner Production 417 (2023) 137810
10
could be used to treat low-strength wastewater with high efciency.
2.2.6. Liquid-liquid extraction
Liquid-liquid extraction (LLE) is another chemical method of sepa-
ration that is based on the compound’s relative solubility. LLE methods
have been used traditionally for the removal of organic compounds from
wastewater. The advantages of these methods are contributed to their
non-destructive nature, high selectivity and solubility, and their ability
to treat different concentrations of pollutants (Mohamad Said et al.,
2021). Organic compounds including phenol can dissolve in organic
solvents such as butanol, octanol, dichloromethane, alcohols, esters,
ethyl acetate, butanone, benzene, and toluene (Cabl´
e et al., 2022; Patel
and Desai, 2022). In addition, a mixture of extraction solvents such as
2-octanone and n-octanol could also efciently recover phenol from
water (Wang et al., 2022). Many studies have shown that different sol-
vents are capable of efcient separation of phenol from wastewater.
Recently, Patel and Desai extracted phenol from synthetic and phar-
maceutical wastewater using toluene, it was found that 20% toluene
removed 60% and 68% of phenol synthetic and real wastewater
respectively, with an initial phenol concentration of 50,000 mg/L and a
pH of 7 (Patel and Desai, 2022). Jiang et al. also extracted phenol with
an initial concentration of 6000 mg/L from wastewater using octanol as
a solvent with an extraction efciency of 99%; moreover, it also can
recycle the extracting solution, leading to reduced operating costs. The
authors noted that phenol extraction efciency was high due to higher
interaction energy between the phenol and the extractant, leading to
stronger hydrogen bond formation and more stable complex formation
(Jiang et al., 2003).
Even though extraction processes have been used efciently for
many years in the recovery of various organic compounds, these pro-
cesses have several disadvantages mainly the toxic nature of the organic
solvents. Therefore, research has been focusing on deep eutectic solvents
(DESs) as a group of solvents that have a lower impact on the environ-
ment. Deep eutectic solvents are considered green solvents that could be
used for the extraction of various compounds including phenol (Cabl´
e
et al., 2022). These solvents are non-volatile with high thermal stability
and can readily dissolve many organic and inorganic compounds. Many
studies investigated deep eutectic solvent (DESs) abilities to extract
phenol as noted in Table 3. Sas et al. recently studied DESs based on
(menthol, thymol, and organic acids) for the extraction of phenolic
compounds. It was noted that DES based on menthol had an extraction
efciency of 70% with various initial phenol concentrations ranging
from 5 mg/L to 1500 mg/L. In this study, it was mentioned that the
higher extraction efciencies were due to higher hydrophobic in-
teractions between the pollutants and the DESs as illustrated in Fig. 2-C
(Sas et al., 2019).
Ionic liquids (ILs) are another group of green solvents that are
composed of ions with melting points below 100 ◦C (Sood et al., 2021).
Ionic liquids are composed of organic cations and inorganic anions
(Shang et al., 2021). Ionic liquids are considered promising green sol-
vents to many organic and inorganic compounds due to their low vapor
pressures, thermal stability, and non-volatile nature (Shang et al., 2021;
Welton, 2004). Due to their benign nature, the usage of ILs in waste-
water treatment is considered viable, environmentally friendly, and an
alternate approach to toxic chemical solvents (Khraisheh et al., 2021).
Over the years, ILs have been used in water treatment as a part of
Liquid-liquid extraction technologies, in membrane technologies, and
nally in adsorption processes. These studies focused on identifying the
effects of ILs in the separation of various compounds including phenolic
compounds.
Many studies have shown that different ILs could be applied for the
extraction of phenolic compounds as noted in Table 2. In one study, ILs
1-ethyl-3-methyl imidazolium cyanoborohydride were used in a liquid-
liquid extraction system to separate phenol at an initial concentration of
100 mg/L. It was found that ILs extracted selectively around 95% of the
phenol from the synthetically prepared organic oil mixture of toluene
and benzene showed acceptable efciencies up to 6 times of reuse
(Mathews et al., 2019). In another study, several types of ILs containing
hydroxyl, benzyl, and dialkyl groups were tested for their ability to
extract phenolic compounds including phenol. The authors noted that
the extraction efciency of these phenols depended on factors such as
pH levels, the ratio of phases, salt contents, and the structure of the IL
used. For instance, around 82% extraction efciency was achieved at pH
9 using 1-butyl-3-(8-hydroxyoctyl) imidazolium hexauorophosphate.
The authors also state that the non-ionized phenols were found to be
more easily transferred into the ILs phases as a result of by
hydrogen-bonding and phenols’ hydrophobicity (Fan et al., 2014).
The supported liquid membrane (SLM) process is a new extraction
technology used for the treatment of wastewater containing organic
compounds. SLM processes have several advantages over solvent
extraction including high selectivity and speed, large mass-transfer
force, minimal extractant loss, and reduced capital and operating costs
(Kocherginsky et al., 2007). In a study, phenols were separated and
recovered from an aqueous solution by a green membrane system where
vegetable oil was used as a green polypropylene-hollow-ber supported
liquid membrane. It was noted that phenols separation with an initial
concentration ranging between 4800 and 5200 mg/L reached 95% (Mei
et al., 2020). In another study, polypropylene hollow ber membrane
(PP-HFM), was improved by grafting heptadecauoro-1,1,2,
2-tetradecyltrimethoxysilane (FAS) and SiO
2
. The modied SLM
showed signicant enhancement in stability and achieved average
phenol removal of 75% in 16 days (Sun et al., 2017).
Solvent extraction is one of the most efcient processes for treating
high-strength wastewater containing up to 50,000 mg/L of phenol, with
Fig. 4. UV photocatalysis of phenol using TiO
2
.
A. Bibi et al.

Journal of Cleaner Production 417 (2023) 137810
11
Table 3
Examples of various types of adsorbents and their phenol adsorption capacities and mechanisms.
Adsorbent Conditions
T (◦C), pH
Adsorption
capacity (mg/
g)
Removal
(%)
BET Surface
area (m
2
/g)
Proposed removal mechanisms Reference
Biopolymers
Magnetic chitosan 20, 3 51.68 97 – At low pH, less repulsion between the negative chitosan
and phenols and more Van der Waals attraction,
hydrophobic interaction, and hydrogen bonding
Salari et al. (2019)
Crab shell chitosan – 59.3 – 191 Mesopore lling and ionic interactions Francis et al.
(2020)
Polyacrylamide/starch hybrid
hydrogels
– 21 – – Interactions between NH/OH of the polymeric structure
and OH of phenol
Dutra et al. (2021)
Cellulose-based hydrogel 25,6 80.71 – – The external layer of
α
-cyclodextrin is enriched for
phenol by inclusion complexation, with hydrogen
bonding as the driving mechanism. The internal
pyridine group has a strong afnity for phenol, which
encourages additional phenol adsorption.
Guo et al. (2022)
Bacterial cellulose nanobers 25,8 146 97 342.1 Electrostatic interactions were found to contribute to
the generation of specic recognition binding sites.
Derazshamshir
et al. (2020)
Carbon-based adsorbent
AC (coconut shell) 55,4 144.93 – 620.49 The lower content of total oxygen-functional groups on
AC strengthens the
π
-
π
dispersive force interaction
between the phenol molecule and AC sample, which
leads to higher phenol adsorption capacity.
Zhang et al. (2016)
AC (coconut shell) 25,7 212.96 – 1025.02 Electrostatic interactions and surface area and pore
diffusion led to higher adsorption
Xie et al. (2020)
GO 30,7 10.23 74 312.00 Interactions between phenoxide ions and GO surface Mukherjee et al.
(2019)
GO-PPY 25,6 201.40 – – Ion exchange,
π
–
π
electron donor-acceptor (EDA)
interaction, hydrophobic interaction, and Lewis’s acid-
base interaction.
Hu et al. (2015)
Nitrogen-doped Reduced
graphene oxide
30,6 155.82 99 245.71 The removal efciency was attributed to
π
-
π
and
hydrophobic interactions
Zhao et al. (2021).
Horseradish Peroxidase
Immobilized on Graphene
Oxide/Fe
3
O
4
(H
2
O
2
)
25,7 – 95 – Peroxidases catalyze the oxidation of organic and the
reduction of (H
2
O
2
) by donating electrons that bind to
substrates to break them into harmless components
Chang et al. (2016)
Carbon nanotubes-PEG 20,6 21.23 100 – The dispersive interactions between the aromatic rings
of phenols and the basal plane of CNT/PEG replenished
with a high
π
-electron density could contribute to
phenol removal
Bin-Dahman and
Saleh (2020).
Polymeric resins 20, - 22.0 – – Ion exchange was the predominant process responsible
for phenols removal
Víctor-Ortega et al.
(2016)
Acrylic ester-based crosslinked
resin
25,7 1000 – – Adsorption by BMS-resin is attributed to the H-bonding
between the ester groups on the surface of resin beads
and the OH groups of phenol
Qiu et al. (2014)
CNT-DESs 25, 7 298 – – The hydrophobic interaction/
π
-
π
interactions resulted
in the adsorption of phenol
Lawal et al. (2019)
Poplar AC 28, - 625.00 – 2711 Surface area and aromaticity could increase
π
–
π
dispersion interaction as well as the reaction between
the active site and phenol. Mesopores also enhance mass
transfer as well as pore accessibility
Hwang et al.
(2017)
Industrial waste materials
AC- oily sludge 25,6 434.78 – 2263
π
-
π
interaction, electron-donor acceptor complex
formation, and the hydrogen bonding
Mojoudi et al.
(2019)
Fly ash- PDDA 25,7 13.05 95 – The coating of cationic polyelectrolyte lm onto FA
introduced surface polarity which enhanced the mass
transfer of phenol to the PDDA-FA surface
Oyehan et al.
(2020)
Sludge Based AC-MgAlFe-LDH 35,6 216.76 – 320.58 The main phenol adsorption mechanisms
π
–
π
interactions because of the phenol aromatic ring
interaction with that of the SBAC-MgAlFe-LDH via
charge transfer, dispersive force, and polar attractions.
OH groups of the adsorbent and hydrophobicity of
phenol also play important roles in the adsorption
process
(Mu’azu et al.,
2021)
Claried sludge 35,7 1.05 63 78.54 The presence of oxides and silica (−) on claried sludge
surface had a high afnity towards adsorption of the OH
group of phenol leading to hydrogen bonding
Mandal and Das
(2019)
Agricultural Waste materials
Activated Biochar 20, - 303.00 95 881 High surface and pore volume and oxygenated
functional groups, especially carbonyl groups were
responsible for phenol removal.
Braghiroli et al.
(2018).
Date palm frond biochar -, 6 17.38 – 245.82 Electrostatic interactions between the biochar and
phenol molecules
Fseha et al. (2023)
Sulfuric acid-treated pea shells 25,7 125.77 – 7.07 The electrostatic interactions were very high between
phenol and the adsorbent surface
Mishra et al.
(2021)
(continued on next page)
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Journal of Cleaner Production 417 (2023) 137810
12
removal efciency ranging from 60 to 99%. Therefore, it is commonly
used as a pretreatment step by petroleum and coke conversion in-
dustries. However, the use of conventional organic solvents in solvent
extraction is associated with the use of toxic and hazardous substances.
This necessitates the search for greener alternatives such as DESs and ILs
as these have shown promise as green solvents. Yet further research is
needed to develop solvents with extraction efciencies that are com-
parable to those of conventional organic solvents.
2.2.7. Adsorption of phenolic compounds
One of the common methods for the removal of phenolic compounds
is adsorption. Adsorption can be dened as the process in which an ion
or a molecule called adsorbate sticks or attaches to the surface of a solid
called adsorbent. The process of adsorption can occur through physical
or chemical mechanisms. Physical adsorption involves the adsorbate
adhering to the surface of the adsorbent through weak physical forces
such as van der Waals interactions. In contrast, chemical adsorption
typically involves stronger covalent bond formation between the
adsorbent and the adsorbate (Akeremale et al., 2023). Like other
chemical reactions, the adsorption processes are inuenced by many
factors as summarized in Fig. 6-B. The main feature that determines the
capacity of an adsorbent is the surface area per volume, thus porous
substances such as activated carbon and clays are considered funda-
mental adsorbents (Artioli, 2008). Over the years, substantial efforts
have been made to produce adsorbents with high selectivity, efciency,
environmental compatibility, and cost-efciency. Many of these adsor-
bents are being used in wastewater treatment due to their broad appli-
cability and benign nature (Pavithra et al., 2017; Thakur and
Kandasubramanian, 2019). Adsorption is an easy and energy-efcient
process that can remove low and high concentrations of numerous
contaminants including phenol from wastewater (Rout et al., 2023).
Moreover, adsorbents could be regenerated/recycled, making this
method more sustainable and cost-effective (Sajid et al., 2018; Wang
et al., 2019). Previous studies stated that activated carbon is an effective
type of adsorbent used for phenol remediation while other adsorbents
used are titanium oxide, activated alumina-modied bentonite, GO, and
biopolymers (Mohamad Said et al., 2021).
It could be noted from Table 3 that phenol can be removed from
wastewater using various types of adsorbents including biopolymers,
carbon-based adsorbents, clays and muds, industrial and agricultural
waste-based adsorbents, and hybrid adsorbents. It was also noticed that
the adsorption capacity of these materials could reach up to 1000 mg/g.
Table 3 (continued )
Adsorbent Conditions
T (◦C), pH
Adsorption
capacity (mg/
g)
Removal
(%)
BET Surface
area (m
2
/g)
Proposed removal mechanisms Reference
Neem leaves 30,3 74.90 97 370 At low pH, the H
+
ions suppress the ionization of
phenol, leading to phenol adsorption
Mandal et al.
(2020a,b)
Ziziphus leaves 25,6 15.00 – – In the acidic medium, the surface of the adsorbent is
dominated by positive charges which increases the
attraction between phenolate and the surface, thus,
enhanced adsorption is observed
Al Bsoul et al.
(2021)
Pomegranate Peel Carbon 25,7 148.38 98 – Phenol interacts with the functional groups of the
adsorbent
Afsharnia et al.
(2016)
AC- Ceiba speciosa wastes 25, 7 156.70 – 842
π
–
π
interactions, hydrogen bonds, surface area, and
pores were responsible for phenol adsorption
Franco et al.
(2021)
AC- lignocellulosic wastes
(Sugarcane bagasse)
25,4 158.96 – 1053 Phenol molecules will interact with the positively
charged surface of carbon through electrostatic
attraction.
El-Bery et al.
(2022)
Clays and muds
Red mud 30,8 49.30 87 300 Pore capture, hydrogen bond formation between the red
mud surface and phenol, and electrostatic interaction
between metal oxides (positive) and phenolate ions all
lead to phenol removal
Mandal et al.
(2020a,b)
Na–bentonite 30,3 8.76 – 2.04 Strong electrostatic interactions between its adsorption
site and phenol.
Asnaoui et al.
(2020)
Clay 30,6.5 30.32 – – Clay particles consist of active sites bearing negative
charge which are neutralized by ions, therefore,
enhancing the diffusion of phenol ions
Nayak and Singh
(2007)
HDTMA-modied clay 40,7 18.8 – – The negatively charged phenolate anion is
electrostatically attracted by the positively charged
HDTMA–bentonite surface.
Gładysz-Płaska
(2017)
HCL activated mud 25,6 8.156 – 20.7 Interaction between metal oxides and polar phenol
molecules
Tor et al. (2009)
Hybrid adsorbents
Chitosan and silica 25,8 149.25 86 1700 Removal was higher because phenol was unionized, and
the dispersion interaction was predominant
Fathy et al. (2020)
MOF/GO 25,7 212.76 – – The presence of GO leads to the adsorption of phenol via
hydrogen-bond and
π
-
π
interactions
Karamipour et al.
(2021)
Synthetic zeolite modied with
chitosan
25,6 5 – 53.5 The hydroxyl groups in the chitosan chain form
hydrogen bond with the –OH group present in phenol
molecules
Bandura et al.
(2020)
Graphene oxide-coated biochar 30,7 23.47 – – Factors like hydrogen bonding, electrostatic interaction,
van der Waals force, and intra-particle diffusion all lead
to phenol adsorption
Manna et al.
(2019)
Horseradish peroxidase
immobilized
Chitosan–halloysite hybrid-
nanotubes
25,7 – 98 CTS–HNT
55.2
Peroxidases catalyze the reduction of H
2
O
2
and the
oxidation of organics by donating electrons that bind to
other substrates such as ferricyanides and ascorbate, to
break them into harmless components
Zhai et al. (2013)
Laccase immobilized on
copolymer nanobers
50,5 40.33 ~90 –
π
-
π
interaction, hydrogen bonding between the electron-
donating atoms and the hydrogen atoms of phenolic
pollutants, and oxidation by laccases, all contribute to
phenol removal
El-Aassar et al.
(2020)
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Journal of Cleaner Production 417 (2023) 137810
13
Biopolymers are polymers that are made of biological monomers that
could degrade in the environment. These materials are produced by
different organisms such as microorganisms, plant biomass, and agri-
cultural wastes, and are composed of proteins, polysaccharides, and fats
as summarized in Fig. 6-A (Pandian et al., 2021). Recently, biopolymeric
composites have gained great attention due to their high surface area
and functionality, structural stability, environmental compatibility,
diverse applicability, and durability leading to superior adsorptive
removal of various pollutants including phenol as noted in Table 3
(Udayakumar et al., 2021; Yaashikaa et al., 2022). Biopolymers could be
prepared either by polymerization or through the termination of raw
material. This process led to the production of good-quality biopolymers
with enhanced features. However, the use of chemicals such as solvents
could be problematic since these chemicals could be hazardous leading
to the production of by-products that would require careful handling
and further treatment. Some studies also suggested the use of green
solvents such as deep eutectic solvents and ionic liquids to avoid the use
of large qualities of organic solvents.
Even though adsorption processes have many signicant advantages
over physiochemical methods of wastewater treatment. However, it is
vital to take into consideration the cost-effectiveness and fate of the
adsorbents after use. In terms of economic value, since biopolymeric
adsorbents could be made from agricultural wastes, the initial costs of
the adsorbents could be acceptable, yet surface modication and labor
costs could lead to an increase in expenses (Yaashikaa et al., 2022). For
instance, a cost analysis was done for a modied pectin composite that
was estimated to reach up to $70/kg. Whereas other studies showed that
membranes made from biopolymeric materials (Arabic gum) were more
cost-effective compared to conventional membranes (Aji et al., 2020).
Additionally, the recyclability and regeneration of these composites
could also signicantly reduce the overall treatment costs. On the other
hand, the environmental fate of spent adsorbents should also be
considered carefully. Most biopolymers could be degraded in the envi-
ronment by the actions of microorganisms and phytoremediators that
oxidize these compounds and use them as carbon and nitrogen sources
for their growth and metabolism. However, the adsorbed pollutants
should be desorbed before the release of these composites into the
environment (Yaashikaa et al., 2022). Several studies have investigated
the use of biopolymers in the removal of phenolic compounds from
aqueous solutions. In one study, researchers used chitosan beads
modied with sodium alginate and CaCl
2
. In this study, the maximum
phenol sorption capacity was found to be 108.7 mg/g. The modication
of the biopolymer has resulted in improved stability and sorption ca-
pacity (Nadavala et al., 2009). In another study, biopolymer-based
biochar was used for the adsorptive removal of phenol and 2-nitrophe-
nol. The adsorptive composite showed monolayer removal of pollut-
ants. In addition, 2-nitrophenol was adsorbed more effectively than
phenol due to the functional group (NO
2
) that had a strong interaction
with the adsorbent. This study also investigated the simultaneous
removal of both pollutants in a binary system, which also showed similar
trends where higher removal was achieved with 2-nitrophenol, indi-
cating that it hindered the adsorption of phenol due to its stronger
attraction to the receptor site (Li et al., 2019).
Activated carbon is one of the most commonly used adsorbents for
the removal of organic pollutants (Allahkarami et al., 2023). However,
its high cost and the environmental problems related to the regeneration
and disposal of its waste dictate the use of more suitable and eco-friendly
alternatives. AC is generated conventionally from coal and petroleum
products and unconventionally from various agricultural and industrial
wastes. These unconventional sources reduced the cost tremendously
and are considered renewable sources of materials, thus, making the
process of adsorption economically feasible (Issabayeva et al., 2018). In
a study, commercial AC-SP1000 was used for phenol removal from
synthetic and real wastewater. At neutral pH, ambient temperatures,
and an initial phenol concentration of 5000 mg/L, the adsorbent was
able to remove about 92% of phenol in the case of real syngas scrubber
wastewater and around 99% of phenol from synthetic wastewater,
demonstrating an adsorption capacity of 270 mg/g (Catizzone et al.,
2021). Recently, several studies were conducted to evaluate low-cost
and environmentally friendly materials for wastewater treatments as
noted in Table 3. Most of these studies considered low-cost materials,
mainly biosorbents that were generated from agricultural wastes, which
were found to be able to reduce the availability and concentration of
some organic compounds mainly phenolic compounds, polycyclic aro-
matic hydrocarbons, and industrial dyes. Additionally, industrial wastes
are also being used as a source of adsorbents with high removal ef-
ciencies (Adeniyi and Ighalo, 2019; Frezzini et al., 2019).
As mentioned, ILs are being studied as part of liquid-liquid extraction
technologies to estimate their efciency in phenol extraction. (Moha-
mad Said et al., 2021). Several recent studies have investigated the use
of ILs for the removal of organic compounds from wastewater. The use of
ILs is environmentally compatible, however, it is not cost-effective
because these solvents are not highly recyclable and thus will be used
in large quantities (Gholami-Bonabi et al., 2020). To overcome this
issue, researchers have investigated the use of supported ILs where
various materials are used to stabilize or immobilize the ILs (Laurent
et al., 2008). For instance, in a study, amine-functionalized IL-modied
graphene oxide was used as an adsorbent for phthalates from water. In
this study, the modied GO was packed into a xed bed column for
solid-phase extraction. The study showed that the modied composite
was able to extract contaminates with a minimum recovery of 95% and
with high reproducibility (reaching up to eight times) without
decreasing the efciency of the absorbent in extraction. This study also
concluded that ILs-modied GO can be used as a high-quality adsorbent
in low-pressure columns (Zhou et al., 2016).
To further improve the adsorptive capabilities and overcome some of
the issues associated with the use of GO in adsorption studies, a mag-
netic nanocomposite made of graphene oxide was modied by an IL (1-
amino-3-methylimidazole chloride (LI-MGO)) and used to adsorb
phenol from contaminated water. GO surface modication with ILs is
possible due to the existence of carboxylic groups on the GO surface. It
was noted in this study at the surface area increased from 64.32 m
2
/g to
110.44 m
2
/g. This is due to the fact that the surface of the magnetic
graphene oxide became rougher with IL modication. Furthermore,
under optimal conditions, around 95% of phenol was removed from the
solution, making this adsorbent a highly efcient and cost-effective
process with high environmental compatibility (Gholami-Bonabi et al.,
2020). Fig. 5 summarizes the adsorption process using M-GO-ILs. It
could be concluded that adsorption is a highly economical and effective
method to remove phenol from wastewater (Melaibari et al., 2023).
It could be noted from these studies that many adsorbents have
extremely high phenol adsorption capacities that range from 1 mg/g to
1000 mg/g and removal efciencies ranging from 50 to 99%. Addi-
tionally, these adsorbents could be prepared from waste materials,
ranking the treatment process as the most economic method of phenol
treatment (Magdy et al., 2021). However, to reduce the environmental
impacts of these adsorbents, further improvements and studies are
required to evaluate the regeneration potential and their disposal.
2.3. Biological degradation of phenolic compounds
Biological remediation of phenol has also been exploited to mitigate
the negative effects associated with physiochemical techniques. These
biological methods have numerous vital properties including their high
specicity, accessibility, and the absence of production of harmful by-
products. Various plant species and microorganisms have been used to
biologically remove pollutants including phenolic compounds from
wastewater. The biological treatment of phenolic compounds in waste-
water can be done in three different ways. First, by phytoremediation
where plants are used to extract, immobilize, contain, or degrade
organic and inorganic pollutants (McCutcheon and Jørgensen, 2008).
Second, by using Enzyme-based methods where enzymes extracted from
A. Bibi et al.

Journal of Cleaner Production 417 (2023) 137810
14
various living organisms such as plants, fungi, and bacteria are used in
water treatment (Alshabib and Onaizi, 2019). Third, bioremediation
where water is treated using microorganisms such as bacteria, fungi, and
yeast that utilize the pollutants as their source of carbon and degrade it
into CO
2
and H
2
O (Anku et al., 2017; Saravanan et al., 2021).
2.3.1. Phytoremediation
Phytoremediation is the process of degradation or accumulation of
harmful pollutants into less harmful substances by plants as illustrated
in Fig. 7-A. It is considered an eco-friendly, cost-effective method to
remove, detoxify or immobilize different organic and inorganic pollut-
ants. In addition to that, phytoremediation usually does not require
costly inputs or expressive operational costs, and can also assist in
enhancing biodiversity, thus is considered a more acceptable and
preferred choice for remediation (Agostini et al., 2011). Phytor-
emediation of organic pollutants is usually done in two ways. First is
phytoextraction and degradation where the pollutants are directly taken
up and sequestered or degraded by plants. Several plant enzymes are
involved in the sequestration and transformation of organic pollutants
such as cytochrome P450 and glutathione-S-transferase. The second
process is through plant-assisted rhizo-remediation. The pollutants are
degraded by enzymes such as laccase, dehalogenase, and
nitro-reductase, which are secreted by the plant or its rhizosphere mi-
crobial community. Soil microbes use plant root exudates (e.g., sugars,
organic acids, etc.) as energy and carbon sources and in turn help in the
degradation of pollutants. In addition, plant exudates can also increase
the bioavailability of pollutants by increasing the solubility of these
pollutants (Chen et al., 2013). To increase the phytoremediation ef-
ciency and achieve higher degradation rates, especially when the con-
taminants are recalcitrant it is recommended to use multiple plants
species, this will be a result of increased microbial functional diversity
and biomass, and higher enzymatic activities (Wei and Pan, 2010).
Fig. 7-A shows the various mechanisms involved in the phytor-
emediation of phenolic compounds. Detoxication of xenobiotics such
as phenol starts with transformation where pollutants become more
soluble. Enzymes such as peroxidases and laccases catalyze these re-
actions. Few studies showed that the crude enzymes of plants were able
to oxidize phenol by cleaving the structural ring and forming muconic
acid and catechol as an intermediate. Further oxidation will lead to the
formation of fumaric acid. Following transformation, pollutants will be
conjugated to the plant’s endogenous compounds. This step is also vital
because it increases pollutants’ mobility and hydrophilicity. Examples of
enzymes used in this step are glutathione s-transferase and N-glucosyl-
transferase. Conjugation can only partially reduce toxicity since the
soluble pollutants will be accumulated in various plant tissues such as
vacuoles with the help of ATP-binding cassette transporters. Following
that, pollutants are either processed or moved out of the cell by exocy-
tosis where they can accumulate in cell walls or apoplasts. These bound
residues cannot be released from the plant matrix by solvent extraction.
Finally, some pollutants including phenols can be excreted by the plant’s
leaves into the surrounding air (Agostini et al., 2011).
Enhanced phytoremediation of phenol could be achieved using
Fig. 5. A-B: An overview of batch adsorption of phenol using magnetic adsorbent modied with ionic liquids (Gholami-Bonabi et al., 2020), and C- Mechanisms of
phenol adsorption onto graphene oxide (Bibi et al., 2022).
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Journal of Cleaner Production 417 (2023) 137810
15
several approaches such as studies that involve screening to identify the
most suitable plant species, the use of transgenic plants with enhanced
capabilities for phenol degradation, increasing soil microbial diversity,
and optimization of agronomic practices to increase biomass production
and subsequently, phenol degradation. Several studies have investigated
the use of plants in the remediation of phenol-contaminated wastewater
as noted in Table 2. Ib´
a˜
nez et al. investigated a legume specie Vicia sativa
L. ability to remediate phenol-contaminated wastewater. Phenol toler-
ance of the plant was assayed at different stages of growth, and it was
noted that 30-day-old planets were able to tolerate phenol concentra-
tions of 100 mg/L and with 60% removal efciencies within 4 days. The
activities of antioxidative enzymes, including peroxidase and ascorbate
peroxidase, signicantly improved with increased phenol concentration,
whereas superoxide dismutase activity, malondialdehyde, and H
2
O
2
levels did not change. The study suggested that Vicia sativa L. could be
considered an exciting tool in the eld of phytoremediation as it has an
efcient protection mechanism against phenol-induced oxidative dam-
age and could tolerate and remove phenol with concentrations up to
100 mg/L without phytotoxic effects (Ib´
a˜
nez et al., 2012). Sosa Alderete
et al. investigated the use of transgenic tobacco hairy roots (HR) system,
which expressed basic peroxidase genes from tomatoes (TPX1 and
TPX2) for phenol removal. The results showed that TPX1 is engaged in
phenol removal not only when it was overexpressed in tomatoes, but
also when it was expressed in other plants including tobacco. The
removal efciency using transgenic HR clones in the presence of H
2
O
2
was optimal for TPX1/TPX2 clone, with a maximum value of about 90%
at an initial phenol concentration of 100 mg/L, which represented an
increment of about 15% compared with the wild type controls. The
greater efciency of TPX2 transgenic hairy root demonstrated that this
peroxidase also participates in the removal of phenol (Sosa Alderete
et al., 2009).
It could be concluded from these studies that phytoremediation
could be used effectively with removal efciencies ranging from 60 to
99%, for large volumes of wastewater with low phenol concentrations
up to 100 mg/L. Further studies are required to investigate the resistant
plant’s ability to degrade contaminants and the possibilities of their
application in constructed wetland systems. It is worth mentioning that
phytoaccumulation is a slower process and it requires further treatment
or disposal of plant materials to prevent any impacts on the
environment.
2.3.2. Enzymes based remediation
Enzymes are a specialized class of proteins (catalysts) responsible for
catalyzing numerous chemical reactions. Compared to inorganic cata-
lysts, enzymes are more effective. In addition, enzymes show a greater
specicity of the effect (Blanco and Blanco, 2017). The use of
enzymatic-based technologies is considered a cost-effective and sus-
tainable approach in the treatment of various pollutants. Microorgan-
isms such as bacteria, cyanobacteria, fungi, and actinomycetes as well as
several plant species are considered sources of different useful enzymes
that can be used in the remediation of pollutants (Singh et al., 2021).
Several groups of enzymes are employed in the degradation of pol-
lutants including hydrolases, oxygenase, oxidoreductases, laccase, and
peroxidases (Singh et al., 2021). These enzymes can selectively and
effectively degrade various pollutants at much faster rates when
compared to other reactions. Another important advantage of enzymatic
systems is the fact that enzymes can remove pollutants even in unfa-
vorable conditions (e.g., temperature, pH, pollutant concentration, etc.)
where bacteria might be inhibited. Additionally, enzyme-based tech-
nologies eliminate the time required for bacterial biomass generation.
Compared to biodegradation, enzymes can be used in various conditions
and will degrade pollutants into harmless products (Anku et al., 2017).
Peroxidases are enzymes that are utilized extensively in phenol reme-
diation due to the presence of heme cofactor or redox-active cys-
teine/selenocysteine residues in their active sites. Several studies
investigated peroxidase from various sources for the removal of phenol
from contaminated water. Kurnik et al. investigated the use of peroxi-
dases produced from potato pulp waste by-products in the removal of
phenol from synthetic and industrial wastewater. Phenol removal ef-
ciency reached 95% with an initial phenol concentration of 94.11 mg/L
where the enzymes maintained their activity at a pH range of 4–8 and
were stable over a wide temperature range. Similar high efciency was
also noted with industrial efuents where 90% removal was achieved
(Kurnik et al., 2015). In another study, peroxidases extracted from an
Fig. 6. A-An overview of the roles of biopolymers in water treatment, and B- Factors inuencing the adsorption process.
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Journal of Cleaner Production 417 (2023) 137810
16
invasive plant (Prosopis juliora) were used for phenol remediation with
an initial phenol concentration of approximately 40 mg/L, where crude
peroxidases were found capable of degrading phenol within 30 min and
with efciencies higher than 90% for textile and leather industrial
wastewaters. This study showed that agricultural waste materials could
be used as an important source of potent enzymes that are highly ef-
cient in the remediation of phenolic compounds (Garg et al., 2020). Such
studies emphasize the importance of nding alternative sources of raw
materials to promote sustainable production practices.
Enzymes could also be immobilized on carbon-based materials such
as activated carbon and graphene oxide, which could be more cost-
effective and can enhance their recovery, stability, and reusability in
water treatment processes. Many studies investigated the use of immo-
bilized enzymes in the removal of phenol from wastewater as noted in
Table 3. In one study, laccase was immobilized on metal-chelated chi-
tosan nanoparticles, and around 82% of phenol with an initial concen-
tration of 20 mg/L was degraded within the initial 4 h compared to free
laccase that degraded 80% of phenol after 12 h. The activity of the
composite increased up to 96% in the presence of a redox mediator
(ABTS). Laccase immobilization preserved the enzymatic activity over a
wider pH range and showed a shift toward higher temperatures
(30–40 ◦C) compared to the free enzymes (30 ◦C). In addition, the
composite retained about 50% of the initial activity after eight reaction
cycles (Alver and Metin, 2017). Besharati Vineh et al. immobilized
horseradish peroxidase (HRP) by covalent bonding to reduced graphene
oxide (RGO). The authors stated that the pH range and temperature
were signicantly improved compared to the free enzyme. Additionally,
the removal efciency was 100% and 55% for the immobilized HRP and
free HRP respectively at an initial phenol concentration of 2500 mg/L
(Besharati Vineh et al., 2018). On the other hand, Abdollahi et al. re-
ported that tyrosinase nano-biocatalyst particles were able to degrade
more than 70% of phenol with an initial concentration of 2500 mg/L
within 4 h. At a lower initial concentration of 250 mg/L, the removal
efciency reached up to 100% for up to 3 cycles after which a decrease
in removal efciency was noted where it reached to about 55% removal
at the 7th cycle (Abdollahi et al., 2018).
As mentioned, enzyme-based processes were developed to overcome
the toxicity issues found in living systems. It was found that these pro-
cesses selectively and effectively treat low concentrations of phenol up
to 2500 mg/L with removal efciencies ranging from 55% to 100%.
Furthermore, the immobilization of enzymes in certain cases signi-
cantly enhanced pollutant degradation process efciency, durability, the
ability to treat higher-strength wastewater, and more importantly cost
efciency.
2.3.3. Microbial bioremediation
Many microorganisms can break down organic compounds into
harmless products. These microorganisms use organic compounds such
Fig. 7. A- Possible mechanisms involved in phenolic compounds phytoremediation, enzymes-based remediation, and bioremediation. B- Pathways of phenol
degradation under aerobic conditions, and C-Graphical representation of granular activated sludge and the role of various microbial populations in the degradation of
phenolic compounds.
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Journal of Cleaner Production 417 (2023) 137810
17
as phenol as their main carbon source and in turn, convert them into CO
2
and H
2
O. The advantages of these methods include their low operational
cost and environmental compatibility (Almasi et al., 2021; Anku et al.,
2017). Numerous gram-negative bacteria can utilize phenol as their sole
carbon source, including members of the three primary genera Pseudo-
monas, Alcaligenes, and Acinetobacter (Tian et al., 2017). Bacteria
belonging to the genus Acinetobacter and Pseudomonas produce
important enzymes such as phenol hydroxylase, which helps in the
degradation of phenolic compounds (Cafaro et al., 2004; Gu et al.,
2016). According to the literature, species belonging to the genus
Pseudomonas and Bacillus are capable of phenol degradation by
meta-cleavage pathway since these organisms have a broad range of
catabolic enzymes involved in the process of phenol degradation (Pan-
igrahy et al., 2022; Sarwade and Gawai, 2014). Both aerobic and
anaerobic microorganisms are capable of phenol degradation (Almasi
et al., 2018; Dargahi et al., 2017). In aerobic degradation, phenol hy-
droxylase enzymes catalyze the oxygenation of phenol to form catechol.
Following that, a ring cleavage adjacent to or in between the two hy-
droxyl groups of catechol is created. Phenol hydroxylases can be simple
avoprotein monooxygenases or multicomponent hydroxylases. Cate-
chol is oxidized via the ortho-cleavage pathway by catechol 1,2-dioxyge-
nase (carbon bond between two hydroxyl groups). The nal product of
the pathway is a molecule that can enter the tricarboxylic acid cycle.
Once the ring is opened, the degradation of the phenol can proceed as
noted in Fig. 7-B (Mahiudddin et al., 2012). In the absence of oxygen,
phenol can be degraded by the carboxylation in the para-position and
the formation of 4-hydroxybenzoate. After that, thioesterication of
4-hydroxybenzoate to co-enzyme A allows consequent ring reduction,
hydration, and ssion. Para-carboxylation appears to be involved in the
anaerobic degradation of several aromatic composites (van Schie and
Young, 2000).
Fig. 7-B illustrates a typical aerobic-activated sludge reactor which is
commonly used for organic waste treatment including phenol. These
technologies use recircled microbial communities to degrade phenol in
aerated systems and are capable of withstanding phenol concentrations
up to approximately 2000 mg/L (Hussain et al., 2015). Many studies
suggest that microbial consortium application for the remediation of
phenolic compounds is a promising technique since mixed microbial
populations can lead to higher tolerance to toxic pollutants, further, it
also increases the efciency of the degradation process by improving the
synergetic activities of various microbial organisms that secrete various
metabolites and enzymes. In such cases of co-metabolism, less toxic
intermediate by-products are accumulated, unlike in the case of pure
cultures (Patel & Kumar, 2016). Poi et al. used bacterial consortia
consisting of 22 cultures including species belonging to Pseudomonas sp.
Bacillus sp., and Acinetobacter sp. It was noted from this study that the
biolm-producing community was capable of remediating
phenol-contaminated wastewater with a concentration of 407 mg/L
using a bio-trickling reactor (Poi et al., 2017). Currently, many studies
are conducted to investigate the ability of sequential anaerobic anoxic
aerobic processes. Studies also show the signicance of moving bed
biolm reactors (MBBRs) that are carbon-based in the fact they provide
a protected surface where diverse groups of microorganisms can accu-
mulate. The coupling of activated carbon with activated sludge has
many benets including the adsorption of pollutants and superior shock
resistance. According to a study, two MBBRs were operated using
different carriers’ lignite-activated coke (LAC) and activated carbon
(AC) to estimate phenol removal. In this system, phenol was used as the
main carbon source in the rst 3 stages of treatment, an initial decrease
in degradation was noticed, however, the performance was stabilized
probably due to the biolm formation and the adsorption capacities of
the used carriers reaching up to 96% removal of phenol. It was noticed
that the LAC-based MBBR performed more efciently when phenol
levels were increased; nonetheless, both reactors had a similar tolerance
limit to phenols (around 450 mg/L). LAC-based MBBR also demon-
strated improved shock loading resistance at higher ammonia levels,
leading to removal that is more efcient on phenols (88.68% vs 94.61%)
when compared to the AC-based MBBR. The higher impact resistance
was formed due to the resilient adsorption capabilities of LAC. In
addition, the sludge ocs were enhanced in terms of compactness, sta-
bility, size, and size distribution, all leading to improved and enhanced
resistance against high-concentration shocks and toxicity. In LAC-based
MBBR, the degradation of phenols was dominated by excellent cooper-
ation among core microbes. Facultative anaerobes Cloacibacterium and
Hydrogenophaga contributed to phenol ring cleavage and enhanced
denitrication. The predatory bacteria Bdellovibrio had a role in nitrogen
xation and biomass conversion by converting complex biopolymers to
extracellular substances leading to more compacted ocs. As anaerobes
on biolm, Tissierella exhibited tolerance to higher levels of ammonia
and stimulated methanogens. For archaea, Thaumarchaeota proportion
of LAC was double the AC, especially for Nitrososphaera, which are
exceptional nitriers. Due to the adaptability of Comamonas belonging
to Burkholderiales, these species were found to be vital because it was
capable of phenol biodegradation, nitrication, and denitrication.
Moreover, they demonstrated interspecic cooperation with other bac-
teria. These species also produced biopolymers and created ocs for
protection from toxicants and predators. Similar to Comamonas, Pseu-
domonadales were also of importance in phenols and ammonia degra-
dation (Zheng et al., 2019). It could be concluded from these studies that
co-metabolism is essential for the complete degradation of phenolics and
for achieving higher toxicity tolerances (Tran et al., 2013).
Another promising technology in the eld of phenolic wastewater
treatments using aerobic bacteria is aerobic granular sludge (AGS)
(Nancharaiah and Sarvajith, 2019). AGS can promote the degradation of
phenolic compounds and increase toxicity tolerance due to the fact that
these granules consist of highly dense, physiologically diverse, and
spatially heterogenic microbial communities. The micro-environment of
AGS is protected by the secretion of polysaccharides on the surface, thus
reducing the toxic effects of pollutants as illustrated in Fig. 7-C.
Exo-polysaccharides have many functional groups such as C=C, OH, and
C=O that reduce phenol toxicity by enhancing the aggregation of AGS.
In addition to that, polysaccharides also help in the biosorption of pol-
lutants by providing efcient electrostatic forces, thus ensuring the
degradation process. He et al. investigated the use of an AGS sequencing
batch reactor for the simultaneous removal of phenol, nitrogen, and
phosphorus from saline wastewater. It was noted that initial phenol
concentration affected the removal of other pollutants by inhibiting the
activities of the heterotrophic denitriers and stimulating phosphorus
removal indicating the importance of co-metabolic activities in the
removal of pollutants. It was noted that the reactor was able to degrade
phenol completely with a concentration of up to 100 mg/L by the action
of multiple bacterial species and their roles in the production of extra-
cellular polymeric substances (He et al., 2021). Ho et al. investigated the
use of AGS for high-strength phenol wastewater. It was noted that
conventional activated sludge’s ability to degrade phenol was inhibited
at phenol concentrations above 3000 mg/L, however, when acclimated
granules were used, effective degradation of phenol was achieved
without severe inhibitory effects at a concentration up to 5000 mg/L.
The authors also noted that by using acclimated granules, a shorter lag
phase or faster degradation rate can be achieved compared to unaccli-
matized sludge (Ho et al., 2010).
It is evident that microbial biodegradation could be used effectively
to treat medium and high-strength wastewater containing up to 5000
mg/L of phenol. Compared to phytoremediation and enzyme-based
remediation this technology is more effective and could achieve
88–100% removal efciencies especially when acclimatized microbes
are used simultaneously in the treatment process.
3. Efciency, sustainability, and economic feasibility of phenol
treatment technologies
Various technologies can be used for the removal of phenol from
A. Bibi et al.

Journal of Cleaner Production 417 (2023) 137810
18
wastewater, each with its unique advantages, implications, and limita-
tions as summarized in Fig. 8-B. For instance, technologies such as
distillation, membrane ltration, chemical oxidation, electrochemical
oxidation, ozonation, and advanced oxidation processes can be highly
effective, but also expensive and energy intensive. On the other hand,
technologies such as liquid-liquid extraction and adsorption can be more
cost-effective but can produce additional waste streams. Additionally,
approaches such as phytoremediation and microbial bioremediation can
be environmentally friendly and sustainable but can be slow and limited
by factors such as pollutant concentration and operating conditions.
Overall, the choice of technology will depend on the specic charac-
teristics of the wastewater, costs, and the desired treatment outcomes.
The sustainability of wastewater treatment technologies is becoming
increasingly important as society continues to recognize the value of
conserving natural resources and protecting the environment. Tradi-
tional wastewater treatment processes often rely on energy-intensive
and resource-consuming methods, which can have negative impacts
on the environment and contribute to climate change. Sustainable
wastewater treatment technologies aim to reduce the environmental
impact of wastewater treatment by utilizing renewable energy sources,
minimizing waste generation, and optimizing resource use (Kadam
et al., 2023). The use of agricultural and industrial waste for adsorption
processes in wastewater treatment is an example of a sustainable tech-
nology that can provide both economic and environmental benets
(Steiger et al., 2023). By reducing the need for costly synthetic adsor-
bents and diverting waste from landlls, the use of waste materials in
wastewater treatment can contribute to a more sustainable future.
The cost of any treatment method is an important factor that de-
termines its feasibility and applicability in environmental applications.
It is also important for decision-making. This review included many
technologies that could be used in the treatment of phenolic wastewater,
each with its cost. For instance, distillation and membrane ltration
require higher capital costs for the equipment, membranes, energy
consumption, and maintenance costs than other conventional treatment
methods. Similarly, the use of ozone for phenol removal can be
considered a relatively expensive treatment option due to the high
capital costs of the ozone generator and the high energy required to
produce ozone. Additionally, ozone treatment requires a high level of
operator expertise and maintenance, which can add to the overall cost.
In comparison, the cost of using biological treatment methods such as
phytoremediation and microbial biodegradation is typically lower due
to lower capital and maintenance costs. However, the cost of operation
may be higher for biodegradation due to the energy required to maintain
the ideal environmental conditions for microorganisms. In contrast to
these technologies, the use of agricultural and industrial wastes as ad-
sorbents for phenol removal is one of the most economical approaches
that should be developed and used for wastewater treatment with costs
Fig. 8. A- An example of converting wastes into adsorbents and their roles in promoting sustainability and circular economy, and B- An illustration of guidelines and
decision criteria used to select treatment technologies and their main implications, limitation, and efciencies.
Table 4
Examples of costs of phenol removal using a few conventional technologies.
Technology Pollutant Cost (unit) Reference
Distillation Phenol 2185 $/(m
3
/day)
Estimate based on multi-
effect distillation plant with
2531 m
3
/day capacity
Fan et al. (2014)
Electro-Fenton Phenol 6.12 (US$/m
3
) Magdy et al.
(2021) Photo-Fenton 1.55 (US$/m
3
)
Photocatalysis 1.66 (US$/m
3
)
Adsorption (AC) 0.74 (US$/m
3
)
Photocatalysis/
Adsorption
2.19 (US$/m
3
)
AOP Phenol 108 (US$/m
3
) Krichevskaya
et al. (2011)
Catalytic wet air
oxidation
Phenol 1.20 (US$/kg phenol) Mohammed et al.
(2017)
A. Bibi et al.

Journal of Cleaner Production 417 (2023) 137810
19
around $0.74/m
3
as noted in Table 4. Many adsorbents could be recy-
cled and regenerated to be used for several cycles before being dis-
carded, making them more cost-effective. In the literature, only a few
reviews have conducted cost analyses for the treatment of phenolic
wastewater. Magdy et al. conducted an economic analysis of ve
different technologies that are used in phenol removal. It stated that
adsorption is the most cost-effective one as noted in Table 4 with costs
less than $1/m
3
. Another recent review has investigated the costs
associated with the various adsorbents for the removal of a wide range of
pollutants. This study stated that there is a broad variation in adsorbent
costs, but the majority of adsorbents fall between 1 and 200 $/mol.
Adsorbents at <1 $/mol threshold can be considered very cheap for the
intended application, while those at >200 $/mol are believed to be
highly costly (Ighalo et al., 2022).
The use of adsorption processes especially in wastewater treatment is
signicant due to their agreement with the concept of cleaner produc-
tion and circular economies mainly when these adsorbents are gener-
ated from waste materials and used again as illustrated in Fig. 8-A. These
materials have been gaining great attention due to their effectiveness in
converting wastes into values, where various raw materials that are
otherwise discarded are being efciently used in the purication pro-
cesses, with high efciencies and low costs. Thus, turning waste into a
renewable source of material for the removal of various pollutants such
as nitrogen and phosphorus, heavy metals, and toxic organic compounds
for effective treatment of wastewater. It is evident from Table 3 that
waste-based adsorbents are highly efcient in removing phenol from
water with adsorption capacities ranging from 13 mg/g and reaching
434 mg/g. Interestingly, waste-based adsorbents could also be used
directly for water treatment without any modication or processing
needed, as in the case of using Ziziphus and Neem leaves with capacities
competing with many expensive carbon-based adsorbents (Sieradzka
et al., 2022).
In conclusion, the choice of technology for the treatment of phenolic
wastewater depends on various factors, including the characteristics of
the wastewater, treatment outcomes, and cost considerations. Sustain-
able wastewater treatment technologies that utilize renewable energy
sources, minimize waste generation, and optimize resource use are
becoming increasingly important to reduce the negative impact of
traditional wastewater treatment processes on the environment. While
some technologies may be highly effective, they may also be expensive
and energy intensive. Therefore, it is essential to conduct further
research and a cost-benet analysis of each technology before deciding
on the appropriate treatment approach.
4. Concluding remarks
To sum up, wastewater contains various recalcitrant pollutants
including phenol with different concentrations. According to the liter-
ature, many methods can be used for the treatment of such inuents.
Physiochemical technologies are very effective; however, these tech-
nologies could be expensive and not compatible with sustainable
development goals. Therefore, green technologies should be developed
to achieve the needed levels of treatment and to accommodate various
types of wastewater.
Depending on the concentration of the pollutants, the type of treat-
ment method can be selected, keeping in mind the cost-effectiveness of
the selected treatment systems. From this review, it is clear that the
concentration of phenol plays an important role in determining the
appropriate treatment method. Consequently, it is recommended that
reverse osmosis/nanoltration chemical, electrochemical, and photo-
catalytic oxidation, ozonation, and biodegradation are used to treat
phenolic wastewaters with low and moderate concentrations, whereas
liquid-liquid extraction, and distillation, are suggested for higher phenol
concentrations. It is worth noting that the majority of these technologies
have demonstrated a remarkable effectiveness of over 90% in removing
phenol from wastewater. Furthermore, with some alterations, many of
these technologies have the potential to become even more environ-
mentally friendly and sustainable by incorporating alternative mate-
rials, selecting non-conventional resources, and optimizing their
recyclability. Integrated water treatment systems where more than one
technology is used are also a great option for sustainable wastewater
treatment since they showed their ability to handle various levels of
concentration. Accordingly, a case-by-case study should be done to
address the various limiting factors and select the appropriate
technology.
Adsorption is an effective method for treating phenolic wastewater
with various initial concentrations. Using adsorbents produced from
agricultural and industrial wastes is a great solution, as they could be
environmentally friendly, biodegradable, cost-effective, and highly
efcient. However, there is still a lack of research on the recovery and
recyclability of these adsorbents, which must be addressed to ensure
cost-effectiveness. Scaling up the use of these adsorbents is also neces-
sary, as most of the research is conducted as batch adsorption studies at
laboratory scales. Additionally, it is important to explore alternative
sustainable and green methods for synthesizing adsorbents with high
removal capacities to promote environmental compatibility. With the
use of low-cost and environmentally friendly adsorbents, adsorption
processes have the potential to become a key solution in addressing
water pollution and ensuring access to clean water.
Declaration of competing interest
The authors declare that they have no known competing nancial
interests or personal relationships that could have appeared to inuence
the work reported in this paper.
Data availability
No data was used for the research described in the article.
Acknowledgment
This report was made possible by Qatar University graduate assis-
tantship program. The statements made herein are solely the re-
sponsibility of the author(s). Certain gures were made with BioRender.
com.
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