2010 bioremediation elements

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: [email protected]

A. Mudhoo

Department of Chemical and Environmental Engineering,

Faculty of Engineering, University of Mauritius, Reduit,

Mauritius

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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


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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


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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


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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


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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


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


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:


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


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)


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


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)


123

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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


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)


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


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


123

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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)


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


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)


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

)


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

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nifi

can

tef

fect

Jay

aku

mar

and

Jale

el

(20

09

)


123

Ta

ble

11

Ap

pli

cati

on

of

ph

yto

rem

edia

tio

nfo

rre

mo

val

of

org

anic

po

llu

tan

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anic

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t(s)

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nt(

s)u

sed

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ov

alp

erfo

rman

ceR

efer

ence

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sole

tep

esti

cid

es—

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pes

tici

des

incl

ud

ing

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lite

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fD

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(dic

hlo

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chlo

roet

han

e)an

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iso

mer

so

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(hex

ach

loro

cycl

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exan

e)

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eK

azak

hst

anm

axim

um

acce

pta

ble

con

cen

trat

ion

for

DD

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dH

CH

met

abo

lite

sin

pla

nt

tiss

ue

is

20

lg/k

g.

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ecie

sin

this

cate

go

ryin

clu

ded

:

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emis

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nn

ua

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chia

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eded

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inth

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no

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ile

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rile

un

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d

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ted

and

ster

ile

un

pla

nte

d

trea

tmen

ts,

resp

ecti

vel

y

Xia

(20

08

)


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

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ov

alp

erfo

rman

ceR

efer

ence

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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.

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per

imen

tal

con

dit

ion

sas

sure

dth

at

ph

oto

deg

rad

atio

n,

adso

rpti

on

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esse

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alls

and

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ial

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atio

nd

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ute

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e

rem

ov

al

Do

rdio

etal

.

(20

09

)

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ng

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es—

dim

eth

om

orp

h

and

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rim

eth

anil

Fiv

em

acro

ph

yte

spec

ies—

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min

or,

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po

lyrh

iza

,

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aq

ua

tica

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stri

san

dE

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na

den

sis

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ere

mo

val

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lds

du

rin

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test

per

iod

sv

arie

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m1

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%an

d7

–1

2%

for

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eth

om

orp

han

d

py

rim

eth

anil

,re

spec

tiv

ely

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he

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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

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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

)


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)


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.


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


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).


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)


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


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


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


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


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’.


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.


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


123

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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

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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


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).


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,


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)


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


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)


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


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


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)


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


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.


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


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.


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


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


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)


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


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.

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2010 bioremediation elements

Documents

environmental engineering

slurryphase bioremediation

bioremediation ofcontaminants

studiesabout bioremediation

environmental problems

environmental protection

faculty of engineering

oil contaminated soils