Post on 30-Dec-2015
description
REVIEWS
A comprehensive overview of elements in bioremediation
Asha A. Juwarkar • Sanjeev K. Singh •
Ackmez Mudhoo
Published online: 29 August 2010
� Springer Science+Business Media B.V. 2010
Abstract Sustainable development requires the
development and promotion of environmental man-
agement and a constant search for green technologies
to treat a wide range of aquatic and terrestrial habitats
contaminated by increasing anthropogenic activities.
Bioremediation is an increasingly popular alternative
to conventional methods for treating waste compounds
and media with the possibility to degrade contami-
nants using natural microbial activity mediated by
different consortia of microbial strains. Many studies
about bioremediation have been reported and the
scientific literature has revealed the progressive
emergence of various bioremediation techniques. In
this review, we discuss the various in situ and ex situ
bioremediation techniques and elaborate on the anaer-
obic digestion technology, phytoremediation, hyper-
accumulation, composting and biosorption for their
effectiveness in the biotreatment, stabilization and
eventually overall remediation of contaminated strata
and environments. The review ends with a note on the
recent advances genetic engineering and nanotechnol-
ogy have had in improving bioremediation. Case
studies have also been extensively revisited to support
the discussions on biosorption of heavy metals, gene
probes used in molecular diagnostics, bioremediation
studies of contaminants in vadose soils, bioremedia-
tion of oil contaminated soils, bioremediation of
contaminants from mining sites, air sparging, slurry
phase bioremediation, phytoremediation studies for
pollutants and heavy metal hyperaccumulators, and
vermicomposting.
Keywords Bioremediation � Green technology �Environmental contaminants � Anaerobic
biotechnology � Composting � Phytoremediation �Biosorption
1 Introduction
The global environment is under great stress due to
urbanization and industrialization as well as popula-
tion pressure on the limited natural resources. The
problems are compounded by drastic changes that
have been taking place in the lifestyle and habits of
people. The environmental problems are diverse and
sometimes specific with reference to time and space.
The nature and the magnitude of the problems are ever
changing, bringing new challenges and creating a
constant need for developing newer and more appro-
priate technologies. In this context, biotechnology has
A. A. Juwarkar (&) � S. K. Singh
Eco-Restoration Division, National Environmental
Engineering Research Institute (NEERI), Council
of Scientific and Industrial Research (CSIR),
Govt. of India, Nehru Marg, Nagpur 440 020,
Maharashtra, India
e-mail: aa_juwarkar@neeri.res.in
A. Mudhoo
Department of Chemical and Environmental Engineering,
Faculty of Engineering, University of Mauritius, Reduit,
Mauritius
123
Rev Environ Sci Biotechnol (2010) 9:215–288
DOI 10.1007/s11157-010-9215-6
tremendous potential to cater for the needs and
holds hope for environmental protection, sustainability
and management (Hatti-Kaul et al. 2007; Azadi and
Ho 2010). While some applications such as bioreme-
diation are direct applications of biotechnology
(Koenigsberg et al. 2005; Dowling and Doty 2009;
Sen and Chakrabarti 2009), there are many which are
indirectly beneficial for environmental remediation,
pollution prevention and waste treatment.
1.1 Environmental pollution and biotreatment
options
The problems of environment can be classified into the
following subheads as most of the problems can be
traced to one or more of the following either directly or
indirectly: Waste generation (sewage, wastewater,
kitchen waste, industrial waste, effluents, agricultural
waste, food waste) and use of chemicals for various
purposes in the form of insecticides, pesticides,
chemical fertilizers, toxic products and by-products
from chemical industries). Waste generation is a side
effect of consumption and production activities and
tends to increase with economic advance. What is of
concern is the increased presence of toxic chemicals
such as halogen aliphatics, aromatics, polychlorinated
biphenyls and other organic and inorganic pollutants
which may reach air, water or soil and affect the
environment in several ways, ultimately threatening
the self-regulating capacity of the biosphere (Sen and
Chakrabarti 2009; Prasad et al. 2010; Beltrame et al.
2010). They may be present in high levels at the points
of discharge or may remain low but can be highly toxic
for the receiving bodies. The underground water
sources are increasingly becoming contaminated. For
example, the underground water sources have been
permanently abandoned in the valley of the River Po in
north Italy due to industrial pollution. Some sub-
stances may reach environment in small concentra-
tions but may be subjected to biomagnification or
bioaccumulation up the food chain, wherein their
concentrations increase as they pass through the food
chain (Davies et al. 2006; Kelly et al. 2007; Fatemi
and Baher 2009; Sharma et al. 2009; Takeuchi et al.
2009). Zhang et al. (2010) have recently detected
legacy pollutants, polychlorinated biphenyls (PCBs),
dichlorodiphenyl trichloroethane and its metabolites
(DDTs), and some emerging organhalogen pollutants,
such as polybrominated diphenyl ethers (PBDEs),
hexabromobenzene (HBB), pentabromotoluene (PBT),
2,3,4,5,6-pentabromoethyl benzene (PBEB), 1,2-
bis (2,4,6-tribromophenoxy) ethane (BTBPE), and
dechlorane plus (DP) in an aquatic food chain
(invertebrates and fish) from an E-waste recycling
region in South China. Polychlorinated biphenyls,
DDTs, PBDEs, and HBB were detected in more than
90% of the samples, with respective concentrations
ranging from not detected (ND)—32,000 ng/g lipid
weight, ND—850 ng/g lipid weight, 8–1,300 ng/g
lipid weight, and 0.28–240 ng/g lipid weight. Pentab-
romotoluene, PBEB, BTBPE, and DP were also
quantifiable in collected samples with a concentration
range of ND—40 ng/g lipid weight. Earlier, Ozkoc
et al. (2007) had detected considerable levels of aldrin,
dieldrin, endrin, heptachlor epoxide, lindane, endo-
sulphan sulphate, and HCB in sediment, mussel, and
seawater samples collected three times during
2001–2003 at nine sampling stations along the mid-
Black Sea coast of Turkey. The highest concentrations
of DDT metabolites measured in the sediment and
mussel samples were 35.9 and 14.0 ng/g wet weight
respectively.
There are three main approaches in dealing with
contaminated sites: Identification of the problem,
assessment of the nature and degree of the hazard,
and the best choice of remedial action. The need to
remediate these sites has led to the development of
new technologies that emphasize the detoxification
and destruction of the contaminants (Wang and Chen
2007; Weber 2007; Kulkarni et al. 2008; Busca et al.
2008) rather than the conventional approach of
disposal. Wang and Chen (2007) recently developed
a novel system of phytoremediation ex planta based on
the overexpression of a secretory laccase (Kunamneni
et al. 2008) that catalyzes the oxidation of various
aromatic compounds, including 2,4,6-trichlorophenol.
All the more, rapid developments in understanding
activated sludge processes and wastewater remedia-
tion warrant exploitation of different strategies for
studying their degradation and some of the biological
remediation terminologies such as bioleaching,
biosorption, bioaugmentation, biostimulation, biopul-
ping, biodeterioration, biobleaching, bioaccumula-
tion, biotransformation and bioattenuation are being
actively researched on (Whiteley and Lee 2006).
Enzyme technology has equally been receiving
increased attention. Hussain et al. (2009) have
216 Rev Environ Sci Biotechnol (2010) 9:215–288
123
reviewed the biotechnological approaches for enhanc-
ing the capability of microorganisms and plants
through the characterization and transfer of pesti-
cide-degrading genes, induction of catabolic path-
ways, and display of cell surface enzymes, while
Theron et al. (2008) have performed a thorough review
of nanotechnology, the engineering and art of manip-
ulating matter at the nanoscale (1–100 nm), and have
highlighted the potential of novel nanomaterials for
treatment of surface water, groundwater, and waste-
water contaminated by toxic metal ions, organic and
inorganic solutes, and microorganisms. Husain et al.
(2009) have analyzed the role of peroxidases in
the remediation and treatment of a wide spectrum
of aromatic pollutants. Peroxidases can catalyze
degradation/transformation of polycyclic aromatic
hydrocarbons (PAHs), PCBs, organochlorines, 2,4,6-
trinitrotoluene, phenolic compounds and dyes. These
enzymes are also capable of treating various types of
recalcitrant aromatic compounds in the presence of
redox mediators.
1.2 Bioremediation: definitions
Bioremediation, which is the use of microorganisms
consortia or microbial processes to degrade and
detoxify environmental contaminants (Margesin et al.
2007; de Lorenzo 2008; Zhao and Poh 2008; Singh
et al. 2008a), is also amongst these new technologies
which derives its scientific justification from the
emerging concept of Green Chemistry and Green
Engineering, and is a fast growing promising remedi-
ation technique increasingly being studied and applied
in practical use for pollutant clean-up. Vidali (2001)
has proposed the following classification of microor-
ganisms involved in bioremediation processes:
Aerobic microbes bring about biodegradation in
the presence of oxygen with Pseudomonas, Alcalig-
enes, Sphingomonas, Rhodococcus, and Mycobacte-
rium being the aerobic bacteria recognized for their
degradative abilities. These microbes have often been
reported to degrade pesticides and hydrocarbons,
both alkanes and polyaromatic compounds. Many of
these bacteria use the contaminant as the sole source
of carbon and energy.
Anaerobic bacteria cause degradation in the
absence of oxygen. There is an increasing interest
in anaerobic bacteria used for bioremediation of
PCBs in river sediments, dechlorination of the
solvent trichloroethylene (TCE), and chloroform.
Ligninolytic fungs are fungi such as the white rot
fungus Phanaerochaete chrysosporium and have the
ability to degrade an extremely diverse range of
persistent or toxic environmental pollutants.
Methylotrophs are aerobic bacteria that grow utiliz-
ing methane for carbon and energy. The initial enzyme
in the pathway for aerobic degradation, methane
monooxygenase, has a broad substrate range and is
active against a wide range of inorganic compounds.
Advances in bioremediation harness molecular,
genetic, microbiology, and protein engineering tools
and rely on identification of novel metal-sequestering
peptides, rational and irrational pathway engineering,
and enzyme design (Singh et al. 2008a). In this
review, the various in situ and ex situ bioremediation
techniques namely anaerobic digestion technology,
phytoremediation, composting, bioaugmentation,
biostimulation and biosorption have been described
and discussed for their effectiveness in the biotreat-
ment, stabilization and eventually overall remediation
of contaminated strata and environments. The last
segment of the review briefly revisits the potential
genetic engineering and nanotechnology have in
enhancing bioremediation. Case studies have also
been extensively revisited to support the discussions
on biosorption of heavy metals, gene probes used in
molecular diagnostics, bioremediation studies of
contaminants in vadose soils, bioremediation of oil
contaminated soils, bioremediation of contaminants
from mining sites, air sparging, slurry phase biore-
mediation, phytoremediation studies for pollutants
and heavy metal hyperaccumulators, and vermicom-
posting. Figure 1 highlights the elements in biore-
mediation that have been addressed and discussed in
this review.
1.3 Characteristics of bioremediation
Bioremediation techniques have been used for decon-
tamination of surface and subsurface soils, freshwater
and marine systems, soils, groundwater and contam-
inated land ecosystems. However, the majority of
bioremediation technologies initially developed were
to treat petroleum hydrocarbon contamination to
immobilize contaminants or to transform them to
chemical products no longer hazardous to human
Rev Environ Sci Biotechnol (2010) 9:215–288 217
123
health and the environment. Where contaminants
pose no significant risk to water supply or surface
water bodies, biodegradation products will include
carbon dioxide, water and other compounds with
little deleterious effects on the environment (Baker
and Herson 1994).
Bioremediation of soils or any site may be enhanced
by fertilizing (adding nutrients such as carbon, nitro-
gen and phosphorous) and/or seeding with suitable
microbial populations. This is enhanced or engineered
bioremediation. Intrinsic bioremediation, which uti-
lizes existing microbial communities, is often the most
cost effective method available for land decontami-
nation. Even in the most contaminated soils, indige-
nous microbial activity can be enough to clean the soil
effectively. Microbial communities within contami-
nated ecosystems tend to be dominated by those
organisms capable of utilizing or surviving toxic
contamination. These communities are typically less
diverse than those in non-stressed systems (Baker and
Herson 1994). Once the soil has been fertilized and/or
seeded, control of temperature, water oxygen content
can be used to speed up the process or reduce the
negative effects of factors such as air pollution. Soil
remediation has suspended the established technolo-
gies of excavation followed by either incineration or
landfilling.
Bioremediation techniques are cost effective as
compared to other technologies as indicated in
Table 1. Biological treatments compare favourably
Fig. 1 Break down of the
elements in bioremediation
discussed in this review
218 Rev Environ Sci Biotechnol (2010) 9:215–288
123
with alternative methods. Treatment periods gener-
ally last from 2 to 48 months, about the same for
chemical or thermal methods. Physical processes
(soil washing and soil vapour extraction) are faster,
rarely lasting more than 1 year. Solidification is
almost instantaneous.
Bioremediation (when used in solution) doses not
require environmentally damaging processes such as
chemicals or heat treatment. It has beneficial effects
upon soil structure and fertility, but with limitation on
its effectiveness. These limitations may be summa-
rized as follows:
• Susceptibility to inhibition by other toxic con-
taminants such as heavy metals
• Low biodegradability of some contaminants such
as chlorinated solvents
• Possible residual contamination after treatment,
such as using hydrogen peroxide as an oxygen
provider
• The potential formation of intermediate com-
pounds which are more toxic than the original
treatment, such as dichlorodiphenyldichloroethy-
lene (DDE) and dichlorodiphenyldichloroethane
(DDD) from the breakdown of DDT (Failey and
Scrivens 1994).
2 Green technology principles
Green technology, emanating directly from Green
Chemistry (or, environmentally benign chemistry)
may be described as the utilization of a set of principles
that reduces or eliminates the use or generation of
hazardous substances in the design, manufacture and
application of chemical products (Kidwai and Mohan
2005). In practice, Green Chemistry is taken to cover
a much broader range of issues than the definition
suggests. As well as using and producing better
chemicals with less waste, Green Chemistry also
involves reducing other associated environmental
impacts, including reduction in the amount of energy
used in chemical processes (Kidwai and Mohan 2005).
Anastas and Warner (1998) have developed ‘The
Twelve Principles of Green Chemistry’ that serve as
valuable and benchmark guidelines for practicing
chemists, researchers and engineers in developing and
assessing how green a synthesis, compound, process or
technology is. These principles are related to the
concepts of prevention, atom economy, less hazardous
chemical syntheses, designing safer chemicals, safer
solvents and auxiliaries, design for energy efficiency,
use of renewable feedstocks, reduce derivatives,
catalysis, design for degradation, real-time analysis
for pollution prevention and inherently safer chemistry
for accident prevention.
Green chemistry is an essential part of green
engineering. The definitions of green chemistry and
green engineering share many commonalities, and the
application of both chemistry and engineering princi-
ples is needed to advance the goals of environmental
sustainability (Kirchhoff 2003). A working definition
of green engineering proposed in Kirchhoff (2003) is
the design, commercialization, and use of processes
and products that are feasible and economical while
minimizing pollution at the source and risk to human
health and the environment. The link between green
chemistry and green engineering is strong in ensuring
that inputs and outputs, both for materials and energy
flows and budgeting, are as inherently safe as possible.
Whilst Green Chemistry focuses on the design of
chemical products and processes that reduce or
eliminate the use and generation of hazardous sub-
stances, it also lays down the ground plan for the
design of the green engineering technologies needed
to implement sustainable products, processes, and
systems (Kirchhoff 2003). The reader is in point of
fact directed to the following excellent publications
which present and discuss the salient aspects of
Green Chemistry and Green Engineering: Anastas
and Kirchhoff (2002), Anastas and Zimmerman
(2003), Anastas and Lankey (2000), Clark (2006),
Hofer and Bigorra (2007), Kirchhoff (2003), Lankey
and Anastas (2002), Ran et al. (2008), Tang et al.
(2008) and Tundo et al. (2000). The subsequent
discussions on bioremediation are contextualized
under the ‘Green Technology’ concept.
Table 1 A comparison of soil remediation treatment costs
Treatment Approximate cost
((£)/tonne soil)
Biological 5–170
Chemical 12–600
Physical 20–170
Solidification/stabilization 17–171
Thermal 30–750
Rev Environ Sci Biotechnol (2010) 9:215–288 219
123
3 Merits and demerits of bioremediation
Although bioremediation is being engineered into a
novel and green technology, microorganisms have
been used routinely for the treatment and transforma-
tion of waste products for at least 100 years so far.
The municipal wastewater treatment industry which is
based on the exploitation of microorganisms in
controlled and engineered systems depends on the
metabolic activities of microorganisms which degrade
the organic matter in wastewaters arriving to waste-
water treatment plants containing selected and accli-
matized populations of microorganisms (Eckenfelder
1989; Vargas et al. 2000; Chen et al. 2005).
3.1 Merits of bioremediation
Bioremediation offers several advantages over the
conventional remediation techniques such as landfill-
ing. Table 2 summarizes the chemical class and their
susceptibility to biodegradation. Often, bioremedia-
tion can be done on site, thereby eliminating trans-
portation costs and liabilities. In many instances,
manufacturing and industrial use of the site can
continue while the bioremediation process is being
implemented. Bioremediation results in the decom-
position of the waste and the long-term liability
associated with non-destructive treatment methods.
Finally, bioremediation can be coupled (i.e., inte-
grated) with other treatment technologies into a
treatment chain allowing for the treatment of mixed
and complex wastes (Yergeau et al. 2009; Goel et al.
2010; McMahon et al. 2008).
The use of renewable (waste) materials has also
boosted the bioremediation of waste streams (Deleu
and Paquot 2004). Residues such us cereals straw,
corn cobs, cotton stalks, various grasses and reed
stems, maize and sorghum stover, vine prunings,
sugarcane and tequila bagasse, coconut and banana
residues, corn husks, coffee pulp and coffee husk,
cottonseed and sunflower seed hulls, peanut shells,
rice husks, sunflower seed hulls, waste paper, wood
sawdust and chips, are some examples of residues
and by-products that can be recovered and upgraded
to higher value and useful products by chemical or
biological processes (Wang 1999; Fan et al. 2000;
Pandey et al. 2000a; Webb et al. 2004). In fact, the
chemical properties of such lignocellulosic agricul-
tural residues make them a substrate of enormous
biotechnological value. They can be converted by
solid state fermentation (SSF) into various different
value-added products including mushrooms, animal
feed enriched with microbial biomass, compost to be
used as biofertilizer or biopesticide, enzymes, organic
acids, ethanol, flavours, biologically active secondary
metabolites and also for bioremediation of hazard-
ous compounds, biological detoxification of agro-
industrial residues and biopulping (Pandey et al.
2000b; Bennet et al. 2002; Sanchez et al. 2002; Kim
and Dale 2004; Nigam et al. 2004; Zervakis et al.
2005; Krishna 2005). SSF has been suggested for
upgrading and valorizing lignocellulosic residues
using basidiomycetous cultures, either through
Table 2 Chemical classes
and their susceptibility to
bioremediation
Chemical class Examples Biodegradability
Aromatic hydrocarbons Benzene, toluene Aerobic and anaerobic
Ketones and esters Acetone, MEK Aerobic and anaerobic
Petroleum hydrocarbons Fuel oil Aerobic
Chlorinated solvents TCE, PCE Aerobic (methanotrophs), anaerobic
(reductive dechlorination)
Polyaromatic
hydrocarbons
Anthracene, benzo
(a)pyrene, creosote
Aerobic
PCBs Arochlors Some evidence; not readily degradable
Organic cyanides Aerobic
Metals Cadmium Not degradable experimental biosorption
Radioactive materials Uranium, plutonium Not biodegradable
Corrosives Inorganic acids, caustics Not biodegradable
Asbestos Not biodegradable
220 Rev Environ Sci Biotechnol (2010) 9:215–288
123
protein enhancement and transformation of residues
into animal feed (Zadrazil 2000), or for enzyme
production (Revankar et al. 2007, Elisashvili et al.
2008). With specific reference to lignocellulolytic,
mushroom fungi like Pleurotus ostreatus and Tra-
metes versicolor have been investigated for bioreme-
diation and biodegradation of toxic and hazardous
compounds like caffeinated residues (Fan et al. 2000)
as well as toxic chemicals such as pesticides, PAHs
and PCBs and chlorinated ethenes (CIUs) in polluted
soils or contaminated groundwater (Perez et al. 2008;
Rigas et al. 2007).
3.2 Demerits of bioremediation
Like most treatment technologies, bioremediation
also has its limitations and disadvantages. Some
chemicals, e.g., highly chlorinated compounds and
heavy metals, are not readily amenable to biological
degradation and stabilization. Table 2 also summa-
rizes the general categories of contaminants and their
relative susceptibility to biodegradation. In addition,
for some chemicals, microbial degradation may lead
to the production of more toxic or mobile substances
than the parent compound(s). For example, under
anaerobic conditions, TCE undergoes a series of
microbiologically mediated reactions resulting in the
sequential removal of chlorine atoms from the
molecule. This process is called reductive dehalo-
genation. The end product of this series of reactions
is vinyl chloride (VC), a known carcinogen. Thus, if
bioremediation is applied without a through under-
standing of the microbial processes involved and the
metabolic and chemical pathways, it could actually
lead to a worse situation than already exists in some
cases. Hence, bioremediation is a scientifically
intensive procedure, which must be tailored to the
site-specific conditions to minimize the effects of
environmental and kinetic constraints (Price et al.
2004; Beck and Jones 1995). Therefore, initial costs
for site assessment, characterization and feasibility
evaluation for bioremediation may be higher than the
costs associated with more conventional technologies
such as air stripping. As with remediation technolo-
gies, there is also the need for extensive monitoring
of the site during implementation of the project
(Sabean et al. 2009) to assess the effectiveness of the
bioremediation technique in its clean-up perfor-
mance. Monitoring requirements may include some
form of microbiological monitoring in addition to the
chemical monitoring associated with physical/chem-
ical remediation techniques. Finally, there are regu-
latory constraints that impact on the implementation
of bioremediation (Talley and Sleeper 2006).
4 Bioremediation technologies
Bioremediation technologies can be broadly classified
as ex situ or in situ (Hatzinger et al. 2002; Talley and
Sleeper 2006). Table 3 summarizes the most com-
monly used bioremediation technologies. Ex situ
technologies are those treatment modalities which
involve the physical removal of the contaminated to
another area (possibly within the site) for treatment.
Bioreactors, landfarming, anaerobic digestion, com-
posting, biosorption and some forms of solid-phase
treatment are all examples of ex situ treatment tech-
niques. In contrast, in situ techniques involve treat-
ment of the contaminated material in place. Bioventing
for the treatment of the contaminated soil and biosti-
mulation of indigenous aquifer microorganisms are
examples of these treatment techniques. Although
some sites may be more easily controlled and main-
tained with ex situ configurations (Talley and Sleeper
2006), others are more effective with in situ treatment.
Table 3 Bioremediation treatment technologies
Bioaugmentation Addition of bacterial cultures to a
contaminated medium; frequently used
in bioreactors and ex situ systems
Biofilters Use of microbial stripping columns to treat
air emission
Biostimulation Stimulation of indigenous microbial
populations in soils and/ or ground water;
may be done in situ or ex situ
Bioreactors Biodegradation in a container or reactor;
may be used to treat liquids or slurries
Bioventing Method of treating contaminated soils
by drawing oxygen through the soil to
stimulate microbial growth and activity
Composting Aerobic, thermophilic treatment process in
which contaminated material is mixed
with a bulking agent; can be done using
static piles, aerated piles, or continuously
fed reactors
Landfarming Solid-phase treatment system for
contaminated soils; may be done in situ
or in a constructed soil treatment cell
Rev Environ Sci Biotechnol (2010) 9:215–288 221
123
For example, many sites are located in industrial/
commercial areas, and these sites normally consist of
numerous structures interconnected by concrete and
asphalt. These physical barriers would make excava-
tion extremely difficult, and if the contamination is
deep in the subsurface, excavation becomes too
expensive. As a result of these physical barriers, the
required excavation efforts may make ex situ biotreat-
ment impracticable. Other factors could also have an
impact on the type of treatment. At a typical site, the
contamination is basically trapped below the surface.
To expose the contamination to the open environment
through excavation can result in potential health and
safety risks (Talley and Sleeper 2006). In addition, the
public’s perception of the excavation of contaminants
could be negative, depending on the situation. All of
these conditions clearly favor in situ biotreatment.
Nonetheless, the key is to carefully consider the
parameters involved with each site before evaluating
which technique to use (Talley and Sleeper 2006).
4.1 Microbial consortia for bioremediation
Regardless of the exact nature of the treatment
technology, all bioremediation techniques depend on
having the right microorganisms in the right place
with the right environmental conditions for degra-
dation to occur (Iranzo et al. 2001; Baxter and
Cummings 2006). The right microorganisms are
those bacteria or fungi that have the physiological
and metabolic capabilities to degrade the contami-
nants. Although it is generally accepted that more
than 80% of the total microorganisms are unknown
(Iranzo et al. 2001), reactions mediated by both the
known and the unknown microorganisms are already
employed in biotreatment and in bioremediation
(Hamer 1993). This consideration, together with the
potential use of engineered microorganisms, offers an
expanded time scale technology (Pieper and Reineke
2000). In many instances, these organisms will
already be present at the (indigenous microorgan-
isms). In other circumstances, such as bioreactors
treating wastes with high concentrations of toxic
material. In order for the microorganisms to degrade
the contaminants, they must be in close proximity to
the contaminants; they must be in the right place.
Thus, the presence of toluene-degrading microorgan-
isms in the surface soils at a site will be of little use
for the remediation of a contaminant which is
biodegradable. If such populations are not present,
then some mechanisms must be engineered to bring
the microorganisms into contact with the contami-
nants. This may involve such techniques as flushing
the system to transport the contaminants to above-
ground bioreactors (Litchfiled 2005), the addition of
surfactants to the subsurface to release adsorbed
contaminants and render them available to the micro-
organisms (Singh et al. 2007), or the introduction
and transport of the microorganisms to the contam-
inated area. Once the right microorganisms are
present in the right place, the environmental condi-
tions must favor the metabolic activities of the
microorganisms. Such environmental factors as tem-
perature, inorganic nutrients (primarily nitrogen and
phosphorus), electron acceptors (oxygen, nitrate, and
sulphate), and pH can be modified to optimize the
environment for bioremediation (Singh et al. 2006a;
Ge et al. 2004).
The objectives of the bioengineered remediation
treatment processes are analogous to those conven-
tional biological treatment operations. With conven-
tional biological treatment systems, a treatment vessel
is ‘‘engineered’’ to provide optimal conditions for the
microorganisms to grow. As a result of their growth,
the microorganisms will metabolize the compound(s)
of interest, usually resulting in the production of
innocuous end products (Ahuja and Kumar 2003). An
example of this concept would be a wastewater
treatment facility. For this process, the conditions in a
treatment vessel (i.e., a large tank) are optimized (pH
is adjusted and provisions to control flow rates to
provide adequate contact time are optimized) to
promote biodegradation of the organic materials in
the wastewater. For a bioengineered treatment system,
instead of utilizing a manufactured container to
accommodate the treatment process, the soil environ-
ment could be ‘‘bioengineered’’ to create an in-place
treatment vessel and to provide optimal growth
conditions for the indigenous microorganisms present.
The effective application of this type of biological
treatment can result in the complete breakdown of the
contaminant(s) to innocuous end products in many
instances (Barnabe et al. 2009). Pyridine and pyridine
based products are of major concern as environmental
pollutants due to their recalcitrant, persistent, toxic
and teratogenic nature. Lodha et al. (2008) have
studied the biodegradation of pyridine by an isolated
222 Rev Environ Sci Biotechnol (2010) 9:215–288
123
consortium/strain under aerobic condition. Batch
experiment results revealed that at lower initial
pyridine concentrations (1–20 mg/l), almost complete
degradation was observed whereas at higher concen-
tration (30–50 mg/l), the degradation efficiency was
dropped significantly. Bioaugmentation of the iso-
lated consortium/strain into the activated sludge
consortium in different quantity had been also done
and the effect of bioaugmentation on degradation has
been studied. This showed that as the quantity of
bioaugmentation increased, the degradation of pyri-
dine also increased. Prasanna et al. (2008) have
studied the bioremediation of soil-bound anthracene
studied in a series of bio-slurry phase reactors
operated in periodic discontinuous/sequencing batch
mode under anoxic–aerobic–anoxic microenviron-
ment using native soil microflora. Five reactors were
operated for a total cycle period of 144 h at soil
loading rate of 16.66 kg soil/m/day at 30 ± 2�C
temperature. The control reactor (without microflora)
showed negligible degradation of anthracene due to
the absence of biological activity, while the perfor-
mance of the bio-slurry system with respect to
anthracene degradation was found to depend on both
substrate loading rate and bioaugmentation. Subse-
quent application of bioaugmentation however
showed positive influence on the rate of degradation
of anthracene. All the more, phytoremediation has
been used as an emerging technology for remediation
of soil contamination with PAHs, ubiquitous persis-
tent environmental pollutants derived from natural
and anthropogenic processes, in the last decade. In
this respect, Xu et al. (2009) carried out a pot
experiment to investigate the potential of phytoreme-
diation of pyrene from spiked soils planted with white
clover (Trifolium repens) in the greenhouse with a
series of pyrene concentrations ranging from 4.22 to
365.38 mg/kg. Their results showed that growth of
white clover on pyrene contaminated soils was not
affected. The removal of pyrene from the spiked soils
planted with white clover was obviously higher than
that from the unplanted soils. At the end of the
experiment (60 days), the average removal ratio of
pyrene in the spiked soils with white clover was 77%,
which was 31 and 57% higher than those of the
controls with or without microbes, respectively,
thereby supporting that the removal of pyrene in the
contaminated soils was feasible using T. repens.
Lately, Osman et al. (2009) investigated the
bioremediation of the nematicide, oxamyl, applied at
6 l/ha in sandy soil cultivated with tomato and
amended with different animal manures at the recom-
mended dose of 2.5 tons/ha. By the end of the exper-
iment (28 days), the dissipation percentage of oxamyl
reached about 99% in the case of bovine manure-
amended soil, and this rate of disappearance was 1.76
times higher than in unamended-soil, while poultry
and sheep manures enhanced the dissipation rate by
1.52 and 1.44 times, respectively. The results of
Osman et al. (2009) demonstrated that animal manures
may offer an efficient, cheap, safe, and friendly
bioremediator for pesticide-polluted soil.
4.2 Approach to bioremediation techniques
The successful implementation of bioengineered
remediation techniques will involve a multidisciplin-
ary approach requiring input from individuals with
expertise in microbiology, chemistry, geology, soil
science, environmental engineering and chemical
engineering. In order to use bioengineering success-
fully for the remediation of environmental contam-
ination problems, the first step is to obtain a through
understanding of the matrix characteristics of the
media to be treated and the properties (physical,
chemical and microbiological) of the contaminant(s).
Lai et al. (2007) stress that the performance moni-
toring of applied remediation technologies is an
important part of site remediation. It involves peri-
odic measurement of site parameters to evaluate
whether the remediation technologies perform as
expected or to determine the termination date of
remediation projects. According to Lai et al. (2007),
performance monitoring can be a difficult undertak-
ing if there are no well-defined and measurable
remediation objectives, such as a reduction in mass
discharge rate from a contaminant source. The
monitoring requirements for a bioremediation treat-
ment system shall most reasonably comprise the
following analyses and inspections.
Daily: Inspection of the system components, i.e.,
piping, pumps and valves; monitoring of pH, dis-
solved oxygen, temperature, and mineral nutrient
levels within the treatment system, and monitoring
flow rates and pumping rates.
Monthly: Monitoring the following parameters
within the treatment system and in the off-site
monitoring wells: contaminant concentration, aerobic
Rev Environ Sci Biotechnol (2010) 9:215–288 223
123
heterotrophic bacterial population density, pH, dis-
solved oxygen, temperature and available mineral
nutrient concentrations.
Quarterly: perform a series of soil boring and
analysis for the following parameters: contaminant
concentration, aerobic heterophic bacterial popula-
tion density, pH, soil moisture and available mineral
nutrient concentrations. Subsequently, any adjust-
ment in the bioremediation technique will be made
accordingly to the treatment system based on the
results from these analyses so as to enhance the
bioremediation performance.
All the more, through the advances in gene technol-
ogy, bioremediation is now in a position to take
advantage of genomic-driven strategies to analyze,
monitor and assess its course by considering multiple
microorganisms with various genomes, expressed tran-
scripts and proteins (Stenuit et al. 2009). High-through-
put methodologies, including microarrays, fingerprinting
(Karpouzas and Singh 2010), real-time polymerase chain
reaction (PCR) (Baek et al. 2009), genotypic profiling,
ultrafast genome pyrosequencing, metagenomics, meta-
transcriptomics, metaproteomics and metabolomics
(Desai et al. 2010; Jerez 2009), show great promise in
environmental interventions against recalcitrant contam-
inants such as 2,4,6-trinitrotoluene (TNT) that have been
studying for many years. The emerging genomic and
metagenomic methodologies now allow environmental
researchers and engineers to promote and restore envi-
ronmental health in impacted sites, monitor remediation
activities, identify key microbial players and processes,
and finally compile an intelligent, site-specific and
pollutant(s)-specific database of genes for targeted use
in bioremediation (Stenuit et al. 2009).
5 Bioremediation techniques for contaminated
sites
Soil is one of the key resources for sustainability and
survival, and its degradation caused by willful or
accidental contamination from industrial sources or
degradation caused by salination and waterlogging is
a great matter of concern. Land degradation is
recognized as the loss of the fertility or potential
utility by changes in irreplaceable features or com-
munities of organisms. Bacteria and fungi are natural
recyclers capable of transforming natural and
synthetic chemicals into sources of energy and raw
materials for their own growth. This implies that
biological processes supplement chemical or physical
remediation processes and that is why bioremediation
is becoming important for the clean-up of contami-
nated soils all around the world.
Contamination of soils can occur through the
accidental release of materials on the surface or
through the direct introduction of contaminants into
the subsurface, as in the case of leaking underground
storage tanks. From the perspective of remediation, the
soil environment can be divided into two zones:
shallow surface soils and subsurface (vadose) soils.
Shallow surface soils usually include the upper 1–3 ft
of the environment. These soils represent the region of
the environment typically included in the agronomic
definition of soils. They are easily modified and are
generally more amenable to remediation than deeper
vadose soils. Operationally, surface soils can be
defined as those soils which can be excavated or
treated by surface amendments not requiring the
installation of wells. Vadose soils are those soils
which lie between the surface soils and the water table
or aquifer. Vadose soils are generally unsaturated,
although there may be pockets of water saturated soil
within the vadose zone, particularly in the area of the
root zone and in the capillary fringe at the surface of
the water table. In addition, there may be inclusions of
low permeability materials such as clay lenses within
the vadose zone, which can become saturated with
water. Unlike surface soils, vadose soils are often not
amenable to excavation or surface treatment; rather,
modification of such soils usually involves the use of
infiltration galleries, injection wells, or other engi-
neered means for introducing materials.
5.1 Characterization of contaminated sites by
molecular diagnostic
A prerequisite for any remediation strategy is the
characterization of the sites with regards to various
factors that may affect bioremediation, particularly
characterization of indigenous bacterial communities
(Hjeitzer and Sayler 1993). Knowledge of the types,
concentrations and activities of biodegradative bac-
teria and of the processes that control their activities
is important in at least two regards:
224 Rev Environ Sci Biotechnol (2010) 9:215–288
123
• as part of site characterization for determining
appropriate remediation strategies; and
• as part of monitoring the progress and effective-
ness of bioremediation.
Although rarely used, concentrations of specific
bacteria may also serve as indicators of remediation
‘‘endpoints’’ i.e., as indicators of remediation suffi-
cient to declare a site ‘‘clean’’. Identification and
enumeration of individual bacterial species are usu-
ally deemed too time consuming and laborious for
routine analysis, and is usually dispensed with in
favor of evaluation of biodegradative potential of
samples in laboratory microcosms.
Among the limitations of site characterization by
bench top microcosms are the lack of sensitivity;
inaccuracy due to an inability to cultivate most soil
bacteria in the laboratory; deviation from what is
actually occurring in the field; and cumbersome
experimental setups. In addition, many sites are
contaminated with mixtures of compounds, thereby
greatly increasing the numbers of analyses required for
sufficient site characterization. A relatively new
approach, molecular diagnostic (Sayler and Layton
1990; Ritalahti et al. 2005; Katsoyiannis et al. 2007),
provides an alternative to traditional laboratory micro-
cosms and has great potential to provide important
information on resident microbial communities that
would be unavailable from traditional laboratory
microcosm studies. The most commonly applied
molecular genetic approach to site characterization is
so-called ‘‘molecular diagnostic,’’ referring to the use
of cloned genes of interest (‘‘gene probes’’) to measure
concentrations of similar genes in microbial DNA or
RNA directly extracted from the soil sample (Yagi and
Madsen 2009; N’Guessan et al. 2010) (Fig. 2).
Molecular diagnostics utilizes information regard-
ing the structure and activities of resident microbial
communities contained in the community’s nucleic
Soil sampleAddition of
bioluminescent reporter strains
Cell lysis and extraction of nucleic acids
Measurement of light production
BIOAVAILABILITY OF SUBSTRATE
Purification of DNA
Purification of RNA
Hybridization with gene probes
Hybridization with gene probes
BIODEGRADATIVE POTENTIAL
IN SITUACTIVITY
Fig. 2 Flow chart of sample characterization by molecular diagnostics
Rev Environ Sci Biotechnol (2010) 9:215–288 225
123
acid to characterize the biodegradative potential of a
site. Much of the information required for accurate
site characterization is encoded in the microbial
community’s DNA, and analyzing DNA directly
isolated from samples may access much of this
information. Utilizing DNA as an indicator molecule
in this way bypasses the need for laboratory cultiva-
tion, and therefore bypasses much of the bias and
uncertainty associated with laboratory incubations. In
addition, molecular diagnostics can be a very rapid
and accurate means of simultaneously screening
numerous samples, thereby increasing the efficiency
with which samples may be processed.
Gene probes cloned from bacterial pathways for
metabolisms of the pollutant of interest may be used
against a large number of samples in a relatively short
period of time, simultaneously providing information
on the specific pathways present and the concen-
trations of bacteria possessing the genes. Selected
examples of gene probes are listed in Table 4.
Information regarding specific pathway(s) involved
in the bioremediation process will increase under-
standing of specific environmental factors that may be
manipulated for optimization of then bioremediation
process. The approach may be quantitative, and the
relative concentration of specific genes in a given
sample is related to the potential for that sample to
biodegrade the compound of interest. This information
will aid scientists and engineers in developing and
monitoring remediation strategies appropriate for the
site and the types of bacteria available for remediation.
The potential for biodegradation to occur is
indicated by the concentration of the gene of interest
within the community DNA, and this information
may be accessed by using DNA as the target for gene
probes. This approach dose not yield information
regarding the activity of these genes, or serve as an
indication of in situ biodegradation rates. The activity
of the genes is expressed as RNA, and using RNA as
a target for gene probes may, in some cases, correlate
with biodegradation rates. Virtually all studies
attempting to correlate biodegradation rates with
mRNA concentrations have involved correlating
nahA (naphthalene biodegradation rates (Fleming
et al. 1993; Sanseverino et al. 1993). Using RNA as
a target in such studies may be limited for some genes
due to the very high turnover rate of some mRNAs,
but nahA transcripts have correlated well with
biodegradation rates in a number of soils.
Some advantages of molecular genetic applications
to site characterization are a rapid characterization of
numerous samples; the simultaneous characterization
of multiple biodegradative pathways from mixed-
waste sites; acquisition of information regarding the
specific biochemical pathway most likely to dominate
remediation process; possibility for a rapid assessment
of the progress of biostimulation strategies; and a rapid
quantitative assessment of indicators of remediation
endpoints.
As with any microbial method, molecular diag-
nostic has limitations, although most of these are
being overcome with time. The greatest limitation to
date is that the gene of interest must be known. Soils
harbor a great diversity of bacteria (perhaps as many
as 10,000 distinct species), most of which have never
been cultivated in the laboratory. Laboratory culti-
vation is an obvious prerequisite for studying the
genetics of biodegradation, and it is likely that we
currently know very little of the diversity of genes
involved in biodegradation of many organic pollu-
tants. This lack of information limits the number of
gene probes available for use, and in many case
Table 4 Example of gene probes used in molecular diagnostics
Target compound Target enzyme Target gene Host strain Reference
Toluene, TCE Toluene-4-monoxygenase Tom A Burkholderia cepacia G4 Shield et al. (1989)
Toluene, TCE Toluene dioxygenase TodC1C2BA Pseudomonas putida F1 Zylstra and Gibson (1989)
Toluene, TCE Toluene-2-monoxygenase tmoABCDE Pseudomonas mendocina KR1 Yen et al. (1991)
BTEX compounds Catechol 2,3 dioxygenase xylE Pseudomonas putida mt-2 Assinder and Williams (1990)
TCE Soluble methane
monooxygenase
mmoB Methylosinus trichsporium0B3b
Tsien and Hanson (1992)
Naphthalene Naphthalene dioxygenase nahAcd Pseudomonas putida G7 Simon et al. (1993)
226 Rev Environ Sci Biotechnol (2010) 9:215–288
123
molecular diagnostic may underestimate both the
numbers of bacteria involved in bioremediation and
types of biodegradative pathways involved.
It is not reasonable to over-emphasize the great
diversity of soil bacteria and our relative ignorance of
the genetic and biochemical diversity present in the
soil. A recent example of the failure of molecular
diagnostics to accurately describe the biodegradative
potential of a site involved a dominant and unde-
scribed toluene pathway. In this case, DNA was
isolated from soil that rapidly mineralized toluene, and
analyzed with gene probes specific to all five toluene
pathways known at that time. No positive signals were
observed with any of the gene probes against the soil
DNA, indicating that these genes were not likely to
represent the dominant toluene degrading bacteria if
they were present at all in the samples. Toluene
metabolizing bacteria were isolated from the soil, and
none of these harbored genes similar to the five known
pathways. The dominant toluene degrading bacterial
strain in this soil was later identified as a Rhodococcus
sp., a gram positive species that shares relatively low
similarity at the DNA level to all cloned genes
involved in toluene metabolism. Virtually all that is
known of the genetics and biochemistry of biodegra-
dation was learned from gram negative species such as
Pseudomonas. These species grow rapidly in standard
enrichment cultures and frequently out grow other
species that may be equally or more important in the
soil. Until more is known of the diversity of biode-
gradative bacteria in soil, the general applicability of
molecular diagnostic will be limited.
5.2 Treatment of contaminated vadose soils
Treatment of contaminated vadose soils several feet
below the surface is a challenging task which usually
involves some types of in situ treatment system(s).
Electronic acceptors, inorganic nutrients, and other
supplements (i.e., bacterial cultures) are introduced,
if necessary, into the subsurface environment to
stimulate microbial degradation of the contaminants.
Proper controls must be included in the system to
ensure that the contaminants do not migrate.
Water was the first medium used to transport
materials throughout the vadose zone. In this approach
to bioremediation, hydraulic control of the site is first
established. This typically involves installation of a
series of injection wells or trenches and recovery
wells. Alteration of the water table level also may be
undertaken. Nutrients and oxygen dissolved in water
are injected into the subsurface environment. As the
water percolates through the subsurface, nutrients and
oxygen are delivered to the microorganisms, stimu-
lating biodegradation of the contaminant. The material
which leaches through the vadose zone into the
saturated zone is captured and pumped to the surface,
where it is treated, if necessary, and recirculated into
the system. This type of treatment system is funda-
mentally the same as treatment systems designed for in
situ bioremediation of contaminated aquifers. Often,
contaminants in the vadose zone and concomitant
contamination of groundwater are treated as a single
unit. Several research works have reported the success
of contaminated vadose soils clean-up by bioremedi-
ation. Table 5 summarizes some of these most recent
examples.
5.3 Bioremediation of oil contaminated media
Large areas of earth surface including land and water
are contaminated with oil-derived compounds and
toxic chemicals. More than 2 million tonnes of oil are
estimated to enter the sea each year. 50% from
industrial effluents, sewage and river overflows and
the rest from non-tanker shipping and natural seepage
from below sea floor. Only about 18% comes from
refineries, off-shore operations and tanker activities.
Oil contamination is easy to detect unlike other
pollutants. Most oils are relatively less toxic to
environment but can affect bird and aquatic animal
life very seriously. Three main types of bioremediation
technologies that are currently being developed for
treatment of oil spills are: (a) addition of nutrients to
open water oil slicks, (b) addition of microbes to oiled
shorelines, and (c) addition of nutrients and/or
microbes to open water slicks. Just adding oil to an
environment will stimulate growth of indigenous
microbes as oil acts as a source of carbon. But there
is a lag period, which can be several days to several
weeks, before they can effectively degrade oil. During
this period, certain fractions initially toxic to microbes
undergo weathering. On March 23, 1989 about 11
million gallons of Prudhoe Bay crude oil were spilled
into Prince William Sound, Alaska from the tanker
Exxon Valdez. The spilled oil spread over 350 miles of
Rev Environ Sci Biotechnol (2010) 9:215–288 227
123
shoreline. About 15–20% was lost by initial weath-
ering due to volatilization. Mostly aliphatic hydrocar-
bons of less than 12 carbon atoms, aromatic
hydrocarbons such as benzene, toluene, xylene and
methyl-substituted naphthalenes were lost this way.
The weathered oil was black in colour. Initial tests
were conducted following which bioremediation was
successfully undertaken to overcome this problem.
Table 6 presents a digest of recent studies conducted
to bioremediate oil contaminated soils.
Table 5 Example of bioremediation studies of contaminants in vadose soils
Contaminant(s) Technology/technique employed Bioremediation performance Reference
Perchlorate Experiments performed in soil slurries
with sediments taken from the
contaminated site with native
microbial communities along
the contaminated vadose zone
With no external carbon source added to
the slurry of soil from land surface, all
perchlorate was removed after 134 days
of incubation
Average perchlorate-reduction rate using
natural organic matter as a carbon
source was 0.45 mg/day, while the
average rate using acetate as an external
carbon source was 7.2 mg/day
Gal et al. (2008)
Perchlorate and nitrate Gaseous electron donor injection
technology
Laboratory microcosm studies
demonstrated that hydrogen and ethanol
promoted nitrate and perchlorate
reduction in vadose zone soil
Nitrate removal in the column studies, up
to 100%, was best promoted by ethyl
acetate
Up to 39% perchlorate removal was
achieved with ethanol and was limited
by insufficient incubation time
Evans and Trute
(2006)
Chromate Test involved injecting hydrogen
sulphide, diluted in air, into
contaminated vadose zone sediment
to reduce Cr(VI) to Cr(III)
All Cr(VI) concentrations measured in the
posttest samples were well below the
remedial goal and regulatory limit of
30 mg/kg
In addition, the field test demonstrated
that vadose zone treatment of
contamination could be safely
conducted using diluted hydrogen
sulphide gas mixtures
Thornton et al.
(2007)
Naphthalene at a creosote-
contaminated
Combined remediation mechanisms
of volatilization and biodegradation
Soil gas sampling showed that more than
90% of the naphthalene vapors were
biodegraded aerobically within
5–10 cm above the water table
Andersen et al.
(2008)
Toluene Radiation-resistant bacterium
Deinococcus radiodurans was
engineered for toluene degradation
by cloned expression of tod and xylgenes of Pseudomonas putida
Complete oxidation of the toluene by the
engineered bacteria under both minimal
and complex nutrient conditions
Brim et al.
(2006)
Atrazine [6-chloro-N-ethyl-
N0-isopropyl-1,3,5-
triazine-2,4-diamine] and
cyanazine {2-[[4-chloro-
6-(ethylamino)-1,3,
5-triazin-2-yl] amino]-2-
methylpropanenitrile}
Combined chemical–biological
approach comprising
Fe0 ? FeSO4 ? emulsified soybean
[Glycine max (L.) Merr.] oil (EOS),
EOS remediation scenarios
Overall temporal sampling (0–342 days)
revealed atrazine and cyanazine
concentrations decreased by 79–91%
Waria et al.
(2009)
228 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Ta
ble
6R
esu
me
of
rece
nt
stu
die
so
nb
iore
med
iati
on
of
oil
con
tam
inat
edso
ils
Oil
/Oil
frac
tio
nT
ech
niq
ue
emp
loy
edB
iore
med
iati
on
per
form
ance
Ref
eren
ce
Cru
de
pet
role
um
-oil
hy
dro
carb
on
s
Bac
teri
ald
egra
dat
ion
insh
ake
flas
kte
sts
usi
ng
Ba
cill
us
sub
tili
sD
M-0
4an
dP
seu
do
mo
na
sa
eru
gin
osa
Man
dN
Mst
rain
s
Bio
aug
men
tati
on
of
TP
Hco
nta
min
ated
mic
roco
smw
ith
P.
aer
ug
ino
saM
and
NM
con
sort
iaan
dB
.su
bti
lis
stra
in
sho
wed
asi
gn
ifica
nt
red
uct
ion
of
TP
Hle
vel
sin
trea
ted
soil
as
com
par
edto
con
tro
lso
ilat
the
end
of
exp
erim
ent
(12
0d
ays)
P.
aer
ug
ino
sast
rain
sw
ere
mo
reef
fici
ent
than
B.
sub
tili
sst
rain
inre
du
cin
gth
eT
PH
con
ten
tfr
om
the
med
ium
Das
and
Mu
kh
erje
e(2
00
7)
Fu
elo
il-c
on
tam
inat
edso
ils
Th
ree-
stag
etr
eatm
ent
trai
nsy
stem
tore
med
iate
fuel
oil
-co
nta
min
ated
soil
s:fi
rst
stag
eo
fb
iod
egra
dab
le
surf
acta
nt
and
gro
un
dw
ater
flu
shin
gfo
llo
wed
by
the
Fen
ton
-lik
eo
xid
atio
n,
end
ing
wit
hap
pli
cati
on
of
enh
ance
db
iore
med
iati
on
for
the
rem
ov
alo
f
resi
du
alto
tal
pet
role
um
Res
ult
sfr
om
the
colu
mn
stu
dy
ind
icat
eth
atap
pro
xim
atel
y
80
%o
fT
PH
inso
il(w
ith
init
ial
con
cen
trat
ion
of
50
,00
0m
g/
kg
)co
uld
be
rem
ov
edaf
ter
the
Sim
ple
Gre
enT
M(S
G)
[50
po
rev
olu
mes
(PV
s)]
foll
ow
edb
yg
rou
nd
wat
er(3
0P
Vs)
flu
shin
g.
Th
eF
ento
n-l
ike
ox
idat
ion
(wit
h6
%o
fH
2O
2
add
itio
n)
was
able
tore
mo
ve
ano
ther
15
%o
fT
PH
.O
bse
rved
firs
t-o
rder
reac
tio
nra
teco
nst
ant
of
TP
Ho
xid
atio
nw
as
2.7
49
10
-2/m
in,
and
the
hal
f-li
few
as2
5m
ind
uri
ng
the
firs
t4
0m
ino
fre
acti
on
Tsa
iet
al.
(20
09)
Oil
con
tam
inat
edan
d
pri
stin
eso
ils
fro
mS
ign
y
Isla
nd
(So
uth
Ork
ney
Isla
nd
s,A
nta
rcti
ca)
Bac
teri
alco
nso
rtiu
mte
sted
for
bio
deg
rad
atio
n
po
ten
tial
Of
the
30
0is
ola
tes
cult
ure
d,P
seu
do
mo
na
sst
rain
ST
41
gre
wo
n
the
wid
est
ran
ge
of
hy
dro
carb
on
sat
4�C
.M
icro
cosm
exp
erim
ents
sho
wed
that
at4
�Cth
ele
vel
so
fo
ild
egra
dat
ion
incr
ease
d,
rela
tiv
eto
the
con
tro
ls,
wit
h(i
)th
ead
dit
ion
of
ST
41
toth
eex
isti
ng
soil
mic
rob
ial
po
pu
lati
on
,(i
i)th
e
add
itio
no
fn
utr
ien
tsan
dto
the
gre
ates
tex
ten
tw
ith
(iii
)
aco
mb
inat
ion
of
bo
thtr
eatm
ents
Ad
dit
ion
of
wat
erto
oil
con
tam
inat
edso
il(h
yd
rati
on
)al
so
enh
ance
do
ild
egra
dat
ion
,al
tho
ug
hle
ssth
anth
eo
ther
trea
tmen
ts
Pse
ud
om
on
as
spec
ies
was
do
min
ant
asso
ilb
acte
ria
inb
oth
bio
aug
men
ted
and
bio
stim
ula
ted
mic
roco
sms
Sta
llw
oo
det
al.
(20
05
)
Die
sel-
con
tam
inat
edso
ilE
ffica
cyo
fC
an
did
aca
ten
ula
taC
M1
,w
asev
alu
ated
du
rin
gco
mp
ost
ing
of
am
ixtu
reco
nta
inin
gfo
od
was
tean
dd
iese
l-co
nta
min
ated
soil
Aft
er1
3d
ays
of
com
po
stin
g,
84
%o
fth
ein
itia
lp
etro
leu
m
hy
dro
carb
on
was
deg
rad
edin
com
po
stin
gm
ixes
con
tain
ing
ap
ow
der
edfo
rmo
fC
M1
(CM
1-s
oli
d),
com
par
edw
ith
48
%
of
rem
ov
alra
tio
inco
ntr
ol
reac
tor
wit
ho
ut
ino
culu
m
Joo
etal
.(2
00
8)
PA
Hin
anag
edcr
eoso
te-
con
tam
inat
edso
il
PA
Hd
egra
dat
ion
effi
cacy
inp
rese
nce
of
Sa
lix
vim
ina
lis
was
inv
esti
gat
edin
ag
reen
ho
use
exp
erim
ent
Ph
enan
thre
ne
and
py
ren
ew
ere
deg
rad
ed1
00
and
80
%,
resp
ecti
vel
y,
inth
ep
rese
nce
of
pla
nts
bu
to
nly
68
and
63
%
wit
ho
ut
pla
nts
Pre
sen
ceo
fS
.vi
min
ali
so
rth
esu
rfac
tan
ten
han
ced
PA
H
deg
rad
atio
n,
pri
mar
ily
by
arh
izo
sph
ere
effe
cto
nth
e
mic
rob
ial
acti
vit
yin
the
form
erca
sean
db
yin
crea
sed
bio
avai
lab
ilit
yin
the
latt
erca
se
Hu
ltg
ren
etal
.(2
01
0)
Rev Environ Sci Biotechnol (2010) 9:215–288 229
123
5.4 Bioremediation of mine spoil dumps
The rapid increase in industrialization in all sectors
has led to degradation of natural resources i.e., air,
water and soil. Mineral exploitation (mining) is
second only to agriculture as the world’s oldest and
most important industry and its operation leads to a
number of environmental problems namely defores-
tation, removal of fertile top soil, unstable slopes
prone to sliding and erosion, siltation of water bodies
due to wash off of mineral overburden dumps, air
pollution due to discharge of dust, ground vibration
and finally the socio-economic status of local people.
As a result of mining activities, significant areas of
land are degraded and undesirable materials in the
form of overburden dumps, tailings and ash dams
replace the existing ecosystems. The overburden
materials (solid wastes) thus produced are physically
and structurally unstable, prone to subsidence and
chemically as well as hydrogically unsuitable for
plant growth. The degraded lands are devoid of
nutritive and supportive capacity for biomass devel-
opment. Gradual increase in such landscapes due to
intensive mining activities adversely impacts aquatic,
land and atmospheric ecosystems. Ecological ame-
lioration of these ecovulnerable systems is a chal-
lenging task as the top soil suitable for plant growth,
which takes a number of decades to be produced have
been disturbed due to mining and buried deep down
the biologically unproductive surface.
Thus, realizing the major physical, chemical and
biological constraints in biorestoration, an Integrated
Biotechnological Approach (IBA) was developed
(Juwarkar et al. 2000) to restore the nature’s pattern
of air, water and land blending stable and diverse
ecosystem comprising of different components of
flora and fauna. An IBA is a biocompatible technol-
ogy which comprised of isolation and inoculation of
site specific specialized nitrogen fixing strains of
Bradyrhizobium and Azotobacter species and nutrient
mobilizing vesicular arbuscular mycorrhizal spores of
Glomus and Gigaspora species in combination with
industrial waste material available near the vicinity of
mine site used as organic amendments to ameliorate
the mine spoil and encourage revegetation. Different
plant species of ecological and economical impor-
tance can be planted on mine spoil dumps using
appropriate blends of organic waste along with site
specific nitrogen fixing bacteria and endomycorrhizal
fungi. Besides this the plants having medicinal value
can be planted on the sites. Different types of grasses
for slope stabilization can be used on large scale to
prevent runoff and erosion of dumps. Thus, IBA is an
ecofriendly technology for holistic restoration of lost
biological diversity of the mined out areas and
commercial utilization of such degraded landscapes
through plantation of ecologically and economically
important plant species. The technology has been
successfully demonstrated and commercialized to
solve problems of different mining sectors and thermal
power (fly ash dump reclamation). Table 7 presents a
summary of some selected studies conducted on the
bioremediation of mine related contaminants.
6 In situ bioremediation
6.1 Bioventing
In bioventing, the aerobic biodegradation of soil
contamination is stimulated by delivery of oxygen to
the subsurface. This is accomplished by injecting or
extracting air through unsaturated soil in a passive
system. This technology is designed primarily to treat
soil contamination by fuels, non-halogenated volatile
organic compounds (VOCs) and semi-volatile organic
compounds (SVOCs), pesticides and herbicides. The
process may be applied to halogeneted organics, but is
less effective. Bioventing typically costs around $15
per cubic yard of soil and uses simple, inexpensive,
low-maintenance equipment that can be left unat-
tended for long periods of time. Also, the technology
tends to enjoy good public acceptance.
The technology requires the presence of indige-
nous organisms capable of degrading the contami-
nants of interest, as well as nutrients necessary for
growth. Also, it is necessary that the contaminants be
available to the organisms, and not tightly sorbed to
soil particles. Bioventing is not as effective in treating
areas where the water table is high, and soils with
very low moisture content. Lastly, the technology is
not applicable in sites where high concentrations of
inorganic salts, heavy metals, or organic compounds
are present, as these hinder microbial growth. How-
ever, some studies have demonstrated the merits of
bioventing as a bioremediation technique. Møller
et al. (1996) have investigated the effects of
230 Rev Environ Sci Biotechnol (2010) 9:215–288
123
bioventing, nutrient addition and inoculation with an
oil-degrading bacterium on biodegradation of diesel
oil in unsaturated soil in a mesocosm system consisting
of 6 soil compartments each containing 6 m3 of
naturally contaminated soil mixed 1:1 with silica sand,
resulting in a diesel oil content of approximately
2,000 mg/kg. Biodegradation was monitored over
112 days by determining the actual diesel oil content
of the soil and by respirometric tests, and it was
observed that the oil composition changed following
degradation resulting in the unresolved complex
mixture constituting up to 96% of the total oil content
at the end of the experimental period. Lately, Shewfelt
et al. (2005) have conducted experiemnets using
small-scale respirometers containing gasoline-con-
taminated soil from an active remediation site to
determine the effects of soil water content, nitrogen
content, nitrogen form, and the composition of the
microbial population on the gasoline biodegradation
rate. Results indicated that optimum bioventing con-
ditions were 18 wt.% soil water content, C:N = 10:1,
using NH4?—Sui et al. (2006) have studied the
cometabolic bioventing for removal of TCE in the
unsaturated zone in a soil column study using methane
as growth substrate, and the experimental data showed
that a total TCE remediation efficiency of over 95%
was obtained with a volatilization -to- biodegradation
ratio of TCE being about 7:1.
6.2 Biostimulation
Biodegradation in the subsurface may be stimulated
by addition of water-based solutions carrying nutri-
ents, electron acceptor or other amendments. These
technologies are designed primarily to treat soil and
groundwater contamination by fuels, non-halogenet-
ed VOCs, SVOCs, pesticides, and herbicides. These
processes may be applied to halogeneted organics,
but are sometimes less effective. Although the costs
of biostimulation technologies vary tremendously
from site to site, these technologies tend to be
amongst the cheapest alternatives when applicable.
The technology requires the presence of indigenous
organisms capable of degrading the contaminants of
interest. Also, it is necessary that the contaminants be
available to the organisms, and not tightly sorbed to
soil particles. With specific reference to chlorine
containing contaminants, the successful application
of bioaugmentation requires consideration of a num-
ber of additional factors including:
Table 7 Studies on remediation of contaminants from mining sites
Contaminants/
contaminated media
Technique employed Bioremediation performance Reference
Fuel oil
contaminated
mixtures of soil
and sawdust
Feasibility of aerated in-vessel
composting at a laboratory scale as
a bioremediation technology to
clean-up contaminated desert
mining soils (fuel
concentration [ 50,000 mg kg-1)
and sawdust (fuel
concentration [ 225,000 mg/kg) in
the Atacama Region
The highest (59%) and the lowest (35%)
contaminant removals were observed in
the contaminated sawdust and
contaminated soil reactors after 56 days
of treatment, respectively
Results of this research indicate that
bioremediation of an aged contaminated
mixture of desert mining soil and sawdust
with fuel oil is feasible
Godoy-Faundez
et al. (2008)
Arsenic and heavy
metals (i.e., Cu, Pb
and Zn) from
oxidized Pb–Zn
mine tailings
samples
Column experiments were carried out
to evaluate the feasibility of using
humic acid (HA) to mobilize arsenic
and heavy metals
It was found that the HA could significantly
enhance the mobilization of arsenic and
heavy metals simultaneously from the
mine tailings. After a 70-pore-volume-
flushing, the mobilization of arsenic,
copper, lead and zinc reached 97, 35, 838
and 224 mg/kg, respectively
Use of HA in arsenic and heavy metal
remediation was deduced to show
promise as an environmentally benign
and possible effective remedial option to
reduce and avoid further contamination
Wang and
Mulligan
(2009)
Rev Environ Sci Biotechnol (2010) 9:215–288 231
123
1. the availability of a sufficient amount of culture
to facilitate complete dechlorination of the target
contaminant;
2. the presence of co-contaminants that may affect
biodegradation; and,
3. the added cost and benefit of adding bacterial
cultures.
Biostimulation is not applicable in sites where high
concentrations of inorganic salts, heavy metals, or
organic compounds are present, as these hinder micro-
bial growth. Lastly, the calculation of water-based
solutions through the soil may increase contaminant
mobility and necessitate treatment of underlying ground-
water. Preferential colonization by microbes may occur
causing clogging of nutrient and water injection wells.
Recent studies on the application of biostimulation
for degrading a variety of contaminants unanimously
advocate the merit of this technique. Krishnani et al.
(2009) have used molecular methods based on
sequencing of clone libraries to provide sequence and
the phylogenetic information of ammonia oxidizing
bacteria (AOB). Ammonia monooxygenase (amoA)
gene, which catalyzed the oxidation of ammonia to
hydroxyl amine in the initial rate-determining step of
nitrification was targeted for detection and character-
ization of AOB using gene-specific primers. The use of
a matrix prepared from abundantly available lignocel-
lulosic agrowaste-bagasse has successfully been dem-
onstrated for biostimulation of AOB in aquaculture
environment by supplementing nutritional require-
ment facilitating the biofilm mode of growth of the
autotrophic consortia, the applicatiom of the results of
this study could be useful in enhanced predictability
and reliability of the treatment of ammonia in brack-
ishwater aquaculture. Dafale et al. (2008) have iden-
tified the strains viz. Pseudomonas aeruginosa and
Bacillus circulans and other unidentified laboratory
isolates (NAD1 and NAD6) to be predominantly
present in a microbial consortium acclimatized from
activated sludge from a textile effluent treatment plant
to effectively decolorize RB5 dye solutions. The
optimum inoculums concentration for maximum
decolorization were found to be 1–5 ml of 1,800 mg/
l MLSS and 37�C, respectively. Overall, the effective-
ness of the acclimatized biomass under optimized
conditions towards decolorization of two types of
synthetic dye bath wastewaters that were prepared
using chosen azo dyes has been demonstrated.
Hirschorn et al. (2007) have reported based on stable
carbon isotope analysis that the dechlorination of TCE
was occurring in in situ biostimulation pilot test areas
during biostimulation. Garcia-Blanco et al. (2007)
have assessed the effectiveness of biostimulation in
restoring an oil-contaminated coastal marsh dominated
by Spartina alterniflora under north-temperate condi-
tions through nitrogen and phosphorus addition for
accelerating oil disappearance, and then have equally
determined the role of nutrients in enhancing restora-
tion in the absence of wetland plants, and the rate at
which the stressed salt marsh recovered. It was
reported that GC–MS resolved alkanes and aromatics
degraded substantially ([90 and[80%, respectively)
after 20 weeks with no loss of TPH. Earlier, Hamdi
et al. (2007) had studied the degradation of spiked
anthracene (ANT), pyrene (PYR) and benzo[a]pyrene
(B[a]P) in soil (3,000 mgP
3 PAHs/kg dry soil) in
aerobically incubated microcosms for 120 days. The
applied treatments aimed at enhancing PAH removal
from the heavily contaminated soils were: (i) bioaug-
mentation by adding aged PAH-contaminated soil
(ACS) containing activated indigenous degraders; and
(ii) combined bioaugmentation/biostimulation by
incorporating sewage sludge compost (SSC) and
decaying rice straw (DRS). Hamdi et al. (2007)
reported that the adopted treatments produced higher
PAH dissipation rates than those observed in una-
mended PAH-spiked soils, especially for ANT and
PYR in the presence of DRS or ACS ([96%). Later,
Salinas-Martınez et al. (2008) have studied the biosti-
mulation of the native microbial consortium as a novel
application of the heap leaching technique to biore-
mediate mining soils contaminated with hydrocarbons.
Two genera, Flavobacterium and Aspergillus, were
identified as the primary microorganisms that degraded
hydrocarbons in the polluted soil. The main outcomes
showed that of the rates tested, biodegradation was
most efficient at a flow rate of 200 ml/h, and the heap
leaching technique demonstrated good efficiency in
the column and pile arrangement, with a 2% soil-sand
mixture lowering the TPH concentration from 61,000
to 1,800 mg/kg (98.5%) in 15 days.
6.3 Air-sparging
Air-sparging stimulates aerobic biodegradation of
contaminated groundwater by delivery of oxygen to
232 Rev Environ Sci Biotechnol (2010) 9:215–288
123
the subsurface (Johnson et al. 2007; Tsai 2008). This
is accomplished by injecting air below the water
table. This technology is designed primarily to treat
groundwater contamination by fuels, non-haloge-
nated VOCs, SVOCs, pesticides, organics, and her-
bicides. Air sparging has also been demonstrated to
be an innovative groundwater remediation technol-
ogy capable of restoring aquifers that have been
polluted by volatile and (or) biodegradable contam-
inants, such as petroleum hydrocarbons (Heron et al.
2002; Gidarakos and Aivalioti 2008). The process
may be applied to halogenated organics, but is less
effective. Air-sparging can cost less than $1 per
1,000 l in favorable situations and tends to be among
the cheapest remedial alternatives when applicable.
The technology uses simple, inexpensive, low-main-
tenance equipment that can be left unattended for
long periods of time. Also, the technology tends to
enjoy good public acceptance. The technology
requires the presence of indigenous organisms capa-
ble of degrading the contaminants of interest, as well
as nutrients necessary for growth. Also, it is neces-
sary that the contaminants be available to the
organisms, and not tightly sorbed to soil particles.
Air sparing is not applicable in sites where high
concentrations of inorganic salts, heavy metals, or
organic compounds are present, as hinder microbial
growth. Table 8 reports some studies on the applica-
tion of air sparging to bioremediate contaminated
media.
6.4 Natural attenuation
Natural attenuation is a proactive approach that
focuses on the verification and monitoring of natural
remediation processes (Khan et al. 2004). Also known
as passive remediation, in situ bioremediation, intrin-
sic remediation, bioattenuation, and intrinsic biore-
mediation, natural attenuation is an in situ treatment
method that uses natural processes to contain the
spread of contamination from chemical spills and to
reduce that concentration and amount of pollutants at
contaminated sites (Boparai et al. 2008; Khan et al.
2004). This means the environmental contaminants are
undisturbed while natural attenuation works on them.
Natural attenuation processes are often categorized as
destructive or non-destructive (Gelman and Binstock
2008). Destructive processes destroy the contaminant,
while non-destructive processes cause a reduction in
contaminant concentrations (Khan et al. 2004).
Target contaminants for natural attenuation include
fuels, non-halogenated VOCs, SVOCs, pesticides and
herbicides. The process may be applied to halogenated
organics, but it requires longer treatment times. Also,
the technology is applicable to especially hydrophobic
contaminants such as high molecular weight PAHs that
tend to sorb very tightly to soil particles and have very
low rates of migration. Often, communities of adopted
degraders will mineralize such contaminants quickly
after they desorb from soil particles. The costs of
natural attenuation are associated with modeling
contaminant migration, degradation rates to determine
its feasibility, and evaluation and monitoring costs to
confirm adequate performance (). Although these costs
tend to be low compared with other remedial alterna-
tives, the public is often suspicious of natural attenu-
ation due to the impression that nothing is being done.
Some very important observations related to the
performance of natural attenuation technology are
(Khan et al. 2004): it is a relatively simple technology
compared to other remediation technologies; it can be
carried out with little or no site disruption; it often
requires more time to achieve clean-up goals than other
conventional remediation methods; it requires a long-
term monitoring program and the program duration
affects the cost; if natural attenuation rates are too
slow, the pollution/contaminant plume could migrate;
and it is difficult to predict with high reliability the
performance of natural attenuation.
6.5 Landfarming
This technology involves the application of contam-
inated material that has been excavated onto the soil
surface and periodically tilled to mix and aerate the
material (Maciel et al. 2009; Harmsen et al. 2007).
The contaminants are degraded, transformed and
immobilized by means of biotic and abiotic reactions
(Rubinos et al. 2007). Sometimes, in cases of very
shallow contamination, the top layer of the site may
simply be tilled without requiring any excavation.
Liners or other methods may be used to control
leachate. This technology is designed primarily to
treat soil contamination by fuels, PAHs, non-haloge-
nated VOCs, SVOCs, pesticides, and herbicides. The
process may be applied to halogenated organics, but
is less effective. Although the technology is very
Rev Environ Sci Biotechnol (2010) 9:215–288 233
123
Ta
ble
8S
tud
ies
on
the
app
lica
tio
no
fai
rsp
arg
ing
for
bio
rem
edia
tio
n
Co
nta
min
ated
med
iaD
escr
ipti
on
of
met
ho
d(s
)em
plo
yed
Bio
rem
edia
tio
np
erfo
rman
ceR
efer
ence
Gro
un
dw
ater
con
tam
inat
edw
ith
tric
hlo
roet
hy
len
eT
CE
ina
san
dy
aqu
ifer
Lab
ora
tory
and
nu
mer
ical
inv
esti
gat
ion
s
usi
ng
ap
uls
edai
rsp
arg
ing
syst
em
Aq
ueo
us
con
cen
trat
ion
sfo
rT
CE
wer
est
ill
mu
chh
igh
erth
an
the
max
imu
mco
nta
min
ant
lev
elin
spit
eo
fsu
cces
sfu
l
rem
ov
alo
f9
5%
of
resi
du
alT
CE
,im
ply
ing
that
itw
ou
ldb
e
mo
reap
pro
pri
ate
toap
ply
air
spar
gin
gco
mb
ined
wit
ho
ther
rem
edia
tio
nte
chn
olo
gie
ssu
chas
bio
rem
edia
tio
nfo
r
rem
edia
tio
no
fT
CE
-co
nta
min
ated
gro
un
dw
ater
Kim
etal
.(2
00
7)
Wat
ersa
tura
ted
con
tam
inat
edso
il
un
der
air
spar
gin
gco
nd
itio
ns
for
dif
fere
nt
NA
PL
san
dso
ilp
rop
erti
es
Air
spar
gin
gco
nd
itio
ns
Sp
arg
ing
air
atam
bie
nt
tem
per
atu
reth
rou
gh
the
con
tam
inat
ed
soil
cou
ldre
mo
ve
NA
PL
s,b
ut
emp
loy
ing
ho
tai
rsp
arg
ing
cou
ldp
rov
ide
hig
her
con
tam
inan
tre
mo
val
effi
cien
cy,
by
abo
ut
9%
Mo
ham
edet
al.
(20
07
)
Rem
ov
alo
fv
ola
tile
con
tam
inan
tfr
om
anaq
uif
erm
od
elan
dm
ass
rem
ov
al
of
den
sen
on
-aq
ueo
us
ph
ase
liq
uid
(DN
AP
L);
Per
chlo
roet
hen
ew
asth
e
test
con
tam
inan
t
Lab
ora
tory
-sca
lesu
rfac
tan
t-en
han
ced
air
spar
gin
g(S
EA
S)
was
eval
uat
edv
ersu
s
con
ven
tio
nal
air
spar
gin
g.
Su
rfac
tan
t
use
dw
asso
diu
md
od
ecy
lben
zen
e
sulf
on
ate
(SD
BS
)
SE
AS
was
sig
nifi
can
tly
mo
reef
fici
ent
than
con
ven
tio
nal
air
spar
gin
gfo
rre
mo
vin
gp
erch
loro
eth
ene.
Fo
rS
EA
S,
abo
ut
78
and
75
%o
fto
tal
per
chlo
roet
hen
em
ass
was
dep
lete
dfr
om
the
flo
wch
amb
erat
35
0an
d1
50
mg
/lS
DB
S
Kim
etal
.(2
00
9)
Rem
ov
alo
fcr
ud
eo
ilfr
om
soil
Air
spar
gin
gas
sist
edst
irre
dta
nk
reac
tors
.
Tw
osu
rfac
tan
ts(r
ham
no
lip
idan
d
sod
ium
do
dec
yl
sulf
ate,
SD
S)
wer
e
test
edan
dth
eef
fect
so
fd
iffe
ren
t
par
amet
ers
(tem
per
atu
re,
surf
acta
nt
con
cen
trat
ion
s,w
ash
ing
tim
e,v
olu
me/
mas
sra
tio
)w
ere
inv
esti
gat
edu
nd
er
var
yin
gw
ash
ing
mo
des
nam
ely
,st
irri
ng
on
ly,
air
spar
gin
go
nly
and
the
com
bin
atio
no
fst
irri
ng
and
air
spar
gin
g
SD
Sre
mo
ved
mo
reth
an8
0%
cru
de
oil
fro
mn
on
-wea
ther
ed
soil
sam
ple
s,w
hil
strh
amn
oli
pid
sho
wed
sim
ilar
oil
rem
ov
al
atth
eth
ird
and
fou
rth
lev
els
of
the
par
amet
ers
test
ed
Th
eap
pro
ach
of
soil
was
hin
gw
asn
ote
dto
be
effe
ctiv
e
inre
du
cin
gth
eam
ou
nt
of
oil
inso
il
Uru
met
al.
(20
05)
Pet
role
um
hy
dro
carb
on
con
tam
inat
ed
aqu
ifer
s
Pu
lsed
op
erat
ion
of
anin
-wel
lai
r
spar
gin
g
Pu
lsed
op
erat
ion
of
the
inst
alle
dsy
stem
last
edab
ou
t
5m
on
ths
and
the
resu
lts
of
freq
uen
tg
rou
nd
wat
ersa
mp
lin
g
and
anal
ysi
sin
dic
ated
anim
po
rtan
td
ecli
ne
into
tal
TP
H
and
BT
EX
con
cen
trat
ion
so
fu
pto
99
%
Gid
arak
os
and
Aiv
alio
ti
(20
08
)
Ch
lori
nat
edal
iph
atic
hy
dro
carb
on
s
(CA
Hs)
ing
rou
nd
wat
er
Co
-met
abo
lic
air
spar
gin
g(C
AS
)
dem
on
stra
tio
nu
sin
gp
rop
ane
asth
e
com
etab
oli
csu
bst
rate
.
TC
E,
cis-
dic
hlo
roet
hen
e(c
-DC
E);
and
dis
solv
edo
xy
gen
con
cen
trat
ion
lev
els
dec
reas
edin
pro
po
rtio
nw
ith
pro
pan
e
usa
ge,
wit
hc-
DC
Ed
ecre
asin
gm
ore
rap
idly
than
TC
E
Inth
ep
rop
ane-
stim
ula
ted
zon
e,c-
DC
Eco
nce
ntr
atio
ns
dec
reas
edb
elo
wth
ed
etec
tio
nli
mit
(1l
g/l
and
TC
E
con
cen
trat
ion
sra
ng
edfr
om
less
than
5–
30
lg
/l,
rep
rese
nti
ng
rem
ov
als
of
90
–9
7%
Inth
eai
rsp
arg
edco
ntr
ol
zon
e,T
CE
was
rem
ov
edat
on
lytw
o
mo
nit
ori
ng
loca
tio
ns
nea
rest
the
spar
ge-
wel
l,to
con
cen
trat
ion
so
f1
5an
d6
0lg
/l
To
van
abo
otr
etal
.(2
00
1)
234 Rev Environ Sci Biotechnol (2010) 9:215–288
123
simple and inexpensive, it does require large space,
and reduction in contaminant concentrations may
sometimes be due to volatilization rather than biodeg-
radation (Sanscartier et al. 2010; Souza et al. 2009).
Souza et al. (2009) have used Allium cepa bioassays
to assess landfarming and landfarming with rice hulls
amendment before and after hydrocarbons biodegra-
dation assay in the laboratory. It has been reported that
after 108 days of biodegradation, the landfarming
reached the rate of 26.30 mmol of CO2 released and
the concentration of TPHs decreased by 27%. Land-
farming treated with rice hulls had the highest release
of CO2, 110.9 mmol, associated with a remarkable
reduction in TPHs concentration, 59%, thereby show-
ing that the use of rice hulls accelerated the biodeg-
radation efficacy of landfarming to improve the
efficiency of bioremediation processes. In their study,
Marın et al. (2005) assessed the ability landfarming to
reduce the total hydrocarbon content added to soil with
refinery sludge in low rain and high temperature
conditions. It was seen that 80% of the hydrocarbons
were eliminated in 11 months, half of this reduction
taking place during the first 3 months. Rubinos et al.
(2007) treated a soil heavily contaminated ([5,000 mg
kg-1) with hexachlorocyclohexane (HCH) isomers
derived from lindane production using the landfarming
technique and observed significant decreases of the
a- and c-HCH isomers woth up to 89 and 82% of
the initial concentration, respectively, at the end of the
11 months. In this respect, the aerobic landfarming
appeared to be a viable and cost effective bioremedi-
ation treatment technology for soils contaminated with
a- and c-HCH isomers on large scales. Clark and
Boopathy (2007) equally concluded from their study
that landfarming from bench-scale studies could be
promisingly transferred to full-scale application.
Lately, Sanscartier et al. (2009) have investigated the
bioremediation of weathered medium- to high-molec-
ular weight petroleum hydrocarbons (HCs) in Polar
regions. Their findings suggested that temperature and
low moisture content had affected the biodegradation
of HCs but with volatilization possibly predominating
in the field.
6.6 Phytoremediation
Using plants in soil and groundwater remediation
(i.e., phytoremediation) is a relatively new concept
and the technology has yet to be extensively proven
in the marketplace. Because of this, most information
about phytoremediation comes mainly from field and
laboratory research (Table 9). However, the potential
of phytoremediation for cheap, simple and effective
soil and groundwater remediation is generating
considerable interest.
Phytoremediation may be used for remediation of
soil and groundwater contaminated with toxic heavy
metals, radionuclides, organic contaminants such as
chlorinated solvents, BTEX compounds, non-aromatic
petroleum hydrocarbons, nitrotoluene ammunition
wastes, and excess nutrients (Schnoor et al. 1995).
Table 10 summarizes some studies which have been
conducted to remove heavy metals from contaminated
media by phtyoremediation and Table 11 presents the
findings of research on the application of phytoreme-
diation for organic pollutants.
Other applications of phytoremediation include
landfill caps, buffer zones for agricultural runoff and
even drinking water and industrial wastewater treat-
ment. Phytoremediation may also be used as a final
polishing step, in conjunction with other treatment
technologies. While indeed promising, the applicabil-
ity of phytoremediation is limited by several factors.
First, it is essential that the contaminated site of
interest is able to support plant growth. This requires
suitable climate, soil characteristics such as pH and
texture, and adequate water and nutrients. Second,
because plant roots only go so deep, phytoremediation
is practical only in situations where contamination is
shallow (less than 5 m), although in some situations
with deeper contamination it may be used in conjunc-
tion with other technologies. Third, since the time
requirements for phytoremediation are sometimes
long relative to some conventional technologies such
as landfilling and incineration, it is not suitable for
situations requiring rapid treatment. Plants facilitate
remediation via several mechanisms:
1. Direct uptake, and incorporation of contaminants
into plant biomass
2. Immobilization, or phytostabilization of contam-
inants in the subsurface
3. Release plant enzymes into the rhizosphere that
act directly on the contaminants
4. Stimulation of microbially mediated degradation
in the rhizosphere
Rev Environ Sci Biotechnol (2010) 9:215–288 235
123
6.6.1 Phytoremediation of inorganic contaminants
Remediation of metal-contaminated soils and ground-
water is another potentially promising application of
phytoremediation. Given that metals cannot be chem-
ically transformed, and they can be toxic to microor-
ganisms, metal contamination is not readily amenable
to in situ treatment strategies such as microbially
mediated remediation. Treatment of metal contami-
nation therefore calls for either extraction or immobi-
lization, and conventional treatment strategies, such as
incineration, landfilling, leaching, and chemical fixa-
tion are often prohibitively expensive. Additionally,
landfilling and incineration are often hampered by
public hostility. Understandably, cost-effective treat-
ments for metal contamination are desperately needed.
This has stimulated considerable interest in using the
natural ability of some plants to accumulate (and
hyperaccumulate) metals in their tissues. Phytoreme-
diation technologies that exploit this trait include
phytoextraction, rhizofiltration, phytostabilization,
phytodegradation and phytovolatilization (Salt et al.
1995).
6.6.2 Phytoextraction
Phytoextraction makes use of metal accumulating
plants to transport and concentrate metals in harvest-
able roots and shoots in order to remove metals from
soil. Typically, multiple crops of metal-accumulators
could be grown, followed by harvest and processing of
the plant material, which could involve reclamation of
Table 9 Plants used in phytoremediation studies for pollutants
Pollutant(s) Plant species used Reference
Nitrogen, phosphorus Thalia geniculata f. rheumoides Shuey, Oenenathejavanica (Blume) DC. ‘Flamingo’, Phyla lanceolata(Michx.) Greene
Polomski et al. (2008)
Cadmium, copper, arsenic Lolium perenne cv Elka Sidoli O’Connor et al.
(2003)
Cadmium, copper, lead, zinc Paulownia tomentosa Doumett et al. (2008)
2,4,6-trinitrotoluene Vetiver grass (Vetiveria zizanioides) Makris et al. (2007)
Anthracene in mycorrhizospheric soil Ryegrass (Lolium multiflorum) Korade and Fulekar (2008)
Phenol Vetiver (Vetiveria zizanoides L. Nash) Singh et al. (2008b)
2,6-dinitrotoluene Arabidopsis thaliana Yoon et al. (2007)
Arsenic species such as arsenate Spider brake (Pteris cretica L.) plants Ebbs et al. (2010)
Arsenic Nugget marigold, a triploid hybrid between American
(Tagetes erecta L.) and French (Tagetes patula)
marigolds
Chintakovid et al. (2008)
Cadmium, chromium, nickel, iron,
arsenic
Helianthus annuus (sunflower) January et al. (2008)
Recalcitrant PAHs Fescue (Festuca arundinacea), switchgrass (Panicumvirgatum), zucchini (Curcubita pepo Raven)
Cofield et al. (2007)
Dibenzofuran-contaminated soil Bermuda grass (Cynodon dactylon), bent grass (Agrostispalustris Huds.), lawn grass (Zoysia japonica), white
clover (Trifolium repens L.)
Wang and Oyaizu (2009)
Selenium-laden drainage sediments Canola (Brassica napus var. Hyola 420), tall fescue
(Festuca arundinacea var. Au Triumph), salado grass
(Sporobulus airoides), cordgrass (Spartina patens var.
Flageo).
Banuelos and Lin (2005)
Soil contaminated with diesel fuel Scots Pine (Pinus sylvestris), Poplar (Populusdeltoides 9 Wettsteinii), Red fescue, Festuca rubra;
Smooth meadowgrass, Poa pratensis, Perennial ryegrass,
Lolium perenne), White clover, Trifolium repens and Pea,
Pisum sativum
Palmroth et al. (2002)
Lead, copper, zinc, cadmium Vetiver grass Vetiveria zizanioides Chen et al. (2004)
236 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Ta
ble
10
Hea
vy
met
alre
mo
val
by
ph
tyo
rem
edia
tio
n:
rem
ov
alco
nd
itio
ns
and
per
form
ance
Hea
vy
met
al(s
)P
lan
t(s)
use
dR
emo
val
per
form
ance
Ref
eren
ce
Ch
rom
ium
-co
nta
min
ated
soil
sF
enu
gre
ek(T
rig
on
ella
foen
um
gra
ecu
mL
.),
spin
ach
(Sp
ina
cia
ole
race
aL
.),
and
ray
a
(Bra
ssic
aca
mp
estr
isL
.)
Th
eC
rco
nce
ntr
atio
nin
fen
ug
reek
,sp
inac
h,
and
ray
ain
crea
sed
wit
hin
crea
sin
g
lev
elo
fad
ded
Cr
inb
oth
soil
s.C
rin
bo
thsh
oo
tan
dro
ot
was
hig
hes
tin
ray
a,
foll
ow
edb
ysp
inac
han
dfe
nu
gre
ek.
Th
eo
ver
all
mea
nu
pta
ke
of
Cr
insh
oo
t
was
alm
ost
fou
rti
mes
and
inro
ot
was
abo
ut
two
tim
esh
igh
erin
ray
aco
mp
ared
tofe
nu
gre
ek.
Th
efi
nd
ing
sin
dic
ated
that
fam
ily
Cru
cife
rae
(ray
a)w
asm
ost
tole
ran
tto
Cr
tox
icit
y,
foll
ow
edb
ych
eno
po
dia
cea
(sp
inac
h)
and
Leg
um
ino
sae
(fen
ug
reek
)
Dh
eri
etal
.(2
00
7)
Cad
miu
man
dzi
nc
Th
lasp
ica
eru
lesc
ens
asa
ph
yto
extr
acti
on
pla
nt
Th
ep
erio
dic
use
of
ph
yto
extr
acti
on
wit
hT
.ca
eru
lesc
ens
tom
ain
tain
soil
s
bel
ow
stat
uto
rym
etal
con
cen
trat
ion
lim
its,
wh
enm
od
ern
sew
age
slu
dg
esar
e
rep
eate
dly
app
lied
,se
ems
ver
yat
trac
tiv
eg
iven
the
no
n-i
ntr
usi
ve
and
cost
-eff
ecti
ve
nat
ure
of
the
pro
cess
Max
ted
etal
.(2
00
7)
Cad
miu
m-c
on
tam
inat
edso
ils
Cd
-hy
per
accu
mu
lato
rR
ori
pp
ag
lob
osa
(Tu
rcz.
)
10
7.0
and
15
0.1
mg
/kg
of
the
Cd
accu
mu
late
din
stem
san
dle
aves
,re
spec
tiv
ely
,
wh
enso
ilC
dad
ded
was
con
cen
trat
edto
25
.0m
g/k
g.
Th
eC
d-r
emo
vin
gra
tio
by
sho
ots
of
R.
glo
bo
sah
arv
este
dat
the
flo
wer
ing
ph
ase
was
up
to7
1.4
%o
f
that
atth
em
atu
rep
has
e
Wei
and
Zh
ou
(20
06
)
Nic
kel
-co
nta
min
ated
soil
sN
ick
elp
hy
toex
trac
tio
nfr
om
fou
r
typ
eso
fN
i-co
nta
min
ated
soil
s
by
Ni-
hy
per
accu
mu
lato
rs
Aly
ssu
mco
rsic
um
,A
lyss
um
mu
rale
,an
dn
on
-
hy
per
accu
mu
lato
rsra
dis
h,
mu
star
d
Ni
con
cen
trat
ion
insh
oo
to
fA
.co
rsic
um
and
A.
mu
rale
was
sig
nifi
can
tly
hig
her
than
rad
ish
or
mu
star
din
all
test
edso
ils.
A.
cors
icu
man
dA
.m
ura
lere
mo
ved
mu
chm
ore
Ni
fro
mM
oji
ang
soil
(to
tal
Ni
1,0
62
mg
/kg
)th
anco
mm
on
veg
etab
les,
bu
tm
ust
ard
extr
acte
dm
ost
Ni
fro
mX
iny
iso
il(t
ota
lN
i1
07
mg
/kg
)
Qiu
etal
.(2
00
8)
Co
pp
erM
aize
(Go
ldD
ent)
,so
yb
ean
(En
rei
and
Su
zuy
uta
ka)
,an
dri
ce
(Nip
po
nb
are
and
Mil
yan
g2
3)
wer
ep
ot-
gro
wn
un
der
aero
bic
soil
wit
hlo
wto
mo
der
ate
Cu
con
tam
inat
ion
Aft
er2
mo
nth
scu
ltiv
atio
n,
the
Go
ldD
ent
mai
zean
dM
ily
ang
23
rice
sho
ots
too
ku
p2
0.2
–2
9.5
and
18
.5–
20
.2%
of
the
0.1
mo
l/l
HC
l-ex
trac
tab
leC
u,
10
.0–
37
.3an
d8
.5–
34
.3%
of
the
DT
PA
-ex
trac
tab
leC
u,
and
2.4
–6
.5an
d
2.1
–5
.9%
of
the
tota
lC
u,
resp
ecti
vel
y,
inth
etw
oso
ils
anal
yze
d
Mu
rak
ami
and
Ae
(20
09
)
Co
nta
min
ants
wer
ead
ded
asle
ad
nit
rate
(Pb
(NO
3) 2
)an
dzi
nc
nit
rate
(Zn
(NO
3) 2
)at
40
0m
g/k
g
wh
ich
rep
rese
nts
up
per
crit
ical
soil
con
cen
trat
ion
for
bo
thP
b
and
Zn
Tw
osp
ecie
so
fsu
nfl
ow
er—
Tit
ho
nia
div
ersi
foli
aan
d
Hel
ian
thu
sa
nn
uu
s
T.
div
ersi
foli
am
op
ped
up
sub
stan
tial
con
cen
trat
ion
so
fP
bin
the
abo
ve-
gro
un
d
bio
mas
sco
mp
ared
toco
nce
ntr
atio
ns
inth
ero
ots
.T
he
con
cen
trat
ion
sin
the
leaf
com
par
tmen
tw
ere
87
.3,
71
.3,
and
71
.5m
g/k
gat
4,
6,
and
8w
eek
saf
ter
pla
nti
ng
(AP
),re
spec
tiv
ely
.In
roo
ts,
itw
as9
9.4
,9
7.4
,an
d7
7.7
mg
/kg
.
Ob
serv
atio
ns
wit
hH
.a
nn
uu
sfo
llo
wed
the
pat
tern
fou
nd
wit
hT
.d
iver
sifo
lia
,sh
ow
ing
sig
nifi
can
tac
cum
ula
tio
no
fP
bin
the
abo
ve-
gro
un
db
iom
ass
Ad
eso
du
net
al.
(20
10
)
Ars
enic
Aru
nd
od
on
ax
for
ph
yto
extr
acti
on
of
arse
nic
fro
msy
nth
etic
was
tew
ater
Incr
easi
ng
As
con
cen
trat
ion
inn
utr
ien
tso
luti
on
cau
sed
anin
crea
sein
sho
ot
and
roo
tb
iom
ass
wit
ho
ut
tox
icit
ysy
mp
tom
sin
A.
do
na
xg
row
ing
un
der
ara
ng
eo
f
As
con
cen
trat
ion
fro
m5
0to
60
0lg
/l.
Th
eA
sd
ose
su
pto
60
0l
g/l
did
no
t
affe
ctth
eg
row
tho
fA
.d
on
ax.
Itw
assu
gg
este
dth
atA
.d
on
ax
pla
nts
may
be
emp
loy
edto
trea
tco
nta
min
ated
wat
ers
con
tain
ing
arse
nic
con
cen
trat
ion
su
pto
60
0lg
/l
Mir
zaet
al.
(20
10
)
Rev Environ Sci Biotechnol (2010) 9:215–288 237
123
Ta
ble
10
con
tin
ued
Hea
vy
met
al(s
)P
lan
t(s)
use
dR
emo
val
per
form
ance
Ref
eren
ce
Mer
cury
Wat
erh
yac
inth
(Eic
ho
rnia
cra
ssip
es),
wat
erle
ttu
ce(P
isti
ast
rati
ote
s),
zeb
raru
sh(S
cirp
us
tab
ern
aem
on
tan
i)an
dta
ro
(Co
loca
sia
escu
len
ta)
Co
ldv
apo
rA
tom
icA
bso
rpti
on
Sp
ectr
osc
op
yco
nfi
rmed
anin
crea
seo
fm
ercu
ry
wit
hin
the
pla
nt
roo
tti
ssu
ean
da
corr
esp
on
din
gd
ecre
ase
of
mer
cury
inth
e
wat
er.
All
spec
ies
of
pla
nts
app
eare
dto
red
uce
mer
cury
con
cen
trat
ion
s
inth
ew
ater
via
roo
tu
pta
ke
and
accu
mu
lati
on
.W
ater
lett
uce
and
wat
er
hy
acin
thap
pea
red
tob
eth
em
ost
effe
ctiv
e,fo
llo
wed
by
taro
and
zeb
raru
sh,
resp
ecti
vel
y
Sk
inn
eret
al.
(20
07
)
Man
gan
ese
(1m
g/l
Mn
fro
m
syn
thet
icw
aste
wat
ers
in
con
stru
cted
wet
lan
ds)
Wat
erh
yac
inth
(Eic
hh
orn
iacr
ass
ipes
(Mar
t.)
So
lms)
Ph
yto
rem
edia
tio
nm
ain
lyd
ue
top
hy
toex
trac
tio
nsu
bst
anti
ally
con
trib
ute
dto
man
gan
ese
rem
ov
al.
Ho
wev
er,
chem
ical
pre
cip
itat
ion
was
abse
nt,
sug
ges
tin
g
that
man
gan
ese
has
ah
igh
erso
lub
ilit
yin
the
giv
enav
erag
ep
H(6
.2–
7.1
)
con
dit
ion
sin
con
stru
cted
wet
lan
ds
Ku
lara
tne
etal
.
(20
09
)
Man
gan
ese
Ph
yto
lacc
aa
mer
ica
na
(po
kew
eed
)P
.a
mer
ica
na
no
tsh
ow
edre
mar
kab
leto
lera
nce
toM
n.
Max
imu
mM
n
con
cen
trat
ion
inth
ele
afd
rym
atte
rw
as8
,00
0m
g/g
on
Xia
ng
tan
Mn
tail
ing
s
was
tela
nd
s.P
.a
mer
ica
na
was
char
acte
rize
db
ya
hig
htr
ansl
oca
tio
nfa
cto
ro
f
mo
reth
an1
0.7
6.
Un
der
nu
trie
nt
solu
tio
ncu
ltu
reco
nd
itio
ns,
man
gan
ese
con
cen
trat
ion
inth
esh
oo
tsin
crea
sed
wit
hin
crea
sin
gex
tern
alM
nle
vel
s,an
d
reac
hed
am
axim
um
con
cen
trat
ion
of
Mn
inle
aves
at4
7.0
6g
/kg
.P
ok
ewee
d
was
thu
scl
assi
fied
asa
new
man
gan
ese
hy
per
accu
mu
lato
rp
lan
t
Min
etal
.(2
00
7)
Co
bal
tC
ob
alt
giv
ento
soy
bea
n(G
lyci
ne
ma
x)p
lan
tsin
po
tcu
ltu
reb
yso
il
dre
nch
ing
met
ho
d
Res
ult
ssh
ow
edh
igh
erco
nce
ntr
atio
n(C
ole
vel
(10
0–
20
0m
g/k
g)
inth
eso
il)
resu
lted
inm
axim
um
accu
mu
lati
on
inal
lp
arts
of
soy
bea
np
lan
ts,w
hil
eth
elo
w
con
cen
trat
ion
so
fco
bal
t(5
0m
g/k
gC
ole
vel
)in
the
soil
did
n’t
sho
wan
y
sig
nifi
can
tef
fect
Jay
aku
mar
and
Jale
el
(20
09
)
238 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Ta
ble
11
Ap
pli
cati
on
of
ph
yto
rem
edia
tio
nfo
rre
mo
val
of
org
anic
po
llu
tan
ts
Org
anic
po
llu
tan
t(s)
Pla
nt(
s)u
sed
Rem
ov
alp
erfo
rman
ceR
efer
ence
Ob
sole
tep
esti
cid
es—
PO
Ps
pes
tici
des
incl
ud
ing
met
abo
lite
so
fD
DT
(dic
hlo
rod
iph
eny
ltri
chlo
roet
han
e)an
d
iso
mer
so
fH
CH
(hex
ach
loro
cycl
oh
exan
e)
Th
eK
azak
hst
anm
axim
um
acce
pta
ble
con
cen
trat
ion
for
DD
Tan
dH
CH
met
abo
lite
sin
pla
nt
tiss
ue
is
20
lg/k
g.
Sp
ecie
sin
this
cate
go
ryin
clu
ded
:
Art
emis
iaa
nn
ua
L.,
Ko
chia
siev
ersi
an
a(P
all.
)C
.A.
Mey
.K
och
iasc
op
ari
a(L
.)S
chra
d.,
and
Xa
nth
ium
stru
ma
riu
mL
Th
ree
spec
ies
exce
eded
the
MA
Cb
yu
pto
90
tim
es
incl
ud
ing
A.
an
nu
a,
Am
bro
sia
art
emis
iifo
lia
L.,
and
Eri
ger
on
can
ad
ensi
sL
.M
ost
pes
tici
des
accu
mu
late
d
inth
ero
ot
syst
ems;
ho
wev
er,
amo
ng
the
spec
ies
inv
esti
gat
ed,
K.
sco
pa
ria
,A
.a
nn
ua
,B
arb
are
avu
lga
ris
W.
T.
Ait
on
,an
dA
.a
rtem
isii
foli
ad
emo
nst
rate
dca
pab
ilit
ies
totr
ansl
oca
tep
esti
cid
es
fro
mro
ots
toab
ov
egro
un
dti
ssu
es
Nu
rzh
ano
va
etal
.(2
01
0)
Her
bic
ides
atra
zin
ean
dm
eto
lach
lor
P4
50
gen
esC
YP
1A
1,
CY
P2
B6
,an
dC
YP
2C
19
inri
ce
pla
nts
(Ory
zasa
tiva
cv.
Nip
po
nb
are)
intr
od
uce
d
usi
ng
the
pla
smid
pIK
BA
CH
Th
etr
ansg
enic
rice
pla
nts
(pIK
BA
CH
rice
pla
nts
)
bec
ame
mo
reto
lera
nt
tow
ard
var
iou
sh
erb
icid
esth
an
no
ntr
ansg
enic
Nip
po
nb
are
rice
pla
nts
Kaw
ahig
ash
i
etal
.(2
00
6)
Co
pp
ersu
lph
ate
(fu
ng
icid
e),
flaz
asu
lfu
ron
(her
bic
ide)
and
dim
eth
om
orp
h(f
un
gic
ide)
Lem
na
min
or
(L.
min
or)
,E
lod
eaca
na
den
sis
(E.
can
ad
ensi
s)an
dC
ab
om
ba
aq
ua
tica
(C.
aq
ua
tica
)
To
xic
ity
of
the
con
tam
inan
tsw
asth
esa
me
for
all
the
aqu
atic
pla
nts
stu
die
dan
do
ccu
rred
inth
isd
esce
nd
ing
ord
ero
fto
xic
ity
:
flaz
asu
lfu
ron
[co
pp
er[
dim
eth
om
orp
h.
L.
min
or
had
the
mo
stef
fici
ent
up
tak
eca
pac
ity
,fo
llo
wed
by
E.
can
ad
ensi
san
dth
enC
.a
qu
ati
ca.
Th
em
axim
um
rem
ov
alra
teo
fco
pp
er,
flaz
asu
lfu
ron
and
dim
eth
om
orp
hw
as3
0,
27
and
11
lg
/gfr
esh
wei
gh
t/
day
,re
spec
tiv
ely
Ole
tte
etal
.
(20
08
)
Sel
ecti
ve
syst
emic
her
bic
ide
2,4
-
dic
hlo
rop
hen
ox
yac
etic
Pea
(Pis
um
sati
vum
),w
ith
ag
enet
ical
lyta
gg
ed
bac
teri
alen
do
ph
yte
Res
ult
ssh
ow
edth
atth
est
rain
test
edh
adac
tiv
ely
colo
niz
edin
ocu
late
dp
lan
tsin
tern
ally
(an
din
the
rhiz
osp
her
e).
Ino
cula
ted
pla
nts
sho
wed
ah
igh
er
cap
acit
yfo
r2
,4-d
ich
loro
ph
eno
xy
acet
icac
idre
mo
val
fro
mso
ilan
dsh
ow
edn
o2
,4-d
ich
loro
ph
eno
xy
acet
ic
acid
accu
mu
lati
on
inth
eir
aeri
alti
ssu
es
Ger
mai
ne
etal
.(2
00
6)
Wea
ther
edp,p0 -
DD
Ein
soil
Th
ree
cult
ivar
so
fzu
cch
ini
(Cu
curb
ita
pep
osp
p.
pep
ocv
Co
stat
aR
om
anes
co,
Go
ldru
sh,
Rav
en)
To
tal
amo
un
to
fco
nta
min
ant
ph
yto
extr
acte
dd
uri
ng
the
62
day
gro
win
gse
aso
nra
ng
edfr
om
0.7
2–
2.9
%
Wh
ite
etal
.
(20
06
)
Dic
ofo
l(a
no
rgan
och
lori
ne
pes
tici
de)
Wat
erh
yac
inth
(Eic
hh
orn
iacr
ass
ipes
)A
fter
10
day
so
fin
cub
atio
nin
nu
trie
nt
solu
tio
nat
25±
1�C
,th
ere
mai
nin
gd
ico
fol
wh
ich
was
spik
ed
init
iall
yat
1m
g/l
was
0.0
5an
d0
.26
mg
/lin
the
no
n-
ster
ile
pla
nte
dan
dn
on
-ste
rile
un
pla
nte
d,
0.0
7an
d
0.3
1m
g/l
inth
est
eril
ep
lan
ted
and
ster
ile
un
pla
nte
d
trea
tmen
ts,
resp
ecti
vel
y
Xia
(20
08
)
Rev Environ Sci Biotechnol (2010) 9:215–288 239
123
metals or simply incineration and disposal. The
mechanisms of root uptake and transport within the
plants are poorly understood. Solubilized metal ions
may enter the root by two pathways: apoplastically
(extracellularly), or symplastically (intracellularly)
(Salt et al. 1995). Metal ions tend to enter plant cells
via metal ion carriers or channels in energy dependent,
saturable process (Clarkson and Luttge 1989). Non-
essential heavy metals may compete with essential
metals for these transmembrane carriers, which may
explain the ability of some non-essential metals to
enter the cell against the concentration gradient (Salt
et al. 1995). Upon entering the roots, metals can either
be stored or transported to the shoot. Xylem transport
is thought to be responsible for transport to the shoot,
although metals may disperse throughout the shoot via
the phloem (Salt et al. 1995). Several wild plants with
the ability to accumulate very large concentrations of
metals in their roots and shoots have been identified
and are termed as hyperaccumulators (Table 12).
Approximately 400 hyperaccumulator species have
been identified, according to the analysis of field-
collected specimens. Metal hyperaccumulators are
interesting model organisms to study for the develop-
ment of a phytoremediation technology, the use of
plants to remove pollutant metals from soils (Kramer
et al. 1997).
Botanists first recognized the metal-accumulating
ability of the genus Thlaspi over a century ago. Some
have suggested this ability may have evolved as a
defense against herbivores (Baker et al. 1994). Baker
and co-workers were able to grow some individual
Thlaspi caerulscens plants that accumulated more
than 30,000 mg of Zn and over 1,000 mg Pb per
gram dried biomass (Baker et al. 1994). They also
showed that the five isolated populations of the plant
were adapted in tolerating and accumulating metals
not present in the parent soil, which suggested the
mechanisms of tolerance and accumulation may be
similar for different metals.
While wild metal-accumulators such as T. cae-
rulescens have shown an impressive ability to
accumulate metals, they tend to be slow growing
and small in size. The use of such plants for
phytoextraction may therefore require an unreason-
able number of harvests to decontaminate a given
site. Ideally, a plant well suited for phytoextraction
should tolerate and accumulate metals, grow rapidly,
and have the potential to produce a high biomass.Ta
ble
11
con
tin
ued
Org
anic
po
llu
tan
t(s)
Pla
nt(
s)u
sed
Rem
ov
alp
erfo
rman
ceR
efer
ence
Clo
fib
ric
acid
(CA
wh
ich
isa
met
abo
lite
of
blo
od
lip
idre
gu
lato
rd
rug
s)
Typ
ha
spp
.A
ta
con
cen
trat
ion
of
20
lg
/l,
Typ
ha
had
rem
ov
ed
[5
0%
of
CA
wit
hin
the
firs
t4
8h
,re
ach
ing
a
max
imu
mo
f8
0%
by
the
end
of
the
assa
y.
Ex
per
imen
tal
con
dit
ion
sas
sure
dth
at
ph
oto
deg
rad
atio
n,
adso
rpti
on
tov
esse
lw
alls
and
mic
rob
ial
deg
rad
atio
nd
idn
ot
con
trib
ute
toth
e
rem
ov
al
Do
rdio
etal
.
(20
09
)
Tw
ofu
ng
icid
es—
dim
eth
om
orp
h
and
py
rim
eth
anil
Fiv
em
acro
ph
yte
spec
ies—
L.
min
or,
S.
po
lyrh
iza
,
C.
aq
ua
tica
,C
.p
alu
stri
san
dE
.ca
na
den
sis
Th
ere
mo
val
yie
lds
du
rin
gth
e4
-day
test
per
iod
sv
arie
d
fro
m1
0to
18
%an
d7
–1
2%
for
dim
eth
om
orp
han
d
py
rim
eth
anil
,re
spec
tiv
ely
.T
he
max
imu
mre
mo
val
rate
du
rin
gth
e4
-day
test
per
iod
was
48
lg
/gfr
esh
wei
gh
t(F
W)
for
dim
eth
om
orp
han
d3
3l
g/g
FW
for
py
rim
eth
anil
.L
.m
ino
ran
dS
.p
oly
rhiz
ash
ow
edth
e
hig
hes
tre
mo
val
effi
cien
cyfo
rth
etw
ofu
ng
icid
es
Do
sno
n-
Ole
tte
etal
.
(20
09
)
240 Rev Environ Sci Biotechnol (2010) 9:215–288
123
In the attempt to find such a plant, Dushenkov and
co-workers have evaluated the heavy-metal accumu-
lating abilities of several high biomass crop species
such as Brassica juncea (Indian mustard) (Dushen-
kov et al. 1995). Although Thlaspi caerulescens has a
higher tolerance for the heavy metals tested and
demonstrated a higher ratio of metal accumulation to
plant mass. Brassica juncea produces 20 times more
biomass. They found Brassica juncea to be especially
adapted in accumulating lead. One strain was able to
accumulate Pb at up to 3.5% dry weight in the shoots
suggesting that a crop of such plants could extract
630 kg ha-1 of Pb in above ground biomass with a
single harvest, more if some root material was
harvested as well (Dushenkov et al. 1995). It is
important to note however, that these experiments
were done hydroponically. In soil, the property of Pb
to bind to clay soil particles and organic matter, and
its inclusion in insoluble precipitates significantly
reduces the bioavailability of Pb to the plant
(Dushenkov et al. 1995). This is true of other metals
as well. For this reason, biological mechanisms that
enhance metal bioavailability are being investigated.
For instance, in response to nutrient deficiencies,
plants can secrete metal-chelating molecules (phyt-
osiderophores) that chelate and solubilize soil-bound
metals such as Fe, Mn, Cu, and Zn (Salt et al. 1995).
Also, plants can reduce soil bound metal ions by
specific plasma membrane-bind metal reductases; and
they can adjust soil pH, which decreases metal
adsorption, by exuding protons (Salt et al. 1995).
Lastly, the presence of some microorganisms in the
rhizosphere has been shown to enhance metal
availability.
Arsenic hyperaccumulators: Arsenic hyperaccu-
mulators bioconcentrate arsenic over 2,000 mg/kg in
plant tissues (Bondada and Ma 2003). In addition, a
good arsenic phytoextraction species should accumu-
late more arsenic in shoots than in roots because for
an easy harvest or removal of arsenic—laden
above—ground biomass. The concentration of the
contaminant is generally very high in these plants
when grown in contaminated media. To compare the
levels of bioconcentation and distribution of arsenic
in plants a bioconcentration factor (BF) and transfer
factor (TF) can be used. The BF of arsenic is the ratio
of the arsenic concentration in plant to the concen-
tration in soil, while the TF is the ratio of the arsenic
concentration in roots to the concentration in shoots.
Ma et al. (2001) discovered an arsenic-hyperaccu-
mulator species-Chinese brake fern (Pteris vittata)
that accumulated arsenic in the shoots to a concen-
tration as high as 22,000 mg/kg (Huang et al. 2004).
Research has demonstrated that other species in the
Pteris genus also hyperaccumulate arsenic in their
shoots. Greenhouse studies (Salido et al. 2003)
indicated that P. vittata accumulated an arsenic
concentration in the above ground plant tissue more
than 200-fold higher than most other plant species
tested using arsenic-contaminated soil. In addition,
this species grows rapidly and generates substantial
amounts of biomass, thus making P. vittata an
excellent candidate to rapidly remove arsenic from
arsenic-contaminated environments. The Chinese
Table 12 Heavy metal hyperaccumulators
Metal Hyperaccumulator plant species References
Arsenic Pteris vittata, Pityrogramma calomelanos Wang et al. (2002), Lombi et al. (2002), Tongbin et al. (2002),
Francesconi et al. (2002), Tu et al. (2003)
Cadmium Thlaspi caerulescens, tumbleweed (Salsola kali),Solanum nigrum L
Pence et al. (2000), de la Rosa et al. (2004), Wei et al. (2006)
Zinc Thlaspi caerulescens, Arabidopsis halleri,Thlaspi praecox Wulfen
Shen et al. (1997), Pence et al. (2000), Zhao et al. (2000),
Sarret et al. (2002), Vogel-Mikus et al. (2006)
Nickel Alyssum lesbiacum, Alyssum bertolonii, Thlaspigoesingense, Berkheya coddii, Sebertia acuminate,
Alyssum murale
Kramer et al. (1997), Kupper et al. (2001), Robinson et al.
(1997), Sagner et al. (1998), Bernal et al. (1994)
Lead Thlaspi praecox Wulfen Vogel-Mikus et al. (2006)
Copper Aeolanthus biformifolius (Labiatae); Commelinacommunis
Malaisse et al. (1997), Wang et al. (2004)
Rev Environ Sci Biotechnol (2010) 9:215–288 241
123
brake fern is a primitive plant which thrives on
arsenic, doubling its biomass in 1 week when
subjected to 100 mg/l arsenic. The striking difference
between P. vittata and arsenic non-accumulators is
the remarkable transport of arsenic from roots to
shoots in P. vittata, accumulating up to 95% of the
arsenic in the above-ground tissue (Doucleff and
Terry 2002). Many other ferns in the Pteris genus like
P. longifolia, P. cretica and P. umbrosa (Zhao et al.
2002), as well as a non-Pteris fern, Pityrogramma
calomelanos (Visoottiviseth et al. 2002) have also
been found to hyperaccumulate arsenic. However, not
all members of the Pteris genus are able to hyper-
accumulate arsenic. Meharg (2003) found that Pteris
tremula and Pteris stramina do not hyperaccumulate
arsenic. To date the only non-Pteris fern to exhibit
this ability is Pityrogramma calomelanos (Frances-
coni et al. 2002). Srivastava et al. (2006) showed
that Pteris biaurita L., P. quadriaurita Retz and
P. ryukyuensis Tagawa could be used as hyperaccu-
mulators of arsenic with the average arsenic concen-
tration ranging from 1,770 to 3,650 mg/kg dry
weight (DW) in the fronds and 182–507 mg/kg DW
in the roots of P. cretica, P. biaurita, P. quadriaurita
and P. ryukyuensis after having been grown in
100 mg As/kg soil.
Soil amendments that can increase metal avail-
ability are also being studied. Chelating agents have
been investigated by Blaylock, Raskin and co-work-
ers for their ability to prevent precipitation and
sorption of metals (Dushenkov et al. 1995). B. juncea
seedling grown for 4 weeks in soil treated with a
chelating agent accumulated 875 mg/kg DW Cd in
the shoot, compared to 164 mg/kg DW Cd by
seedlings grown in untreated soils. Additionally,
numerous studies have demonstrated that lowering
soil pH increases metal availability.
6.6.3 Rhizofiltration
Rhizofiltration is defined as the use of plant roots to
absorb, precipitate, and concentrate toxic metals from
water (Salt et al. 1995). In engineered systems, metal-
contaminated water may be passed through a network
of roots, which can then be harvested, dried and either
combusted and discarded or subjected to process to
reclaim metals from plant biomass. Given its potential
for treating high volumes of water contaminated with
low concentration of metals, rhizofiltration enthusiasts
assert that it will be a cost-effective treatment
for everything from industrial wastewater, to agricul-
tural runoff, to contaminated surface water and
groundwater.
The mechanisms for root accumulation include: (1)
surface sorption, in which physical/chemical processes
such as chelation and ion exchange lead to sorption by
the root; (2) biological processes, including intracel-
lular uptake, vacuolar deposition, and translocation to
the shoots (Chaney 1983); and (3) root remediated
precipitation, which probably involves the release of
root exudates (Dushenkov et al. 1995). Surface sorp-
tion tends to be the fastest of these, especially in the
case of Pb, and root—remediated precipitation the
slowest, although the relative importance of these
different mechanisms is dependent on concentration.
At low concentration surface sorption dominates, at
higher concentrations, however, when sites for sorp-
tion are saturated, biological processes and precipita-
tion assume more importance. These mechanisms also
differ between metals and plant species.
Dushenkov and co-workers have tested the ability
of hydroponically grown plants to remove toxic
metals from aqueous waste streams (Dushenkov et al.
1995). The ability of the high biomass crop plant
Brassica juncea to accumulate Pb in the roots was
compared to 24 other plant species such as Helanthus
annus (sunflower), and various grasses such as
colonial bentgrass and Poapratensis (Kentucky blue-
grass). While all the species tested demonstrated
significant root accumulation of Pb, Brassica juncea
(14% dry weight Pb in the roots) possessed the most
favorable combination of metal accumulating ability
and high-biomass production. Additionally, Brassica
juncea was shown to accumulate significant amounts
of Cu2? Cd2?, Cr 2?, Ni 2? and Zn2? as well. Given
that the magnitude of metals accumulated by way of
surface sorption is proportional to root mass, the
ability of Brassica juncea to generate a large mass of
roots quickly and economically makes it a promising
candidate for rhizofiltration.
6.6.4 Phytostabilization
Phytostabilization of inorganics involve the use
of metal tolerant plants to reduce the mobility of
metals in the subsurface. Soils contaminated with
toxic metals lack vegetation due to either physical
disturbance or toxic effects of the contamination.
242 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Metal contamination of exposed soils is often more
mobile due to leaching and transport by wind and
water. Metal tolerant plants may be useful in reducing
metal mobility that results from these mechanisms.
This strategy has been used successfully to stabilize
metalliferous mine-wastes by a group in Liverpool
(Cunningham and Berti 1993). Also, Salt and
co-workers were able to demonstrate that seedlings
of B. juncea were able to reduce the level of Pb
leached from contaminated soils into the groundwater
(Salt et al. 1995).
At the NEERI, Nagpur, extensive work on phyto-
stabilization of coal mine dumps, manganese mine
dumps, fly ash dumps and metalliferous mine wastes
has been carried out using IBA (Juwarkar et al. 2000).
Legumes are especially well suited for reestablishing
vegetation and stabilizing degraded and metal con-
taminated soils. This is due to several factors: first,
they accumulate nitrogen in a mineralizable form in
symbiosis with rhizobia, which then becomes avail-
able to non-leguminous plants; second, they are able to
grow in low nutrient conditions; and third, they are
able to colonize barren habitats which are subject to
strong winds and flooding (Jha et al. 1995). Jha and
co-workers studied how this rhizobial symbiosis
influences the ability of several wild legumes to
revegetate an unreclaimed limestone quarry. They
found that seeds encapsulated with polyacrylamide-
entrapped rhizobia showed higher establishment, sur-
vival, and subsequent growth than uninnoculated
seedlings. Mycorrhizae inoculation to leguminous
and nonleguminous tree species of resulted in rapid
reclamation of mine spoil dumps (Juwarkar et al. 1992,
1997, 2000).
6.6.5 Phytodegradation
Phytodegradation is the breakdown of organics, taken
up by the plant to simpler molecules that are incorpo-
rated into the plant tissues (Chaudhry et al. 1998).
Plants contain enzymes that can breakdown and
convert ammunition wastes, chlorinated solvents
(such as trichloroethylene) and other herbicides. The
enzymes include usually dehalogenases, oxygenases
and reductases (Black 1995). Rhizodegradation is the
breakdown of organics in the soil through microbial
activity of the root zone (rhizosphere). Soil microor-
ganisms can utilize organic pollutants as their carbon
and energy sources. Indeed, all phytoremediation
processes or technologies are not exclusive and may
be used simultaneously. For example, a constructed
treatment wetland may involve all the phytoremedia-
tion processes for the cleanup of wastewaters contam-
inated with both metals and organic compounds.
6.6.6 Phytovolatilization
Phytovolatilization involves the use of plants and
plant-associated soil microbes to take up contaminants
from the soil, transform them into volatile forms and
release them into the atmosphere (Lin 2008). Phyto-
volatilization occurs as growing trees and other plants
take up water and the organic and inorganic contam-
inants. Metalloids, such as selenium, arsenic, and tin,
can be methylated to volatile compounds or mercury
that can be biologically transformed to elemental Hg.
Phytovolatilization has been primarily used for the
removal of mercury and selenium.
7 Ex situ bioremediation
7.1 Composting
Composting is the biochemical degradation of organic
materials to a sanitary, nuisance-free, humus-like
material (Kulcu and Yildiz 2004). Composting has
been defined as a controlled microbial aerobic
decomposition process with the formation of stabi-
lized organic materials that may be used as soil
conditioner (Negro et al. 1999). The main factors in
the control of a composting process include environ-
mental parameters (temperature, moisture content, pH
and aeration) and substrate nature parameters (C/N
ratio, particle size, and nutrient content) (Diaz et al.
2002; Artola et al. 2009). Aerobic composting is the
decomposition of organic substrates in the presence of
sufficient oxygen (Agnew and Leonard 2003). The
main products of the biological metabolism are
carbon dioxide, water and considerable amounts of
heat (Ghaly et al. 2006). Various factors correlate with
each other physically, chemically and biologically in
complicated composting processes (Agnew and Leon-
ard 2003). A slight change in a single factor may cause
a drastic avalanche of metabolic and physical changes
in the overall process. In other words, there may be
extremely strong non-linearities involved in these
processes (Seki 2000). These processes occur in
Rev Environ Sci Biotechnol (2010) 9:215–288 243
123
matrix of organic particles and interconnected pores,
and the pores are partially filled with air, aqueous
solution, or a combination of the two (Richard et al.
2006). A multitude of microorganisms and their
enzymes is responsible for the biodegradation process
(Fogarty and Tuovinen 1991), resulting in a complex
biochemical–microbial system.
Because of its complicated and dynamic nature, the
composting process is one of the most intractable
processes from an engineering point of view although
the macroscopic process kinetics have been well
engineered to date to remediate a wide variety of
organic wastes namely municipal solids wastes, poul-
try litter, wastes vegetables, food processing residuals,
and sludge from wastewater treatment plants and other
sludge generating processes. Under optimal condi-
tions, composting proceeds from the psychrophilic
state through three phases: (a) the mesophilic or
moderate-temperature phase, (b) the thermophilic or
high temperature phase, and (c) the cooling and
maturation phase which lasts for several months
(Mohee et al. 2008). The first, second and third phases
are referred to as the active stage in which heat is
produced (Ghaly et al. 2006). This active stage is
governed by the basic principles of heat and mass
transfer (Keener et al. 1993; Mudhoo and Mohee
2008) and by the biological constraints of living
microorganisms.
Many anthropogenic organic contaminants entering
the environment are not fully degraded during treat-
ment and eventually accumulate in biosolids (Ang
et al. 2005; Bhandari and Xia 2005). Due their
relatively low water solubility and high lipophilicity
(Bhandari and Xia 2005), organic contaminants easily
partition into biosolids resulting in their accumulation
in biosolids at concentrations several orders of mag-
nitude greater than influent concentrations (Govind
et al. 1991). The following sections now present and
discuss selected research findings for the application of
composting in bioremediating such organic contami-
nants, namely, PAHS, petroleum-based hydrocarbons,
phenol derivatives, polychlorinated biphenyls, phtha-
lic acid esters (PAEs) and pesticides.
7.1.1 Vermistabilisation
Vermicomposting is the term given to the process of
conversion of biodegradable matter by earthworms
into vermicompost (Garg et al. 2006; Tognetti et al.
2005). In the process, a major fraction of the nutrients
contained in the organic matter is converted to more
bioavailable forms. The first step in vermicomposting
occurs when earthworms break the substrate down to
small fragments prior to ingesting the substrate
(Gajalakshmi and Abbasi 2008). This increases the
surface area of the substrate, facilitating microbial
and enzymatic actions. The substrate is then ingested
and goes through a process of ‘‘enzymatic digestion’’
brought about by numerous species of bacteria and
enzymes present in the worms’ gut (Gajalakshmi and
Abbasi 2008).
Due to their biological, chemical and physical
actions, earthworms can be directly employed to
promote biodegradation of organic contaminants in
bioremediation processes (Hickman and Reid 2008).
Earthworms have been shown to aerate and biotur-
bate soils and thence improve their nutritional status
and fertility, which are normally variables known to
limit bioremediation rates (Hickman and Reid 2008).
Earthworms also hinder processes during which
organic contaminants bind to soils, and thus promote
the dispersion and bioavailablity of organic contam-
inants to the degrading microorganisms (Hickman
and Reid 2008). Earthworms in general are tolerant to
many chemical contaminants including heavy metals
and organic pollutants in soil and can bio-accumulate
them in their tissues (Sinha et al. 2008). Earthworms
species like Eisenia fetida, Eisenia tetraedra, Lum-
bricus terrestris, Lumbricus rubellus and Allobopho-
ra chlorotica have been found to remove heavy
metals (Cd, Pb, Cu and Hg) pesticides and lipophilic
organic micropollutants like the PAH from the soil
(Sinha et al. 2008). Therefore, by using these
excellent properties of earthworms, the vermicompo-
sting process has been employed to degrade organic
pollutants like PAHs, PCBs (Contreras-Ramos et al.
2009), atrazine and metamitron (Forouzangohar et al.
2005). Table 13 summarizes a few studies where
vermicomposting has been employed to bioremediate
such contaminants.
7.1.2 PAHs remediation by composting
Polycyclic aromatic hydrocarbons are a class of
organic compounds that have accumulated in the
natural environment mainly as a result of anthropo-
genic activities such as the combustion of fossil fuels
(Bamforth and Singleton 2005; Johnsen et al. 2005).
244 Rev Environ Sci Biotechnol (2010) 9:215–288
123
The increasing use of fossil fuels and their combus-
tion products by human beings during the two past
centuries raises several questions about PAHs haz-
ards for living organisms. First, apart from accidental
oil spills leading to massive pollutions, the precise
origin of trace PAHs, e.g., natural versus anthropo-
genic, has rarely been clear traced. Second, the
toxicity of PAHs, like other hazardous chemicals,
requires their bioavailability. And since most PAHs
are highly hydrophobic (Wild and Jones 1992), their
pathways of transfer through geological and biolog-
ical media are far from being comprehensively
understood. Third, explicit correlations between
PAH sources and carcinogenic effects have been
reported only for intense exposure to PAHs such as
for coal–mine workers. PAHs structure and stability
stand in the way of their biodegradation by micro-
organisms (fungi and bacteria). Biodegradation is
slow and is a function of environmental parameters
such as oxygen, water and nutriment contents.
Interest has continuously surrounded the occurrence
and distribution of PAHs for many decades due to
their potentially harmful effects to human health
(Juhasz and Naidu 2000). Although various physico-
chemical methods have been used to remove these
compounds from our environment, they have many
limitations (Samanta et al. 2002). This concern has
prompted researchers to address ways to detoxify and
remove these organic compounds from the natural
environment (Bamforth and Singleton 2005). Biore-
mediation is one approach that has been used to
remediate contaminated land and waters, and that has
promoted the natural attenuation of the contaminants
using the in situ microbial community of the site.
PAHs are recalcitrant and can persist in the
environment for long periods, but are conducive to
biodegradation by certain enzymes found in bacteria
and fungi (Juhasz and Naidu 2000; Ang et al. 2005).
In the past several years, several oxidoreductases
such as laccases and cytochrome P450 monooxyge-
nases have been exploited for the enzymatic degra-
dation of PAHs. Composting has been applied as a
Table 13 Degradation of chemical contaminants through vermicomposting
Contaminant(s)/contaminate
media
Earthworm
species
Vermistabilizaion performnace References
Polychlorinated biphenyls Eiseniafetida
Results demonstrated that earthworms survived and reproduced
in the presence of contaminated media
Biomass increase decreased rapidly with increasing mass
fraction of sludge, and biomass increased ranged from 103%
in the negative control to biomass reduction of 54 with 75%
sludge
Gas chromatography results demonstrated an 80% reduction
in PCB level in all vermicomposting bioreactors
Tharakan et al.
(2004)
Beverage industry bio sludge Eiseniafetida
Degradation of 50:50 mixture of bio sludge and cattle dung
could be achieved in 75 days when worms were inoculated
at 25 g kg-1 feed mixture
Singh et al.
(2010)
Distillery industry sludge mixed
with a bulking agent (cow
dung)
Perionyxexcavatus
Inoculated earthworms could maximize the decomposition and
mineralization rate when sludge was used with appropriate
bulking material for earthworm feed
Vermicomposting also caused significant reduction in total
concentration of metals: Zn (15.1–39.6%), Fe (5.2–29.8%),
Mn (2.6–36.5%) and Cu (8.6–39.6%) in sludge
Suthar and
Singh (2008)
Phenanthrene, anthracene and
benzo(a)pyrene
Eiseniafetida
Average anthracene removal by the autochthonous
microorganisms was 23, 77% for phenanthrene and 13%
for benzo(a)pyrene, while it was 51% for anthracene, 47%
for benzo(a)pyrene and 100% for phenanthrene in soil with
earthworms. At 50 and 100 mg phenanthrene/kg E. fetidasurvival was 91 and 83%, but at 150 mg/kg all died within
15 days. Survival of E. fetida in soil amended with
anthracene B 1,000 mg/kg and benzo(a)pyrene B 150 mg/kg
was higher than 80% and without weight loss compared to the
untreated soil
Contreras-
Ramos et al.
(2006)
Rev Environ Sci Biotechnol (2010) 9:215–288 245
123
bioremediation technique for degrading toxic organic
compounds and perhaps lowering their persistence
and toxicity in organic residues and wastes (Barker
and Bryson 2002). The biochemical and physico-
chemical processes of remediation in composts are
similar to those that usually occur biologically in soil.
However, composting may accelerate the destruction
of organic contaminants since metabolic temperatures
developed are generally higher in composts than in
soils. This remediation characteristic of composts and
composting matrices has been successfully explored
and exploited for the degradation of PAHs.
Al-Daher et al. (2001) selected the bioremediation
technique involving the use of composting soil piles
from among the most appropriate methods and
evaluated its performance to remediate PAHs on a
pilot scale. Soil piles were constructed from the
contaminated soil after amendment with necessary
soil additives and the piles were subjected to regular
irrigation and turning, and a monitoring program was
carried out, including monthly soil sample collection
from each pile for the measurement of petroleum
hydrocarbon PAHs, soil microbial counts, mineral
and metal concentrations. Al-Daher et al. (2001)
found that the composting soil pile treatment resulted
in the reduction of up to 59% total extractable matter
of oil contamination within 8 months of the com-
posting process. More interestingly, Reid et al. (2002)
studied the catabolism of phenanthrene within mush-
room compost resulting from its incubation with (1)
phenanthrene, and (2) PAH-contaminated soil. Res-
pirometers measuring mineralization of freshly added14C-9-phenanthere were used to evaluate the induc-
tion of phenanthrene-catabolism. Where pure phen-
anthrene spiked at a concentration of 400 mg/kg wet
weight was used to induce phenanthrene-catabolism
in compost, induction was measurable, with maximal
mineralization observed after 7 weeks phenanthrene-
compost contact time. Where PAH-contaminated soil
was used to induce phenanthrene-catabolism in un-
induced compost, induction was observed after
5 weeks soil-compost contact time. Microcosm-scale
amelioration of soil contaminated with 14C-phenan-
threne (aged in soil for 516 days prior to incubation
with compost) indicated that both induced (using pure
phenanthrene) and uninduced mushroom composts
were equally able to promote degradation of this soil-
associated contaminant. After 111 days incubation
time, 42.7% loss of soil-associated phenanthrene was
observed in the induced-compost soil mixture, while
36.7% loss of soil-associated phenanthrene was
observed in the uninduced-compost soil mixture.
Antizar-Ladislao et al. (2005) investigated the bio-
degradation of 16 United States Environmental Pro-
tection Agency (USEPA)—listed PAHs (Fu et al.
2003) present in contaminated soil from a manufac-
tured gas plant site using laboratory-scale in—vessel
composting—bioremediation reactors over 8 weeks.
Antizar-Ladislao et al. (2005) found that temperature
and amendment ratio were important operating param-
eters for PAH removal for in—vessel composting—
bioremediation of aged coal tar-contaminated soil and
thereafter recommended that when conventional com-
posting processes using temperature profiles to meet
regulatory requirements for pathogen control need to
be used, these should be preferably started with a
prolonged mesophilic stage followed by thermophilic,
cooling, and maturation stages. More recent studies on
the application of composting to degrade PAHs have
been conclusive and in concert with the findings of
earlier studies. Plaza et al. (2009) have investigated the
binding of phenanthrene and pyrene, by humic acids
(HAs) isolated from an organic substrate at different
stages of composting and a soil using a batch
fluorescence quenching method and the modified
Freundlich model. With respect to soil HA, the organic
substrate HA fractions were characterized by larger
binding affinities for both phenanthrene and pyrene.
Further, Plaza et al. (2009) found that the isotherm
deviation from linearity was larger for soil HA than for
organic substrate HAs, indicating a larger heteroge-
neity of binding sites in the former. The composting
process decreased the binding affinity and increased
the heterogeneity of binding sites of HAs and hence
Plaza et al. (2009) inferred that the changes undergone
by the HA fraction during composting may be
expected to contribute to facilitate microbial accessi-
bility to PAHs. The results obtained also suggested
that bioremediation of PAH-contaminated soils with
matured compost, rather than with fresh organic
amendments, may result in faster and more effective
clean-up. The beneficial use of compost to bioreme-
diate PAHs was further evidenced from the findings of
Yuan et al. (2009). Yuan et al. (2009) have studied the
biodegradation of phenanthrene and pyrene in com-
post and compost-amended soil. The degradation rate
of phenanthrene was found to be more than that of
pyrene. The degradation of the PAHs was enhanced
246 Rev Environ Sci Biotechnol (2010) 9:215–288
123
when the two species were present simultaneously in
the soil, thereby suggesting some kind of mutually
supported synergistic effect which favored their indi-
vidual degradation rate since the addition of either of
the two types of compost (straw and animal manure)
individually enhanced PAH degradation. Further to
analyze the effect of compost size, compost samples
were separated into fractions with various particle size
ranges, which spanned 2–50, 50–105, 105–500 and
500–2,000 lm. Yuan et al. (2009) observed that the
compost fractions with smaller particle sizes demon-
strated higher PAH degradation rates but the when the
different compost fractions were added to soil, com-
post particle size had no significant effect on the rate of
PAH degradation. Of the microorganisms isolated
from the soil-compost mixtures, Arthrobacter nicoti-
anae, Pseudomonas fluorescens, and Bordetella Petrii,
respectively, demonstrated the best degradation ability
for the PAHs studied.
7.1.3 Petroleum-based hydrocarbons remediation
by composting
‘Total Petroleum Hydrocarbons’ is a term used to
describe a broad family of several hundred chemical
compounds that originally come from crude oil. Crude
oils can vary in how much of each chemical they
contain, and so can the petroleum products that are
made from crude oils. Some are clear or light-colored
liquids that evaporate easily, and others are thick, dark
liquids or semi-solids that do not evaporate. Many of
these products have characteristic gasoline, kerosene,
or oily odors. Because modern society uses so many
petroleum-based products (gasoline, kerosene, fuel
oil, mineral oil, and asphalt), contamination of the
environment by them is potentially widespread. Con-
tamination caused by petroleum products contains a
variety of these hydrocarbons. Because they are found
in a complex mixture, it is not usually practical to
measure each one individually and treat them sepa-
rately with complete remediation. The amount of TPH
found in a sample is useful as a general indicator of
petroleum contamination at that site.
Composting of contaminated soil in biopiles is an
ex situ technology, where OM such as bark chips are
added to contaminated soil as a bulking agent.
Composting of lubricating oil-contaminated soil was
performed in field scale (5 9 40 m3) using bark chips
as the bulking agent, and two commercially available
mixed microbial inocula as well as the effect of the
level of added nutrients (nitrogen, potassium and
phosphorus) were tested by (Jørgensen et al. 2000).
Jørgensen et al. (2000) also performed the composting
of diesel oil-contaminated soil at one level of nutrient
addition and with no inoculum. Jørgensen et al. (2000)
noted that the mineral oil degradation rate was most
rapid during the first months of the composting
process, and it followed a typical first order degrada-
tion curve. During these 5 months, composting of the
mineral oil had decreased in all piles with lubrication
oil from approximately 2,400–700 mg/kg dry weight,
which was about 70% of the mineral oil content.
Correspondingly, the mineral oil content in the pile
with diesel oil-contaminated soil decreased with 71%
from 700 to 200 mg/kg dry weight. In this type of
treatment with addition of a large amount of OM, the
general microbial activity as measured by soil respi-
ration was enhanced and no particular effect of added
inocula was observed, thereby advocating the suit-
ability of composting to bioremediate diesel oil-
contaminated soil. Namkoong et al. (2002) conducted
research to find the appropriate mix ratio of organic
amendments for enhancing diesel oil degradation
during contaminated soil composting by adding sew-
age sludge or compost as an amendment for supple-
menting OM for composting of the contaminated soil.
Namkoong et al. (2002) thereafter found that the
degradation of diesel oil was significantly enhanced by
the addition of these organic amendments relative to
straight soil. The degradation rates of TPH and n-
alkanes were found to be greatest at the ratio of 2:1 of
contaminated soil to organic amendments on wet
weight basis. The work of Marın et al. (2006)
ascertains the efficacy of composting as a low cost
technology bioremediation technique for reducing the
hydrocarbon content of oil refinery sludge with a large
total hydrocarbon content of 250–300 g/kg, in semi-
arid conditions. The composting system designed by
Marın et al. (2006), which involved open air piles
turned periodically over a period of 3 months, proved
to be inexpensive and reliable. The influence on
hydrocarbon biodegradation of adding wood shavings
as bulking agent and inoculation of the composting
piles with pig slurry was also studied. Marın et al.
(2006) determined that the most effective composting
treatment was the one in which the bulking agent was
added, where the initial hydrocarbon content was
Rev Environ Sci Biotechnol (2010) 9:215–288 247
123
reduced by 60% in 3 months as compared with the
32% reduction achieved without bulking agent.
Although, spiking the piles with an organic fertilizer
did not significantly improve the degree of hydrocar-
bon degradation, Marın et al. (2006) concluded that the
composting process without doubt led to the biodeg-
radation of toxic compounds.
Mihial et al. (2006) determined that bioremediation
by composting was a suitable alternative for the
remediation of soil in and around a pit contaminated
with petroleum waste comprising used oil, gasoline,
diesel fuel and paint thinners. Mihial et al. (2006)
conducted a bench scale treatability study to assess the
potential for successful bioremediation of the site
using composting. They set up two reactors each with
ammonium phosphate fertilizer as the nutrient amend-
ment using a mixture of grass clippings and sheep
manure in one reactor to determine if the composting
process could be accelerated by the addition of these
abundantly available waste materials. Based on the
results of the treatability study, the half-life of the
petroleum hydrocarbons at the subject site was esti-
mated to be 36.3 and 121.6 days with and without the
addition of grass clippings and sheep manure, respec-
tively. It was estimated that it would take approxi-
mately 192 and 643 days to remediate the soil and
lower reduce the TPH to 1,000 mg/l using the
amendments of the reactors, respectively. Atagana
(2008) inoculated the contaminated soil with sewage
sludge and incubated for the mix for 19 months.
Compost heaps were set up in triplicates on wood
pallets covered with double layers of nylon straw
sheets and control experiments which contained
contaminated soil and wood chips but without sewage
sludge were set up in triplicate, and the concentrations
of selected hydrocarbons in the contaminated soil were
measured monthly during the incubation period.
Atagana (2008) noted a typical composting perfor-
mance through the temperature rise up to about 58�C
in the sewage sludge compost within 60 days of
incubation, while temperature in the control fluctuated
between 15 and 35�C throughout the incubation
period. All the more, TPH was reduced by 17% in
the control experiments and up to 99% in the sewage
sludge compost at the end of the incubation period.
Much promisingly as being a reliable bioremediation
technique, the composting process reduced the con-
centrations of the TPH by up to 100% within the same
period.
7.1.4 Phenol derivatives
Phenol is both a synthetically and naturally produced
aromatic compound. Microorganisms capable of
degrading phenol are common and include both
aerobes and anaerobes (van Schie and Young 2000).
Many aerobic phenol-degrading microorganisms have
been isolated and the pathways for the aerobic degra-
dation of phenol are now established (van Schie and
Young 2000). The first steps include oxygenation of
phenol by phenol hydroxylase enzymes to form
catechol, followed by ring cleavage adjacent to or in
between the two hydroxyl groups of catechol. Phenol
can also be degraded under anaerobic conditions, but
this process is less well understood and documented,
and only a few anaerobic phenol-degrading bacteria
have been isolated to date (van Schie and Young 2000).
A number of practical applications exist for microbial
phenol degradation and these comprise the exploitation
of anaerobic phenol-degrading bacteria in the in situ
bioremediation of creosote-contaminated subsurface
environments, and the use of phenol as a co-substrate
for aerobic phenol-degrading bacteria to enhance in situ
biodegradation of chlorinated solvents. Chlorophenols
have been introduced into the environment through
their use as biocides and as by-products of chlorine
bleaching in the pulp and paper industry (Field and
Sierra-Alvarez 2008). Chlorophenols are subject to
both anaerobic and aerobic metabolism (Antizar-
Ladislao and Galil 2003). Under anaerobic conditions,
chlorinated phenols can undergo reductive dechlorina-
tion when suitable electron-donating substrates are
available (Field and Sierra-Alvarez 2008). Under
aerobic conditions, both lower and higher chlorinated
phenols can serve as sole electron and carbon sources
supporting growth. Two main strategies are used by
aerobic bacteria for the degradation of chlorophenols.
Lower chlorinated phenols are initially attacked by
monooxygenases yielding chlorocatechols as the first
intermediates whilst polychlorinated phenols are
converted to chlorohydroquinones as the initial inter-
mediates. Fungi and some bacteria are additionally
known that co-metabolize chlorinated phenols (Field
and Sierra-Alvarez 2008). These microbial degrada-
tion mechanisms have gradually been put to use for
the remediation of phenol- and phenol derivatives-
contaminated soil and other strata through composting.
Some of the most conclusive studies where
composting has been applied to bioremediate phenol
248 Rev Environ Sci Biotechnol (2010) 9:215–288
123
contaminated strata are now discussed. Laine et al.
(1997) have studied the fate of chlorophenols during
the composting of sawmill soil and impregnated wood
to see whether chlorophenols, in addition to mineral-
ization, would form any harmful metabolites. The
toxicity assessed by luminescent bacteria tests
decreased during the composting, and it followed the
chlorophenol concentrations in the compost piles. The
threshold value for chlorophenol toxicity appeared to
be 200 mg of total chlorophenols/kg dry weight.
Based on the results obtained, Laine et al. (1997)
deduced that the toxicity tests were a quick and
promising tool for assessing the toxicity changes in
chlorophenol-contaminated soil but were not sensitive
enough to detect the concentrations that would meet
the remediation criteria for clean-up of chlorophenol-
contaminated soil, which in this case was 10 mg/
kg dry weight total chlorophenols. In conclusion,
Laine et al. (1997) found that no harmful metabolites
were formed during composting of chlorophenol-
contaminated soil, but the existing ones such as
polychlorinated dibenzo-p-dioxins and dibenzofurans
(PCDD/Fs) compounds were not removed during the
biological treatment. The results of biotransformation
studies suggested that the 30–40% of the carbon in
chlorophenols that disappeared but did not mineralize
during the composting process most likely was built
into the bacterial biomass. Das and Xia (2008)
characterized the transformation kinetics of 4-NP
and its isomers during biosolids composting. Five
distinctive 4-NP isomer groups with structures relative
to a- and b-carbons of the alkyl chain were identified in
biosolids. Composting biosolids mixed with wood
shaving at a dry weight percentage ratio of 43:57 (C:N
ratio of 65:1) removed 80% of the total 4-NP within
2 weeks of the composting experiments. Das and Xia
(2008) have also found that isomers with a-methyl-
a-propyl structure transformed significantly slower
than those with less branched tertiary a-carbon and
those with secondary a-carbon, suggesting isomer-
specific degradation of 4-NP during biosolids
composting.
7.1.5 PCBs
Polychlorinated biphenyls (PCBs), that can be mix-
tures of up to 209 congeners, were first manufactured
in 1929 (Bhandari and Xia 2005) and these are
among the most widely detected chemicals in waste-
water residual biosolids. Although PCBs are no
longer produced in the United States because they
build up in the environment and can cause harmful
health effects, they are still in use in many other
countries. Polychlorinated dibenzodioxins and poly-
chlorinated dibenzofurans (dioxins) (Fu et al. 2003)
consist of 210 different compounds which have
similar chemical properties (Bhandari and Xia
2005). This class of compounds is persistent, toxic,
and bioaccumulative. They are generated as byprod-
ucts during incomplete combustion of chlorine con-
taining wastes like municipal solid waste, sewage
sludge, and hospital and hazardous wastes (Bhandari
and Xia 2005). PCBs were widely used in the past
and now contaminate many industrial and natural
areas.
PCBs can be degraded by microorganisms via a
metacleavage pathway to yield tricarboxylic acid
cycle intermediate and (chloro)benzoate (CBA). The
initial step in the aerobic biodegradation of PCBs is
the dioxygenation of PCB congeners by the biphenyl
dioxygenase enzyme (Ang et al. 2005). In this step,
the enzyme catalyzes the incorporation of two
hydroxyl groups into the aromatic ring of a PCB
congener, which increases the reactivity of the PCBs,
rendering them more susceptible to enzymatic ring
fission reactions (Bruhlmann and Chen 1999).
Only one research has been reported in the literature
where composting has been applied for bioremediating
PCBs. Michel et al. (2001) determined the effects of
soil to amendment ratio on PCB degradation when a
PCB-contaminated soil from a former paper mill was
mixed with a yard trimmings amendment and com-
posted in field scale piles. Temperature, oxygen
concentrations, and a number of other environmental
parameters that usually influence microbial activity
during composting were monitored. The PCBs in the
contaminated soil had a concentration of 16 mg/kg dry
weight and an average of 4 chlorines per biphenyl. The
soil was composted with five levels of yard trimmings
amendment (14–82% by weight) in pilot scale com-
post piles of volume 25 m3 and turned once every
month. Michel et al. (2001) observed that up to a 40%
loss of PCBs with amendment levels of 60 and 82%.
Also, congener specific PCB analysis indicated that
less chlorinated PCB congeners (1–3 chlorines per
biphenyl) were preferentially degraded during the
composting process. On the other hand, bench-scale
Rev Environ Sci Biotechnol (2010) 9:215–288 249
123
studies indicated that less than 1% of the PCBs in the
contaminated soil were volatilized from composts
during incubation with forced aeration at 55�C. In
conclusion, Michel et al. (2001) observed PCB loss
during the composting of the PCB-contaminated soil
and this appeared to be for the most part due to
biodegradation, rather than volatilization.
7.1.6 Phthalic acid esters
Phthalic acid esters (PAEs) or phthalate esters are
manufactured in large quantities and have been used in
the production of plastics. Di-(2-ethylhexyl) phthalate
(DEHP), the most widely used phthalate ester, is
persistent during sewage treatment and readily accu-
mulates in sediments and lipid tissues in aquatic
organisms (Bhandari and Xia 2005). DEHP, a sus-
pected endocrine disruptor (Hoyer 2001) has been
reported in a variety of media including water,
atmospheric deposition, sediments, soil, biosolids,
biota, and food products (Fu et al. 2003; Bluthgen
2000). Among the PAEs targeted by the USEPA as
priority pollutants, DEHP is the major pollutant
identified at high concentrations level in lagooning
sludge at about 28.67 mg/kg and in activated sludge at
about 6.26 mg/kg. Other PAEs, such as di-butyl
phthalate (DBP) and di-methyl phthalate (DMP) show
very low concentrations (Amir et al. 2005).
Several studies have been carried out to assess the
biodegradability and bioremediation of phthalate
esters by composting and results have so far been
promising. Marttinen et al. (2004) studied the potential
of composting and aeration to remove DEHP from
municipal sewage sludge with raw sludge and anaer-
obically digested sludge. They found that composting
removed 58% of the DEHP of the raw sludge and 34%
of that of the anaerobically digested sludge during
85 days stabilization in compost bins, while a compa-
rable removal for the anaerobically digested sludge
was achieved in a rotary drum composter in 28 days.
Although DEHP removal was greater from raw sludge
compost than anaerobically digested sludge compost,
the total and volatile solids removals were similar in
the two composts. Moreover, Marttinen et al. (2004)
determined that in the aeration process mode of raw
sludge at 20�C, the DEHP removals were 33–41 and
50–62% in 7 and 28 days, respectively. The pool of
results hence collected by Marttinen et al. (2004)
indicated that both composting and aeration have the
potential to reduce the DEHP contents typically found
in sewage sludges to levels acceptable for agricultural
use. On a similar note, in assessing sludge composting
as a bioremediation approach for DEHP, Gibson et al.
(2007) investigated the impact of pilot-scale compost-
ing and drying of sludge on the physicochemical
characteristics and on the concentrations of some
organic contaminants. During the 143-day composting
experiments, OM content fell by 22% and moisture by
50%. Concentrations of 4-nonylphenols fell by 88%
and DEHP by 60%, and these losses continued
throughout the procedure. The drying process was
much shorter and lasted only 40 days, yet OM content
decreased by 27% and moisture by 85%. Losses of
4-NPs (39%) and DEHP (22%) were less than in
composting and stopped when the moisture content
quasi stabilized. Gibson et al. (2007) concluded that
composting would be the method of choice for
reducing organic contaminants but this bioremediation
technique requires much longer times than drying.
Cheng et al. (2008) also came to similar inferences as
Amir et al. (2005) when investigating the potential
degradation of DEHP and OM of sewage sludge by
composting using laboratory reactors at different
operating conditions. At the end of composting, Cheng
et al. (2008) observed that the total DEHP degradation
was more than 85% in all conditions and the total
carbon reduction varied from 7.6 to 11.8%. Cheng et al.
(2008) deduced that the degradation kinetics of DEHP
in thermophilic phase and the phase thereafter were
modeled by first order and fractional power kinetics,
respectively.
7.1.7 Bioremediation of pesticides
Chemical pesticides1 have consistently demonstrated
their merit by increasing the global agricultural
productivity (Ecobichon 2001), reducing insect-
borne, endemic diseases and protecting plantations,
forests and harvested wood (Ecobichon 2000). As of
date, pesticides are more valued in developing
1 According to the United States Environmental Protection
Agency (US EPA) (1999), the term pesticide is a broad
nonspecific term covering a large number of substances
including, insecticides, herbicides and fungicides, ‘though
often misunderstood to refer only to insecticides’.
250 Rev Environ Sci Biotechnol (2010) 9:215–288
123
countries, particularly those in tropical regions seek-
ing to enter the global economy by providing off-
season fresh fruits and vegetables to countries in
more temperate climates (Ecobichon 2001). How-
ever, the continuous use of pesticides has caused
severe irreversible damage to the environment,
caused human ill-health, negatively impacted on
agricultural production and reduced agricultural sus-
tainability (Wilson and Tisdell 2001).
Traditional methods of pesticide remediation
which are however relatively costly include excava-
tion and/or chemical oxidation processes (for exam-
ple, photocatalysis, ozonation and iron-catalyzed
Fenton’s reaction) or thermal processes (for example
low temperature themal desorption, incineration). On
the other hand, bioremediation and phytoremediation
are the biotic processes that are sometimes employed
for the remediation of pesticides contaminated sites
(Lynch and Moffat 2005). The use of phytotechnol-
ogies to remediate these more persistent pesticides is
only emerging (Chaudhry et al. 2002; Zhuang et al.
2007). Still, difficulties persist, including the poten-
tial phytotoxicity of some herbicides (Eullaffroy and
Vernet 2003; Van Eerd et al. 2003) that were
originally developed but destroyed plant material.
Typically the mechanisms involved in pesticide
phytoremediation are phytodegradation, rhizodegra-
dation, and phytovolatilization. As a form of low cost
clean-up bioremediation option, composting and
biobeds2 are increasingly being assessed as an
approach to remediate pesticides. Some studies have
been carried out to this end and they unanimously are
in favor of composting. The fate of the widely used
lawn care herbicide 2,4-dichlorophenoxyacetic acid
(2, 4-D) during the composting of yard trimmings
consisting of primarily leaves and grass is an
important unexplored question. In their study, Michel
et al. (1995) determined the extent of 2, 4-D
mineralization, incorporation into humic matter,
volatilization, and sorption during the composting
of yard trimmings. Yard trimmings (2:1 [wt/wt]
leaves–grass) were amended with 14C-ring-labeled 2,
4-D (17 mg/kg dry weight) and composted in a
temperature-controlled laboratory scale compost sys-
tem. During composting, thermophilic microbes were
numerically dominant, reaching a maximum of
2 9 1011/g. At the end of composting, 46% of the
OM present in the yard trimmings was lost and the
compost was stable, with an oxygen uptake rate of
0.09 mg O2/g OM/h, and was well humified. Michel
et al. (1995) also observed that the mineralization of
the OM temporally paralleled the mineralization of
2,4-D. In the final compost, 47% of the added 2,4-D
carbon was mineralized, about 23% was complexed
with high-molecular-weight humic acids while about
20% remained bound. With very little volatilization of
2,4-D occurred during the composting process, Michel
et al. (1995) noted with interest that their results
indicated an active mineralization of 2,4-D at com-
posting temperatures of 60�C. To elucidate the hazard
potential of compost application, Hartlieb et al. (2003)
amended municipal biowaste with 14C labelled pyrene
and simazine, which they incubated in a pilot-scale
composting simulation system. A mass balance incor-
porating the mineralization, metabolism and sorption
of the two model substances was then established over
a period of 370 days. Hartlieb et al. (2003) found that
the results wee quite different for the two chemicals
thereby reflecting their intrinsic properties during their
degradation in the composting environment. Ghaly
et al. (2007) have evaluated the effectiveness of in-
vessel thermophilic composting on the destruction of
pirimiphos-methyl (O-(2-diethylamine-6-methylpyri-
midin-4-yl) O,O-dimethyl phosphorothioate). Pirimi-
phos-methyl is an insecticide with both contact and
fumigant action and shows activity against a wide
variety of insects including ants, beetles, caterpillars,
cockroaches, fleas, flies, mites, mosquitoes and moths.
With a half-life of 117 days in water, 180–270 days on
greens and seeds, pirimiphos-methyl has been reported
to cause cholinesterase inhibition in humans which at
high dose rates results in nausea, dizziness, and
confusion and at high exposure due to accidents and
major spills results in respiratory paralysis and death.
The bioreactor for the composting process studied by
Ghaly et al. (2007) was operated on a mixture of
tomato plant residues, wood shavings and municipal
solid compost. Ghaly et al. (2007) found that the
composting process successfully destroyed 81–89% of
pirimiphos-methyl within the first 54 h of the com-
posting process, while the complete destruction of the
2 A biobed in its simplest form is a rectangular lined pit,
1–1.3 m deep, filled with a mixture of topsoil, peat-free
compost and straw in a ratio of 1:1:2, respectively and turfed
over. Biobeds filter out pesticides and use enhanced microbial
activity to break them down.
Rev Environ Sci Biotechnol (2010) 9:215–288 251
123
pesticide required approximately 440 h. Ghaly et al.
(2007) also inferred that a number of physical,
chemical and biological mechanisms contribute to
the degradation of pirimiphos-methyl in the environ-
ment and these consist of mineralization, abiotic
transformations, adsorption, leaching, humification,
and volatization. During composting of greenhouse
wastes, in particular, the degradation of pirimiphos-
methyl is accelerated by high temperatures developed
during the thermophilic stage of the process, OM
content, moisture of the compost matrix and level of
biological activity. Delgado-Moreno and Pena (2009)
amended a typical calcareous agricultural soil of the
Mediterranean area contaminated with four triazine
herbicides with olive cake, compost and vermicom-
post of olive cake at rates four times higher than the
agronomic dose in order to stimulate the biodegrada-
tion of simazine, terbuthylazine, cyanazine and prom-
etryn, and thereafter observed that the residual
herbicide concentrations at the end of the degradation
assay showed no significant differences between non
amended and amended soil. However, interestingly,
Delgado-Moreno and Pena (2009) found that the
addition of compost and vermicompost had enhanced
the biological degradation rate of triazines during the
first week of incubation, with half-lives ranging from 5
to 18 days for the amended soils.
7.2 Controlled solid phase biotreatment
These processes include prepared treatment beds,
biotreatment cells, and soil piles, biopiles or com-
posting matrices. Moisture, heat, nutrients, oxygen,
and pH can be controlled to enhance biodegradation.
These technologies differ from landfarming in that
treatment processes are often enclosed to control off-
gases. Typically, excavated material is mixed with
soil amendments and placed on a treatment area that
includes leachate collection systems and some of
aeration. The costs of these techniques vary widely,
but are among the expensive ones when applicable.
Some prepared bed bioremediation techniques
involved the continuous spray application of a
nutrient solution into the soil and collection and
recycle of the drainage from the soil pile. The
drainage itself may be treated in a slurry-phase
bioreactor before recycling. Vendors have developed
proprietary nutrient and additive formulations and
methods for incorporating the formulation into the
soil to stimulate biodegradation. Target contaminants
include non-halogenated VOCs and SVOCs. Pesti-
cides also can be treated, but the process may be less
effective and may be applicable only to some
compounds within these contaminant groups. Like
landfarming, these technologies require a lot of space,
and excavation of contaminated material is required.
One advantage, however, of contained ex situ methods
is that toxic byproducts or metabolites formed during
the biodegradation process (e.g., vinyl chloride from
TCE) are contained.
7.3 Slurry phase bioremediation
These technologies involve the treatment of excavated
contaminated soils in the controlled environment of a
bioreactor. Excavated soil is processed to separate
stones and rubble, then mixed with water to a
predetermined concentration dependent upon the
concentration of the contaminants, the rate of biodeg-
radation, and the physical nature of the soils. Usually
slurries contain from 10 to 40% solids. Electron
acceptors and nutrients are added to the reactor, and
parameters such as pH and temperature are controlled
to optimize biological processes. Also, the reactor may
be inoculated with specialized organisms if a suitable
population is not present. Both aerobic and anaerobic
reaction environments may be used. Target contam-
inants include petrochemicals, solvents, pesticides,
wood preservatives, explosives, petroleum hydrocar-
bons and other organic chemicals. Bioreactors are
favored over in situ biological techniques for hetero-
geneous soils, low permeability soils, areas where
underlying groundwater would be difficult to capture,
or when faster treatment times are required. Like solid
phase ex situ treatments, they have the advantage of
containing toxic metabolites such as vinyl chloride.
Slurry phase treatment tends to be faster, but more
expensive, than controlled solid phase treatment.
Table 14 highlights some findings of recent studies
which have demonstrated the promise of slurry phase
bioremediation of organic contaminants and soils.
8 Anaerobic digestion biotechnology
Anaerobic processes are defined as biological pro-
cesses in which organic matter is metabolized in an
environment free of dissolved oxygen or its precursors
252 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Ta
ble
14
Ex
amp
leo
fsl
urr
yp
has
eb
iore
med
iati
on
:P
roce
ssfe
atu
res
and
rem
edia
tio
np
ote
nti
al
Co
nta
min
ated
med
ia/
con
tam
inan
ts
Bio
pro
cess
feat
ure
(s)
Bio
rem
edia
tio
np
erfo
rman
ceR
efer
ence
Sim
ula
ted
py
ren
e-
con
tam
inat
edso
il
Bio
-slu
rry
ph
ase
reac
tors
op
erat
edin
per
iod
ic
dis
con
tin
uo
us
bat
chm
od
eu
nd
eran
ox
ic–
aero
bic
–an
ox
ic–
ano
xic
mic
roen
vir
on
men
t
Co
ntr
ol
reac
tor
(kil
led
con
tro
l)sh
ow
edo
nly
6%
of
py
ren
ed
egra
dat
ion
wh
ile
the
no
n-a
ug
men
ted
reac
tor
sho
wed
anef
fici
ency
of
34
%
(su
bst
rate
deg
rad
atio
nra
te(S
DR
)—0
.01
65
gp
yre
ne/
kg
soil
/day
).
Inth
eca
seo
fau
gm
ente
dre
acto
rs,
the
syst
emo
per
ated
wit
hlo
w
sub
stra
telo
adin
gra
te(S
LR
)sh
ow
eda
py
ren
ed
egra
dat
ion
effi
cien
cy
of
alm
ost
90
%(S
DR
—0
.04
gp
yre
ne
/kg
soil
/day
)an
dth
ere
acto
r
wit
hh
igh
SL
Rsh
ow
ed5
0%
(SD
R—
0.0
25
gp
yre
ne
/kg
soil
/day
)
of
py
ren
ed
egra
dat
ion
Ven
kat
a
Mo
han
etal
.
(20
08
)
Pen
dim
eth
alin
con
tam
inat
edso
il
Bio
-slu
rry
ph
ase
reac
tor
op
erat
edin
the
seq
uen
cin
gb
atch
mo
de
un
der
ano
xic
–ae
rob
ic–
ano
xic
mic
ro-e
nv
iro
nm
ent
Eff
ect
of
aug
men
tati
on
wit
hef
flu
ent
trea
tmen
tp
lan
t(E
TP
)
mic
ro-fl
ora
on
the
pro
cess
per
form
ance
of
the
reac
tor
was
also
eval
uat
ed
Th
ere
acto
rw
aso
per
ated
wit
ha
tota
lcy
cle
per
iod
of
12
0h
com
pri
sin
g3
ho
ffi
ll,
11
4h
of
reac
tan
d3
ho
fd
ecan
t
ph
ases
Co
ntr
ol
reac
tor
(wit
ho
ut
ET
Pm
icro
-flo
ra)
sho
wed
23
%o
fsu
bst
rate
rem
ov
alef
fici
ency
.T
he
reac
tor
aug
men
ted
wit
hm
ixed
ET
Pm
icro
-
flo
rash
ow
eden
han
ced
per
form
ance
wit
hm
ore
than
90
%o
f
sub
stra
tere
mo
val
effi
cien
cyw
ith
in7
2h
of
the
cycl
ep
erio
d
Ram
aK
rish
na
etal
.(2
00
6)
Ch
lorp
yri
fos
con
tam
inat
ed
soil
usi
ng
nat
ive
mix
ed
mic
rofl
ora
Bio
slu
rry
reac
tor
ado
pti
ng
seq
uen
cin
gb
atch
mo
de
(an
ox
ic–
aero
bic
–an
ox
ic)
op
erat
ion
Rea
cto
ro
per
atio
nw
asm
on
ito
red
for
ato
tal
cycl
ep
erio
do
f
72
hco
nsi
stin
go
f3
ho
fF
ILL
,6
4h
RE
AC
T,
2h
of
SE
TT
LE
,an
d3
ho
fD
EC
AN
Tw
ith
chlo
rpy
rifo
s
con
cen
trat
ion
so
f3
,00
0,
6,0
00
and
12
,00
0lg
/g
At
3,0
00
lg/g
of
chlo
rpy
rifo
sco
nce
ntr
atio
n,
91
%w
asd
egra
ded
afte
r7
2h
of
the
cycl
ep
erio
d,
wh
erea
sin
the
case
of
6,0
00
lg/g
of
chlo
rpy
rifo
s,8
2.5
%w
asd
egra
ded
.H
ow
ever
,fo
r1
2,0
00
lg/g
of
chlo
rpy
rifo
s,o
nly
14
.5%
deg
rad
atio
nw
aso
bse
rved
Th
ed
egra
dat
ion
rate
was
rap
idat
low
ersu
bst
rate
con
cen
trat
ion
and
12
,00
0lg
/go
fsu
bst
rate
con
cen
trat
ion
was
fou
nd
tob
ein
hib
ito
ry
Ven
kat
a
Mo
han
etal
.
(20
04
)
Bio
slu
rry
cult
ure
ssp
iked
wit
hn
apro
xen
e(N
AP
)
and
carb
amaz
epin
e(C
BZ
)
Gro
wth
and
acti
vit
yo
fth
ew
hit
e-ro
tfu
ng
us
Tra
met
esve
rsic
olo
ro
nse
wag
esl
ud
ge
wer
eas
sess
edin
bio
slu
rry
and
soli
d-p
has
esy
stem
s
Dep
leti
on
of
aro
un
d4
7an
d5
7%
of
NA
Pan
dC
BZ
wit
hin
24
h,
resp
ecti
vel
y.
Co
mp
lete
dep
leti
on
of
NA
Pan
dar
ou
nd
48
%fo
rC
BZ
wer
eac
hie
ved
wit
hin
72
hin
slu
dg
eso
lid
cult
ure
sw
ith
38
%b
ulk
ing
mat
eria
l
Ro
drı
gu
ez-
Ro
drı
gu
ez
etal
.(2
01
0)
Car
bo
fura
n(2
0m
g/k
gso
il)
con
tam
inat
edso
il
Bio
slu
rry
ph
ase
seq
uen
cin
gb
atch
reac
tor.
A2
-Lla
bo
rato
ry
gla
ssb
ott
lew
asu
sed
asa
bio
reac
tor
wit
ha
wo
rkin
g
vo
lum
eo
f1
.5l
atro
om
tem
per
atu
re(2
7±
2�C
).O
ne
tota
lcy
cle
per
iod
of
the
SB
Rw
asco
mp
rise
do
f1
ho
ffi
ll
ph
ase,
82
ho
fre
act
ph
ase,
and
1h
of
dec
ant
ph
ase
Car
bo
fura
nd
egra
der
iso
late
dfr
om
carb
ofu
ran
ph
yto
rem
edia
ted
soil
,B
urk
ho
lder
iace
pa
cia
PC
L3
(PC
L3
)im
mo
bil
ized
on
corn
cob
use
das
the
ino
culu
m
Bio
aug
men
tati
on
trea
tmen
t(a
dd
itio
no
fP
CL
3)
gav
eth
eh
igh
est
per
cen
tag
eo
fca
rbo
fura
nre
mo
val
(96
.97
%),
foll
ow
edb
y
bio
aug
men
tati
on
tog
eth
erw
ith
bio
stim
ula
tio
n(a
dd
itio
no
fm
ola
sses
)
trea
tmen
t(8
8.2
3%
),su
gg
esti
ng
that
bio
rem
edia
tio
nw
asan
effe
ctiv
e
tech
no
log
yfo
rre
mo
vin
gca
rbo
fura
nin
con
tam
inat
edso
il
Pla
ng
kla
ng
and
Reu
ng
san
g
(20
10
)
TN
T(2
,4,6
-tri
nit
roto
luen
e)-
con
tam
inat
edso
il
Mo
lass
esan
dst
arch
use
dto
esti
mat
eth
eef
fect
of
co-
sub
stra
teo
nan
aero
bic
slu
rry
ph
ase
bio
rem
edia
tio
n
Deg
rad
atio
nef
fici
ency
of
TN
Tin
mo
lass
esan
dst
arch
add
itio
nw
as
app
rox
imat
ely
97
and
87
%,
wh
ich
is5
0–
60
%h
igh
erth
an3
8%
wit
ho
ut
co-s
ub
stra
tead
dit
ion
.M
ola
sses
and
star
chad
dit
ion
enh
ance
dT
NT
deg
rad
atio
n
Inet
al.
(20
08
)
Rev Environ Sci Biotechnol (2010) 9:215–288 253
123
(Khanal 2008). The anaerobic process is classified as
either anaerobic fermentation (Valdez-Vazquez et al.
2005; Ren et al. 2006) or anaerobic respiration
(Rhoads et al. 2005) depending on the type of electron
acceptors (Khanal 2008).
In an anaerobic fermentation, organic matter is
catabolized in the absence of an external electron
acceptor by facultative anaerobes through internally
balanced oxidation–reduction reactions under dark
conditions (Khanal 2008; Vatsala et al. 2008). The
product generated during the process accepts the
electrons released during the breakdown of organic
matter. Thus, organic matter acts as both electron
donor and acceptor. During the fermentation reactions,
the substrate is only partially oxidized, and therefore,
only a small amount of the energy stored in the
substrate is conserved (Khanal 2008). The major
portion of the adenosine triphosphate (ATP) or energy
is generated by substrate-level phosphorylation (Sgar-
bi et al. 2009; Lemire et al. 2009; Atlante et al. 2005).
Anaerobic respiration on the other hand requires
external electron acceptors for the disposal of elec-
trons released during the degradation of organic
matter. The electron acceptors in this case could be
CO2, SO42- or NO3
-. Both substrate-level phosphor-
ylation and oxidative phosphorylation generate energy
(or ATP) (Khanal 2008). The energy released under
such a condition is much greater than anaerobic
fermentation (Skoog et al. 2007). Skoog et al. (2007)
have reported that at in situ geochemical conditions
where large numbers of heterotrophic microorganisms
inhabit hydrothermal systems, for aldose being reacted
upon by these microbial populations, fermentation
yields 220–420 kJ/mol of energy while anaerobic
respiration releases 500–2,400 kJ/mol.
Anaerobic biotechnology is becoming widely
popular due to its potential to produce renewable
biofuels and value-added products from low-value
feedstock such as waste streams (Khanal 2008). In
addition, it provides an opportunity for the removal of
pollutants from liquid and solid wastes more eco-
nomically than the aerobic processes (Marttinen et al.
2003; Khanal 2008). The merits of anaerobic diges-
tion technology are a recovery of bioenergy and
biofuels, recovery of value-added products and waste
treatment. Although the anaerobic process has many
inherent benefits, it is not a panacea for the treatment
of all types of wastewaters and sludges (Khanal
2008). Some of the limitations of anaerobic treatmentTa
ble
14
con
tin
ued
Co
nta
min
ated
med
ia/
con
tam
inan
ts
Bio
pro
cess
feat
ure
(s)
Bio
rem
edia
tio
np
erfo
rman
ceR
efer
ence
So
ilco
nta
min
ated
wit
h
pen
dim
eth
alin
Slu
rry
ph
ase
bio
reac
tor
op
erat
edin
seq
uen
cin
gb
atch
mo
de
(an
ox
ic–
aero
bic
–an
ox
icm
icro
env
iro
nm
ents
)
At
1:2
0so
il–
wat
erra
tio
,th
esl
urr
yp
has
esy
stem
sho
wed
enh
ance
d
deg
rad
atio
no
fsu
bst
rate
(62
9lg
pen
dim
eth
alin
/gso
il).
Th
ere
mo
val
effi
cien
cyo
fp
end
imet
hal
inin
the
reac
tors
was
dep
end
ent
on
the
mas
s-tr
ansf
erra
tes
of
the
sub
stra
tefr
om
the
soil
toth
eaq
ueo
us
ph
ase
Ven
kat
a
Mo
han
etal
.
(20
07
)
So
il-b
ou
nd
anth
race
ne
Ser
ies
of
bio
-slu
rry
ph
ase
reac
tors
op
erat
edin
per
iod
ic
dis
con
tin
uo
us/
seq
uen
cin
gb
atch
mo
de
un
der
ano
xic
–
aero
bic
–an
ox
icm
icro
env
iro
nm
ent
usi
ng
nat
ive
soil
mic
rofl
ora
Fiv
ere
acto
rsw
ere
op
erat
edfo
ra
tota
lcy
cle
per
iod
of
14
4h
(6d
ays)
atso
illo
adin
gra
teo
f1
6.6
6k
gso
il/m
/day
at
30
±2�C
tem
per
atu
re
Th
eco
ntr
ol
reac
tor
(wit
ho
ut
mic
rofl
ora
)sh
ow
edn
egli
gib
le
deg
rad
atio
no
fan
thra
cen
ed
ue
toth
eab
sen
ceo
fb
iolo
gic
alac
tiv
ity
Th
ep
erfo
rman
ceo
fth
eb
io-s
lurr
ysy
stem
wit
hre
spec
tto
anth
race
ne
deg
rad
atio
nw
asfo
un
dto
dep
end
on
bo
thsu
bst
rate
load
ing
rate
and
bio
aug
men
tati
on
Ap
pli
cati
on
of
bio
aug
men
tati
on
sho
wed
po
siti
ve
infl
uen
ceo
nth
era
te
of
deg
rad
atio
no
fan
thra
cen
e
Pra
san
na
etal
.
(20
08
)
254 Rev Environ Sci Biotechnol (2010) 9:215–288
123
system are: long start-up time, long recovery time,
specific nutrients and trace metal requirements, more
susceptible to changes in environmental conditions,
treatment of high-sulphate wastewater and constant
meticulous operational attention.
8.1 Anaerobic digestion chemistry
The anaerobic digestion process is characterized by
a series of biochemical transformations brought on
by different consortia of bacteria (Fantozzi and
Buratti 2009). The anaerobic digestion of organic
matter basically follows the following stages: hydro-
lysis, acidogenesis, acetogenesis and methanogenesis
(Appels et al. 2008; Vavilin et al. 2008; Fountoulakis
et al. 2008; Fantozzi and Buratti 2009). Despite the
successive steps, hydrolysis is generally considered as
rate limiting (Appels et al. 2008) and the rate of
hydrolysis depends on the pH, temperature, composi-
tion and concentration of intermediate compounds
(Fantozzi and Buratti 2009). The hydrolysis step
degrades both insoluble organic material and high
molecular weight compounds such as lipids, polysac-
charides, proteins and nucleic acids, into soluble
organic substances (e.g., amino acids and fatty acids)
(Appels et al. 2008) by extracellular hydrolytic
enzymes produced by hydrolytic bacteria and then
dissolved into solution. The components formed
during hydrolysis are further split during acidogenesis,
the second step. Volatile fatty acids, alcohols (Fantozzi
and Buratti 2009) are produced by acidogenic bacteria
(Bengtsson et al. 2008) along with ammonia, carbon
dioxide, hydrogen sulpide and other by-products
(Goblos et al. 2008). This phase is accompanied by
decrease of pH due to production of acids and protonic
acidification. If the reactor is overloaded, low pH value
may inhibit the process (Chen et al. 2008b). The main
species identified as responsible for the biological
hydrogen production during the acidogenesis of the
carbohydrates are Enterobacter, Bacillus and Clos-
tridium (Davila-Vazquez et al. 2008; Cai et al. 2009).
The third stage in anaerobic digestion is acetogenesis,
where the higher organic acids and alcohols produced
by acidogenesis (Shida et al. 2009) are further digested
by acetogens to produce mainly acetic acid as well as
CO2 and H2. This conversion is controlled to a large
extent by the partial pressure of H2 in the mixture
(Appels et al. 2008). The final stage of methanogenesis
produces methane (Tatsuzawa et al. 2006) by two
groups of methanogenic bacteria (Narihiro and Sekig-
uchi 2007): the first group splits acetate into methane
and carbon dioxide and the second group uses hydro-
gen as electron donor and carbon dioxide as acceptor to
produce methane. The bacteria involved in the meth-
anogenesis stage are sensitive to low as well as to high
pH, which must be kept within a range of 6.5–8.
8.2 Sludge digestion
Sludge treatment has long become the one of the most
challenging problems in wastewater treatment plants
(Zhang et al. 2007; Yu et al. 2008). As a result of the
wide application and utilization of the waste activated
sludge process, excess sludge presents a serious
disposal problem (Neyens and Baeyens 2003; Hao
et al. 2007). The management of excess activated
sludge also imposes great economic costs on the
operation and maintenance of wastewater treatment
plants and hence represents in itself significant
technical challenges (Li et al. 2008) as a results of
environmental, economic, social and legal factors
(Chu et al. 2009). Many efforts have been devoted to
reduce the excess sludge burden (Naddeo et al. 2009)
by treatments such as digestion and dewatering. Some
sludge treatment technologies include pre-treatment
and sludge minimization, anaerobic digestion, aerobic
digestion, alkaline stabilization, composting, dewa-
tering, drying and innovative technologies (Fitzmorris
et al. 2009). Anaerobic digestion has now become a
commonly applied biological process for stabilization
of sewage sludges (Arnaiz et al. 2006; Aitken et al.
2005). The process is more beneficial among several
sludge stabilization methods by reason of it having be
able to produce a net energy gain (Mao et al. 2004; Lu
et al. 2008; Bohn et al. 2007) in the form of methane
gas leading to cost-effectiveness (Mao et al. 2004).
The biodegradability of waste sludge can be
improved by using thermal energy (Bougrier et al.
2008), enzymes and bacteria (Li et al. 2009), ozonation
(Zhang et al. 2009; Dytczak et al. 2007), acidification,
alkaline addition (Lopez Torres and Espinosa Llorens
2008), high pressure homogenization (Kidak et al.
2009), mechanical disintegration and ultrasound (Chu
et al. 2001) pre-treatments. Some investigations have
discussed the combined treatment of alkaline addition
and ultrasound. Among these processes of physical
pre-treatments, ultrasonication is viewed as an envi-
ronmentally and economically sound pretreatment
Rev Environ Sci Biotechnol (2010) 9:215–288 255
123
(Show et al. 2007; Mao and Show 2007) which exhibits
the benefit of not being hazardous to the environment,
and hence being ‘green’ (Chu et al. 2001; Cintas and
Luche 1999; Nikolopoulos et al. 2006).
8.3 Anaerobic biotechnology and pollutant
remediation
During the last recent years, research interest has also
been growing in the study and application of anaer-
obic digestion for the degradation and elimination of
pollutants such as dyes (Senthilkumar et al. 2009),
PAHs (Bernal-Martinez et al. 2007), highly chlori-
nated hydrocarbons and xenobiotics (Zhang and
Bennett 2005), adsorbable organic halides (Savant
et al. 2006) and pesticides (dos Santos et al. 2007).
Very recently, Baczynski and Pleissner (2010) have
used methanogenic granular sludge and wastewater
fermented sludge as inocula for batch tests of anaer-
obic bioremediation of chlorinated pesticide contam-
inated soil, and their results obtained for both types
of biomass were similar wit 80 to over 90% of c-
hexachlorocyclohexane (c-HCH), 1,1,1-trichloro-2,2-
bis-(4-methoxyphenyl)ethane (methoxychlor) and 1,1,
1-trichloro-2,2-bis-(4-chlorophenyl)ethane (DDT)
removed in 4-6 weeks. Bernal-Martinez et al. (2009)
have assessed the removal of PAH naturally present in
sludge by continuous anaerobic digestion with recir-
culation of ozonated sludge. Recirculation of ozonat-
ed digested sludge allowed enhancing PAH removals
with the highest efficiency obtained with the highest
ozone dose (0.11 g O3/g TS). In another study,
Bernal-Martınez et al. (2005) investigated the com-
bined effects of anaerobic digestion and ozonation in
reducing sewage sludge production, and it was
ozonation of anaerobically digested sludge improved
the PAH removal rate up to 61%. In their study,
Fuchedzhieva et al. (2008) have tested Bacillus cereus
isolated from municipal wastewater treatment plant to
assess the efficiency of two anionic surfactants, a
chemical surfactant and a biosurfactant during fluo-
ranthene biodegradation under anaerobic methano-
genic conditions (linear alkyl benzene sulphonates
(LAS) and rhamnolipid-biosurfactant complex from
Pseudomonas sp. PS-17, respectively). Biodegrada-
tion of fluoranthene was monitored by GC/MS for a
period up to 12th day. No change in the fluoranthene
concentration was registered after 7th day. While it
was reported that the rhamnolipid-biosurfactant had
inhibited the cell growth and had no effect on the
biodegradation rate, LAS enhanced the cell growth as
well as the fluoranthene biodegradation, thereby
demonstrating the latter surfactant’s promise as an
agent for facilitating the process of anaerobic PAH
biodegradation under methanogenic conditions.
9 Biosorption of heavy metals
Wastewaters from various industries, such as metal
finishing, electroplating, plastics, pigments and min-
ing, contain several heavy metals of health and
environmental concern, such as cadmium, copper,
chromium, zinc and nickel (Dang et al. 2009).
Industrial wastewater containing heavy metals is a
threat to the public health because of the accumulation
of the heavy metals in the aquatic life which is
transferred to human bodies through the food chain.
All the more, nowadays, an increasing number of
hazardous organic compounds together with variable
levels of heavy metals ions are also being discharged
into the environment (Aksu 2005). Most of the organic
pollutants are degraded or detoxificated by physical,
chemical and biological treatments before released
into the environment. Although the biological treat-
ments are a removal process for some organic com-
pounds, their products of biodegradation may also be
hazardous. Moreover, some non-degradable com-
pounds like the heavy metals ions discharged into the
environment along with the treated compounds can
cause problems due to non-degradability, bioaccumu-
lation, biomagnification and transport to long dis-
tances. As a result, some organic molecules and the
heavy metals ions are not biodegradable and persist in
the environment.
9.1 Heavy metals removal
Conventional methods for the removal of the heavy
metals ions from wastewaters include chemical
precipitation, electroflotation, ion exchange, reverse
osmosis and adsorption onto activated carbon
(Cimino et al. 2005). But due to operational demerits,
high cost of the treatment and the generation of
toxic chemical sludges, some new technologies have
been tried for a long time (Elouear et al. 2008).
256 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Among them less expensive non-conventional adsor-
bents like apple waste (Maranon and Sastre 1991),
peanut hull carbon (Periasamy and Namasivayam
1995), agricultural wastes (Azab and Peterson 1989)
and red mud (Apak et al. 1998) are being investigated
for the removal of ions like the Cd and Ni ions. Sud
et al. (2008) propose the use of agricultural waste
materials as bioadsorbents of heavy metals as a low
cost and highly efficient technology, because the
functional groups present in agricultural waste bio-
mass (acetamido, alcoholic, carbonyl, phenolic,
amido, amino and sulphydryl groups) have affinity
for heavy metals ions to form metal complexes or
chelates that immobilize the contaminants through
reactions of chemisorption, complexation, adsorption
on surface, diffusion through pores and ion exchange.
As a result, researchers and engineers, all alike, have
been oriented toward the practical use of adsorbents
for the treatment of wastewater polluted by heavy
metals (Kocasoy and Guvener 2009).
Many agricultural wastes, including barks,
manures, and composts, contain high levels of ligno-
cellulosic materials. Harman et al. (2007) have
hypothesized that the lignin fraction, which contains
numerous reactive groups, would be highly effective
in binding and removing heavy metals ions from
contaminated water, and, further, that the absorptive
capabilities of the materials would be strongly affected
by the pH of the solution. A series of materials have
been tested by Harman et al. (2007), and, at pH levels
above about 5.5, these materials were highly effective
in removing heavy metals ions, generally as large or
larger than nickel, but ineffective in removal of lighter
ions such as sodium or magnesium. Various barks
were generally observed to be the most effective and
were capable of removing more than 90% of iron,
copper, or lead from solutions in simple shake flask
experiments. Harman et al. (2007) also highlighted
that materials that retain cellular structures and that
have high lignin contents were highly effective with
barks possessing these properties. At alkaline pH
levels, many heavy metals ions precipitate, but three
separate lines of evidence from the extensive study of
Harman et al. (2007) indicate that ions were removed
from aqueous solutions by absorption to barks rather
than by precipitation. At acidic pH levels, they also
were partially effective in removal of the oxyanion
chromate. The study of Harman et al. (2007) hence
underpinned that biosorption is becoming a promising
alternative to replace or supplement the present
removal processes of pollutants from wastewaters
and other contaminated aqueous media.
9.2 Removal of heavy metals by biosorption
Among these pollutants consisting of dyes, phenolics,
herbicides, hormones and pesticides, heavy metals
ions have recently been of great and renewed concern
because of the extreme toxicity and/or persistency in
the environment. Biosorption is the binding and
concentration of adsorbate(s) from aqueous solutions
(even very dilute ones) by certain types of inactive,
dead, microbial biomass. The major advantages of
biosorption over conventional treatment methods
include: low cost, high efficiency, minimization of
chemical or biological sludge, regeneration of biosor-
bents and possibility of metal recovery (Sud et al.
2008). Another powerful technology is adsorption of
heavy metals by activated carbon for treating domestic
and industrial wastewater. However the high cost of
activated carbon and its loss during the regeneration
restricts its application. Since the 1990s the adsorption
of heavy metals ions by low cost renewable organic
materials has gained momentum. Recently attention
has been diverted towards the biomaterials which are
byproducts or the wastes from large scale industrial
operations and agricultural waste materials.
Hence, research on biosorption of heavy metals,
intrinsically guided by Green Chemistry, has led to the
identification of a number of microbial biomass types
that are extremely effective in concentrating metals.
Some types of biomass are waste byproducts of large-
scale industrial fermentations while other metal-bind-
ing biomass types can be readily harvested from the
oceans. These biomass types can accumulate in excess
of 25% of their dry weight in deposited heavy metals:
Pb, Cd, U, Cu, Zn, Cr and others. Some biosorbents
can bind and collect a wide range of heavy metals with
no specific priority, whereas others are specific for
certain types of metals. When choosing the biomass
for metal biosorption experiments, its origin is a major
factor to be considered. In general terms, biomass can
come from industrial wastes which should be obtained
free of charge, organisms that can be obtained easily in
large amounts in nature (e.g., bacteria, yeast, algae) or
fast-growing organisms that are specifically cultivated
or propagated for biosorption purposes (crab shells,
Rev Environ Sci Biotechnol (2010) 9:215–288 257
123
seaweeds). Research on biosorption (examples of
which are given in Table 15) is revealing that it is
sometimes a complex phenomenon where the metallic
species could be deposited in the solid biosorbent
through various sorption processes, such as ion
exchange, complexation, chelation, microprecipita-
tion and oxidation/reduction.
9.3 Scientific basis of biosorption
Several important aspects of biosorption for heavy
metal removal need to be considered when exploring
this emerging bioremediation technique for optimi-
zation purposes. The underlying principles of bio-
sorption for removal of metal ions, the kinetics of
mass transfer during the process of biosorption
of metal ions, the theory and models that can be
used to describe the mass transfer process and the
thermodynamics of biosorption of heavy metals onto
biomass and the models which can be used to
quantify metal-biomass interactions at equilibrium,
all are key knowledge areas in biosorption science
which have been hence so far relatively well
presented, dicussed and reviewed in the literature.
The reader is earnestly directed to more comprehen-
sive and extensive reviews by Davis et al. (2003),
Figueira et al. (2000), Loukidou et al. (2004), Naja
and Volesky (2006), Vijayaraghavan and Yun (2008)
and Volesky (2001) where these scientific aspects of
biosorption have excellently and extensively been
reported.
A biosorption process can be performed via
several modes (Vijayaraghavan and Yun 2008); of
which, batch and continuous modes of operation are
frequently employed to conduct laboratory scale
biosorption processes. Although most industrial
applications prefer a continuous mode of operation,
Table 15 Biosorption studies for heavy metals
Heavy metal Biosorbents References
Cadmium Black gram husk (Cicer arientinum), Rice polish
agricultural waste, Orange wastes from orange juice
production processes, Wheat bran, Pretreated
rice husk (RRH), Red algae (Ceramium virgatum)
Saeed and Iqbal (2003), Singh et al. (2006b),
Perez-Marın et al. (2007), Kumar and Bandyopadhyay
(2006), Sarı and Tuzen (2008)
Chromium Neurospora crassa fungal biomass, Mucilaginous
seeds of Ocimum basilicum, Sargassum sp. algae,
Turbinaria ornata seaweed, Helianthus annuus(sunflower) stem waste
Tunali et al. (2005), Melo and D’Souza (2004),
Vieira et al. (2008), Aravindhan et al. (2004),
Jain et al. (2009)
Copper Lichen biomass of Cladonia rangiformis hoffm.,
Sphaerotilus natans immobilised in polysulfone
matrices, Marıne alga Sargassum sp., Spent-grain,
Grape stalks
Marıne alga Gracilaria Corticata
Ekmekyapar et al. (2006), Beolchini et al. (2003),
Da Silva et al. (2002), Lu and Gibb (2008),
Machado et al. (2003), Esmaeili et al. (2008)
Nickel Loofa sponge-immobilized biomass of Chlorellasorokiniana, Sargassum wightii seaweed, Cone
biomass of Thuja orientalis, Marıne green alga
Ulva reticulata, Biomass and silica-immobilized
biomass of Medicago sativa (alfalfa)
Akhtar et al. (2004), Vijayaraghavan et al. (2005a, b),
Malkoc (2006), Gardea-Torresdey et al. (1996)
Lead Candida albicans, Rhodotorula glutinis yeast,
Powder of mature leaves of the Neem
(Azadirachta indica) tree, Green algae
Cladophora fascicularis, Formaldehyde
polymerized banana stem
Baysal et al. (2009), Cho and Kim (2003),
Bhattacharyya and Sharma (2004),
Deng et al. (2007), Noeline et al. (2005)
Zinc Azadirachta indica bark, Orange peel cellulose with
Phanerochaete chrysosporium immobilized
Ca-alginate beads, Gossypium hirsutum (Cotton)
waste biomass, Mature leaves and stem bark
of the Neem (Azadirachta indica) tree
King et al. (2008), Lai et al. (2008),
Riaz et al. (2009), Arshad et al. (2008)
258 Rev Environ Sci Biotechnol (2010) 9:215–288
123
batch experiments have to be used to evaluate the
required fundamental information, such as biosorbent
efficiency, optimum experimental conditions, bio-
sorption rate and possibility of biomass regeneration.
The factors influencing bacterial batch biosorption
are solution pH, temperature, ionic strength, biosor-
bent dosage, biosorbent size, initial solute concentra-
tion and agitation rate (Vijayaraghavan and Yun
2008).
9.4 Models for biosorption
Within the literature, the Langmuir and Freundlich
models (two-parameter models) have been used to
describe biosorption isotherm. The models are sim-
ple, well-established and have physical meaning and
are easily interpretable, which are some of the
important reasons for their frequent and extensive
use (Vijayaraghavan and Yun 2008). Some other two-
parameter models widely used for describing bio-
sorption isotherms include the Temkin isotherm, the
Dubinin–Radushkevich model, the Redlich–Peterson
model, the Sips model, the Khan model, the Radke–
Prausnitz model and the Toth model. Of these three-
parameter models, the Redlich–Peterson and Sips
models have been used with most success. Preetha
and Viruthagiri (2007) have repoted the biosorption
of chromium using suspended and immobilized cells
of Rhizopus arrhizus by evaluating the physicochem-
ical parameters of the solution such as initial
chromium ion concentration in both batch and packed
bed reactor. Besides the Langmuir, Freundlich and
Redlich–Peterson adsorption isotherm models which
fitted accurately with the experimental data, the
Thomas model, Adams–Bohart and Wolborska mod-
els were also used to represent the dynamic sorption
of chromium using immobilized beads. Preetha and
Viruthagiri (2007) deduced that the Thomas model
represented well the sorption of chromium at differ-
ent residence times whilst the Adams–Bohart model
was fitted better at the initial part of the breakthrough,
with the Wolborska model also representing the
sorption of chromium accurately.
Mechanistic models have been proposed to describe
solute adsorption onto the surfaces of biomass. The
development of a mechanistic model is usually based
on preliminary biomass characterization, with the
formulation of a set of hypothesized reactions between
the sorbent sites and solutes, which also considers the
particular solution chemistry of the solutes. Mecha-
nistic models can often be characterized by the
different degrees of complexity or accuracy in a
system description to account for the surface hetero-
geneity and other factors that contribute to non-ideal
adsorption phenomena. Mechanistic modeling of bio-
sorption has been attempted in several investigations,
with significant success.
Mathematical models that can describe the behav-
iour of a batch biosorption process operated under
different experimental conditions are very useful for
scale up studies or process optimization (Loukidou
et al. 2004). Over 20 models have been reported in
the literature, all of which have attempted to quan-
titatively describe the kinetic behavior during the
adsorption process. Each adsorption kinetic model
has its own limitations, which are derived according
to specific experimental and theoretical assumptions.
Even though they violate the fundamental assump-
tions, many adsorption models have been used to
successfully test experimental biosorption data. Of
these, pseudo-first and pseudo-second order models
(Eqs. 1, 2, respectively) have often been used to
describe biosorption kinetic data.
Qt ¼ Qe 1� e�Kt� �
ð1Þ
Qt ¼ Qe 1� 1
1þ Qe Pt
� �
ð2Þ
where Qe is the amount of solute sorbed at equilib-
rium (mg/g); Qt the amount of solute sorbed at time
t (mg/g); K the first order equilibrium rate constant
(min-1) and P the second order equilibrium rate
constant (g/mg/min). In most published cases involv-
ing biosorption, the pseudo-first order equation was
found to not fit well over the entire contact time
range, but was generally applicable over the initial
periods of the sorption process.
9.5 Mechanisms of biosorption
Different metal-binding mechanisms have been pos-
tulated to be active in biosorption metal uptake such
as chemisorption by ion-exchange, complexation,
coordination, chelation; physical adsorption and
microprecipitation (Volesky 2001). There are also
possible oxidation–reduction reactions taking place
in the biosorbent. Due to the complexity of bioma-
terials and biosorbents, it is also plausible that at
Rev Environ Sci Biotechnol (2010) 9:215–288 259
123
least some of these mechanisms are acting simulta-
neously to varying extents depending on the biosorbent
composition, surface properties and functional chem-
ical groups, and the solution environment (Volesky
2001). Biomass materials offer several molecular
groups that are known to offer ion exchange sites,
carboxyl, sulphate, phosphate, and amine, could be the
main ones (Volesky 2001). Ion-exchange is an impor-
tant concept in biosorption, because it explains many
of the observations made during heavy metal uptake
experiments (Davis et al. 2003). It should be pointed
out that the term ion-exchange does not explicitly
identify the binding mechanism, rather it is used here
as an umbrella term to describe the experimental
observations (Davis et al. 2003). The precise binding
mechanism(s) may range from physical (i.e., electro-
static or London–van der Waals forces) to chemical
binding (i.e., ionic and covalent).
10 Factors influencing bioremediation
The microbial population follows a growth cycle
comprising the three distinct phases namely the lag
phase, exponential phase, stationary phase, and death
phase (Brul et al. 2008; Chong et al. 2008; Akerlund
et al. 1995). In the lag phase there is a delay in the
microbial population growth until the microbes have
become acclimatized to the substrate(s) (Bai et al.
2009a, b; Saravanan et al. 2008), which in many
instances are the contaminants/pollutants under reme-
diation, and surrounding conditions. The microbes
cannot consume the food source until they have
developed the required enzymes and metabolites
necessary to break down the contaminant (Talley and
Sleeper 2006). After the necessary enzymes and
metabolites have been produced, the microbes enter
the exponential phase of growth (Rahman et al. 2006).
The rate of exponential growth is influenced by
environmental conditions as well as by characteristics
of the organism itself. However, exponential growth
cannot occur indefinitely. Generally, either an essen-
tial nutrient for growth is used up or some waste
product of the organism builds up to such a level that
the exponential growth is inhibited and ultimately
ceases (Mulchandani et al. 1989; Okpokwasili and
Nweke 2006). At this point the microbial population
reached the stationary phase, where there is no net
increase or decrease in microbial cell populations.
Dependent on the possible build up of environmental
toxins or depletion of bioavailable substrates (Yates
and Smotzer 2007), the microbial population may
enter the death phase and the viable number of
microbes will decrease. Based on this growth cycle,
almost all organic compounds are degradable given
the proper environmental, physicochemical and time
conditions (Talley and Sleeper 2006). However, a
range of physical, chemical and biochemical condi-
tions or materials can interfere with bioremediation
rates. Some of them can be controlled or modified
while some are difficult to control. The most salient
factors are discussed below.
10.1 pH
pH values\3 and[9 or 10 as well as sudden changes
in the pH of the waste/treatment system matrix can
significantly inhibit microbial growth by interfering
with the microbial metabolism, gas solubility in soil
water, nutrients availability and bioavailability in soil
water, and heavy metal solubilities (Agarry et al.
2008). Most natural environments have values of pH
between 5.0 and 9.0, and as a result this range is
optimal for microbial enhanced biodegradation of
waste contamination. This pH range is maintained by a
natural buffering capacity that exists in most fertile
native soils due to the presence of carbonates and other
minerals (Robinson et al. 2009). However, this
buffering capacity can be depleted over time as a
result of acidic byproducts of degradation (Komnitsas
et al. 2004). The majority of bacteria exhibit growth
optima at or near neutral pH (Andreas and Ekelund
2005) whilst most soils are acidic throughout the
world. Treatment commonly known as liming involves
the addition of finely ground agricultural limestone,
calcium hydroxide, calcium carbonate or magnesium
carbonate during tilling and mixing of the upper layers
to keep the pH in a favorable range for optimal
microbial metabolism. This treatment may affect the
solubility, bioavailability and the chemical form of the
organic pollutants and of soil macro- and micro-
nutrients. Fungi are generally more resistant to acidic
soils than soil and aquifer bacteria.
10.2 Temperature
Temperature affects (a) the bacterial metabolism (b)
microbial growth rates (c) the soil matrix and (d)
260 Rev Environ Sci Biotechnol (2010) 9:215–288
123
physic-chemical state of the contaminants. Generally
in situ bioremediation is carried under mesophilic
condition (20–40�C). Even for laboratory studies,
bacteria with potential remediation value have
focused on mesophilic species because of the species,
of cultivation and relatively short doubling times
(Kuntz et al. 2008; Abid et al. 2007). The rate of
biochemical reactions in cells increases with temper-
ature up to a maximum, above which the rate of
activity declines as enzyme denaturation occurs and
organisms either die or become less active (Trasar-
Cepeda et al. 2007). Low temperatures seldom kill the
microbes and with warming the microbes typically
recover. Temperature also affects gas solubilities and
must be taken into account when designing a reme-
diation system. In compost heaps or biopiles, the
temperature in the center of the soil/sediment may
reach 70�C or higher during the initial active phase
and thermophilic bacteria can be seen performing
under such conditions (Bongochgetsakul and Ishida
2007; Eklind et al. 2007). In situ technology for soils
in tropical countries where the soil temperature may
exceed 50�C merits further investigation. Even at
0.2–8.3�C, purging of contaminated groundwater,
enrichment with nutrients and hydrogen peroxide
have been carried successfully. Even modest increases
in temperature can significantly increase bioremedi-
ation rates. Melin et al. (1998) have mineralized
groundwater contaminants including 2,4,6-trichloro-
phenol (TCP), 2,3,4,6-tetrachlorophenol (TeCP), and
pentachlorophenol (PCP) in three aerobic fluidized-
bed reactors (FBRs) employing sand, volcanite, and
diatomaceous earth as biomass carriers. The effect of
temperature on the chlorophenol degradation kinetics
was studied in FBR batch tests at temperatures
ranging from 4 to 16.5�C. Melin et al. (1998) reported
that the specific maximum degradation rates for TCP
and TeCP varied with temperature from 0.46 9 10-3
to 31 9 10-3 mg/mgVS/h and Ks varied from zero to
7.1 mg/l, while the specific degradation rates for PCP
degradation varied with temperature from 0.24 9
10-3 to 1.7 9 10-3 mg/mgVS/h and were always
lower than for other chlorophenols. Use of the
Arrhenius equation described the temperature effects
on biodegradation of chlorophenols, and in the studied
temperature range, Melin et al. (1998) deduced that a
10�C increase in temperature generally resulted in
over seven times higher degradation rates. Ferguson
et al. (2003) studied the effects of temperature on the
hydrocarbon mineralisation rate in Antarctic terrestrial
sediments. 14C-labelled octadecane was added to
nutrient amended microcosms that were incubated
over a range of temperatures between -2 and 42�C.
Ferguson et al. (2003) found a positive correlation
between temperature and mineralisation rate, with the
fastest rates occurring in samples incubated at the
highest temperatures. The main implications for
bioremediation in Antarctica from the study of Fergu-
son et al. (2003) have been that a high-temperature
treatment would yield the most rapid biodegradation
of the contaminant. Still, Coulon et al. (2005) have
conducted mesocosm studies using sub-Antarctic soil
artificially contaminated with diesel or crude oil in
Kerguelen Archipelago in an attempt to evaluate the
potential of a bioremediation approach in high latitude
environments. All mesocosms were sampled on a
regular basis over 6 months period, and it was found
that soils responded positively to temperature increase
from 4 to 20�C, and to the addition of a commercial
oleophilic fertilizer containing N and P. Both factors
were seen to have increased the hydrocarbon-degrad-
ing microbial abundance and total petroleum hydro-
carbons (TPH) degradation. The major inferences
from the study of Coulon et al. (2005) were that the
bioremediation of hydrocarbon-contaminated sub-
Antarctic soil appeared to be feasible, and various
engineering strategies, such as heating or amending
the soil could accelerate hydrocarbon degradation.
Still, a number of techniques are used to increase in
situ soil remediation applications. These include use of
mulches, plastic covers, vegetation cover to moderate
fluctuations in soil temperature. In cold climates,
steam may be injected to raise the soil temperature or
heating return water. Temperatures of compost heaps
can be increased by irrigation with heated waters.
For very cold environments, above ground liquid
and slurry bioreactors, where the temperature can be
optimized, are the only choices.
10.3 Metals
Metals can inhibit various cellular processes and their
effects are often concentration-dependent (Salanitro
et al. 1997; Sani et al. 2001; Alisi et al. 2009). Metal
toxicity for microbes will usually involve specific
chemical reactivity. Metals such as copper, silver,
and mercury are typically very toxic particularly as
Rev Environ Sci Biotechnol (2010) 9:215–288 261
123
ions, while metals such as lead, barium and iron are
usually benign to the microbes at levels typically
encountered. The nutrient metals are usually found
naturally in the necessary amounts for plants and
microbes in fertile soils (Khan 2005). The principal
inorganic nutrients are nitrogen and phosphorus;
however, trace amounts of potassium, calcium, sul-
phur, magnesium, iron, and manganese are also
required for optimum biological growth (Rajeshwari
et al. 2000). The availability and/or toxicity of these
metals to the microbes is usually dependent on pH,
with the metals becoming more mobile/available at
higher values of pH.
Metals can be actively accumulated by certain
microorganisms and plant species (Kamal et al. 2004).
Living cells can adsorb metals and concentrate inor-
ganics within the cell and although heavy metals may
not be metabolically essential, they are taken up by the
biomass as a side effect of the normal metabolic
activity of the cell (Teng et al. 2008; Azcon et al.
2009). Activated biomass removes metals from solu-
tion by a variety of mechanisms which include ion
exchange at the cell walls, complexation reactions at
the cell walls, and intra- and extra-cellular complex-
ation reactions. Inactivated biomass removes metals
primarily by adsorbing metals to the ionic groups
either on the cell surface (Powell et al. 1999;
Ahluwalia and Goyal 2007) or in the polysaccharide
coating found on most forms of bacteria (Das et al.
2007). The metal ions are bound by exchange of
functional groups or by sorption on polymers. Adsorp-
tion is therefore a procedure of choice for treating
industrial effluents, and a useful tool for protecting the
environment when used to bioremediate contaminated
aqueous media (Crini 2005). Adsorption on natural
polymers and their derivatives are known to remove
pollutants from water. The increasing number of
publications on adsorption of toxic compounds by
modified polysaccharides (over 3,500 since 2000) has
shown the recent increasing interest in the synthesis of
new low-cost adsorbents used in wastewater treat-
ment. In this context, Crini (2005) has performed an
excellent review of the latest developments in the
synthesis of adsorbents containing polysaccharides, in
particular modified biopolymers derived from chitin,
chitosan, starch and cyclodextrin. New polysaccharide
based-materials have described and their advantages
for the removal of pollutants from the wastewater
thoroughly discussed.
10.4 Toxic compounds
Just as contaminant concentrations that are too low can
complicate bioremediation (Sikdar et al. 1998), high
aqueous-phase concentrations of some contaminants
can create problems (Volkering et al. 1997). At high
concentrations, some chemicals are toxic to microbes,
even if the same chemical is readily degraded at lower
concentrations (Ramos et al. 2009). Toxicity prevents
or slows metabolic reactions and often prevents the
growth of new biomass needed to stimulate rapid
contaminant removal (Agarry et al. 2008). The degree
and mechanisms of toxicity vary with specific toxi-
cants, their concentration, and the exposed microor-
ganisms. Microbial cells cease to function when at
least one of the essential steps in their numerous
physiological processes is blocked. The blockage may
result from gross physical disruption of the cell
structure or competitive binding of a single enzyme
essential for metabolizing the toxicant (Agarry et al.
2008; Talley and Sleeper 2006). By design, some
organic compounds are toxic to targeted life forms
such as insects and plants, and may also be toxic to
microbes. These compounds include herbicides, pes-
ticides, rodenticides, fungicides and insecticides.
In addition, some classes of inorganic compounds
such as cyanides and azides are toxic to many
microbes (Talley and Sleeper 2006; Gijzen et al.
2000); however, these compounds may be degraded
following a period of microbial adaption (Marsolek
et al. 2007; Kwon and Yeom 2009). In this respect,
certain studies have indeed shown the fungal biodeg-
radation of cyanide and microbial adaption to such
toxic compounds. Dumestre et al. (1997) identified a
fungus identified as Fusarium solani IHEM 8026 as a
good potential for cyanide biodegradation under
alkaline conditions (pH 9.2–10.7). Results of K14 CN
biodegradation studies had showed that the fungal
metabolism seemed to proceed by a two-step hydro-
lytic mechanism with the first reaction involving the
conversion of cyanide to formamide by a cyanide-
hydrolyzing enzyme, cyanide hydratase (EC 4.2.1.66),
and thereafter a second reaction consisting of the
conversion of formamide to formate, which was
ociated with fungal growth. Earlier, Shah and Aust
(1993) had demonstrated the mineralization of potas-
sium cyanide and various other cyanide salts (Fe, Cu,
Zn, Cd and Cr) by the white rot fungus Phanerochaete
chrysosporium with a 1.5 mmol/L potassium cyanide
262 Rev Environ Sci Biotechnol (2010) 9:215–288
123
solution having a rate of mineralization was about
0.17 mmol/L/day. P. chrysosporium also mineralized
[14C]–cyanide contaminated soil (3,000 mmol/L/day)
using ground corn cobs as nutrient (10 mg/l/day).
Cyanide was oxidized to the cyanyl radical by a lignin
peroxidase from P. chrysosporium. Lately, Gurbuz
et al. (2009) have examined cyanide effluent degrada-
tion by Scenedesmus obliquus. Gold mill effluents
containing cyanide concentration of 77.9 mg/l was fed
to batch unit to examine the ability of S. obliquus for
degrading cyanide. Cyanide was reduced down to
6 mg/l in 77 h. Gurbuz et al. (2009) reported that the
cells had well adapted to high pH and the effluent
contained cyanide and the metals. All the more,
chlorinated aromatic compounds are biorecalcitrant,
and in particular, 2,4,5-trichlorophenol demonstrates
greater resistance to biodegradation than other tri-
chlorophenols and is also a known uncoupler of the
electron transport chain (Marsolek et al. 2007). In this
respect, Marsolek et al. (2007) have investigated the
biorecalcitrance, inhibition, and more specifically the
adaptation to 2,4,5-trichlorophenol by aerobic mixed
microbial communities. Although it was initially
observed that 2,4,5-trichlorophenol was strongly
resistant to biodegradation at concentrations greater
than 40 lmol/L, and inhibited to respiration in direct
proportion to 2,4,5-trichlorophenol concentration, the
microbial communities later showed consistent adap-
tation patterns to 2,4,5-trichlorophenol at concentra-
tions of 10 and 20 lmol/L.
10.5 Water content and geological characters
Water contents, and especially water availability,
influences bioremediation rates (Boopathy 2000). Li
et al. (1997) have reported that a lack in the effect
from bioremediation could be attributed to poor soil
water sorption, which was negatively influenced by
hydrocarbon residuals. This study hence supported
that the soil-water relation is one of the most
important factors in assessing endpoint of bioreme-
diated soils for plant growth. Water in soils or
sediments may not be available to microorganisms
because it is absorbed by solid substances or tied up
as water of hydration to dissolved solutes. This can be
solved by irrigating the contaminated soils, compost
heaps and/or biopiles. In general, in situ degradation
rates are enhanced when the soil is granular or porous
with a relatively high permeability and uniform
mineralogy. Rocky conditions, low permeability,
complex mineralogy and water logged or arid con-
ditions are not favourable to bioremediation (Bali
et al. 2002). Vinas et al. (2005) have examined the
bacterial community dynamics and biodegradation
processes in a highly creosote-contaminated soil
undergoing a range of laboratory-based bioremedia-
tion treatments. The dynamics of the eubacterial
community, the number of heterotrophs and PAH
degraders, and the TPH and PAH concentrations were
monitored during the bioremediation process. While
TPH and PAHs were significantly degraded in all
treatments, Vinas et al. (2005) maintained the mois-
ture content and aeration were the key factors
associated with the PAH bioremediation. Holden
et al. (1997) had earlier quantified the effects of
matric and solute waterpotential on toluene biodeg-
radation by Pseudomonasputida mt-2, a bacterial
strain originally isolated from soil. Across the matric
potential range of 0–1.5 MPa, growth rates were
maximal for P. putida at -0.25 MPa and further
reductions in the matric potential resulted in con-
comitant reductions in growth rates. Growth rates
were constant over the solute potential range
0–1.0 MPa and lower at -1.5 MPa. This specific
study revealed that for P. putida, slightly negative
matric potentials facilitate faster growth rates on
toluene but more negative water potentials result in
slower growth. Also, the toluene utilization rate per
cell mass was observed to be highest without matric
water stress and was unaffected by solute potential.
10.6 Nutrient availability
Nutrients are generally supplemented in both in situ
and ex situ bioremediation of soils, sediments, ground
and surface waters for the promoting the bioremedi-
ation rates (Aspray et al. 2007; Liu et al. 2009).
Nutrient requirement depends on the nature of
contaminants and the extent to which the polluted
site has been subjected to agricultural use. Remedi-
ation of petroleum hydrocarbon contaminated sites
typically requires nitrogen, phosphorus. Chen et al.
(2008a) have reported that addition of excess ferric
iron combined with limited nitrate could promote the
in situ bioremediation of benzene, toluene, ethylben-
zene and xylene isomers and trimethylbenzene iso-
mers in the Borden aquifer and possibly for other sites
contaminated by hydrocarbons. Zhou et al. (2009)
Rev Environ Sci Biotechnol (2010) 9:215–288 263
123
have investigated the effect of phosphorus concentra-
tion on PAH dissipation in the rhizosphere of
mycorrhizal plants in a pot experiment using two
plant species, alfalfa (Medicago sativa) and tall fescue
(Festuca arundinacea), The major finding was the
significant positive impact of mycorrhizal plants on
the dissipation of high molecular weight PAH in high-
water low-phosphorus treatment. Earlier, El-Bestawy
and Albrechtsen (2007) investigated the mineraliza-
tion and/or degradation of the phenoxy herbicide
mecoprop (MCPP) by a group of soil bacteria under
the effects of nutrient amendments. Five different
species of Pseudomonas (P. paucimobilis, P. aeru-
ginosa, P. mallei, P. pseudomallei, and P. pickettii)
were isolated for the MCPP mineralization and/or
removal. Significant variations in the removal per-
centages of MCPP by either mineralization or bio-
degradation were observed. Also, the highest MCPP
mineralization and degradation by the selected Pseu-
domonas spp. were achieved by their inactive (dead)
followed by active-rich cultures with both inoculated
in nutrient-rich medium, confirming the positive
effects of nutrient amendments and sterilization on
MCPP decontamination. Børresen and Rike (2007)
have studied the effects of increased salinity (ionic
strength) and varying concentrations of nutrient and
soil moisture on hexadecane mineralization in a
hydrocarbon contaminated and nutrient deficient high
Arctic soil were assessed. Ammonium chloride
(NH4Cl) was added to give nitrogen concentrations
ranging from 0 to 1,000 mg NH4-N/kg soil, corre-
sponding to molar cation concentrations of NH4?
from 0 to 71 mmol/kg soil. Soil samples with
combinations of NH4? and Na? were also included,
and the soil moisture content varied from 10 to 20%. It
was found that the fertilizing with NH4-N had
increased the total hexadecane mineralization com-
pared to unfertilized soil at all concentrations inves-
tigated, and that the highest mineralization rates were
found in soil samples added 50–200 mg NH4-N/kg at
10% moisture, where 50–58 mg hexadecane/kg/day
had been mineralized.
10.7 External electron availability
Biostimulation through substrate addition is com-
monly practiced to support co-metabolic biodegrada-
tion processes. Addition of stimulatory substrates to
enhance bacterial growth and metabolic activity
through enhanced electron transfer processes of
between electron donor and electron acceptors has
also been used in bio-augmentation experiments
involving both environmental clean-up and agricul-
tural applications (Aboul-Kassim and Simoneit
2001).
Oxygen is used as an electron acceptor to increase
bioremediation activity (Boopathy 2000). A number
of anaerobic bacteria can break down a variety of
aliphatic and aromatic organic compounds both of
natural and anthropogenic origin wholly or partially by
denitrifying bacteria by sulphate, iron, and molybde-
num reducers and by methanogenic consortia. Efforts
are being made to use anaerobic bacteria for breaking
down petroleum contaminated groundwater in oil
refinery sites in the presence of nitrates. Benzene,
toluene, ethyl benzene, xylene (BTEX) and chlori-
nated aliphatic and aromatic compounds have suc-
cessfully removed. Methanogenic bacteria can degrade
chlorinated ethanes such as tetrachloroethane, trichlo-
roethane, dichloroethane, perchloroethylene, carbon
tetrachloride, chloroform, tetrachloromethane, alkyl-
benzenes and a number of chlorinated aromatic
compounds. BTEX bioremediation projects often
focus on overcoming limitations to natural degradative
processes associated with the insufficient supply of
inorganic nutrients and electron acceptors. However,
other limitations associated with the presence and
expression of appropriate microbial catabolic capaci-
ties may also hinder the effectiveness of bioremedia-
tion. Thus, while subsurface addition of oxygen or
nitrate has proven sufficient to remove BTEX below
detection levels it has been only marginally effective at
some sites (Aboul-Kassim and Simoneit 2001). Dou
et al. (2008) have reported an effective anaerobic
BTEX biodegradation under nitrate and sulphate
reducing conditions by the mixed bacterial consortium
that were enriched from gasoline contaminated soil.
Under the conditions of using nitrate or sulphate as
reducing acceptor, the degradation rates of the six
tested substrates decreased with toluene [ ethylben-
zene [ m-xylene [ o-xylene [ benzene [ p-xylene.
Drzyzga et al. (2002) carried out a sediment column
study to demonstrate the bioremediation of chloroeth-
ene- and nickel-contaminated sediment in a single
anaerobic step under sulfate-reducing conditions. By
stimulating the activity of sulphate-reducing bacteria
by the addition of sulphate as supplementary electron
264 Rev Environ Sci Biotechnol (2010) 9:215–288
123
acceptor, complex anaerobic communities were main-
tained with lactate as electron donor (with or without
methanol), which achieved complete dehalogenation
of tetra- and tri-chloroethenes (PCE and TCE) to
ethene and ethane. A few weeks after sulphate
addition, production of sulphide had increased, indi-
cating an increasing activity of sulphate-reducing
bacteria. Hence, it may be deduced that microbial
activity stimulated under sulphate-reducing conditions
can have a beneficial effect on both the precipitation of
heavy metals and the complete dechlorination of
organochlorines as a result of the strongly negative
redox potential created by the activity of sulphate-
reducing bacteria. Regarding nitrate as a stimulant in
bioremediation, Lee et al. (2007) reported that triethyl
phosphate (TEP) treated along with NO3-, was most
effective for the biodegradation of diesel, this being
possible since TEP could be delivered more efficiently
to the target zones and with less phosphorus loss than
KH2PO4.
10.8 Bioavailability of pollutants
Biology in regards to bioremediation refers to the
intrinsic ability of the biota to assimilate and metab-
olize the contaminant (Pignatello 2009), and matrix
effects include the ways in which the biodegradation
is influenced by the interactions of the soil with the
biota and the contaminants. Bacteria in soils are
predominantly attached to soil particles, and so will
be constrained by this attachment and by the physico-
chemical properties of the surface (Pignatello 2009).
Contaminants interact with soils in complex ways
through sorption and mass transfer resistance that
generally impede their availability to organisms. For
example, anthropogenic organic polymers such as
polystyrene and polyvinyl chloride are highly
recalcitrant because of their insolubility and the lack
of extracellular microbial enzymes capable of cata-
lyzing depolymerization. However, non-polymer
degrading bacteria and actinomycetes are able to
degrade oligomeric polystyrene fragments and low
molecular weight fragments of lignin resulting from
fungal attacks on the lignin polymer.
Raj et al. (2007) have isolated eight bacterial
strains on kraft lignin (KL) containing mineral salt
medium (L-MSM) agar with glucose and peptone
from the sludge of pulp and paper mill. Out of these,
ITRC-S8 was selected for KL degradation, because
of its fast growth at the highest tested KL concen-
tration and use of various lignin-related low molec-
ular weight aromatic compounds (LMWACs) as sole
source of carbon and energy. Significant reduction in
colour and KL content in subsequent incubations
have been reported and the degradation of KL by
bacterium was confirmed by gas chromatography–
mass spectrometry (GC–MS) analysis indicating
formation of several LMWACs such as t-cinnamic
acid, 3,4,5-trimethoxy benzaldehyde and ferulic acid
as degradation products, which were not present in
the uninoculated sample, which led to think of the
biochemical modification of the KL polymer to a
single monomer unit. Recently, Trinh Tan et al.
(2008) have investigated the aerobic biological
degradation of the synthetic aliphatic–aromatic
co-polyester EcoflexTM (BASF) by 29 strains of
enzyme-producing soil bacteria, fungi and yeasts at
moderate environmental conditions. It was found
that the aliphatic–aromatic co-polyester could be
degraded by a number of different microorganisms
and the bacteria studied preferentially degraded the
bonds between aliphatic components of the copoly-
mer and the rate of biodegradation of oligomers was
appreciably faster than that for the polymer chains.
Using GC–MS techniques, Trinh Tan et al. (2008)
identified the degradation intermediates as the
monomers of the co-polyester, and gel permeation
chromatography was able to suggest exo-enzyme
type degradation, whereby the microbes had hydro-
lysed the ester bonds at the termini of the polymeric
chains preferentially.
All the more, low water solubility and a tendency
to adsorb to particulate matter in soils and sediments
are factors that can severely limit in situ biodegra-
dation of many sediments contaminated with organic
contaminants, polychlorinated and polycyclic aro-
matic hydrocarbons. Consequently, the rates of
desorption and dissolution of contaminants in the
water phase can be improved by adding surfactants
(either biosurfactants or synthetic detergents) to the
contaminated zone (Mulligan 2009). Biosurfactants
are surface active compounds having a wide range of
industrial applications such as enhanced oil recovery,
lubricants, bioremediation of pollutants and food
processing. The structures of these complex mole-
cules include lipopeptides, glycolipids, polysaccha-
ride protein complexes, fatty acids and phospholipids.
Rev Environ Sci Biotechnol (2010) 9:215–288 265
123
Optimal production of biosurfactant (glycolipid) by
Bacillus megaterium was obtained in 3L laboratory
scale fermenter when peanut oil cake (2%) was used
as carbon source (Thavasi et al. 2008). Carotenoids
are important natural pigments with a range of
applications as colorants, feed supplements and
neutraceuticals. Lycopene is a red coloured interme-
diate of the b-carotene biosynthetic pathway and is an
important dietary carotenoid. It is reported to inhibit
the harmful effect of ferric nitrilotriacetate on DNA
in rats and prevents liver necrosis. Lopez-Nieto et al.
(2004) reported the development of a semi-industrial
process (800 l fermentor) for lycopene production by
mated fermentation of Blakeslea trispora plus (?)
and minus (-) strains. This process describes the
critical requirement of soybean cake (44 g/l) as
nitrogen source for optimal lycopene production.
Mustard oil cake (6%) in the presence of Mg2? ions
is reported to improve lactic acid production ability
of agar-gel immobilized Lactobacillus casei after
48 h, when further addition of the substrate (whey
lactose) failed to maintain the process efficiency (Tuli
et al. 1985). Biosurfactants produced by microorgan-
isms within soils and sediments have been shown to
enhance biodegradation rates. In a laboratory study to
assess the effect of adding either Pseudomonas
aeruginosa UG2 cells or the biosurfactants produced
by this microorganism on the biodegradation of a
hydrocarbon mixture in soil at 20�C over a 2-month
incubation period, Jain et al. (1992) had observed that
the addition of 100 lg of UG2 biosurfactants g-1 soil
significantly enhanced the degradation of tetradecane,
hexadecene and pristane but not 2-methylnaphtha-
lene, the most water-soluble of the hydrocarbons. The
effect of two nonionic surfactants of the alkylpheno-
lethoxylate type, Arkopal N-300 and Sapogenat
T-300, on the bioavailability of PAH in manufactured
gas plant soil was evaluated in soil columns perco-
lated by recirculating flushing water by Tiehm et al.
(1997). It was observed that both surfactants had
enhanced the mass transfer rate of sorbed PAH into
the aqueous phase due to solubilization and made it
more bioavailable for biodegradation. Bardi et al.
(2000) have analyzed the in vitro effect of cyclodex-
trins on the biodegradative activity of a microbial
population isolated from a petroleum-polluted soil, as
shown by the decrease of dodecane (C12), tetraco-
sane (C24) anthracene and naphthalene added indi-
vidually as the sole carbon source to mineral medium
liquid cultures. B-cyclodextrin was seen to have
accelerated the degradation of all four hydrocarbons,
particularly naphthalene, and influenced the growth
kinetics as shown by a higher biomass yield and
better utilization of hydrocarbon as a carbon and
energy source.
Lately, Whang et al. (2008) have investigated the
potential application of two biosurfactants, surfactin
(SF) and rhamnolipid (RL), for the biodegradation of
diesel-contaminated water and soil with a series of
bench-scale experiments. The rhamnolipid used in this
study was a commonly isolated glycolipid biosurfac-
tant produced by Pseudomonas aeruginosa J4, while
the surfactin, was a lipoprotein type biosurfactant
produced by Bacillus subtilis ATCC 21332. It was
deduced that both biosurfactants had been able in
increasing the diesel solubility with increased biosur-
factant addition. In the diesel/water batch experiments,
an addition of 40 mg l-1 of surfactin significantly
enhanced biomass growth (2,500 mg VSS l-1) as well
as increased the diesel biodegradation percentage to
94%, compared to batch experiments with no surfactin
addition (1,000 mg VSS/l and 40% biodegradation).
The addition of rhamnolipid to diesel/water systems
from 0 to 80 mg/l substantially increased biomass
growth and diesel biodegradation percentage from
1,000 to 2,500 mg VSS/l and 40–100%, respectively.
Hence, the enhancing capability on both efficiency and
rate of diesel biodegradation in diesel/soil systems of
surfactin and rhamnolipid was clearly demonstrated by
Whang et al. (2008). Last but not least, Lai et al. (2009)
have recently developed a screening method to eval-
uate the oil removal capability of biosurfactants for oil-
contaminated soils collected from a heavy oil-polluted
site using two biosurfactants (rhamnolipids and surf-
actin) and synthetic surfactants (Tween 80 and Triton
X-100). Their results have convincingly shown that
biosurfactants exhibited much higher TPH removal
efficiency than the synthetic ones examined. Lai et al.
(2009) also reported that the TPH removal efficiency
had increased with an increase in biosurfactant
concentration (from 0 to 0.2 mass %). Yet another
promising approach to improve bioremediation rates is
to add biodegradable solvents to assist desorption and
dissolution rates with consequent increase in the
biodegradation of the adsorbed pollutants (Ludmer
et al. 2009).
Zoller and Reznik (2006) have developed a
surfactant/surfactant-nutrient mix (SSNM) for
266 Rev Environ Sci Biotechnol (2010) 9:215–288
123
enhanced bioremediation methodologies for sustain-
able, in situ bioremediation of fuel-contaminated
aquifers. The major findings of this study were the
kerosene’s maximum enhanced mobilization com-
pared with that of deionized water when using SSNM
having composition of linear alkylbenzene sulphonate
(LABS): coco-amphodiacetate (containing N): surfac-
tant-nutrient X (containing both N and P) at 0.15: 0.15:
0.05 g/l, respectively. The major effects of the SSNM
addition were reported to be the enhanced mobiliza-
tion of the bulky nonaqueous phase liquid (NAPL) and
the enhanced desorbtion/ solubilization/dispersion of
the entrapped NAPL which, in turn, facilitated their
enhanced biodegradation. Sun et al. (2009) have
examined the laboratory use of aqueous ethyl lacta-
teodified [S,S]-ethylenediaminedisuccinic acid (EDDS)
washing solutions for the simultaneous removal of
phenanthrene, pyrene, and Cu from contaminated soils.
Ethyl lactate demonstrated greater solubilization effi-
ciency for phenanthrene and pyrene than ethanol. Thus,
ethyl lactate is believed to have a greater potential for
extracting PAHs from contaminated soils. Sun et al.
(2009) highlight that the addition of ethyl lactate in
EDDS solution (EDDS/Cu molar ratio = 2) efficiently
enhanced the extraction of the PAHs and also signif-
icantly increased the Cu removal from 34.8 to 42.9%.
The latter was mainly attributed to the fact that ethyl
lactate increased the stability constant for Cu-EDDS
complexes, hence shifting the degree of desorption of
Cu from soil.
10.9 Co-metabolism
Co-metabolism is a process whereby microorganisms
involved in the metabolism of a growth promoting
substrate also transform other organic contaminants
which can be called as co-substrates (Prince 2010).
The latter are however not growth supporting if they
are provided as the only sources of carbon and
energy. This is one of the most important mecha-
nisms involved in the transformation of chlorinated
organic aliphatic and aromatic contaminants by
microbes. Such co-metabolic transformation of
organic pollutants is an important process in both
aerobic and anaerobic environments namely bacterial
transformation of dichlorodiphenyl trichloroethane
DDT, PCBs (Abraham et al. 2002) and trichloroeth-
ylene. Volpe et al. (2007) have used batch reactors
and microcosms to evaluate groundwater bioremedi-
ation potential of tetrachloroethene (PCE) in the
presence of additional pollutants present. Reductive
dechlorination of PCE was studied under anaerobic
and aerobic conditions. It was observed that the
consortia derived from anaerobic sludge and amended
with electron donors quantitatively and incompletely
degraded PCE to cis-dichloroethylene, whereas in
reactors augmented with a dehalogenating culture
complete dechlorination of PCE occurred even in the
presence of additional toxic contaminants. Adebusoye
et al. (2008) have observed substantial metabolism of
2,3,4,5-tetrachlorobiphenyl (2,3,4,5-tetraCB) and
2,30,40,5-tetraCB by axenic cultures of Ralstonia sp.
SA-5 and Pseudomonas sp. SA-6 in the presence of
biphenyl supplementation, although, the strains were
unable to utilize tetrachlorobiphenyls as growth sub-
strate. Ziagova et al. (2009) have reported a compar-
ison of the ability of Staphylococcus xylosus to
degrade 2,4-dichlorophenol and 4-Cl-m-cresol in sep-
arate cultures. In this study, bacterial adaptation and
the continuous presence of glucose, as a conventional
carbon source, were found to stimulate the degrading
efficiency of S. xylosus. All the more, microbes can
sequentially remove halogen atoms from polluting
halogenated compounds wherein halogen atoms are
replaced by hydrogen under anaerobic conditions.
Here, halogen atoms serve as hydrogen acceptors and
hence dehalogenation involves co-metabolism and
provision of a growth promoting substance.
10.10 Gene expression
The ability of indigenous microorganisms to degrade
organic pollutants is dependent on the expression of
the genes encoding the required enzymes. These genes
may not express if these substances are available in
very low concentrations. This can be overcome by
adding substances that are structurally related to the
organic pollutants which will act as inducers. Simi-
larly, the presence of alternate carbon or energy source
may repress the expression of the degradative enzyme
needed to transform the target pollutant. For example,
addition of glucose or amino acids to aquifer samples
contaminated with toluene, ethylenedibromide, phe-
nol and p-nitrophenol inhibits the degradation of these
contaminants because the microbes will prefer the
more easily degradable substrate.
Rev Environ Sci Biotechnol (2010) 9:215–288 267
123
10.11 Bioaugmentation
Where degradative microbes do not exit or where the
process is too slow, microbial inoculates may be added
to enhance bioremediation rates. This technique is
known as bioaugmentation (Lima et al. 2009) and may
involve (a) an addition of natural isolates of bacteria or
(b) genetically engineered organisms (GEMs). There
are rigid rules governing the release of GEMs as there
is concern about their potential negative impacts on the
environment. The genetic patterns have evolved over
several decades and they are relatively stable. It is
believed that altered genomes have greater instability
and increase the chances of mutations, some of which
may not be safe. Bioaugmentation has met with
varying degrees of success. Gertler et al. (2009) have
applied an experimental prototype oil boom including
oil sorbents, slow-release fertilizers and biomass of the
Marıne oil-degrading bacterium, Alcanivorax bor-
kumensis, for sorption and degradation of heavy fuel
oil in a 500-L mesocosm experiment, and it was found
that growth of this obligate oil-degrading bacterium on
immobilized oil coincided with a 30-fold increase in
total respiration. Earlier, Bento et al. (2005) evaluated
the effect of bioaugmentation on the degradation of
TPH in soil. It was reported that bioaugmentation of
the contaminated soil showed the greatest degradation
in the light (72.7%) and heavy (75.2%) fractions of
TPH since the greatest microbial activity (dehydroge-
nase activity) had occurred with bioaugmentation up
to 3.3-fold. Jacques et al. (2008) have evaluated the
capacity of a defined microbial consortium (five
bacteria: Mycobacterium fortuitum, Bacillus cereus,
Microbacterium sp., Gordonia polyisoprenivorans,
Microbacteriaceae bacterium, Naphthalene-utilizing
bacterium; and a fungus identified as Fusarium
oxysporum) isolated from a PAHs contaminated
landfarm site to degrade and mineralize different
concentrations (0, 250, 500 and 1,000 mg/kg) of
anthracene, phenanthrene and pyrene in soil, and it
was found that the microbial consortium had degraded
on average, 99, 99 and 96% of the different concen-
trations of anthracene, phenanthrene and pyrene in the
soil, in 70 days, respectively. Domde et al. (2007)
equally reported a 52.2% removal of chemical oxygen
demand (COD) in a bioaugmented reactor while only
15.1% reduction of COD was observed in the reactor
without bioaugmentation. Domde et al. (2007) have
suggested that the gene pool of the bioaugmented
reactor had catabolic loci that could degrade accumu-
lated intermediates, thereby improving the efficiency
of the oevrall system. Much recently, Teng et al.
(2010) have conducted a microcosm study to test the
bioremediation potential of Paracoccus sp. strain
HPD-2 on an aged PAH-contaminated soil. The
bioaugmented microcosms showed (a) a 23.2%
decrease in soil total PAH concentrations after
28 days, with a decline in average concentration from
9,942 to 7,638 lg/kg dry soil, and (b) higher counts of
culturable PAH-degrading bacteria, microbial bio-
mass and enzyme activities were observed in bioaug-
mented soil.
11 Novel research trends in bioremediation
Bioremediation, an intimate branch of biotechnology,
in principle includes the use of microorganisms in
improving the condition of a contaminated site, with
most commonly bacteria being the degraders and other
organisms, such as soil animals or plant roots, playing
a role in disseminating the bacteria and, in providing
nutrients and co-substrates for the bacteria active in the
degradation processes (Romantschuk et al. 2000).
Bioremediation has, in principle, considerable public
support because it aims to enhance natural processes
and it is generally seen as ‘‘environmentally appropri-
ate.’’ However, bioremediation rates are often consid-
erably slower than physical methods such as removing
the contaminated material to a secure landfill (Prince
2010). In this respect, and in pursuit to improve the
performance of bioremediation processes, there have
been a number of different procedures that have been
tested more-or-less successfully with a view to
improve reliability, cost efficiency and bioremediation
rates. These methods range from minimal interven-
tion, such as mere monitoring of intrinsic bioremedi-
ation, through in situ introduction of nutrients and/or
bacterial inocula or improvement of physico-chemical
conditions, or still excavation followed by on site or ex
situ composting in its different varieties.
However, modern biotechnology including genetic
engineering; culture of recombinant microorganisms,
cells of animals and plants; metabolic engineering;
hybridoma technology; bioelectronics; nanobiotech-
nology; protein engineering; transgenic animals and
plants; tissue and organ engineering; immunological
assays; genomics and proteomics; bioseparations and
268 Rev Environ Sci Biotechnol (2010) 9:215–288
123
bioreactor technologies (Gavrilescu and Chisti 2005)
have been gaining momentum in research and showing
much promise to improve bioremediation rates. Strat-
egies for improving bioremediation efficiency using
genetic engineering consist in improving strains and
chemotactic ability, the use of mixed population
biofilms and optimization of physico-chemical condi-
tions. For example, biofilms are assemblages of single
or multiple populations that are attached to abiotic or
biotic surfaces through extracellular polymeric sub-
stances (Singh et al. 2006a). Gene expression in
biofilm cells differs from planktonic stage expression
and these differentially expressed genes regulate
biofilm formation and development. Biofilm systems
have been shown to be especially suitable for the
treatment of recalcitrant compounds because of their
high microbial biomass and ability to immobilize
compounds (Singh et al. 2006a). All the more,
bioremediation is also facilitated and bioremediation
rates enhanced by gene transfer among biofilm organ-
isms and by the increased bioavailability of pollutants
for degradation as a result of bacterial chemotaxis
(Singh et al. 2006a). Table 16 presents some of the
novel research trends and/or advances depicted in
bioremediation.
11.1 Genetically engineered microorganisms
(GEMs) and microbial systems
As in practically all microbial applications, the use of
genetic engineering to improve microbial capacities
opens many interesting possibilities to obtain new
species that are able to use or to degrade different
contaminants with high efficiency (Iranzo et al.
2001). In the case of bioremediation there is much
scientific work suggesting that engineered microor-
ganisms have greater potential for environmental
clean-up than natural ones (Raskin 1996; Pieper and
Reineke 2000; Iranzo et al. 2001). Particular attention
is also being given to the genetic engineering of
bacteria using bacterial hemoglobin (VHb) for the
treatment of aromatic organic compounds under
hypoxic conditions (Urgun-Demirtas et al. 2006).
The application of VHb technology may advance
treatment of contaminated sites, where oxygen
availability limits the growth of aerobic bioremedi-
ating bacteria, as well as the functioning of oxygen-
ases required for mineralization of many organic
pollutants (Urgun-Demirtas et al. 2006). However,
despite the many advantages of GEMs, there are still
concerns that their introduction into polluted sites to
enhance bioremediation may have adverse environ-
mental effects, such as gene transfer.
A number of new recombinant DNA techniques
have been developed for genetically engineered
microorganisms for the biodegradation of environ-
mental contaminants or for the synthesis of small
molecules (Keasling and Bang 1998). These tech-
niques include new expression vectors to carry the
heterologous genes into the host organism, new
mechanisms to control gene expression, containment
mechanisms to control persistence of genetically-
engineered microorganisms, application of site-direc-
ted and random mutagenesis to increase the substrate
range or activity of biodegradative enzymes, and
methods to track genetically-engineered microorgan-
isms (Keasling and Bang 1998). The application of
culture-independent molecular biological techniques
also offers new opportunities to better understand the
dynamics of microbial communities (Iwamoto and
Nasu 2001). Fluorescence in situ hybridization (FISH),
in situ PCR, and quantitative PCR are expected to be
powerful tools for bioremediation to detect and
enumerate the target bacteria that are directly related
to the degradation of contaminants, and thence better
engineer these for enhanced metabolisms related to
pollutants degradation (Iwamoto and Nasu 2001).
Nucleic acid based molecular techniques for finger-
printing the 16S ribosomal DNA (rDNA) of bacterial
cells, i.e., denaturing gradient gel electrophoresis
(DGGE) and terminal restriction fragment length
polymorphism (T-RFLP), have enabled the monitoring
of the changes in bacterial community in detail, and
such advanced molecular microbiological techniques
will definitely provide new insights into bioremedia-
tion in terms of process optimization, validation, and
the impact on the ecosystem, which are indispensable
data to make the technology reliable and safe (Iwamoto
and Nasu 2001).
Although plants have the inherent ability to detoxify
some xenobiotic pollutants, they generally lack the
catabolic pathway for complete degradation/miner-
alization of these compounds compared to micro-
organisms. Hence, transfer of genes involved in
xenobiotic degradation from microbes/other eukary-
otes to plants may further enhance their potential
for remediation of such dangerous groups of com-
pounds. Transgenic plants with enhanced potential for
Rev Environ Sci Biotechnol (2010) 9:215–288 269
123
detoxification of xenobiotics such as trichloro ethyl-
ene, pentachlorophenol, trinitro toluene, glycerol trin-
itrate, atrazine, ethylene dibromide, metolachlor and
hexahydro-1,3,5-trinitro-1,3,5-triazine are a few suc-
cessful examples of utilization of transgenic technol-
ogy (Eapen et al. 2007). Trees are already being used
for wastewater clean-up, for site stabilization, and as
barriers to subsurface flow of contaminated ground-
water. Clonal propagation and the genetic tools of both
classical breeding and genetic engineering exist for a
number of both angiosperm and gymnosperm species,
opening the door to creation of tree ‘‘remediation’’
cultivars (Stomp et al. 1993). Active research is also
underway to screen tree and plant species for their
enhanced ability to tolerate, take up, translocate,
sequester, and degrade organic compounds and heavy
metal ions. Chen and Wilson (1997) have evaluated
cells of a genetically engineered Escherichia coli
Table 16 Recent research trends and advances reported in bioremediation
Bioremediation method Outline of novel finding(s) Reference
Treatment of sites contaminated
with chlorinated solvents
Results suggest that the reductive treatment of chlorinated
solvent sites with nano-scale zero-valent iron particles
might be enhanced by the concurrent or subsequent
participation of bacteria that exploit cathodic
depolarization and reductive dechlorination as metabolic
niches
Xiu et al. (2010)
Bioremediation process by biosorption
of effluents of wash process of the cotton
fabric by silver nanoparticles with the
bacterium Chromobacterium violaceum
The bacteria after biosorption were morphologically
transformed, but the normal morphology after a new
culture was completely restored. The process also allowed
the recovery of silver material that was leached into the
effluent for a reutilization avoiding any effect to the eco-
environment
Duran et al. (2010)
Reduction and adsorption of Pb2? in aqueous
solutions
Nano-zero-valent iron was produced by a reduction method
and compared with commercial available zero-valent iron
powder for Pb2? removal from aqueous phase. In
comparison with Fluka zero-valent iron, nano-zero-valent
iron has much higher reactivity towards Pb2? and within
just 15 min 99.9% removal can be reached. Nano-zero-
valent iron material has thus been demonstrated to have
great potential for heavy metal immobilization from
wastewater
Xi et al. (in press)
Bacterial degradation of organophosphates
(OPs)
Stenotrophomonas sp. strain YC-1, a native soil bacterium
that produces methyl parathion hydrolase (MPH), was
genetically engineered to possess a broader substrate
range (OPs). Results indicate that the broader substrate
specificity in combination with the rapid degradation rate
makes this engineered strain a promising candidate for in
situ remediation of OP-contaminated sites
Yang et al. (2010)
Fungal degradation of oily
sludge-contaminated soil
A novel yeast strain Candida digboiensis TERI ASN6 was
developed and could degrade 40 mg of eicosane in 50 ml
of minimal salts medium in 10 days and 72% of
heneicosane in 192 h at pH 3. The degradation of alkanes
yielded monocarboxylic acid intermediates while the
polycyclic aromatic hydrocarbon pyrene found in the
acidic oily sludge yielded the oxygenated intermediate
pyrenol
The strain C. digboiensis could efficiently degrade the
acidic oily sludge on site because of its robust nature,
probably acquired by prolonged exposure to the
contaminants. Hence, the potential of Candidadigboiensis TERI ASN6 to bioremediate hydrocarbons
at low pH under field conditions has been demonstrated
Sood et al. (2010)
270 Rev Environ Sci Biotechnol (2010) 9:215–288
123
strain, JM109, which expresses metallothionein and a
Hg2? transport system after induction for their selec-
tivity for Hg2? accumulation in the presence of
sodium, magnesium, or cadmium ions and their
sensitivity to pH or the presence of metal chelators
during Hg2? bioaccumulation. The genetically engi-
neered E.coli cells in suspension were observed to have
accumulated Hg2? effectively at low concentrations
(0–20 lmol/l) over a broad range of pH (3–11). These
results suggested that the E. coli strain JM109 could be
used for selective removal of Hg2? from wastewater or
from contaminated solutions which are normally
resistant to common treatments. In a attempt to further
enhance the efficiency and potential of plants for
phytoremediation of mercury pollution, Nagata et al.
(2009) constructed a genetically engineered tobacco to
simultaneously express mercury transporter, mercury
transporter (MerT) and mercury chelator (Kiyono and
Pan-Hou 2006), polyphosphate (polyP) by integrating
bacterial merT gene in polyphosphate kinase gene
(ppk)-transgenic tobacco to evaluate its ability to
phytoremediate mercury. It was observed that the
integration of the merT gene into the ppk-transgenic
tobacco did not significantly affect the mercury
resistant phenotypes and polyp production but the
transgenic expression of MerT in ppk-transgenic
tobacco had resulted in an accelerated and enhanced
mercury uptake into tobacco. In addition, tobacco
expressing MerT and polyP accumulated significantly
more mercury than the ppk-transgenic tobacco from
medium containing a wide range of low concentrations
of Hg2?. Later, Deng et al. (2005) constructed a
genetically engineered E. coli SE5000 strain simulta-
neously expressing nickel transport system and metal-
lothionein to accumulate Ni2? from aqueous solution.
Compared with 1.62 mg/g of Ni2? uptake capacity by
original host E. coli cells, the genetically engineered E.
coli could remarkably bind 7.14 mg/g Ni2?, and it
accumulated Ni2? effectively over a broad range of pH
(4–10), with an optimal pH at 8.6.
However, the vast majority of studies pertaining to
genetically engineered microbial bioremediation are
mostly supported by laboratory-based experimental
data (Sayler and Ripp 2000). In general, relatively few
examples of GEM applications in environmental
ecosystems exist, and unfortunately, the only manner
in which to fully address the competence of GEMs in
bioremediation efforts is through long-term field scale
studies whereby a reasonable pool of requisite
information for determining the overall effectiveness
and risks associated with GEM introduction into
natural ecosystems is acquired (Sayler and Ripp
2000).
11.2 Nanotechnology and bioremediation
Nanotechnology has contributed to the development
of a great diversity of materials as those used in
electronic, optoelectronic, biomedical, pharmaceuti-
cal, cosmetic, energy, catalytic, and materials appli-
cations. As a general definition, nanotechnology is
involved with objects on the nano scale, or materials
measuring between 1 and 100 nm (Duran 2008). In
future, modification and adaptation of nanotechnol-
ogy will extend the quality and length of life
(Rajendran and Gunasekaran 2007). The social
benefits are significant from nanomaterials and the
new products are applicable to information technol-
ogy, medicine, energy, and environment.
The emergence of nanotechnology presents a
number of potential environmental benefits. Most
environmental applications of nanotechnology fall
into three categories: (i) environmentally-benign and/
or sustainable products (e.g., green chemistry or
pollution prevention), (ii) remediation of materials
contaminated with hazardous substances, and (iii)
sensors for environmental agents (Tratnyek and
Johnson 2006). Some nanoparticles destroy contam-
inants, for instance, while others sequester them (Rao
and Murthy 2007, Telling et al. 2009). Carbon
nanotubes, for example, have been recognized for
their ability to adsorb dioxin much more strongly
than traditional activated carbon (Duran 2008). All
the more, the utilization of microbes for intracellular/
extracellular synthesis of nanoparticles with different
chemical composition, size/shapes and controlled
monodispersity can be a novel, economically viable
and eco-friendly strategy that can reduce toxic
chemicals in the conventional protocol.
Mace et al. (2006) have studied the assessment of
remediation of soil heavy metals with nano-particle
hydroxyapatite by the Toxicity Characteristic Leaching
Procedure by cultivation experiment. Ther results
indicated that nano-particle hydroxyapatite significantly
reduced the bioavailability of soil Cu and Zn when
compared with the control. The more nano-particle
hydroxyapatite were added, the more was the increasing
equilibrium time, and the more was the decreased
Rev Environ Sci Biotechnol (2010) 9:215–288 271
123
bioavailability of soil Cu and Zn, since soil pH was
significantly increased after the addition of nano-
particle hydroxyapatite, and heavy metals could adsorb
on nano-particle hydroxyapatite. In their study, Vara-
nasi et al. (2007) have used nano-particles to remediate
PCB contaminated soil and an attempt was made to
maximize PCB destruction in each treatment step. Their
results showed that nano-particles did aid in the
dechlorination process and high PCB destruction
efficiencies could be achieved, with a minimum total
PCB destruction efficiency reported at 95%. Kanel et al.
(2007) have synthesized, characterized and tested
surface-modified iron nanoparticles (S-INP) for the
remediation of arsenite (As(III)), a well known toxic
groundwater contaminant of concern. The results using
S-INP pretreated 10 cm sand-packed columns contain-
ing *2 g of S-INP showed that 100% of As(III) was
removed from influent solutions at a flow rate 1.8 ml/
min containing 0.2, 0.5 and 1.0 mg/l As(III) for 9, 7 and
4 days providing 23.3, 20.7 and 10.4 l of arsenic free
water, respectively. In addition, it was found that 100%
of As(III) in 0.5 mg/l solution for the same flow rate was
removed by S-INP pretreated 50 cm sand packed
column containing 12 g of S-INP for more than
2.5 months providing 194.4 l of arsenic free water.
These results hence suggested that S-INP have great
potential to be used as a mobile, injectable reactive
material for in situ sandy groundwater aquifer treatment
of As(III). In their recent study, Elliott et al. (2008)
exposed groundwater and aquifer samples from a site
contaminated by hexachlorocyclohexanes (totaling
1,500 lg l-1)) to nanoscale iron particles to evaluate
the technology as a potential remediation method. Batch
experiments with 2.2–27.0 g/l iron nanoparticles
showed that more than 95% of the HCHs were removed
from solution within 48 h. Based on a survey of
literature of previously published work on a wide
variety of chlorinated organic solvents, the work of
Elliott et al. (2008) additionally demonstrated the
potential of zerovalent iron nanoparticles for treatment
and remediation of persistent organic pollutants (POPs).
12 Concluding remarks
Anthropogenic activities have caused widespread
pollution of the natural environment. A number of
organic pollutants, such as PAHs, PCBs and pesticides,
and inorganic pollutants (heavy metals like arsenic,
cadmium, chromium, lead and zinc) are resistant to
degradation and represent an ongoing toxicological
threat to both wildlife and human beings. Bioremedi-
ation has grown into a green, attractive and promising
alternative to traditional physico-chemical techniques
for the remediation of these POPs at a contaminated
site, as it can be more cost-effective and it can
selectively degrade the pollutants without damaging
the site or its indigenous flora and fauna. However,
bioremediation technologies have had limited appli-
cations due to the constraints imposed by substrate and
environmental variability, and the limited biodegrada-
tive potential and viability of naturally occurring
microorganisms.
This review was not intended to address the much
voluminous literature on bioremediation, but rather to
revisit the basic of bioremediation and demonstrate
that the application of biotreatment is growing
rapidly due to its merits which outweigh the demerits.
The application of diverse bioremediation technolo-
gies must be based on sound and relilable scientific
data obtained in both fundamental as well as research
environmental laboratories. For the development of
bioremedial processes to succeed commercially, it is
essential to link different disciplines such as micro-
bial ecology, biochemistry and microbial physiology,
together with biochemical and bioprocess engineer-
ing. In short, the key to successful bioremediation
resides in continuing to develop the scientific and
engineering work that provides the real bases for both
the technology and its evaluation; and simultaneously
in explaining and justifying the valid reasons which
allow scientists and engineeres to actually use these
technologies for the welfare and safety of a public
which is more and more concerned about the
environment and its protection.
Acknowledgments We wish to express our deepest gratitude
to all the researchers whose valuable data as reported in their
respective publications and cited in this review have been of
considerable significance in adding substance to this review.
We are also grateful to our other colleagues and the anonymous
reviewers whose constructive criticisms have benefited the
manuscript, and brought it to its present form.
References
Abid N, Chamkha M, Godon JJ, Sayadi S (2007) Involvement
of microbial populations during the composting of olive
mill wastewater sludge. Environ Technol 28:751–760
272 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Aboul-Kassim TAT, Simoneit BRT (2001) Microbial trans-
formations at aqueous-solid phase interfaces: a bioreme-
diation approach. In: Aboul-Kassim TAT, Simoneit BRT
(eds) The handbook of environmental chemistry, vol. 5,
part E, pollutant-solid phase interactions: mechanism,
chemistry and modeling. Springer, Berlin
Abraham W-R, Nogales B, Golyshin PN, Pieper DT, Timmis
KN (2002) Polychlorinated biphenyl-degrading microbial
communities in soils and sediments. Curr Opin Microbiol
5:246–253
Adebusoye SA, Ilori MO, Picardal FW, Amund OO (2008)
Cometabolic degradation of polychlorinated biphenyls
(PCBs) by axenic cultures of Ralstonia sp. strain SA-5
and Pseudomonas sp. strain SA-6 obtained from Nigerian
contaminated soils. World J Microbiol Biotechnol
24:61–68
Adesodun JK, Atayese MO, Agbaje TA, Osadiaye BA, Mafe
OF, Soretire AA (2010) Phytoremediation potentials of
sunflowers (Tithonia diversifolia and Helianthus annuus)
for metals in soils contaminated with zinc and lead
nitrates. Water Air Soil Pollut 207:195–201
Agarry SE, Durojaiye AO, Solomon BO (2008) Microbial
degradation of phenols: a review. Int J Environ Pollut
32:12–28
Agnew JM, Leonard JJ (2003) Literature review—the physical
properties of compost. Compost Sci Util 11:238–264
Ahluwalia SS, Goyal D (2007) Microbial and plant derived
biomass for removal of heavy metals from wastewater.
Bioresour Technol 98:2243–2257
Ahuja R, Kumar A (2003) Metabolism of DDT [1, 1, 1-Tri-
chloro-2, 2-bis(4-chlorophenyl)ethane] by Alcaligenes
denitrificans ITRC-4 under aerobic and anaerobic condi-
tions. Curr Microbiol 46:65–69
Aitken MD, Walters GW, Crunk PL, Willis JL, Farrell JB,
Schafer PL, Arnett C, Turner BG (2005) Laboratory
evaluation of thermophilic-anaerobic digestion to produce
Class A biosolids. 1. Stabilization performance of a con-
tinuous-flow reactor at low residence time. Water Environ
Res 77:3019–3027
Akerlund T, Nordstrom K, Bernander R (1995) Analysis of cell
size and DNA content in exponentially growing and sta-
tionary-phase batch cultures of Escherichia coli. J Bacte-
riol 177:6791–6797
Akhtar N, Iqbal J, Iqbal M (2004) Removal and recovery of
nickel(II) from aqueous solution by loofa sponge-immo-
bilized biomass of Chlorella sorokiniana: characterization
studies. J Hazard Mater 108:85–94
Aksu Z (2005) Application of biosorption for the removal
of organic pollutants: a review. Process Biochem 40:
997–1026
Al-Daher R, Al-Awadhi N, Yateem A, Balba MT, ElNawawy
A (2001) Compost soil piles for treatment of oil-con-
taminated soil. Soil Sediment Contam An Int J 10:
197–209
Alisi C, Musella R, Tasso F, Ubaldi C, Manzo S, Cremisini C,
Sprocati AR (2009) Bioremediation of diesel oil in a co-
contaminated soil by bioaugmentation with a microbial
formula tailored with native strains selected for heavy
metals resistance. Sci Total Environ 407:3024–3032
Amir S, Hafidi M, Merlina G, Hamdi H, Jouraiphy A, El
Gharous M, Revel JC (2005) Fate of phthalic acid esters
during composting of both lagooning and activated slud-
ges. Process Biochem 40:2183–2190
Anastas PT, Kirchhoff MM (2002) Origins, current status, and
future challenges of green chemistry. Acc Chem Res
35:686–694
Anastas PT, Lankey RL (2000) Life-cycle assessment and
green chemistry: the yin and yang of industrial ecology.
Green Chem 6:289–295
Anastas PT, Warner JC (1998) Green chemistry, theory and
practice. Oxford University Press, Oxford
Anastas PT, Zimmerman JB (2003) Design through the 12
Principles of green engineering. Environ Sci Technol
37:94–101
Andersen RG, Booth EC, Marr LC, Widdowson MA, Novak JT
(2008) Volatilization and biodegradation of naphthalene
in the Vadose Zone impacted by phytoremediation.
Environ Sci Technol 42(7):2575–2581
Andreas KA, Ekelund NGA (2005) Effects on motile fac-
tors and cell growth of euglena gracilis after exposure to
wood ash solution; assessment of toxicity, nutrient avail-
ability and pH-dependency. Water Air Soil Pollut 162:
353–368
Ang EL, Zhao H, Obbard JP (2005) Recent advances in the
bioremediation of persistent organic pollutants via bio-
molecular engineering. Enzym Microbiol Technol
37:487–496
Antizar-Ladislao B, Galil NI (2003) Simulation of bioremedi-
ation of chlorophenols in a sandy aquifer. Water Res
37:238–244
Antizar-Ladislao B, Lopez-Real J, Beck AJ (2005) In-vessel
composting-bioremediation of aged coal tar soil: effect of
temperature and soil/green waste amendment ratio.
Environ Int 31:173–178
Apak R, Tutem E, Hugul M, Hizal J (1998) Heavy metal cation
retention by unconventional sorbents (red muds and fly
ashes). Water Res 32:430–440
Appels L, Baeyens J, Degreve J, Dewil R (2008) Principles and
potential of the anaerobic digestion of waste-activatedsludge. Prog Energy Combust Sci 34:755–781
Aravindhan R, Madhan B, Raghava Rao J, Unni Nair B (2004)
Recovery and reuse of chromium from tannery waste-
waters using Turbinaria ornata seaweed. J Chem Technol
Biotechnol 79:1251–1258
Arnaiz C, Gutierrez JC, Lebrato J (2006) Biomass stabilization
in the anaerobic digestion of wastewater sludges. Biore-
sour Technol 97:1179–1184
Arshad M, Zafar MN, Younis S, Nadeem R (2008) The use of
Neem biomass for the biosorption of zinc from aqueous
solutions. J Hazard Mater 157:534–540
Artola A, Barrena R, Font X, Gabriel D, Gea T, Mudhoo A,
Sanchez A (2009) Composting from a sustainable point of
view: respirometric indices as a key parameter. In: Mar-
tın-Gil J (ed) Compost II, dynamic soil dynamic plant, vol
3, pp 1–16
Aspray TJ, Carvalho DJC, Philip JC (2007) Application of soil
slurry respirometry to optimise and subsequently monitor
ex situ bioremediation of hydrocarbon-contaminated soils.
Int Biodeterior Biodegrad 60:279–284
Assinder SJ, Williams PA (1990) The TOL plasmids: deter-
minants of the catabolism’s of toluene and the xylenes.
Adv Microb Physiol 31:1–69
Rev Environ Sci Biotechnol (2010) 9:215–288 273
123
Atagana HI (2008) Compost bioremediation of hydrocarbon-
contaminated soil inoculated with organic manure. Afr J
Biotechnol 7:1516–1525
Atlante A, Giannattasio S, Bobba A, Gagliardi S, Petragallo V,
Calissano P, Marra E, Passarella S (2005) An increase in
the ATP levels occurs in cerebellar granule cells en route
to apoptosis in which ATP derives from both oxidative
phosphorylation and anaerobic glycolysis. Biochim Bio-
phys Acta (BBA)—Bioenergetics 1708:50–62
Azab MS, Peterson PJ (1989) The removal of Cd from
wastewater by the use of biological sorbent. Water Sci
Technol 21:1705–1706
Azadi H, Ho P (2010) Genetically modified and organic crops
in developing countries: a review of options for food
security. Biotechnol Adv 28:160–168
Azcon R, del Carmen Peralvarez M, Biro B, Roldan A, Ruız-
Lozano JM (2009) Antioxidant activities and metal
acquisition in mycorrhizal plants growing in a heavy-
metal multicontaminated soil amended with treated lig-
nocellulosic agrowaste. Appl Soil Ecol 41:168–177
Baczynski TP, Pleissner D (2010) Bioremediation of chlori-
nated pesticide-contaminated soil using anaerobic slud-
ges and surfactant addition. J Environ Sci Health B
45:82–88
Baek KH, Yoon BD, Cho DH, Kim BH, Oh HM, Kim HS
(2009) Monitoring bacterial population dynamics using
real-time PCR during the bioremediation of crude-oil-
contaminated soil. J Microbiol Biotechnol 19:339–345
Bai M-D, Chao Y-C, Lin Y-H, Lu W-C, Lee H-T (2009a)
Immobilized biofilm used as seeding source in batch
biohydrogen fermentation. Renew Energy 34:1969–1972
Bai FW, Zhang W, Zhong J-J (2009b) Special section on
biotechnology for the sustainability of human society.
Biotechnol Adv 26:939
Baker KH, Herson DS (1994) Bioremediation. McGraw-Hill
Inc, New York
Baker DB, Conradi MS, Norberg RE (1994) Explanation of the
high-temperature relaxation anomaly in a metal-hydrogen
system. Phys Rev B 49:11773–11782
Bali G, Rallapalli R, Sullia SB, Shiralipour A, Kasturi S (2002)
Environmental biotechnology: concepts, definitions and
criteria. In: Nangia SB (ed) Environmental biotechnology.
A.P.H. Publishing Corporation, New Delhi, pp 1–29
Bamforth SM, Singleton I (2005) Bioremediation of polycyclic
aromatic hydrocarbons: current knowledge and future
directions. J Chem Technol Biotechnol 80:723–736
Banuelos G, Lin Z-Q (2005) Phytoremediation management of
selenium-laden drainage sediments in the San Luis Drain:
a greenhouse feasibility study. Ecotoxicol Environ Saf
62:309–316
Bardi L, Mattei A, Steffan S, Marzona M (2000) Hydrocarbon
degradation by a soil microbial population with b-cyclo-
dextrin as surfactant to enhance bioavailability. Enzym
Microb Technol 27:709–713
Barker AV, Bryson GM (2002) Bioremediation of heavy
metals and organic toxicants by composting. Sci World J
2:407–420
Barnabe S, Brar SK, Tyagi RD, Beauchesne I, Surampalli RY
(2009) Pre-treatment and bioconversion of wastewater
sludge to value-added products—Fate of endocrine dis-
rupting compounds. Sci Total Environ 407:1471–1488
Baxter J, Cummings SP (2006) The current and future
applications of microorganism in the bioremediation of
cyanide contamination. Antonie van Leeuwenhoek 90:
1–17
Baysal Z, Cinar E, Bulut E, Alkan H, Dogru M (2009) Equi-
librium and thermodynamic studies on biosorption of
Pb(II) onto Candida albicans biomass. J Hazard Mater
161:62–67
Beck AJ, Jones KC (1995) Kinetic constraints on the In–situremediation of soils contaminated with organic chemicals.
Environ Sci Pollut Res 2:244–252
Beltrame MO, De Marco SG, Marcovecchio JE (2010) Effects
of zinc on molting and body weight of the estuarine crab
Neohelice granulata (Brachyura: Varunidae). Sci Total
Environ 408:531–536
Bengtsson S, Hallquist J, Werker A, Welander T (2008) Aci-
dogenic fermentation of industrial wastewaters: effects of
chemostat retention time and pH on volatile fatty acids
production. Biochem Eng J 40:492–499
Bennet JW, Wunch KG, Faison BD (2002) Use of fungi bio-
degradation. In: Hurst CC (ed) Environmental microbi-
ology, 2nd edn. ASM Press, Washington
Bento FM, Camargo FAO, Okeke BC, Frankenberger WT
(2005) Comparative bioremediation of soils contaminated
with diesel oil by natural attenuation, biostimulation and
bioaugmentation. Bioresour Technol 96:1049–1055
Beolchini F, Pagnanelli F, Toro L, Veglio F (2003) Biosorption of
copper by Sphaerotilus natans immobilised in polysulfone
matrix: equilibrium and kinetic analysis. Hydrometall
70:101–112
Bernal MP, McGrath SP, Miller AJ, Baker AJM (1994)
Comparison of the chemical changes in the rhizosphere of
the nickel hyperaccumulator Alyssum murale with the
non-accumulator Raphanus sativus. Plant Soil 164:
251–259
Bernal-Martinez A, Patureau D, Delgenes J-P, Carrere H
(2009) Removal of polycyclic aromatic hydrocarbons
(PAH) during anaerobic digestion with recirculation of
ozonated digested sludge. J Hazard Mater 162:1145–1150
Bernal–Martinez A, Carrere H, Patureau D, Delgenes J–P
(2007) Ozone pre-treatment as improver of PAH removal
during anaerobic digestion of urban sludge. Chemosphere
68:1013–1019
Bernal-Martınez A, Carrere H, Patureau D, Delgenes J-P
(2005) Combining anaerobic digestion and ozonation to
remove PAH from urban sludge. Process Biochem
40:3244–3250
Bhandari A, Xia K (2005) Hazardous organic chemicals in
biosolids recycled as soil amendments. Handbook Environ
Chem 5(Part F 1):217–239
Bhattacharyya KG, Sharma A (2004) Adsorption of Pb(II)
from aqueous solution by Azadirachta indica (Neem) leaf
powder. J Hazard Mater 113:97–109
Black H (1995) Absorbing possibilities: phytoremediation.
Environ Health Persp 103:1106–1108
Bluthgen A (2000) Organic migration agents into milk at farm
level (illustrated with di-ethylhexyl phthalate). Bull Int
Dairy Fed 356:39–42
Bohn I, Bjornsson L, Mattiasson B (2007) The energy balance
in farm scale anaerobic digestion of crop residues at
11–37�C. Process Biochem 42:57–64
274 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Bondada B, Ma LQ (2003) Tolerance of heavy metals in
vascular plants: arsenic hyperaccumulation by Chinese
brake fern (Pteris vittata L.). In: Chandra S, Srivastava M
(eds) Pteridology in the new millennium. Kluwer, The
Netherlands, pp 397–420
Bongochgetsakul N, Ishida T (2007) A new analytical
approach to optimizing the design of large-scale com-
posting systems. Bioresour Technol 99:1630–1641
Boopathy R (2000) Factors limiting bioremediation technolo-
gies. Bioresour Technol 74:63–67
Boparai HK, Shea PJ, Comfort SD, Machacek TA (2008)
Sequencing zerovalent iron treatment with carbon
amendments to remediate agrichemical-contaminated soil.
Water Air Soil Pollut 193:189–196
Børresen MH, Rike AG (2007) Effects of nutrient content,
moisture content and salinity on mineralization of
hexadecane in an Arctic soil. Cold Reg Sci Technol 48:
129–138
Bougrier C, Delgenes JP, Carrere H (2008) Effects of thermal
treatments on five different waste activated sludge sam-
ples solubilisation, physical properties and anaerobic
digestion. Chem Eng J 139:236–244
Brim H, Osborne JP, Kostandarithes HM, Fredrickson JK,
Wackett LP, Daly MJ (2006) Deinococcus radioduransengineered for complete toluene degradation facilitates
Cr(VI) reduction. Microbiology 152:2469–2477
Bruhlmann F, Chen W (1999) Tuning biphenyl dioxygenase
for extended substrate specificity. Biotechnol Bioeng
63:544–551
Brul S, Mensonides FIC, Hellingwerf KJ, Teixeira de Mattos
MJ (2008) Microbial systems biology: new frontiers open
to predictive microbiology. Int J Food Microbiol 128:
16–21
Busca G, Berardinelli S, Resini C, Arrighi L (2008) Technol-
ogies for the removal of phenol from fluid streams: a short
review of recent developments. J Hazard Mater 160:
265–288
Cai JL, Wang GC, Li YC, Zhu DL, Pan GH (2009) Enrichment
and hydrogen production by Marıne anaerobic hydrogen-
producing microflora. Chin Sci Bull 54:2656–2661
Chaney RL (1983) In: Parr J, Marsh EM (eds) Land treatment
of hazardous wastes. Noyes Data Corp., Park Rdge,
pp 50–76
Chaudhry TM, Hayes WJ, Khan AG, Khoo CS (1998) Phy-
toremediation-focusing on accumulator plants that reme-
diate metal-contaminated soils. Australas J Ecotoxicol
4:37–51
Chaudhry Q, Schroder P, Werck-Reichhart D, Grajek W,
Marecik R (2002) Prospects and limitations of phyto-
remediation for the removal of persistent pesticides in the
environment. Environ Sci Pollut Res 9:4–17
Chen S, Wilson DB (1997) Genetic engineering of bacteria and
their potential for Hg2? bioremediation. Biodegradion
8:97–103
Chen Y, Shen Z, Li X (2004) The use of vetiver grass (Vetiveriazizanioides) in the phytoremediation of soils contaminated
with heavy metals. Appl Geochem 19:1553–1565
Chen Y, Liu Y, Zhou Q, Gu G (2005) Enhanced phosphorus
biological removal from wastewater—effect of microor-
ganism acclimatization with different ratios of short-chain
fatty acids mixture. Biochem Eng J 27:24–32
Chen YD, Barker JF, Gui L (2008a) A strategy for aromatic
hydrocarbon bioremediation under anaerobic conditions
and the impacts of ethanol: a microcosm study. J Contam
Hydrol 96:17–31
Chen Y, Cheng JJ, Creamer KS (2008b) Inhibition of anaer-
obic digestion process: a review. Bioresour Technol 99:
4044–4064
Cheng H–F, Kumar M, Lin J–G (2008) Degradation kinetics of
di-(2-ethylhexyl) phthalate (DEHP) and organic matter of
sewage sludge during composting. J Hazard Mater 154:
55–62
Chintakovid W, Visoottiviseth P, Khokiattiwong S, Lau-
engsuchonkul S (2008) Potential of the hybrid marigolds
for arsenic phytoremediation and income generation of
remediators in Ron Phibun District, Thailand. Chemo-
sphere 70:1532–1537
Cho DH, Kim EY (2003) Characterization of Pb2? biosorption
from aqueous solution by Rhodotorula glutinis. Bioproc-
ess Biosyst Eng 25:271–277
Chong SL, Mou DG, Ali AM, Lim SH, Tey BT (2008) Cell
growth, cell-cycle progress, and antibody production in
hybridoma cells cultivated under mild hypothermic con-
ditions. Hybridoma 27:107–111
Chu CP, Chang BV, Liao GS, Jean SD, Lee DJ (2001)
Observations on changes in ultrasonically treated waste-
activated sludge. Water Res 35:1038–1046
Chu L, Yan S, Xing X-H, Sun X, Jurcik B (2009) Progress and
perspectives of sludge ozonation as a powerful pretreat-
ment method for minimization of excess sludge produc-
tion. Water Res 43:1811–1822
Cimino G, Cappello RM, Caristi C, Toscano G (2005) Char-
acterization of carbons from olive cake by sorption of
wastewater pollutants. Chemosphere 61:947–955
Cintas P, Luche JL (1999) Green chemistry: the sonochemical
approach. Green Chem 1:115–125
Clark JH (2006) Green chemistry: today (and tomorrow).
Green Chem 8:17–21
Clark B, Boopathy R (2007) Evaluation of bioremediation
methods for the treatment of soil contaminated with
explosives in Louisiana Army Ammunition Plant, Min-
den, Louisiana. J Hazard Mater 143:643–648
Clarkson DT, Luttge V (1989) Mineral intrusion: divalent
cations transport and compartmentalization. Prog Bot 51:
93–112
Cofield N, Schwab AP, Banks MK (2007) Phytoremediation of
polycyclic aromatic hydrocarbons in soil: part I. Dissipa-
tion of target contaminants. Int J Phytoremed 9:355–370
Contreras-Ramos SM, Alvarez-Bernal D, Dendooven L (2006)
Eisenia fetida increased removal of polycyclic aromatic
hydrocarbons from soil. Environ Pollut 141:396–401
Contreras-Ramos SM, Alvarez-Bernal D, Dendooven L (2009)
Characteristics of earthworms (Eisenia fetida) in PAHs
contaminated soil amended with sewage sludge or ver-
micompost. Appl Soil Ecol 41:269–276
Coulon F, Pelletier E, Gourhant L, Delille D (2005) Effects of
nutrient and temperature on degradation of petroleum
hydrocarbons in contaminated sub-Antarctic soil. Che-
mosphere 58:1439–1448
Crini G (2005) Recent developments in polysaccharide-based
materials used as adsorbents in wastewater treatment.
Prog Polym Sci 30:38–70
Rev Environ Sci Biotechnol (2010) 9:215–288 275
123
Cunningham SD, Berti WR (1993) Remediation of contami-
nated soils with green plants: An overview. In vitro Cell
Dev Biol 29:207–212
Da Silva EA, Cossich ES, Tavares CRG, Filho LC, Guirardello
R (2002) Modeling of copper(II) biosorption by marine
alga Sargassum sp. in fixed-bed column. Process Biochem
38:791–799
Dafale N, Nageswara Rao N, Meshram SU, Wate SR (2008)
Decolorization of azo dyes and simulated dye bath
wastewater using acclimatized microbial consortium-
Biostimulation and halo tolerance. Bioresour Technol
99:2552–2558
Dang VBH, Doan HD, Dang-Vu T, Lohi A (2009) Equilibrium
and kinetics of biosorption of cadmium(II) and copper(II)
ions by wheat straw. Bioresour Technol 100:211–219
Das K, Mukherjee AK (2007) Crude petroleum-oil biodegra-
dation efficiency of Bacillus subtilis and Pseudomonasaeruginosa strains isolated from a petroleum-oil contam-
inated soil from North-East India. Bioresour Technol
98:1339–1345
Das KC, Xia K (2008) Transformation of 4-nonylphenol isomers
during biosolids composting. Chemosphere 70:761–768
Das SK, Das AR, Guha AK (2007) A study on the adsorption
mechanism of mercury on Aspergillus versicolor biomass.
Environ Sci Technol 41:8281–8287
Davies OA, Allison ME, Uyi HS (2006) Bioaccumulation of
heavy metals in water, sediment and periwinkle (Tympa-notonus fuscatus var radula) from the Elechi Creek, Niger
Delta. Afr J Biotechnol 5:968–973
Davila-Vazquez G, Arriaga S, Alatriste-Mondragon F, de
Leon-Rodrıguez A, Rosales-Colunga LM, Razo-Flores E
(2008) Fermentative biohydrogen production: trends and
perspectives. Rev Environ Sci Biotechnol 7:27–45
Davis TA, Volesky B, Mucci A (2003) A review of the bio-
chemistry of heavy metal biosorption by brown algae.
Water Res 37:4311–4330
de la Rosa G, Peralta-Videa JR, Montes M, Parsons JG, Cano-
Aguilera I, Gardea-Torresdey JL (2004) Cadmium uptake
and translocation in tumbleweed (Salsola kali), a potential
Cd-hyperaccumulator desert plant species: ICP/OES and
XAS studies. Chemosphere 55:1159–1168
de Lorenzo V (2008) Systems biology approaches to biore-
mediation. Curr Opin Biotechnol 19:579–589
Deleu M, Paquot M (2004) From renewable vegetables
resources to microorganisms: new trends in surfactants.
C R Chimie 7:641–646
Delgado-Moreno L, Pena A (2009) Compost and vermicom-
post of olive cake to bioremediate triazines-contaminated
soil. Sci Total Environ 407:1489–1495
Deng X, Li QB, Lu YH, He N, Jiang J (2005) Genetic engi-
neering of E. coli SE5000 and its potential for Ni2? bio-
remediation. Process Biochem 40:425–430
Deng L, Su Y, Su H, Wang X, Zhu X (2007) Sorption and
desorption of lead (II) from wastewater by green algae
Cladophora fascicularis. J Hazard Mater 143:220–225
Desai C, Pathak H, Madamwar D (2010) Advances in molec-
ular and ‘‘-omics’’ technologies to gauge microbial com-
munities and bioremediation at xenobiotic/anthropogen
contaminated sites. Bioresour Technol 101:1558–1569
Dheri GS, Brar MS, Malhi SS (2007) Comparative phytoreme-
diation of chromium-contaminated soils by fenugreek,
spinach, and raya. Commun Soil Sci Plant Anal 38:
1655–1672
Diaz MJ, Madejon E, Lopez F, Lopez R, Cabrera F (2002)
Optimization of the rate vinasse/grape marc for co-com-
posting process. Process Biochem 37:1143–1150
Domde P, Kapley A, Purohit HJ (2007) Impact of bioaug-
mentation with a consortium of bacteria on the remedia-
tion of wastewater-containing hydrocarbons. Environ Sci
Pollut Res Int 14:7–11
Dordio AV, Duarte C, Barreiros M, Palace Carvalho AJ, Pinto
AP, Teixeira da Costa C (2009) Toxicity and removal
efficiency of pharmaceutical metabolite clofibric acid by
Typha spp.—potential use for phytoremediation? Biore-
sour Technol 100:1156–1161
dos Santos AB, Cervantes FJ, van Lier JB (2007) Review paper
on current technologies for decolourisation of textile
wastewaters: Perspectives for anaerobic biotechnology.
Bioresour Technol 98:2369–2385
Dosnon–Olette R, Couderchet M, Eullaffroy P (2009) Phyto-
remediation of fungicides by aquatic macrophytes: tox-
icity and removal rate. Ecotoxicol Environ Saf 72:
2096–2101
Dou J, Liu X, Hu Z, Deng D (2008) Anaerobic BTEX bio-
degradation linked to nitrate and sulfate reduction.
J Hazard Mater 151:720–729
Doucleff M, Terry N (2002) Pumping out the arsenic. Nat
Biotechnol 20:1094–1095
Doumett S, Lamperi L, Checchini L, Azzarello E, Mugnai S,
Mancuso S, Petruzzelli G, Del Bubba M (2008) Heavy
metal distribution between contaminated soil and Pau-lownia tomentosa, in a pilot-scale assisted phytoremedi-
ation study: Influence of different complexing agents.
Chemosphere 72:1481–1490
Dowling DN, Doty SL (2009) Improving phytoremediation
through biotechnology. Curr Opin Biotechnol 20:204–206
Drzyzga O, El Mamouni R, Agathos SN, Gottschal JC (2002)
Dehalogenation of chlorinated ethenes and immobiliza-
tion of nickel in anaerobic sediment columns under sul-
fidogenic conditions. Environ Sci Technol 36:2630–2635
Dumestre A, Chone T, Portal J, Gerard M, Berthelin J (1997)
Cyanide degradation under alkaline conditions by a strain
of fusarium solani isolated from contaminated soils. Appl
Environ Microbiol 63:2729–2734
Duran N (2008) Use of nanoparticles in soil-water bioreme-
diation processes. J Soil Sci Plant Nutrit 8:33–38
Duran N, Marcato PD, Alves OL et al (2010) Ecosystem
protection by effluent bioremediation: silver nanoparticles
impregnation in a textile fabrics process. J Nanoparticle
Res 12:285–292
Dushenkov V, Kumar PB, Motto AN, Raskin H (1995) Phy-
toextraction: the use of plants to remove heavy metals
from soils. Environ Sci Technol 29:1232–1238
Dytczak MA, Londry KL, Siegrist H, Oleszkiewicz JA (2007)
Ozonation reduces sludge production and improves
denitrification. Water Res 41:543–550
Eapen S, Singh S, D’Souza SF (2007) Advances in develop-
ment of transgenic plants for remediation of xenobiotic
pollutants. Biotechnol Adv 25:442–451
Ebbs S, Hatfield S, Nagarajan V, Blaylock M (2010) A
Comparison of the dietary arsenic exposures from inges-
tion of contaminated soil and hyperaccumulating Pteris
276 Rev Environ Sci Biotechnol (2010) 9:215–288
123
ferns used in a residential phytoremediation project. Int J
Phytoremed 12:121–132
Eckenfelder WW Jr (1989) Industrial water pollution control.
McGraw-Hill, New York
Ecobichon DJ (2000) Our changing perspectives on benefit and
risks of pesticides: a historical overview. Neurotoxicol
21:211–218
Ecobichon DJ (2001) Pesticide use in developing countries.
Toxicol 160:27–33
Eklind Y, Sundberg C, Smars S, Steger K, Sundh I, Kirchmann
H, Jonsson H (2007) Carbon turnover and ammonia
emissions during composting of biowaste at different
temperatures. J Environ Qual 36:1512–1520
Ekmekyapar F, Aslan A, Kemal Bayhan Y, Cakici A (2006)
Biosorption of copper(II) by nonliving lichen biomass of
Cladonia rangiformis hoffm. J Hazard Mater 137:
293–298
El-Bestawy E, Albrechtsen H-J (2007) Effect of nutrient
amendments and sterilization on mineralization and/or
biodegradation of 14C-labeled MCPP by soil bacteria
under aerobic conditions. Int Biodeterior Biodegrad
59:193–201
Elisashvili V, Penninckx M, Kachlishvili E et al (2008)
Lentinus edodes and Pleurotus species lignocellulolytic
enzymes activity in submerged and solid-state fermenta-
tion of lignocellulosic wastes of different composition.
Bioresour Technol 99:457–462
Elliott DW, Lien H-L, Zhang W (2008) Zerovalent iron
nanoparticles for treatment of ground water contami-
nated by hexachlorocyclohexanes. J Environ Qual 37:
2192–2201
Elouear Z, Bouzid J, Boujelben N, Feki M, Montiel A (2008)
The use of exhausted olive cake ash (EOCA) as a low cost
adsorbent for the removal of toxic metal ions from
aqueous solutions. Fuel 87:2582–2589
Esmaeili A, Beirami P, Rustaiyan A, Rafiei F, Ghasemi S,
Assadian F (2008) Evaluation of the marine alga Graci-laria Corticata for the biosorption of Cu (II) from
wastewater in a packed column. J Mar Environ Eng
9:65–73
Eullaffroy P, Vernet G (2003) The F684/F735 chlorophyll
fluorescence ratio: a potential tool for rapid detection and
determination of herbicide phytotoxicity in algae. Water
Res 37:1983–1990
Evans PJ, Trute MM (2006) In Situ bioremediation of nitrate
and perchlorate in vadose zone soil for groundwater
protection using gaseous electron donor injection tech-
nology. Water Environ Res 78:2436–2446
Failey RA, Scrivens AJ (1994) Contaminated land, assessment and
redevelopment. Stanley Thrones Publishers, Cheltenham
Fan L, Pandey A, Mohan R et al (2000) Use of various coffee
industry residues for the cultivation of Pleurotus ostreatusin solid state fermentation. Acta Biotechnol 20:41–52
Fantozzi F, Buratti C (2009) Biogas production from different
substrates in an experimental continuously stirred tank reac-
tor anaerobic digester. Bioresour Technol 100:5783–5789
Fatemi MH, Baher E (2009) A novel quantitative structure-
activity relationship model for prediction of biomagnifi-
cation factor of some organochlorine pollutants. Mol
Divers 13:343–352
Ferguson SH, Franzmann PD, Snape I, Revill AT, Trefry MG,
Zappia LR (2003) Effects of temperature on mineralisa-
tion of petroleum in contaminated Antarctic terrestrial
sediments. Chemosphere 52:975–987
Field JA, Sierra-Alvarez R (2008) Microbial degradation
of chlorinated phenols. Rev Environ Sci Biotechnol 7:
211–241
Figueira MM, Volesky B, Ciminelli VST, Roddick FA (2000)
Biosorption of metals in brown seaweed biomass. Water
Res 34:196–204
Fitzmorris KB, Sarmiento F, O’Callaghan P (2009) Biosolids
and sludge management. Water Environ Res 81:1376–
1393
Fleming J, Sanseverino J, Sayler G (1993) Quantitative rela-
tionship between naphthalene catabolic gene frequency
and expression in predicting PAH degradation in soils at
town gas manufacturing sites. Environ Sci Technol
27:1068–1074
Fogarty AM, Tuovinen OH (1991) Microbiological degrada-
tion of pesticides in yard waste composting. Microbiol
Rev Am Soc Microbiol 55:225–233
Forouzangohar M, Haghnia GH, Koocheki A (2005) Organic
amendments to enhance atrazine and metamitron degra-
dation in two contaminated soils with contrasting textures.
Soil Sediment Contam 14:345–355
Fountoulakis MS, Stamatelatou K, Lyberatos G (2008) The
effect of pharmaceuticals on the kinetics of methanogen-
esis and acetogenesis. Bioresour Technol 99:7083–7090
Francesconi K, Visoottiviseth P, Sridokchan W, Goessler W
(2002) Arsenic species in an arsenic hyperaccumulating
fern, Pityrogramma calomelanos: a potential phytoreme-
diator of arsenic-contaminated soils. Sci Total Environ
284:27–35
Fu J, Mai B, Sheng G, Zhang G, Wang X, Peng P, Xiao X, Ran
R, Cheng F, Peng X, Wang Z, Tang UW (2003) Persistent
organic pollutants in environment of the Pearl River
Delta, China: an overview. Chemosphere 52:1411–1422
Fuchedzhieva N, Karakashev D, Angelidaki I (2008) Anaero-
bic biodegradation of fluoranthene under methanogenic
conditions in presence of surface-active compounds.
J Hazard Mater 153:123–127
Gajalakshmi S, Abbasi SA (2008) Solid waste management by
composting: state of the art. Crit Rev Environ Sci Technol
38:311–400
Gal H, Ronen Z, Weisbrod N, Dahan O, Nativ R (2008)
Perchlorate biodegradation in contaminated soils and
the deep unsaturated zone. Soil Biol Biochem 40:1751–
1757
Garcia-Blanco S, Venosa AD, Suidan MT, Lee K, Cobanli S,
Haines JR (2007) Biostimulation for the treatment of an
oil-contaminated coastal salt marsh. Biodegradation
18:1–15
Gardea-Torresdey JL, Tiemann KJ, Gonzalez JH, Cano-
Aguilera I, Henning JA, Townsend MS (1996) Removal
of nickel ions from aqueous solution by biomass and sil-
ica—immobilized biomass of Medicago sativa (alfalfa).
J Hazard Mater 49:205–216
Garg P, Gupta A, Satya S (2006) Vermicomposting of different
types of waste using Eisenia foetida: a comparative study.
Bioresour Technol 97:391–395
Rev Environ Sci Biotechnol (2010) 9:215–288 277
123
Gavrilescu M, Chisti Y (2005) Biotechnology—a sustainable
alternative for chemical industry. Biotechnol Adv
23:471–499
Ge Y, Yan L, Qing K (2004) Effect of environment factors on
dye decolorization by P. sordida ATCC90872 in a aerated
reactor. Process Biochem 39:1401–1405
Gelman F, Binstock R (2008) Natural attenuation of MTBE
and BTEX compounds in a petroleum contaminated
shallow coastal aquifer. Environ Chem Lett 6:259–262
Germaine KJ, Liu X, Cabellos GG et al (2006) Bacterial endo-
phyte-enhanced phytoremediation of the organochlorine
herbicide 2, 4-dichlorophenoxyacetic acid. FEMS Micro-
biol Ecol 57:302–310
Gertler C, Gerdts G, Timmis KN, Golyshin PN (2009)
Microbial consortia in mesocosm bioremediation trial
using oil sorbents, slow-release fertilizer and bioaug-
mentation. FEMS Microbiol Ecol 69:288–300
Ghaly AE, Alkoaik F, Snow A (2006) Thermal balance of
invessel composting of tomato plant residues. Can Biosyst
Eng 48:1–11
Ghaly AE, Alkoaik F, Snow A (2007) Degradation of pirimi-
phos-methyl during thermophilic composting of green-
house tomato plant residues. Can Biosyst Eng 49:1–11
Gibson RW, Wang M-J, Padgett E, Lopez-Real JM, Beck AJ
(2007) Impact of drying and composting procedures on
the concentrations of 4-nonylphenols, di-(2-ethyl-
hexyl)phthalate and polychlorinated biphenyls in anang-
aerobically digested sewage sludge. Chemosphere
68:1352–1358
Gidarakos EL, Aivalioti MV (2008) In-well air sparging effi-
ciency in remediating the aquifer of a petroleum refinery
site. J Environ Eng Sci 7:71–82
Gijzen HJ, Bernal E, Ferrer H (2000) Cyanide toxicity and
cyanide degradation in anaerobic wastewater treatment.
Water Res 34:2447–2454
Goblos Sz, Portor}o P, Bordas D, Kalman M, Kiss I (2008)
Comparison of the effectivities of two-phase and single-
phase anaerobic sequencing batch reactors during dairy
wastewater treatment. Renew Energy 33:960–965
Godoy-Faundez A, Antizar-Ladislao B, Reyes-Bozo L, Cam-
ano A, Saez-Navarrete C (2008) Bioremediation of con-
taminated mixtures of desert mining soil and sawdust with
fuel oil by aerated in-vessel composting in the Atacama
Region (Chile). J Hazard Mater 151:649–657
Goel M, Chovelon J-M, Ferronato C, Bayard R, Sreekrishnan
TR (2010) The remediation of wastewater containing
4-chlorophenol using integrated photocatalytic and bio-
logical treatment. J Photochem Photobiol B Biol 98:1–6
Govind R, Lai L, Dobbs R (1991) Integrated model for pre-
dicting the fate of organics in wastewater treatment plants.
Environ Prog 10:13–23
Gurbuz F, Ciftci H, Akcil A (2009) Biodegradation of cyanide
containing effluents by Scenedesmus obliquus. J Hazard
Mater 162:74–79
Hamdi H, Benzarti S, Manusadzianas L, Aoyama I, Jedidi N
(2007) Bioaugmentation and biostimulation effects on
PAH dissipation and soil ecotoxicity under controlled
conditions. Soil Biol Biochem 39:1926–1935
Hamer G (1993) Bioremediation: a response to gross envi-
ronmental abuse. Trends Biotechnol 11:317–319
Hao X-D, Zhang L-P, Li L (2007) Global overview of excess
sludge treatment and disposal methods. China Water
Wastewater 23:1–5
Harman G, Patrick R, Spittler T (2007) Removal of heavy
metals from polluted waters using lignocellulosic agri-
cultural waste products. Ind Biotechnol 3:366–374
Harmsen J, Rulkens WH, Sims RC, Rijtema PE, Zweers AJ
(2007) Theory and application of landfarming to reme-
diate polycyclic aromatic hydrocarbons and mineral oil-
contaminated sediments; beneficial reuse. J Environ Qual
36:1112–1122
Hartlieb N, Ertunc T, Schaeffer A, Klein W (2003) Minerali-
zation, metabolism and formation of non-extractable
residues of 14C-labelled organic contaminants during
pilot-scale composting of municipal biowaste. Environ
Pollut 126:83–91
Hatti-Kaul R, Tornvall U, Gustafsson L, Borjesson P (2007)
Industrial biotechnology for the production of bio-based
chemicals—a cradle-to-grave perspective. Trends Bio-
technol 25:119–124
Hatzinger PB, Whittier MC, Arkins MD, Bryan CW, Guarini
WJ (2002) In-situ and ex-situ bioremediation options for
treating perchlorate in groundwater. Remed J 12:69–86
Heron G, Gierke JS, Faulkner B, Mravik S, Wood L, Enfield
CG (2002) Pulsed air sparging in aquifers contaminated
with dense nonaqueous phase liquids. Ground Water
Monit Remed 22:73–82
Hickman ZA, Reid BJ (2008) Earthworm assisted bioremedi-
ation of organic contaminants. Environ Int 34:1072–1081
Hirschorn SK, Grostern A, Lacrampe-Couloume G, Edwards
EA, MacKinnon L, Repta C, Major DW, Sherwood Lollar
B (2007) Quantification of biotransformation of chlori-
nated hydrocarbons in a biostimulation study: added value
via stable carbon isotope analysis. J Contam Hydrol
94:249–260
Hjeitzer A, Sayler G (1993) Monitoring the efficacy of biore-
mediation. Trends Biotechnol 11:334–343
Hofer R, Bigorra J (2007) Green chemistry—a sustainable
solution for industrial specialties applications. Green
Chem 9:203–212
Holden PA, Halverson LJ, Firestone MK (1997) Water stress
effects on toluene biodegradation by Pseudomonas putida.
Biodegradation 8:143–151
Hoyer PB (2001) Reproductive toxicology: current and future
directions. Biochem Pharmacol 62:1557–1564
Huang JW, Poynton CY, Kochian LV, Elless MP (2004)
Phytofiltration of arsenic from drinking water using
arsenic-hyperaccumulating ferns. Environ Sci Technol
38:3412–3417
Hultgren J, Pizzul L, del Pilar Castillo M, Granhall U (2010)
Degradation of PAH in a creosote-contaminated soil: a
comparison between the effects of willows (Salix Vimi-nalis), wheat straw and a nonionic surfactant. Int J Phy-
toremed 12:54–66
Husain Q, Husain M, Kulshrestha Y (2009) Remediation and
treatment of organopollutants mediated by peroxidases: a
review. Crit Rev Biotechnol 29:94–119
Hussain S, Siddique T, Arshad M, Saleem M (2009) Biore-
mediation and phytoremediation of pesticides: recent
advances. Crit Rev Environ Sci Technol 39:843–907
278 Rev Environ Sci Biotechnol (2010) 9:215–288
123
In B-H, Park J-S, Namkoong W, Hwang E-Y, Kim J-D (2008)
Effect of co-substrate on anaerobic slurry phase biore-
mediation of TNT-contaminated soil. Korean J Chem Eng
25:102–107
Iranzo M, Sainz-Pardo I, Boluda R, Sanchez J, Mormeneo S
(2001) The use of microorganisms in environmental
remediation. Ann Microbiol 51:135–143
Iwamoto T, Nasu M (2001) Current bioremediation practice
and perspective. J Biosci Bioeng 92:1–8
Jacques RJS, Okeke BC, Bento FM, Teixeira AS, Peralba
MCR, Camargo FAO (2008) Microbial consortium bio-
augmentation of a polycyclic aromatic hydrocarbons
contaminated soil. Bioresour Technol 99:2637–2643
Jain DK, Lee H, Trevors JT (1992) Effect of addition of
Pseudomonas aeruginosa UG2 inocula or biosurfactants
on biodegradation of selected hydrocarbons in soil. J Ind
Microbiol Biotechnol 10:87–93
Jain M, Garg VK, Kadirvelu K (2009) Chromium(VI) removal
from aqueous system using Helianthus annuus (sun-
flower) stem waste. J Hazard Mater 162:365–372
January MC, Cutright TJ, Van Keulen H, Wei R (2008)
Hydroponic phytoremediation of Cd, Cr, Ni, As, and Fe:
Can Helianthus annuus hyperaccumulate multiple heavy
metals? Chemosphere 70:531–537
Jayakumar K, Jaleel CA (2009) Uptake and accumulation of
cobalt in plants: a study based on exogenous cobalt in
soybean. Botany Res Int 2:310–314
Jerez CA (2009) Biomining microorganisms: molecular
aspects and applications in biotechnology and bioreme-
diation. In: Advances in applied bioremediation. Springer,
Berlin, pp 239–256
Jha KP, Nair S, Gopinathan MC, Babu CR (1995) Suitability of
rhizobia-inoculated wild legumes Argyolobium flaccidum,Astagalus graveolens, Indigo gangetica, and Lespedezastenocarpa in providing a vegetational cover in an unre-
claimed limestone quarry. Plant Soil 177:139–149
Johnsen AR, Wick LY, Harms H (2005) Principles of micro-
bial PAH-degradation in soil. Environ Pollut 133:71–84
Johnson RL, Johnson PC, McWhorter DB, Hinchee RE,
Goodman I (2007) An overview of in situ air sparging.
Ground Water Monit Remed 13:127–135
Joo H-S, Ndegwa PM, Shoda M, Phae C-G (2008) Bioreme-
diation of oil-contaminated soil using Candida catenulataand food waste. Environ Pollut 156:891–896
Jørgensen KS, Puustinen J, Suortti A–M (2000) Bioremedia-
tion of petroleum hydrocarbon-contaminated soil by
composting in biopiles. Environ Pollut 107:245–254
Juhasz AL, Naidu R (2000) Bioremediation of high molecular
weight polycyclic aromatic hydrocarbons: a review of the
microbial degradation of benzo[a]pyrene. Int Biodeterior
Biodegrad 45:57–88
Juwarkar AS, Juwarkar A, Pande VS, Bal AS (1992) Resto-
ration of manganese mine spoil dump productivity using
pressmud. In: Singhal RK, Mehrotra AK, Kostas, Jeanlue
Collines AA (eds) Environmental issues and management
of waste in energy and mineral production. Balekema
Roterdam, Brookfield
Juwarkar AA, Juwarkar AS, Mowade S, Jambhulkar H,
Shrivastava A, Kulkarni A, Amte P, Khanna P (1997)
Role of biofertilizers in reclamation of manganese mine
spoil dumps. Biofertil Newslett (July & December) 18–24
Juwarkar AA, Dubey K, Khobragade R, Nimje M (2000)
Phytoremediation of mine spoil dump using integrated
biotechnological approach. In Proceedings of interna-
tional symposium on geo environmental reclamation,
November 19–22. A. D., Nagpur, pp 425–429
Kamal M, Ghaly AE, Mahmoud N, Cote R (2004) Phytoac-
cumulation of heavy metals by aquatic plants. Environ Int
29:1029–1039
Kanel S, Nepal D, Manning B, Choi H (2007) Transport of
surface-modified iron nanoparticle in porous media and
application to arsenic(III) remediation. J Nanoparticle Res
9:725–735
Karpouzas DG, Singh BK (2010) Application of fingerprinting
molecular methods in bioremediation studies. Bioreme-
diation 599:69–88
Katsoyiannis A, Terzi E, Cai Q–Y (2007) On the use of PAH
molecular diagnostic ratios in sewage sludge for the
understanding of the PAH sources. Is this use appropriate?
Chemosphere 69:1337–1339
Kawahigashi H, Hirose S, Ohkawa H, Ohkawa Y (2006)
phytoremediation of the herbicides atrazine and metola-
chlor by transgenic rice plants expressing human
CYP1A1, CYP2B6 and CYP2C19. J Agric Food Chem
54:2985–2991
Keasling JD, Bang S-W (1998) Recombinant DNA techniques
for bioremediation and environmentally-friendly synthe-
sis. Curr Opin Biotechnol 9:135–140
Keener HM, Marugg C, Hansen RC, Hoitink H (1993) Opti-
mizing the efficiency of the compost process. Science and
Engineering of Composting. The Ohio State University,
Columbus, pp 59–94
Kelly BC, Ikonomou MG, Blair JD, Morin AE, Gobas FAPC
(2007) Food web-specific biomagnification of persistent
organic pollutants. Science 317:236–239
Khan AG (2005) Role of soil microbes in the rhizospheres of
plants growing on trace metal contaminated soils in
phytoremediation. J Trace Elements Med Biol 18:355–
364
Khan FI, Husain T, Hejazi R (2004) An overview and analysis of
site remediation technologies. J Environ Manag 71:95–122
Khanal SK (2008) Overview of anaerobic biotechnology.
Chapter 1, anaerobic biotechnology for bioenergy pro-
duction: principles and applications. Wiley and Black-
well, pp 1–27
Kidak R, Wilhelm A-M, Delmas H (2009) Effect of process
parameters on the energy requirement in ultrasonical
treatment of waste sludge. Chem Eng Process Process
Intensif 48:1346–1352
Kidwai M, Mohan R (2005) Green chemistry: an innovative
technology. Found Chem 7:269–287
Kim S, Dale BE (2004) Global potential bioethanol production
from wasted crops and crop residues. Biomass Bioenergy
26:361–375
Kim H-M, Hyun Y, Lee K-K (2007) Remediation of TCE-
contaminated groundwater in a sandy aquifer using pulsed
air sparging: laboratory and numerical studies. J Environ
Eng 133:380–388
Kim H, Annable MD, Rao PS, Cho J (2009) Laboratory evalu-
ation of surfactant-enhanced air sparging for perchloro-
ethene source mass depletion from sand. J Environ Sci
Health A, Toxicol Hazard Subst Environ Eng 44:406–413
Rev Environ Sci Biotechnol (2010) 9:215–288 279
123
King P, Anuradha K, Lahari SB, Kumar YP, Prasad VSRK
(2008) Biosorption of zinc from aqueous solution using
Azadirachta indica bark: Equilibrium and kinetic studies.
J Hazard Mater 152:324–329
Kirchhoff MM (2003) Promoting green engineering through
green chemistry. Environ Sci Technol 37:5349–5353
Kiyono M, Pan-Hou H (2006) Genetic engineering of bacteria
for environmental remediation of mercury. J Health Sci
52:199–204
Kocasoy G, Guvener Z (2009) Efficiency of compost in the
removal of heavy metals from the industrial wastewater.
Environ Geol 57:291–296
Koenigsberg SS, Hazen TC, Peacock AD (2005) Environ-
mental biotechnology: a bioremediation perspective. Re-
med J 15:5–25
Komnitsas K, Bartzas G, Paspaliaris I (2004) Efficiency of
limestone and red mud barriers: laboratory column stud-
ies. Miner Eng 17:183–194
Korade DL, Fulekar MH (2008) Remediation of anthracene in
mycorrhizospheric soil using ryegrass. Afr J Environ Sci
Technol 2:249–256
Kramer U, Smith RD, Wenzel WW, Raskin I, Salt DE (1997)
The role of metal transport and tolerance in nickel hy-
peraccumulation by Thlaspi goesingense Halacsy. Plant
Physiol 115:1641–1650
Krishna C (2005) Solid-state fermentation systems-an over-
view. Crit Rev Biotechnol 25:1–30
Krishnani KK, Shekhar MS, Gopikrishna G, Gupta BP (2009)
Molecular biological characterization and biostimulation
of ammonia—oxidizing bacteria in brackishwater aqua-
culture. J Environ Sci HealthPart A 44:1598–1608
Kularatne RKA, Kasturiarachchi JC, Manatunge JMA, Wij-
eyekoon SL (2009) Mechanisms of manganese removal
from wastewaters in constructed wetlands comprising
water hyacinth (Eichhornia crassipes (Mart.) Solms)
grown under different nutrient conditions. Water Environ
Res 81:165–172
Kulcu R, Yildiz O (2004) Determination of aeration rate and
kinetics of composting some agricultural wastes. Biore-
sour Technol 93:49–57
Kulkarni PS, Crespo JG, Afonso CAM (2008) Dioxins sources
and current remediation technologies—a review. Environ
Int 34:139–153
Kumar U, Bandyopadhyay M (2006) Sorption of cadmium
from aqueous solution using pretreated rice husk. Biore-
sour Technol 97:104–109
Kunamneni A, Plou FJ, Ballesteros A, Alcalde M (2008)
Laccases and their applications: a patent review. Recent
Patents Biotechnol 2:10–24
Kuntz J, Nassr–Amellal N, Lollier M, Schmidt JE, Lebeau T
(2008) Effect of sludges on bacteria in agricultural soil.
Analysis at laboratory and outdoor lysimeter scale. Eco-
toxicol Environ Saf 69:277–288
Kupper H, Lombi E, Zhao F-J, Wieshammer G, McGrath SP
(2001) Cellular compartmentation of nickel in the hy-
peraccumulators Alyssum lesbiacum, Alyssum bertoloniiand Thlaspi goesingense. J Exp Bot 52:2291–2300
Kwon KH, Yeom SH (2009) Optimal microbial adaptation
routes for the rapid degradation of high concentration of
phenol. Bioprocess Biosyst Eng 32:435–442
Lai KCK, Surampalli RY, Tyagi RD, Lo IMC, Yan S (2007)
Performance monitoring of remediation technologies for
soil and groundwater contamination: review. Pract Per-
iod Hazard Toxicol Radioactive Waste Manag 11:
132–157
Lai Y-L, Annadurai G, Huang F-C, Lee J-F (2008) Biosorption
of Zn(II) on the different Ca-alginate beads from aqueous
solution. Bioresour Technol 99:6480–6487
Lai C-C, Huang Y-C, Wei Y-H, Chang J-S (2009) Biosurfac-
tant-enhanced removal of total petroleum hydrocarbons
from contaminated soil. J Hazard Mater 167:609–614
Laine MM, Ahtiainen J, Wagman N, Oberg LG, Jørgensen KS
(1997) Fate and toxicity of chlorophenols, polychlorinated
dibenzo-p-dioxins, and dibenzofurans during composting
of contaminated sawmill soil. Environ Sci Technol 31:
3244–3250
Lankey RL, Anastas PT (2002) Life-cycle approaches for
assessing green chemistry technologies. Ind Eng Chem
Res 41:4498–4502
Lee G-T, Ro H-M, Lee S-M (2007) Effects of triethyl phos-
phate and nitrate on electrokinetically enhanced biodeg-
radation of diesel in low permeability soils. Environ
Technol 28:853–860
Lemire J, Mailloux R, Puiseux-Dao S, Appanna VD (2009)
Aluminum-induced defective mitochondrial metabolism
perturbs cytoskeletal dynamics in human astrocytoma
cells. J Neurosci Res 87:1474–1483
Li X, Feng Y, Sawatsky N (1997) Importance of soil-water
relations in assessing the endpoint of bioremediated soils.
Plant Soil 192:219–226
Li W, Zhang G, Zhang P, Liu H (2008) Waste activated sludge
reduction using sonication and cryptic growth. Int J Bio-
technol 10:64–72
Li X, Ma H, Wang Q, Matsumoto S, Maeda T, Ogawa HI
(2009) Isolation, identification of sludge-lysing strain and
its utilization in thermophilic aerobic digestion for waste
activated sludge. Bioresour Technol 100:2475–2481
Lima D, Viana P, Andre S, Chelinho S, Costa C, Ribeiro R,
Sousa JP, Fialho AM, Viegas CA (2009) Evaluating a
bioremediation tool for atrazine contaminated soils in open
soil microcosms: the effectiveness of bioaugmentation and
biostimulation approaches. Chemosphere 74:187–192
Lin Z-Q (2008) Ecological process: volatilization. In: Jorgen-
sen SE, Fath B (eds) Encyclopedia of ecology. Elsevier,
Oxford, pp 3700–3705
Litchfiled C (2005) Thirty years and counting: bioremediation
in its prime? Bioscience 55:273–279
Liu CW, Chang WN, Liu HS (2009) Bioremediation of n-
alkanes and the formation of biofloccules by Rhodococcuserythropolis NTU-1 under various saline conditions and
sea water. Biochem Eng J 45:69–75
Lodha B, Bhadane R, Patel B, Killedar D (2008) Biodegra-
dation of pyridine by an isolated bacterial consortium/
strain and bio–augmentation of strain into activated
sludge to enhance pyridine biodegradation. Biodegrada-
tion 19:717–723
Lombi E, Zhao F-J, Fuhrmann M, Ma LQ, McGrath SP
(2002) Arsenic distribution and speciation in the fronds
of the hyperaccumulator Pteris vittata. New Phytol 156:
195–203
280 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Lopez Torres M, Espinosa Llorens MC (2008) Effect of
alkaline pretreatment on anaerobic digestion of solid
wastes. Waste Manag 28:2229–2234
Lopez-Nieto MJ, Costa J, Peiro E et al (2004) Biotechnological
lycopene production by mated fermentation of Blakesleatrispora. Appl Microbiol Biotechnol 66:153–159
Loukidou MX, Zouboulis AI, Karapantsios TD, Matis KA
(2004) Equilibrium and kinetic modeling of chro-
mium(VI) biosorption by Aeromonas caviae. Colloids
Surf A Physicochem Eng Aspects 242:93–104
Lu S, Gibb SW (2008) Copper removal from wastewater
using spent-grain as biosorbent. Bioresour Technol 99:
1509–1517
Lu J, Gavala HN, Skiadas IV, Mladenovska Z, Ahring BK
(2008) Improving anaerobic sewage sludge digestion by
implementation of a hyper-thermophilic prehydrolysis
step. J Environ Manag 88:881–889
Ludmer Z, Golan T, Ermolenko E, Brauner N, Ullmann A
(2009) Simultaneous removal of heavy metals and organic
pollutants from contaminated sediments and sludges by a
novel technology, sediments remediation phase transition
extraction. Environ Eng Sci 26:419–430
Lynch JM, Moffat AJ (2005) Bioremediation-prospects for the
future application of innovative applied biological
research. Ann Appl Biol 146:217–221
Ma LQ, Komar KM, Tu C, Zhang W, Cai Y, Kennelley ED
(2001) A fern that hyperaccumulates arsenic. Nature
(London) 409:579
Mace C, Desrocher S, Gheorghiu F, Kane A, Pupeza M et al
(2006) Nanotechnology and groundwater remediation: a
step forward in technology understanding. Remed J
16:23–33
Machado RM, Correia MJN, Carvalho JMR (2003) Integrated
process for biosorption of copper from liquid effluents
using grape stalks. Sep Sci Technol 38:2237–2254
Maciel BM, Santos ACF, Dias JCT, Vidal RO, Dias RJC,
Gross E, Cascardo JCM, Rezende RP (2009) Simple DNA
extraction protocol for a 16S rDNA study of bacterial
diversity in tropical landfarm soil used for bioremediation
of oil waste. Genet Mol Res 8:375–388
Makris KC, Shakya KM, Datta R, Sarkar D, Pachanoor D
(2007) High uptake of 2, 4, 6-trinitrotoluene by vetiver
grass—potential for phytoremediation? Environ Pollut
146:1–4
Malaisse F, Gregoire J, Brooks RR, Morrison RS, Reeves RD
(1997) Aeolanthus biformifolius De Wild.: a hyperaccu-
mulator of copper from Zaire. Science 199:887–888
Malkoc E (2006) Ni(II) removal from aqueous solutions using
cone biomass of Thuja orientalis. J Hazard Mater 137:
899–908
Mao T, Show K-Y (2007) Influence of ultrasonication on
anaerobic bioconversion of sludge. Water Environ Res
79:436–441
Mao T, Hong S-Y, Show K-Y, Tay J-H, Lee D-J (2004) A
comparison of ultrasound treatment on primary and sec-
ondary sludges. Water Sci Technol 50:91–97
Maranon E, Sastre H (1991) Heavy metal removal in packed
beds using apple wastes. Bioresour Technol 38:39–43
Margesin R, Hammerle M, Tscherko D (2007) Microbial
activity and community composition during bioremedia-
tion of diesel-Oil-contaminated soil: Effects of
hydrocarbon concentration, fertilizers, and incubation.
Microb Ecol 53:259–269
Marın JA, Hernandez T, Garcia C (2005) Bioremediation of oil
refinery sludge by landfarming in semiarid conditions:
influence on soil microbial activity. Environ Res 98:
185–195
Marın JA, Moreno JL, Hernandez T, Garcıa C (2006) Biore-
mediation by composting of heavy oil refinery sludge in
semiarid conditions. Biodegradation 17:251–261
Marsolek MD, Kirisits MJ, Rittmann BE (2007) Biodegrada-
tion of 2, 4, 5-trichlorophenol by aerobic microbial
communities: biorecalcitrance, inhibition, and adaptation.
Biodegradation 18:351–358
Marttinen SK, Kettunen RH, Sormunen KM, Rintala JA (2003)
Removal of bis(2-ethylhexyl) phthalate at a sewage
treatment plant. Water Res 37:1385–1393
Marttinen SK, Hanninen K, Rintala JA (2004) Removal of
DEHP in composting and aeration of sewage sludge.
Chemosphere 54:265–272
Maxted AP, Black CR, West HM, Crout NMJ, McGrath SP,
Young SD (2007) Phytoextraction of cadmium and zinc
from arable soils amended with sewage sludge using
Thlaspi caerulescens: Development of a predictive model.Environ Pollut 150:363–372
McMahon V, Garg A, Aldred D, Hobbs G, Smith R, Tothill IE
(2008) Composting and bioremediation process evalua-
tion of wood waste materials generated from the con-
struction and demolition industry. Chemosphere 71:1617–
1628
Meharg AA (2003) Variation in arsenic accumulation–hyper-
accumulation in ferns and their allies. New Phytol
157:25–31
Melin ES, Jarvinen KT, Puhakka JA (1998) Effects of tem-
perature on chlorophenol biodegradation kinetics in flu-
idized-bed reactors with different biomass carriers. Water
Res 32:81–90
Melo JS, D’Souza SF (2004) Removal of chromium by
mucilaginous seeds of Ocimum basilicum. Bioresour
Technol 92:151–155
Michel FC Jr, Reddy CA, Forney LJ (1995) Microbial degra-
dation and humification of the lawn care pesticide 2, 4-
dichlorophenoxyacetic acid during the composting of yard
trimmings. Appl Environ Microbiol 61:2566–2571
Michel FC Jr, Quensen J, Reddy CA (2001) Bioremediation of
a PCB—contaminated soil via composting. Compost Sci
Util 9:274–283
Mihial DJ, Viraraghavan T, Jin Y-C (2006) Bioremediation of
petroleum–contaminated soil using composting. Pract
Period Hazard Toxicol Radioactive Waste Manag 10:
108–115
Min Y, Boqing T, Meizhen T, Aoyama I (2007) Accumulation
and uptake of manganese in a hyperaccumulator Phy-tolacca Americana. Miner Eng 20:188–190
Mirza N, Mahmood Q, Pervez A, Ahmad R, Farooq R, Shah
MM, Azim MR (2010) Phytoremediation potential of
Arundo donax in arsenic-contaminated synthetic waste-
water. Biores Technol 101:5815–5819
Mohamed AMI, El-Menshawy N, Saif AM (2007) Remedia-
tion of saturated soil contaminated with petroleum prod-
ucts using air sparging with thermal enhancement.
J Environ Manag 83:339–350
Rev Environ Sci Biotechnol (2010) 9:215–288 281
123
Mohee R, Mudhoo A, Unmar GD (2008) Windrow co-com-
posting of shredded office paper and broiler litter. Special
issue on solid waste management—Part 1. Int J Environ
Waste Manag 2:3–23
Møller J, Winther P, Lund L, Kirkebjerg K, Westermann P
(1996) Bioventing of diesel oil-contaminated soil: com-
parison of degradation rates in soil based on actual oil
concentration and on respirometric data. J Ind Microbiol
Biotechnol 16:110–116
Mudhoo A, Mohee R (2008) Modeling heat loss during self–
heating composting based on combined fluid film theory
and boundary layer concepts. J Environ Inf 11:74–89
Mulchandani A, Luong JHT, Groom C (1989) Substrate inhi-
bition kinetics for microbial growth and synthesis of poly-
b-hydroxybutyric acid by Alcaligenes eutrophus ATCC
17697. Appl Microbiol Biotechnol 30:11–17
Mulligan CN (2009) Recent advances in the environmental
applications of biosurfactants. Curr Opin Colloid Interface
Sci 14:372–378
Murakami M, Ae N (2009) Potential for phytoextraction of
copper, lead, and zinc by rice (Oryza sativa L.), soybean
(Glycine max [L.] Merr.), and maize (Zea mays L.).
J Hazard Mater 162:1185–1192
N’Guessan AL, Elifantz H, Nevin KP, Mouser PJ, Methe B,
Woodard TL, Manley K, Williams KH, Wilkins MJ,
Larsen JT, Long PE, Lovley DR (2010) Molecular anal-
ysis of phosphate limitation in Geobacteraceae during the
bioremediation of a uranium-contaminated aquifer. ISME
J 4:253–266
Naddeo V, Belgiorno V, Landi M, Zarra M, Napoli RMA
(2009) Effect of sonolysis on waste activated sludge sol-
ubilisation and anaerobic biodegradability. Desalination
249:762–767
Nagata T, Nakamura A, Akizawa T, Pan-Hou H (2009)
Genetic engineering of transgenic tobacco for enhanced
uptake and bioaccumulation of mercury. Biol Pharm Bull
32:1491–1495
Naja G, Volesky B (2006) Behavior of the mass transfer zone in
a biosorption column. Environ Sci Technol 40:3996–4003
Namkoong W, Hwang E-Y, Park J-S, Choi J-Y (2002) Biore-
mediation of diesel-contaminated soil with composting.
Environ Pollut 119:23–31
Narihiro T, Sekiguchi Y (2007) Microbial communities in
anaerobic digestion processes for waste and wastewater
treatment: a microbiological update. Curr Opin Biotech-
nol 18:273–278
Negro MJ, Solano PC, Carasco J (1999) Composting of sweet
sorghum bagasse with other wastes. Bioresour Technol
67:89–92
Neyens E, Baeyens J (2003) A review of thermal sludge pre-
treatment processes to improve dewaterability. J Hazard
Mater 98:51–67
Nigam P, Robinson T, Singh D (2004) Solid-state fermenta-
tion: an overview. In: Arora D (ed) Handbook of fungal
biotechnology, mycology, vol 20. CRC Press, London
Nikolopoulos AN, Igglessi-Markopoulou O, Papayannakos N
(2006) Ultrasound assisted catalytic wet peroxide oxida-
tion of phenol: kinetics and intraparticle diffusion effects.
Ultrason Sonochem 13:92–97
Noeline BF, Manohar DM, Anirudhan TS (2005) Kinetic and
equilibrium modelling of lead(II) sorption from water and
wastewater by polymerized banana stem in a batch reac-
tor. Sep Purif Technol 45:131–140
Nurzhanova A, Kulakow P, Rubin E et al (2010) Obsolete
pesticides pollution and phytoremediation of contami-
nated soil in Kazakhstan. In: Application of phytotech-
nologies for cleanup of industrial, agricultural, and
wastewater contamination. Springer, The Netherlands,
pp 87–111
Okpokwasili GC, Nweke CO (2006) Microbial growth and
substrate utilization kinetics. Afr J Biotechnol 5:305–317
Olette R, Couderchet M, Biagianti S, Eullaffroy P (2008)
Toxicity and removal of pesticides by selected aquatic
plants. Chemosphere 70:1414–1421
Osman KA, Al-Rehiayani SM, Al-Deghairi MA, Salama AK
(2009) Bioremediation of oxamyl in sandy soil using ani-
mal manures. Int Biodeterior Biodegradation 63:341–346
Ozkoc HB, Bakan G, Ariman S (2007) Distribution and bio-
accumulation of organochlorine pesticides along the
Black Sea coast. Environ Geochem Health 29:59–68
Palmroth MRT, Pichtel J, Puhakka JA (2002) Phytoremedia-
tion of subarctic soil contaminated with diesel fuel.
Bioresour Technol 84:221–228
Pandey A, Soccol CR, Nigam P et al (2000a) Biotechnological
potential of agro-industrial residues: II—Cassava bagasse.
Bioresour Technol 74:81–87
Pandey A, Soccol CR, Mitchell D (2000b) New developments
in solid state fermentation: I—bioprocesses and products.
Process Biochem 35:1153–1169
Pence NS, Larsen PB, Ebbs SD, Letham DLD, Lasat MM,
Garvin DF, Eide D, Kochian LV (2000) The molecular
physiology of heavy metal transport in the Zn/Cd hyper-
accumulator Thlaspi caerulescens. Proc Natl Acad Sci U
S A 97:4956–4960
Perez SR, Garcıa ON, Bermudez RC et al (2008) Decolouri-
sation of mushroom farm wastewater by Pleurotus os-treatus. Biodegradation 19:519–526
Perez-Marın AB, Meseguer ZV, Ortuno JF, Aguilar M, Saez J,
Llorens M (2007) Removal of cadmium from aqueous
solutions by adsorption onto orange waste. J Hazard
Mater 139:122–131
Periasamy K, Namasivayam C (1995) Removal of nickel(II)
from aqueous solution and plating industry wastewater
using an agriculture waste peanut hulls. Waste Manag
15:63–68
Pieper DH, Reineke W (2000) Engineering bacteria for bio-
remediation. Curr Opin Biotechnol 11:262–270
Pignatello JJ (2009) Bioavailability of contaminants in soil. In:
Advances in applied bioremediation, vol 17. Springer,
Berlin, pp 35–71
Plangklang P, Reungsang A (2010) Bioaugmentation of car-
bofuran by Burkholderia cepacia PCL3 in a bioslurry
phase sequencing batch reactor. Process Biochem 45:
230–238
Plaza C, Xing B, Fernandez JM, Senesi N, Polo A (2009)
Binding of polycyclic aromatic hydrocarbons by humic.
Biodegradation 21:345–356
Polomski RF, Bielenberg DG, Whitwell T, Taylor MD,
Bridges WC, Klaine SJ (2008) Differential nitrogen and
phosphorus recovery by five aquatic garden species in
laboratory-scale subsurface-constructed wetlands. Hort
Sci 43:868–874
282 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Powell JJ, Jugdaohsingh R, Thompson RPH (1999) The reg-
ulation of mineral absorption in the gastrointestinal tract.
Proc Nutr Soc 58:147–153
Prasad MNV, Freitas H, Fraenzle S, Wuenschmann S,
Markert B (2010) Knowledge explosion in phytotech-
nologies for environmental solutions. Environ Pollut
158:18–23
Prasanna D, Venkata Mohan S, Purushotham Reddy B, Sarma
PN (2008) Bioremediation of anthracene contaminated
soil in bio-slurry phase reactor operated in periodic dis-
continuous batch mode. J Hazard Mater 153:244–251
Preetha B, Viruthagiri T (2007) Batch and continuous bio-
sorption of chromium(VI) by Rhizopus arrhizus. Sep Purif
Technol 57:126–133
Price ND, Reed JL, Palsson BØ (2004) Genome-scale models
of microbial cells: evaluating the consequences of con-
straints. Nat Rev 2:886–897
Prince RC (2010) Can we improve bioremediation? Handbook
of hydrocarbon and lipid microbiology, Part 30. Springer,
Berlin, pp 3351–3355
Qiu R-L, Liu F-J, Wan Y-B et al (2008) Phytoremediation on
nickel–contaminated soils by hyperaccumulators Alyssumcorsicum and Alyssum murale. China Environ Sci
28:1026–1031
Rahman M, Hasan MR, Oba T, Shimizu K (2006) Effect of
rpoS gene knockout on the metabolism of Escherichiacoli during exponential growth phase and early stationary
phase based on gene expressions, enzyme activities and
intracellular metabolite concentrations. Biotechnol Bio-
eng 94:585–595
Raj A, Krishna Reddy MM, Chandra R, Purohit MJ, Kapley A
(2007) Biodegradation of kraft-lignin by Bacillus sp.
isolated from sludge of pulp and paper mill. Biodegra-
dation 18:783–792
Rajendran P, Gunasekaran P (2007) Nanotechnology for bio-
remediation of heavy metals. Environmental bioremedia-
tion technologies. Springer, Berlin, pp 211–221
Rajeshwari KV, Balakrishnan M, Kansal A, Lata K, Kishore
VVN (2000) State-of-the0art of anaerobic digestion
technology for industrial wastewater treatment. Renew
Sustain Energy Rev 4:135–156
Rama Krishna M, Shailaja S, Sirisha K, Venkata Mohan S,
Sarma PN (2006) Bio-remediation of pendimethalin
contaminated soil by bio-slurry phase reactor: bio-aug-
menting with ETP micro-flora. Int J Environ Poll
27:373–387
Ramos JL, Krell T, Daniels G, Segura A, Duque E (2009)
Responses of Pseudomonas to small toxic molecules by a
mosaic of domains. Curr Opin Microbiol 12:215–220
Ran N, Zhao L, Chen Z, Tao J (2008) Recent applications of
biocatalysis in developing green chemistry for chemical
synthesis at the industrial scale. Green Chem 10:
361–372
Rao TK, Murthy YLN (2007) Role of nano-science and tech-
nology for environmental protection. Nat Environ Pollut
Technol 6:665–672
Raskin I (1996) Plant genetic engineering may help with
environmental cleanup. Proc Natl Acad Sci U S A
93:3164–3166
Reid JB, Botwright NA, Smith JJ, O’Neill DP, Kerckhoffs
LHJ (2002) Control of gibberellin levels and gene
expression during de-etiolation in pea. Plant Physiol
128:734–741
Ren N, Li N, Li B, Wang Y, Liu S (2006) Biohydrogen pro-
duction from molasses by anaerobic fermentation with a
pilot-scale bioreactor system. Int J Hydrogen Energy
31:2147–2157
Revankar MS, Desai KM, Lele SS (2007) Solid-state fermen-
tation for enhanced production of Laccase using indige-
nously isolated Ganoderma sp. Appl Biochem Biotechnol
143:16–26
Rhoads A, Beyenal H, Lewandowski Z (2005) Microbial fuel
cell using anaerobic respiration as an anodic reaction and
biomineralized manganese as a cathodic reactant. Environ
Sci Technol 39:4666–4671
Riaz M, Nadeem R, Hanif MA, Ansari TM, Khalil-ur-Rehman
(2009) Pb(II) biosorption from hazardous aqueous streams
using Gossypium hirsutum (Cotton) waste biomass.
J Hazard Mater 161:88–94
Richard TL, Walker PL, Gossett JM (2006) Effects of oxygen
on aerobic solid-state biodegradation kinetics. Biotechnol
Prog 22:60–69
Rigas F, Papadopoulou K, Dritsa V et al (2007) Bioremediation
of a soil contaminated by lindane utilizing the fungus
Ganoderma australe via response surface methodology.
J Hazard Mater 140:325–332
Ritalahti KM, Loffler FE, Rasch EE, Koenigsberg SS (2005)
Bioaugmentation for chlorinated ethene detoxification:
bioaugmentation and molecular diagnostics in the biore-
mediation of chlorinated ethene-contaminated sites. Ind
Biotechnol 1:114–118
Robinson BH, Brooks RR, Howes AW, Kirkman JH, Gregg
PEH (1997) The potential of the high-biomass nickel
hyperaccumulator Berkheya coddii for phytoremediation
and phytomining. J Geocheml Exploration 60:115–126
Robinson C, Barry DA, McCarty PL, Gerhard JI, Kouznetsova
I (2009) pH control for enhanced reductive bioremedia-
tion of chlorinated solvent source zones. Sci Total Envi-
ron 407:4560–4573
Rodrıguez-Rodrıguez CE, Marco-Urrea E, Caminal G (2010)
Degradation of naproxen and carbamazepine in spiked
sludge by slurry and solid-phase Trametes versicolorsystems. Bioresour Technol 101:2259–2266
Romantschuk M, Sarand I, Petanen T, Peltola R, Jonsson-
Vihanne M, Koivula T, Yrjala K, Haahtela K (2000)
Means to improve the effect of in situ bioremediation of
contaminated soil: an overview of novel approaches.
Environ Pollut 107:179–185
Rubinos DA, Villasuso R, Muniategui S, Barral MT, Dıaz-
Fierros F (2007) Using the landfarming technique to
remediate soils contaminated with hexachlorocyclohexane
isomers. Water Air Soil Pollut 181:385–399
Sabean JAR, Scott DB, Lee K, Venosa AD (2009) Monitoring
oil spill bioremediation using marsh foraminifera as
indicators. Mar Pollut Bull 59:352–361
Saeed A, Iqbal M (2003) Bioremoval of cadmium from
aqueous solution by black gram husk (Cicer arientinum).Water Res 37:3472–3480
Sagner S, Kneer R, Wanner G, Cosson JP, Deus-Neumann B,
Zenk MH (1998) Hyperaccumulation, complexation and
distribution of nickel in Sebertia acuminate. Phytochem-
istry 47:339–347
Rev Environ Sci Biotechnol (2010) 9:215–288 283
123
Salanitro JP, Dorn PB, Huesemann MH, Moore KO, Rhodes
IA, Jackson LMR, Vipond TE, Western MM, Wisniewski
HL (1997) Crude oil hydrocarbon bioremediation and soil
ecotoxicity assessment. Environ Sci Technol 31:1769–
1776
Salido AL, Hasty KL, Lim J–M, Butcher DJ (2003) Phyto-
remediation of arsenic and lead in contaminated soil using
chinese brake ferns (Pteris vittata) and Indian mustard
(Brassica juncea). Int J Phytoremed 5:89–103
Salinas-Martınez A, de los Santos-Cordova M, Soto-Cruz O,
Delgado E, Perez-Andrade H, Hauad-Marroquın LA,
Medrano-Roldan H (2008) Development of a bioremedi-
ation process by biostimulation of native microbial con-
sortium through the heap leaching technique. J Environ
Manag 88:115–119
Salt DE, Blaylock M, Kumar N, Dushenkov V, Ensley B, Chet
I, Raskin I (1995) Phytoremediation: a novel strategy for
the removal of toxic metals from the environment using
plants. Biotechnology 13:468–474
Samanta SK, Singh OV, Jain RK (2002) Polycyclic aromatic
hydrocarbons: environmental pollution and bioremedia-
tion. Trends Biotechnol 20:243–248
Sanchez A, Ysunza F, Beltran-Garcıa MJ (2002) Biodegrada-
tion of viticulture wastes by pleurotus: a source of
microbial and human food and its potential use in animal
feeding. J Agric Food Chem 50:2537–2542
Sani RK, Peyton BM, Brown LT (2001) Copper-induced
inhibition of growth of Desulfovibrio desulfuricans G20:
assessment of its toxicity and correlation with those of
zinc and lead. Appl Environ Microbiol 67:4765–4772
Sanscartier D, Laing T, Reimer K, Zeeb B (2009) Bioreme-
diation of weathered petroleum hydrocarbon soil con-
tamination in the Canadian High Arctic: laboratory and
field studies. Chemosphere 77:1121–1126
Sanscartier D, Reimer K, Zeeb B, George K (2010) Manage-
ment of hydrocarbon—contaminated soil through biore-
mediation and landfill disposal at a remote location in
Northern Canada. Can J Civil Eng 37:147–155
Sanseverino J, Applegate BM, Henry King JM, Sayler GS
(1993) Plasmid-mediated mineralization of napthalene,
phenanthrene and anthracene. App and Environ Micro-
biol 59(6):1931–1937
Saravanan P, Pakshirajan K, Saha P (2008) Growth kinetics of
an indigenous mixed microbial consortium during phenol
degradation in a batch reactor. Bioresour Technol 99:
205–209
Sarı A, Tuzen M (2008) Biosorption of cadmium(II) from
aqueous solution by red algae (Ceramium virgatum):
equilibrium, kinetic and thermodynamic studies. J Hazard
Mater 157:448–454
Sarret G, Saumitou-Laprade P, Bert V, Proux O, Hazemann JL,
Traverse A, Marcus MA, Manceau A (2002) Forms of
zinc accumulated in the hyperaccumulator Arabidopsishalleri. Plant Physiol 130:1815–1826
Savant DV, Abdul–Rahman R, Ranade DR (2006) Anaerobic
degradation of adsorbable organic halides (AOX) from
pulp and paper industry wastewater. Bioresour Technol
97:1092–1104
Sayler GS, Layton AC (1990) Environmental application
of nucleic acid hybridization. Annu Rev Microbiol 44:
625–648
Sayler GS, Ripp S (2000) Field applications of genetically
engineered microorganisms for bioremediation processes.
Curr Opin Biotechnol 11:286–289
Schnoor JL, Licht LA, Mc Cutcheon SC, Wolf NL, Carreira
LH (1995) Phytoremediation of organic and nutrient
contaminants. Environ Sci Technol 29:317–323
Seki H (2000) Stochastic modeling of composting process with
batch operation by the Fokker–Planck equation. Trans
ASAE 43:169–179
Sen R, Chakrabarti S (2009) Biotechnology-applications to
environmental remediation in resource exploitation. Curr
Sci 97:768–775
Senthilkumar M, Arutchelvan V, Kanakasabai V, Venkatesh
KR, Nagarajan S (2009) Biomineralisation of dye waste in
a two-phase hybrid UASB reactor using starch effluent as
a co-substrate. Int J Environ Waste Manag 3:354–365
Sgarbi G, Casalena GA, Baracca A, Lenaz G, DiMauro S,
Solaini G (2009) Human NARP mitochondrial mutation
metabolism corrected with a-ketoglutarate/aspartate a
potential new therapy. Arch Neurol 66:951–957
Shah MM, Aust SD (1993) Degradation of cyanides by the
white rot fungus phanerochaete chrysosporium. In:
Emerging technologies for hazardous waste management
III, Chapter 10, pp 191–202
Sharma CM, Rosseland BO, Almvik M, Eklo OM (2009)
Bioaccumulation of organochlorine pollutants in the fish
community in Lake Arungen, Norway. Environ Pollut157:2452–2458
Shen ZG, Zhao FJ, Mcgrath SP (1997) Uptake and transport of
zinc in the hyperaccumulator Thlaspi caerulescens and the
non-hyperaccumulator Thlaspi ochroleucum. Plant Cell
Environ 20:898–906
Shewfelt K, Lee H, Zytner RG (2005) Optimization of nitrogen
for bioventing of gasoline contaminated soil. J Environ
Eng Sci 4:29–42
Shida GM, Barros AR, Marques dos Reis C, Cavalcante de Am-
orim EL, Damianovic MHRZ, Silva EL (2009) Long-term
stability of hydrogen and organic acids production in an
anaerobic fluidized-bed reactor using heat treated anaerobic
sludge inoculum. Int J Hydrogen Energy 34:3679–3688
Shield MS, Montgomery SO, Cuskey SM, Chapman PJ, Prit-
chard PH (1989) Novel pathway of toluene catabolism in
the trichloroethylene—degrading bacterium 04. Appl
Environ Microbiol 55:1624–1629
Show K–Y, Mao T, Lee D–J (2007) Optimisation of sludge
disruption by sonication. Water Res 41:4741–4747
Sidoli O’Connor C, Lepp NW, Edwards R, Sunderland G
(2003) The combined use of electrokinetic remediation
and phytoremediation to decontaminate metal-polluted
soils: A laboratory-scale feasibility study. Environ Monit
Assess 84:141–158
Sikdar SK, Grosse D, Rogut I (1998) Membrane technologies
for remediating contaminated soils: a critical review.
J Membr Sci 151:75–85
Simon MJ, Osslund TD, Saunders R, Esley B, Suggs S, Har-
court A, Suen WC, Cruden DL, Gibson DT, Zylstra GJ
(1993) Sequences of genes encoding naphthalene dioxy-
genase in Pseudomonasputida strains G7 and NCIB 9816-
4. Gene 127:31–37
Singh R, Paul D, Jain RK (2006a) Biofilms: implications in
bioremediation. Trends Microbiol 14:389–397
284 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Singh KK, Singh AK, Hasan SH (2006b) Low cost bio-sorbent
‘wheat bran’ for the removal of cadmium from waste-
water: kinetic and equilibrium studies. Bioresour Technol
97:994–1001
Singh A, Van Hamme JD, Ward OP (2007) Surfactants in
microbiology and biotechnology: part 2. Application
aspects. Biotechnol Adv 25:99–121
Singh S, Kang SH, Mulchandani A, Chen W (2008a) Biore-
mediation: environmental clean–up through pathway
engineering. Curr Opin Biotechnol 19:437–444
Singh S, Melo JS, Eapen S, D’Souza SF (2008b) Potential of
vetiver (Vetiveria zizanoides L. Nash) for phytoremedia-
tion of phenol. Ecotoxicol Environ Saf 71:671–676
Singh J, Kaur A, Vig AP, Rup PJ (2010) Role of Eisenia fetidain rapid recycling of nutrients from bio sludge of beverage
industry. Ecotoxicol Environ Saf 73:430–435
Sinha RK, Bharambe G, Ryan D (2008) Converting wasteland
into wonderland by earthworms-a low-cost nature’s
technology for soil remediation: a case study of vermi-
remediation of PAHs contaminated soil. Environmentalist
28:466–475
Skinner K, Wright N, Porter–Goff E (2007) Mercury uptake
and accumulation by four species of aquatic plants.
Environ Pollut 145:234–237
Skoog A, Vlahos P, Rogers KL, Amend JP (2007) Concen-
trations, distributions, and energy yields of dissolved
neutral aldoses in a shallow hydrothermal vent system of
Vulcano, Italy. Org Geochem 38:1416–1430
Sood N, Patle S, Lal B (2010) Bioremediation of acidic oily
sludge–contaminated soil by the novel yeast strain Can-dida digboiensis TERI ASN6. Environ Sci Pollut Res
17:603–610
Souza TS, Hencklein FA, Angelis DF, Goncalves RA, Fonta-
netti CS (2009) The Allium cepa bioassay to evaluate
landfarming soil, before and after the addition of rice hulls
to accelerate organic pollutants biodegradation. Ecotoxi-
col Environ Saf 72:1363–1368
Srivastava M, Ma LQ, Santos JAG (2006) Three new arsenic
hyperaccumulating ferns. Sci Total Environ 364:24–31
Stallwood B, Shears J, Williams PA, Hughes KA (2005) Low
temperature bioremediation of oil-contaminated soil using
biostimulation and bioaugmentation with a Pseudomonas sp.
from maritime Antarctica. J Appl Microbiol 99:794–802
Stenuit B, Eyers L, Schuler L, George I, Agathos SN (2009)
Molecular tools for monitoring and validating bioreme-
diation. In: Advances in applied bioremediation, vol 17.
Springer, Berlin, pp 339–353
Stomp A-M, Han K-H, Wilbert S, Gordon MP (1993) Genetic
improvement of tree species for remediation of hazardous
wastes. In Vitro Cell Dev Biol Plant 29:227–232
Sud D, Mahajan G, Kaur MP (2008) Agricultural waste
material as potential adsorbent for sequestering heavy
metal ions from aqueous solutions—a review. Bioresour
Technol 99:6017–6027
Sui H, Li X, Huang G, Jiang B (2006) A study on cometabolic
bioventing for the in situ remediation of trichloroethylene.
Environ Geochem Health 28:147–152
Sun Y, Ji L, Wang W, Wang X, Wu J, Li H, Guo H (2009)
Simultaneous removal of polycyclic aromatic hydrocar-
bons and copper from soils using ethyl lactate—amended
EDDS solution. J Environ Qual 38:1591–1597
Suthar S, Singh S (2008) Feasibility of vermicomposting in
biostabilization of sludge from a distillery industry. Sci
Total Environ 394:237–243
Takeuchi I, Miyoshi N, Mizukawa K, Takada H, Ikemoto T,
Omori K, Tsuchiya K (2009) Biomagnification profiles of
polycyclic aromatic hydrocarbons, alkylphenols and
polychlorinated biphenyls in Tokyo Bay elucidated by
d13C and d15N isotope ratios as guides to trophic web
structure. Mar Pollut Bull 58:663–671
Talley WF, Sleeper PM (2006) Roadblocks to the imple-
mentation of biotreatment strategies. Ann NY Acad Sci
16–29
Tang SY, Bourne RA, Smith RL, Poliakoff M (2008) The 24
principles of green engineering and green chemistry:
improvements productively. Green Chem 10:268–269
Tatsuzawa T, Hao L, Ayame S, Shimomura T, Kataoka N, Miya
A (2006) Population dynamics of anaerobic microbial
consortia in thermophilic methanogenic sludge treating
paper-containing solid waste. Water Sci Technol 54:
113–119
Telling ND, Coker VS, Cutting RS, van der Laan G, Pearce CI,
Pattrick RAD, Arenholz E, Lloyd JR (2009) Remediation
of Cr(VI) by biogenic magnetic nanoparticles: an X-ray
magnetic circular dichroism study. Appl Phys Lett
95:163701–163703
Teng Y, Luo Y-M, Huang C-Y, Long J, Li Z-G, Christie P
(2008) Tolerance of grasses to heavy metals and microbial
functional diversity in soils contaminated with copper
mine tailings. Pedosphere 18:363–370
Teng Y, Luo Y, Sun M, Liu Z, Li Z, Christie P (2010) Effect of
bioaugmentation by Paracoccus sp. strain HPD-2 on the
soil microbial community and removal of polycyclic
aromatic hydrocarbons from an aged contaminated soil.
Bioresour Technol 101:3437–3443
Tharakan J, Addagada A, Tomlinson D, Shafagati A (2004)
Vermicomposting for the bioremediation of PCB cong-
eners in SUPERFUND site media. In: Waste management
and the environment II: international conference on waste
management and the environment No. 2, Rhodes,
pp 117–124
Thavasi R, Jayalakshmi S, Balasubramanian T et al (2008)
Production and characterization of aglycolipid biosurfac-
tant from Bacillus megaterium using economically cheaper
sources. World J Microbiol Biotechnol 24:917–925
Theron J, Walker JA, Cloete TE (2008) Nanotechnology and
water treatment: applications and emerging opportunities.
Crit Rev Microbiol 34:43–69
Thornton EC, Gilmore TJ, Olsen KB, Giblin JT, Phelan JM
(2007) Treatment of a chromate-contaminated soil site by
in situ gaseous reduction. Ground Water Monit Remed
27:56–64
Tiehm A, Stieber M, Werner P, Frimmel FH (1997) Surfactant-
enhanced mobilization and biodegradation of polycyclic
aromatic hydrocarbons in manufactured gas plant soil.
Environ Sci Technol 31:2570–2576
Tognetti C, Laos F, Mazzarino MJ, Hernandez MT (2005)
Composting vs. Vermicomposting: a comparison of end
product quality. Compost Sci Util 13:6–13
Tongbin C, Chaoyang W, Zechun H, Qifei H, Quanguo L,
Zilian F (2002) Arsenic hyperaccumulator Pteris VittataL. and its arsenic accumulation. Chin Sci Bull 47:902–905
Rev Environ Sci Biotechnol (2010) 9:215–288 285
123
Tovanabootr A, Semprini L, Dolan ME, Azizian M, Magar VS,
Debacker D, Leeson A, Kempisty CD (2001) Cometabolic
air sparging field demonstration with propane to remediate
trichloroethene and cis-dichloroethene. In: 6th Interna-
tional in situ and on site bioremediation symposium, San
Diego, pp 145–153
Trasar-Cepeda C, Gil-Sotres F, Leiros MC (2007) Thermody-
namic parameters of enzymes in grassland soils from
Galicia, NW Spain. Soil Biol Biochem 39:311–319
Tratnyek PG, Johnson RL (2006) Nanotechnologies for envi-
ronmental cleanup. Nanotoday 1:44–48
Trinh Tan F, Cooper DG, Maric M, Nicell JA (2008) Biodeg-
radation of a synthetic co-polyester by aerobic mesophilic
microorganisms. Polym Degrad Stabil 93:1479–1485
Tsai Y-J (2008) Air distribution and size changes in the re-
mediated zone after air sparging for soil particle move-
ment. J Hazard Mater 158:438–444
Tsai T-T, Kao C-M, Yeh T -Y, Liang S-H, Chien H-Y (2009)
Remediation of fuel oil-contaminated soils by a three-
stage treatment system. Environ Eng Sci 26:651–659
Tsien HC, Hanson RS (1992) Soluble methane monooxygenase
component trichloroethylene. Appl Environ Microbiol
58:953–960
Tu C, Ma LQ, Zhang W, Cai Y, Harris WG (2003) Arsenic spe-
cies and leachability in the fronds of the hyperaccumulator
Chinese brake (Pteris vittata L.). Environ Pollut
124:223–230
Tuli A, Sethi RP, Khanna PK et al (1985) Lactic acid pro-
duction from whey permeate by immobilized Lactoba-cillus casei. Enzym Microbiol Technol 7:164–168
Tunali S, Kiran I, Akar T (2005) Chromium(VI) biosorption
characteristics of Neurospora crassa fungal biomass.
Miner Eng 18:681–689
Tundo P, Anastas P, StC Black D, Breen J, Collins T, Memoli
S, Miyamoto J, Polyakoff M, Tumas W (2000) Synthetic
pathways and processes in green chemistry. Introductory
overview. Pure Appl Chem 72:1207–1228
Urgun-Demirtas M, Stark B, Pagilla K (2006) Use of
genetically engineered microorganisms (GEMs) for the
bioremediation of contaminants. Crit Rev Biotechnol
26:145–164
Urum K, Pekdemir T, Ross D, Grigson S (2005) Crude oil con-
taminated soil washing in air sparging assisted stirred tank
reactor using biosurfactants. Chemosphere 60:334–343
Valdez-Vazquez I, Sparling R, Risbey D, Rinderknecht-Seijas
N, Poggi-Varaldo HM (2005) Hydrogen generation via
anaerobic fermentation of paper mill wastes. Bioresour
Technol 96:1907–1913
Van Eerd LL, Hoagland RE, Zablotowicz RM, Hall JC (2003)
Pesticide metabolism in plants and microorganisms. Weed
Sci 51:472–495
van Schie PM, Young LY (2000) Biodegradation of phenol:
mechanisms and applications. Bioremed J 4:1–18
Varanasi P, Fullana A, Sidhu S (2007) Remediation of PCB
contaminated soils using iron nano-particles. Chemo-
sphere 66:1031–1038
Vargas A, Soto G, Moreno J, Buitron G (2000) Observer-based
time-optimal control of an aerobic SBR for chemical and
petrochemical wastewater treatment. Water Sci Technol
42:163–170
Vatsala TM, Mohan Raj S, Manimaran A (2008) A pilot-scale
study of biohydrogen production from distillery effluent
using defined bacterial co-culture. Int J Hydrogen Energy
33:5404–5415
Vavilin VA, Fernandez B, Palatsi J, Flotats X (2008) Hydro-
lysis kinetics in anaerobic degradation of particulate
organic material: an overview. Waste Manag 28:939–951
Venkata Mohan S, Sirisha K, Chandrasekhara Rao N, Sarma
PN, Jayarama Reddy S (2004) Degradation of chlorpyri-
fos contaminated soil by bioslurry reactor operated in
sequencing batch mode: bioprocess monitoring. J Hazard
Mater 116:39–48
Venkata Mohan S, Ramakrishna M, Shailaja S, Sarma PN
(2007) Influence of soil–water ratio on the performance of
slurry phase bioreactor treating herbicide contaminated
soil. Bioresour Technol 98:2584–2589
Venkata Mohan S, Prasanna D, Purushotham Reddy B, Sarma
PN (2008) Ex situ bioremediation of pyrene contaminated
soil in bio-slurry phase reactor operated in periodic dis-
continuous batch mode: Influence of bioaugmentation. Int
Biodeterior Biodegrad 62:162–169
Vidali M (2001) Bioremediation. An overview. Pure Appl
Chem 73:1163–1172
Vieira MGA, Oisiovici RM, Gimenes ML, Silva MGC (2008)
Biosorption of chromium(VI) using a Sargassum sp.
packed-bed column. Bioresour Technol 99:3094–3099
Vijayaraghavan K, Yun Y-S (2008) Bacterial biosorbents and
biosorption. Biotechnol Adv 26:266–291
Vijayaraghavan K, Jegan J, Palanivelu K, Velan M (2005a)
Biosorption of cobalt(II) and nickel(II) by seaweeds:
batch and column studies. Sep Purif Technol 44:53–59
Vijayaraghavan K, Jegan J, Palanivelu K, Velan M (2005b)
Biosorption of copper, cobalt and nickel by marine green
alga Ulva reticulata in a packed column. Chemosphere
60:419–426
Vinas M, Sabate J, Espuny MJ, Solanas AM (2005) Bacterial
community dynamics and polycyclic aromatic hydrocar-
bon degradation during bioremediation of heavily creo-
sote-contaminated soil. Appl Environ Microbiol 71:8–18
Visoottiviseth P, Francesconi K, Sridokchan W (2002) The
potential of Thai indigenous plant species for the phyto-
remediation of arsenic contaminated land. Environ Pollut
118:453–461
Vogel-Mikus K, Pongrac P, Kump P, Necemer M, Regvar M
(2006) Colonisation of a Zn, Cd and Pb hyperaccumulator
Thlaspi praecox Wulfen with indigenous arbuscular
mycorrhizal fungal mixture induces changes in heavy
metal and nutrient uptake. Environ Pollut 139:362–371
Volesky B (2001) Detoxification of metal–bearing effluents:
biosorption for the next century. Hydrometall 59:203–216
Volkering F, Breure AM, Rulkens WH (1997) Microbiological
aspects of surfactant use for biological soil remediation.
Biodegradation 8:401–417
Volpe A, Del Moro G, Rossetti S, Tandoi V, Lopez A (2007)
Remediation of PCE-contaminated groundwater from an
industrial site in southern Italy: a laboratory-scale study.
Process Biochem 42:1498–1505
Wang HH (1999) Development and/or reclamation of biore-
sources with solid state fermentation. Proc Natl Sci
Council 23:45–61
286 Rev Environ Sci Biotechnol (2010) 9:215–288
123
Wang G-D, Chen X-Y (2007) Detoxification of soil phenolic
pollutants by plant secretory enzyme, phytoremedation.
Humana Press, Totowa, pp 49–57
Wang S, Mulligan CN (2009) Enhanced mobilization of
arsenic and heavy metals from mine tailings by humic
acid. Chemosphere 74:274–279
Wang Y, Oyaizu H (2009) Evaluation of the phytoremediation
potential of four plant species for dibenzofuran-con-
taminated soil. J Hazard Mater 168:760–764
Wang J, Zhao F-J, Meharg AA, Raab A, Feldmann J, McGrath
SP (2002) Mechanisms of arsenic hyperaccumulation in
Pteris Vittata. uptake kinetics, interactions with phosphate,
and arsenic speciation. Plant Physiol 130:1552–1561
Wang H, Shan X, Wen B, Zhang S, Wang Z (2004) Responses
of antioxidative enzymes to accumulation of copper in a
copper hyperaccumulator of Commoelina communis. Arch
Environ Contamin Toxicol 47:185–192
Waria M, Comfort SD, Onanong S, Satapanajaru T, Boparai H,
Harris C, Snow DD, Cassada DA (2009) Field-scale
cleanup of atrazine and cyanazine contaminated soil with
a combined chemical–biological approach. J Environ
Qual 38:1803–1811
Webb C, Koutinas AA, Wang R (2004) Developing a
sustainable bioprocessing strategy based on a generic
feedstock. Adv Biochem Eng Biotechnol 86:195–268
Weber R (2007) Relevance of PCDD/PCDF formation for the
evaluation of POPs destruction technologies—review on
current status and assessment gaps. Chemosphere 67:109–
117
Wei S, Zhou QX (2006) Phytoremediation of cadmium-con-
taminated soils by Rorippa globosa using two-phase
planting. Environ Sci Pollut Res 13:151–155
Wei S, Zhou Q, Koval PV (2006) Flowering stage character-
istics of cadmium hyperaccumulator Solanum nigrum L.
and their significance to phytoremediation. Sci Total
Environ 369:441–446
Whang L-M, Liu P-WG, Ma C-C, Cheng S-S (2008) Appli-
cation of biosurfactants, rhamnolipid, and surfactin, for
enhanced biodegradation of diesel-contaminated water
and soil. J Hazard Mater 151:155–163
White JC, Ross DW, Gent MPN et al (2006) Effect of mycor-
rhizal fungi on the phytoextraction of weathered p, p-DDE
by Cucurbita pepo. J Hazard Mater 137:1750–1757
Whiteley CG, Lee D-J (2006) Enzyme technology and bio-
logical remediation. Enzym Microb Technol 38:291–316
Wild SR, Jones KC (1992) Organic chemicals entering agri-
cultural soils in sewage sludges: screening for their
potential to transfer to crop plants and livestock. Sci Total
Environ 119:85–119
Wilson C, Tisdell C (2001) Why farmers continue to use
pesticides despite environmental, health and sustainability
costs. Ecol Econ 39:449–462
Xi Y, Mallavarapu M, Naidu R (in press) Reduction and
adsorption of Pb2? in aqueous solution by nano-zero-
valent iron—a SEM, TEM and XPS study. Mater Res Bull
Xia H (2008) Enhanced disappearance of dicofol by water
hyacinth in water. Environ Technol 29:297–302
Xiu Z, Jin Z, Li T, Mahendra S et al (2010) Effects of nano-scale
zero-valent iron particles on a mixed culture dechlorinating
trichloroethylene. Biores technol 101:1141–1146
Xu S-Y, Chen Y-X, Lin K-F, Chen X-C, Lin Q, Li F, Wang Z-
W (2009) Removal of pyrene from contaminated soils by
white clover. Pedosphere 19:265–272
Yagi JM, Madsen EL (2009) Diversity, abundance, and con-
sistency of microbial oxygenase expression and biodeg-
radation in a shallow contaminated aquifer. Appl Environ
Microbiol 75:6478–6487
Yang C, Song C, Mulchandani A, Qiao C (2010) Genetic
engineering of Stenotrophomonas Strain YC-1 to possess
a broader substrate range for organophosphates. J Agric
Food Chem 58:6762–6766
Yates GT, Smotzer T (2007) On the lag phase and initial
decline of microbial growth curves. J Theor Biol 244:
511–517
Yen KM, Karl MR, Blatt LM, Simon MJ, Winter RB, Fausset
PR, Lu HS, Harcourt AA, Chen KK (1991) Cloning and
characterization of Pseudomonas mendocina KRI gene
cluster encoding touene -4- monooxygenase. J Bacteriol
173:5315–5327
Yergeau E, Arbour M, Brousseau R, Juck D, Lawrence JR,
Masson L, Whyte LG, Greer CW (2009) Microarray and
real-time PCR analyses of the responses of high-arctic soil
bacteria to hydrocarbon pollution and bioremediation
treatments. Appl Environ Microbiol 75:6258–6267
Yoon JM, Oliver DJ, Shanks JV (2007) Phytotoxicity and
phytoremediation of 2, 6-dinitrotoluene using a model
plant, Arabidopsis thaliana. Chemosphere 68:1050–1057
Yu J, Tian N-N, Wang K-J et al (2008) New thought on
treatment and disposal of sludge of municipal sewage
treatment plants. China Wat Wastewat 24:11–14
Yuan SY, Su LM, Chang BV (2009) Biodegradation of
phenanthrene and pyrene in compost-amended soil.
J Environ Sci Health Part A 44:648–653
Zadrazil F (2000) Is conversion of lignocellulosics into feed with
white-rot fungi realizable? Practical problems of scale-up
and technology. In: Van Griensven LJLD (ed) Science and
cultivation of edible fungi. Balkema, Rotterdam
Zervakis G, Papadopoulou K, Ehaliotis C et al (2005) Use of
composts deriving from Mediterranean agro-industrial
wastes in vegetable crops: effects on disease suppression
and plant growth. In: de Kreij C, Warmenhoven M (eds)
Proceedings of the international symposium on the use of
composted organic wastes in horticulture, Wageningen
Zhang C, Bennett GN (2005) Biodegradation of xenobiotics by
anaerobic bacteria. Appl Microbiol Biotechnol 67:
600–618
Zhang G, Zhang P, Yang J, Chen Y (2007) Ultrasonic reduc-
tion of excess sludge from the activated sludge system.
J Hazard Mater 145:515–519
Zhang G, Yang J, Liu H, Zhang J (2009) Sludge ozonation:
disintegration, supernatant changes and mechanisms. Bi-
oresour Technol 100:1505–1509
Zhang Y, Luo X-J, Wu J-P, Liu J, Wang J, Chen S-J, Mai B-X
(2010) Contaminant pattern and bioaccumulation of leg-
acy and emerging organhalogen pollutants in the aquatic
biota from an e-waste recycling region in South China.
Environ Toxicol Chem 24(4):852–859
Zhao B, Poh CL (2008) Insights into environmental bioreme-
diation by microorganisms through functional genomics
and proteomics. Proteomics 8:874–881
Rev Environ Sci Biotechnol (2010) 9:215–288 287
123
Zhao FJ, Lombi E, Breedon T et al (2000) Zinc hyperaccu-
mulation and cellular distribution in Arabidopsis halleri.Plant Cell Environ 23:507–514
Zhao FJ, Dunham SJ, McGrath SP (2002) Arsenic hyperac-
cumulation by different ferns species. New Phytol
156:27–31
Zhou XB, Cebron A, Beguiristain T, Leyval C (2009) Water
and phosphorus content affect PAH dissipation in spiked
soil planted with mycorrhizal alfalfa and tall fescue.
Chemosphere 77:709–713
Zhuang X, Chen J, Shim H, Bai Z (2007) New advances in
plant growth-promoting rhizobacteria for bioremediation.
Environ Int 33:406–413
Ziagova M, Kyriakou G, Liakopoulou–Kyriakides M (2009)
Co-metabolism of 2, 4-dichlorophenol and 4-Cl-m-cresol
in the presence of glucose as an easily assimilated carbon
source by Staphylococcus xylosus. J Hazard Mater
163:383–390
Zoller U, Reznik A (2006) In-situ surfactant/surfactant-nutrient
mix-enhanced bioremediation of NAPL (fuel)-contaminated
sandy soil aquifers. Environ Sci Pollut Res 13:392–397
Zylstra GJ, Gibson DT (1989) Toluene degradation by
Pseudomonasputida FI, nucleotide sequence of the tod
CICBADE genes and their expression in E. coli. J Biol
Chem 264:149400–149446
288 Rev Environ Sci Biotechnol (2010) 9:215–288
123