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Review

Bioremediation approaches for organic pollutants: A critical perspective

Mallavarapu Megharaj a,e, Balasubramanian Ramakrishnan a,b,e, Kadiyala Venkateswarlu a,c,e,⁎,Nambrattil Sethunathan d, Ravi Naidu a,e

a Centre for Environmental Risk Assessment and Remediation, University of South Australia, SA5095, Australiab Division of Microbiology, Indian Agricultural Research Institute, New Delhi 110012, Indiac Department of Microbiology, Sri Krishnadevaraya University, Anantapur 515055, Indiad Flat No. 103, Ushodaya Apartments, Sri Venkateswara Officers Colony, Ramakrishnapuram, Secunderabad 500056, Indiae Cooperative Research Centre for Contamination Assessment and Remediation of Environment, PO Box 486 Salisbury South, SA5106, Australia

a b s t r a c ta r t i c l e i n f o

Article history:Received 24 November 2010Accepted 7 June 2011Available online 1 July 2011

Keywords:Bioremediation approachesOrganic pollutantsElectrobioremediationGEMsRhizoremediationLimitations

Due tohumanactivities toagreater extent andnatural processes to someextent, a largenumberof organic chemicalsubstances such as petroleumhydrocarbons, halogenated andnitroaromatic compounds, phthalate esters, solventsand pesticides pollute the soil and aquatic environments. Remediation of these polluted sites following theconventional engineering approaches based on physicochemical methods is both technically and economicallychallenging. Bioremediation that involves the capabilities of microorganisms in the removal of pollutants isthe most promising, relatively efficient and cost-effective technology. However, the current bioremediationapproaches suffer from a number of limitations which include the poor capabilities of microbial communities inthe field, lesser bioavailability of contaminants on spatial and temporal scales, and absence of bench-mark valuesfor efficacy testing of bioremediation for their widespread application in the field. The restoration of all naturalfunctions of somepolluted soils remains impractical and, hence, the applicationof theprinciple of function-directedremediation may be sufficient to minimize the risks of persistence and spreading of pollutants. This reviewselectively examines and provides a critical view on the knowledge gaps and limitations in field applicationstrategies, approaches such as composting, electrobioremediation andmicrobe-assisted phytoremediation, and theuse of probes and assays for monitoring and testing the efficacy of bioremediation of polluted sites.

© 2011 Elsevier Ltd. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13632. Bioavailability: what fraction of pollutants is available to microorganisms? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1363

2.1. Surfactants: bioavailability enhancers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13643. Biodegradation: cooperation and networking . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13654. In situ and ex situ bioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1366

4.1. Bioattenuation: the natural way . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13664.2. Biostimulation: importance of correct nutrient ratios . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13664.3. Bioaugmentation: when the locals aren't up to the task? . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1366

4.3.1. Biosurfactant-producing and pollutant-degrading microbial inoculants: dual role . . . . . . . . . . . . . . . . . . . . . 13674.3.2. Genetic engineering of microorganisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1367

5. Bioremediation technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13685.1. Composting and addition of composted material . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13685.2. Electrobioremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13695.3. Microbe-assisted phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1370

6. Bioremediation monitoring and efficacy testing: relevance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13717. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1371Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1372References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1372

Environment International 37 (2011) 1362–1375

⁎ Corresponding author at: Department of Microbiology, Sri Krishnadevaraya University, Anantapur 515055, India. Tel.: +91 8554 255760; fax: +91 8554 255805.E-mail address: v_kadiyala@hotmail.com (K. Venkateswarlu).

0160-4120/$ – see front matter © 2011 Elsevier Ltd. All rights reserved.doi:10.1016/j.envint.2011.06.003

Contents lists available at ScienceDirect

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1. Introduction

Any unwanted substance introduced into the environment isreferred to as a ‘contaminant’. Deleterious effects or damages bythe contaminants lead to ‘pollution’, a process by which a resource(natural or man-made) is rendered unfit for use, more often thannot, by humans. Pollutants are present since time immemorial, andlife on the earth as we define now has always evolved amongst them.With pollutant analogues from geothermal and volcanic activities,comets, and space dust which are about 100 t of organic dust per day,the earth is forever a polluted planet (Marcano et al., 2003). Relativeto the pre-industrialization era, industrialization and intensive useof chemical substances such as petroleum oil, hydrocarbons (e.g.,aliphatic, aromatic, polycyclic aromatic hydrocarbons (PAHs), BTEX(benzene, toluene, ethylbenzene, and xylenes), chlorinated hydro-carbons like polychlorinated biphenyls (PCBs), trichloroethylene(TCE), and perchloroethylene, nitroaromatic compounds, organo-phosphorus compounds) solvents, pesticides, and heavy metals arecontributing to environmental pollution. Large-scale pollution dueto man-made chemical substances and to some extent by naturalsubstances is of global concern now. Seepage and run-offs due to themobile nature, and continuous cycling of volatilization and conden-sation of many organic chemicals such as pesticides have even ledto their presence in rain, fog and snow (Dubus et al., 2000). Everyyear, about 1.7 to 8.8 million metric tons of oil is released into theworld's water. More than 90% of this oil pollution is directly related toaccidents due to human failures and activities including deliberatewaste disposal (Zhu et al., 2001).

PAHs are present at levels varying from 1 μg to 300 g kg−1 soil,depending on the sources of contamination like combustion of fossilfuels, gasification and liquefaction of coal, incineration of wastes, andwood treatment processes (Bamforth and Singleton, 2005). Incom-plete combustion of organic substances gives out about 100 differentPAHs which are the ubiquitous pollutants. Except for a few PAHs usedin medicines, dyes, plastics and pesticides, they are rarely of industrialuse (US EPA, 1998). Some PAHs and their epoxides are highly toxic,and mutagenic even to microorganisms. About six specific PAHs arelisted among the top 126 priority pollutants by the US EnvironmentalProtection Agency. PCBs, used in hydraulic fluids, plasticizers, ad-hesives, lubricants, flame retardants and dielectric fluids in trans-formers are toxic, carcinogenic, and degrade slowly. Polychlorinateddibenzodioxins and dibenzofurans are recalcitrant chemicals andsome of the congeners with lateral chlorine substitutions at positions2,3,7 and 8 are carcinogenic to humans (Kaiser, 2000). Many solventssuch as TCE and carbon tetrachloride pollute the environments dueto large-scale industrial production and anthropogenic uses. Pesti-cides are regularly used in agricultural- and public health-programsworldwide. In many cases, the environmental effects of these chem-ical substances outweigh the benefits they accrue to humans andnecessitate the need of their degradation after the intended uses.

The microbial transformation may be driven by energy needs, ora need to detoxify the pollutants, or may be fortuitous in nature(cometabolism). Because of the ubiquitous nature of microorganisms,their numbers and large biomass relative to other living organisms inthe earth (Curtis et al., 2002), wider diversity and capabilities in theircatalytic mechanisms (Chen et al., 1999; Paul et al., 2005), and theirability to function even in the absence of oxygen and other extremeconditions (Mishra et al., 2001; Watanabe, 2001), the search forpollutant-degrading microorganisms, understanding their geneticsand biochemistry, and developing methods for their application inthe field have become an important human endeavor. The recentadvances in metagenomics and whole genome sequencing haveopened up new avenues for searching the novel pollutant degradativegenes and their regulatory elements from both culturable and non-culturable microorganisms from the environment (Golyshin et al.,2003; Zhao and Poh, 2008). Compared to other living organisms

which can degrade organic pollutants as well as the cost-intensivephysical and chemical methods for the cleanup, microorganisms arepreferred agents. Their capabilities to degrade organic chemical com-pounds can be made use of to attenuate the polluted sites.

Bioremediation, which is defined as a process that uses microor-ganisms, green plants or their enzymes to treat the polluted sitesfor regaining their original condition (Glazer and Nikaido, 1995), hasconsiderable strength and certain limitations. Remediation, whetherby biological, chemical or a combination of both means, is the onlyoption as the problem of pollution has to be solved without trans-ferring to the future. As the knowledge demand and complexities varyfor different bioremediation treatments, a better understanding ofthe premises together with the limitations of bioremediation aids inmaximizing the benefits and minimizing the cost of treatments. In thepresent review, we examine critically and present (i) the advancesmade thus far and the requisite foci of research on bioavailability oforganic chemical pollutants, (ii) the search, identification, stimulationand augmentation of pollutant-degrading microorganisms, (iii) theapplication of innovative approaches such as electrobioremediationand microbe-assisted phytoremediation, and (iv) an assessment onthe use of probes and assays that are required for monitoring andtesting the efficacy of bioremediation of contaminated sites.

2. Bioavailability: what fraction of pollutants is availableto microorganisms?

The process of bioremediation depends on the metabolic potentialof microorganisms to detoxify or transform the pollutant molecule,which is dependent on both accessibility and bioavailability (Antizar-Ladislao, 2010). There is a considerable debate in the literature on“what constitutes the bioavailable fraction” and the methods of itsmeasurements (Alexander, 2000; Vasseur et al., 2008). Followingentry into the soil environment, pollutants rapidly bind to the mineraland organic matter (solid phases) via a combination of physicaland chemical processes. Sorption, complexation and precipitationconstitute the pollutant–soil interaction. The ability of soils to release(desorb) pollutants determines its susceptibility to microbial degra-dation, thereby influencing effectiveness of the bioremediationprocess. In soil aggregates which are the smallest ‘composite units’in the heterogeneous soil environment, bioavailability is limited bytransport of the pollutant molecule to a microbial cell, i.e., diffusion ofpollutant out of a soil aggregate to the cell attached to the externalsurface of the aggregate.

Sorption which influences the bioavailability of a contaminant is acritical factor, yet a poorly understood process in bioremediation.There are two schools of thought concerning bioavailability and theconsequent biodegradation of organic contaminants (Singh et al.,2008): (i) the pre-requisite release of contaminant from sorbed phaseto aqueous phase for its degradation by microorganisms (Harms andZehnder, 1994; Shelton and Doherty, 1997), and (ii) biodegradation ofthe contaminant in the sorbed phase, without being desorbed, by theenzymes (Singh et al., 2003). The degradation of sorbed contaminantscan presumably occur via microbially-mediated desorption of con-taminants through production of biosurfactants and the developmentof a steep gradient between solid phase and interfacial contaminant(Tang et al., 1998). Thus, these reports suggest that bioavailability iseven species specific (i.e., the ability of certain species to desorb thecontaminant and then degrade). The organic contaminants can also bedegraded without prior desorption. Singh et al. (2003) demonstratedthat a soil bacterium, Brevibacterium sp. degraded the pesticidefenamiphos which was intercalated into the cationic-surfactantmodified montmorillonite clay (CTMA–Mt–fenamiphos complex).The interlayer space is otherwise inaccessible to the bacterium dueto its size of several orders lower than that of the bacteria. Thescanning electron microscope analysis showed the surface attach-ment of bacteria to the surface of the CTMA–Mt–fenamiphos complex,

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suggesting the involvement of extracellular enzyme in the degrada-tion of fenamiphos, without its prior desorption. The degradationof sorbed contaminants depends on the enrichment and isolationprocedures used for obtaining the culturable bacteria. As against theconventional approach of providing the contaminant as a sole carbonsource in aqueous medium, the provision of phenanthrene sorbedon a polyacrylic porous resin to the bacterial cultures led to fasterdegradation of phenanthrene than those isolated by the conventionaltechnique (Grosser et al., 2000; Tang et al., 1998).

Aqueous solubility, volatility or reactivity of organic pollutantsvaries greatly, and all of them may influence their bioavailability inwater and soils. On a mass basis, no relationship exists between thechemical pollutant in soil and its biological effect. The dissolved formof contaminants in pore water is considered to be bioavailable, com-pared to the bound chemical which does not exert direct biologicaleffects. This has led to the ‘pore water hypothesis.’ The equilibriumpartitioning theory is applied to estimate the dissolved fraction ofpollutant in pore water and to remove the soil to soil differences intoxicological effects (EFSA, 2009; Ronday et al., 1997). The basicassumption of equilibrium partitioning theory is that the partitioningof an ionic chemical between the mineral and organic matter in soilor sediment and the pore water is at equilibrium, and in each phasethe chemical potential which controls its biological activity isthe same. The performance of chemical extraction data of nonionicorganic chemicals can be improved by organic matter normalizationin order to predict the occurrence of toxicity effects.

For highly hydrophobic chemical pollutants which have higheroctanol–water partition coefficient (Kow) with log Kow values morethan 4, the measured concentration in the pore water is the sum ofthe free chemical and the fraction sorbed to dissolved organic matter(DOM). To account for the sorbed fraction to DOM, the separationmethods for DOM are required (Landrum et al., 1984). The soil–chemical contact time determines the usefulness of pore waterhypothesis in measuring bioavailability and predicting the biologicaleffects or the fraction which can be degraded, but not immediatelyafter contamination. There are also variations in bioavailability dueto the nature of chemical pollutants, soil types, and other factors suchas water content and temperature. Toxicity testing of a pollutant tomicroorganisms (Ronday et al., 1997) or the use of extracts such as themild hydroxypropyl-β-cyclodextrin for PAHs (Ling et al., 2010) or thematrix solid-phase microextraction for DDTs (1,1,1-trichloro-2,2-bis(p-chlorophenyl) ethane and its metabolites) (Fang et al., 2010) canprovide direct measures of bioavailability. Cornelissen et al. (1998)demonstrated that microbial factors, not bioavailability, were respon-sible for the persistence of rapidly desorbing fractions of the non-degraded PAHs, and these fractions were found to be substantial (upto 55%) and remained unchanged during remediation. For the purposeof bioremediation and regulatory measures, the bioavailability inthe initial rapid phase and the ensuing slow phase in the biphasicdegradation profile of an organic pollutant is to be monitored.

The sequestration of pollutants over time may occur due to thecontact and interaction of soil with pollutant molecules. Factors suchas organic matter, cation exchange capacity, micropore volume, soiltexture and surface area affect the pollutant sequestration (Chungand Alexander, 2002). Sequestration and reduced bioavailability ofphenanthrene were reported for a Gram-negative bacterial isolate(strain PS5-2) when the hydrophobic compound entered intonanopores having hydrophobic surfaces (Nam and Alexander,1998). Sharer et al. (2003) observed that aging caused an increasein sorption for some organic compounds (e.g., 2,4-dichlorophenox-yacetic acid) but not for others (chlorobenzene, ethylene dibromide)on a common soil type. Even a weakly sorbed and easily degradedcarbamate insecticide, carbaryl, can be effectively sequestrated in soilwith aging, thereby rendering it partly inaccessible tomicroorganismsand affecting the bioavailability (Ahmad et al., 2004). Hence, thegeneralizations about the effects of aging on the sorption–desorption

behavior of different organic chemicals are difficult to achieve. Somepertinent issues that need to be considered include: (a) bioavail-ability and toxicity of parent molecules and their residues in soils,(b) standardized protocols for different pollutants and their use acrossthe sites, (c) assessment on remobilization of pollutants during thepost-remediation period, and (d) determination of environmentallyacceptable pollutant end-points in the bioremediated soils. The‘pollutant (or contaminant) sequestration’ due to the prolonged con-tact between soil particles and chemical molecules, however, poses lessrisk and threat to the environmental health. In general, difficulties withanalytical measurements for determining low levels of new organicpollutants in soils, the absence of base-line values related to theircompositional, geographical and distribution patterns, and the com-plexities in their toxicological interactions (Mas et al., 2010) make thebioavailability measurements of organic pollutants exigent.

2.1. Surfactants: bioavailability enhancers

Application of surfactants to polluted soils has been used as one ofthe treatment strategies for increasing the mass transfer of hydro-phobic organic contaminants (Laha et al., 2009; Rosen, 1989). Thesurfactants are amphiphilic molecules that contain hydrophilic andhydrophobic moieties; hydrophilic groups can be anionic, cationic,zwitter ionic, and nonionic. The synthetic surfactants contain sulfate,sulfonate or carboxylate group (anionic); quaternary ammoniumgroup (cationic); polyoxyethylene, sucrose, or polypeptide (nonionic)and the hydrophobic parts of paraffins, olefins, alkylbenzenes,alkylphenols, or alcohols. The common chemical surfactants such asTriton X-100, Tween 80 and sodium dodecyl sulphate are petroleum-derived products. The zwitter ionic surfactants (e.g., N-dodecylbetaine) which contain both anionic and cationic groups have lowcritical micelle concentration (CMC) values, more surface active,and high solubilization capacity. Increased desorption rates of sorbedpollutants from soils by the application of surfactants make thepollutants available for remediation (Fu and Alexander, 1995). Solu-bilization of hydrophobic contaminants is attributed to the incorpo-ration of the molecule into the hydrophobic core of micelles insolution (Guha and Jaffe, 1996). The salient mechanisms which areinvolved in the surfactant-amended remediation are: (i) loweringof interfacial tension, (ii) surfactant solubilization of hydrophobicorganic compounds, and (iii) the phase transfer of organic compoundsfrom soil-sorbed to pseudo-aqueous phase (Laha et al., 2009).

Surfactants enhance mobilization and biodegradation of PAHs insoils (Tiehm et al., 1997). Enhanced rates of degradation of naph-thalene and phenanthrene in the presence of some nonionic sur-factants at applications below their CMC were observed by Aronsteinet al. (1991). Similarly, significant solubility enhancements of DDTin Triton and Brij 35 surfactants were noticed by Kile and Chiou(1989) below their CMC. Factors such as cost, effectiveness at con-centrations lower than 3%, low toxicity to humans, animals and plants,low adsorption to soil, low soil dispersion, and low surface tensiondetermine the selection of surfactants for field application (Mulliganet al., 2001). Toxicities of surfactants to soil biota can prevent thebiodegradation of pollutants and disturb the balanced ecologicalfunctions (Rosal et al., 2010).

The food-grade surfactants (T-MAZ 28, T-MAZ 10, and T-MAZ 60)(Shiau et al., 1995), the plant-based surfactants (e.g., fruit pericarpfrom Sapindus mukurossi) (Roy et al., 1997) or the natural surfactantssuch as humic acids (Conte et al., 2005) may be preferred to thesynthetic surfactants due to high biodegradability, low toxicity, andhigher public acceptance. Microorganisms also produce surfactants(surface-active amphiphilic metabolites such as glycolipids, phos-pholipids, lipopeptides, lipoproteins, and lipopolysaccharides). Theselow- and high-molecular weight biosurfactants find their uses in foodprocessing, cosmetic and pharmaceutical industries, in addition tobioremediation efforts (Christofi and Ivshina, 2002). The classes of

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biosurfactant and microbial species which can produce them arenumerous, leading to continuous search for the novel biosurfactants(Satpute et al., 2010). However, the in situ application of surfactantsto enhance bioavailability of persistent organic pollutants requirescareful planning and selection based on the prior information aboutthe fate and behavior of the surfactant and the target pollutant.Caution is required to prevent groundwater contamination vialeaching and consequent toxicity to microorganisms. Hence, a goodstrategy will be to select bacteria that are capable of not onlycatabolizing the target contaminant but also producing surfactant.More knowledge on the mechanisms of pollutant–surfactant in-teractions with regard to diffusion, in and out of the micelles, andmodeling of pollutant's transport at the field site can help to designefficient remediation strategy.

3. Biodegradation: cooperation and networking

As much as the diversity in sources and chemical complexities inorganic pollutants exists, there is probablymore diversity in microbialmembers and their capabilities to synthesize or degrade organiccompounds (Ramakrishnan et al., 2010, 2011; Watanabe, 2001).Microbial populations even contribute to naturally-occurring hydro-carbons by diagenesis of bacteriohopanetetrol (a membrane constit-uent) into the formation of hopanoic acids and hydrocarbons such ashopane (Stout et al., 2001). Themicrobial diversity is larger than whatis known from the cultured members (Curtis et al., 2002). However,the metabolic diversity of culturable microorganisms for degradingorganic pollutants may be insufficient to protect the earth from theanthropogenic pollution. This is largely due to recalcitrant chemicalswith substituent or structural elements, which seldom occur in nature(Pieper and Reineke, 2000). But, Singer et al. (2004) were of theopinion that the naturally-occurring tritrophic trinity of microbe–plant–insect interactions has capabilities to produce hundreds ofthousands of different chemicals to attract, defend, antagonize,monitor and misdirect one another among these members and onlynegligible numbers are truly novel chemicals of anthropogenic origin.Hence, there is a fortuitous evolution of xenobiotic-degradingenzymes from the interactions of microbe–plant–insect.

The potential to degrade organic pollutants varies amongmicrobialgroups or different guilds (group of species that exploit the same classof environmental resources in a similar way) and is dose-dependent.For example, mycobacteria are excellent candidates for remediatingaged PAH-polluted sites as these organisms have lipophilic surfaces,suitable for uptake of bound pollutants from soil particles and havecatabolic efficiency towards PAHs up to five benzene rings (Boganet al., 2003). Higher doses of PAHs are phytotoxic to algae includingmicroalgae. The inhibitory or toxic components of the pollutantmixture can attenuate the potential of microbial degradation and areimportant stressors. Organic compounds such as toluene are toxic tomicroorganisms because they disrupt cell membranes. Providentially,several bacteria develop resistance to solvents by the cis to transisomerization of fatty acids, increased synthesis of phospholipids,low cell-surface hydrophobicity (modification of the lipopolysaccha-ride or porines of the outer membrane), and the presence of solventefflux pump. In the soils polluted by aromatic hydrocarbons, thesolvent-tolerant microorganisms are the first to colonize and becomepredominant in the removal of pollutants (Huertas et al., 1998). Eitherbioaugmentation with the solvent-tolerant bacteria or modifyingthese bacteria with an appropriate catabolic potential will provideadvantages in bioremediation programs.

The microbial populations of soil or aquatic environments arecomposed of diverse, synergistic or antagonistic communities ratherthan a single strain. In the natural environments, biodegradationinvolves transferring the substrates and products within a well co-ordinated microbial community, a process referred to as metaboliccooperation (Abraham et al., 2002). It still remains very challenging to

introduce all the genes required for degradation for many organicpollutants or stable maintenance of even a single gene or a desiredtrait such as enhanced degradative capacity in a single organism.Hence, the microbial consortia of ecologically relevant candidate taxawhich are known to degrade the chemical pollutants and respond todifferent environmental stimuli are desired, rather than the singleisolate for augmentation (Supaphol et al., 2006).

The reductionist approach to studying biodegradation processeshas been very useful so far for understanding individual genes,enzymes and organisms, but the systems biology approach is nec-essary to examine the complex web of metabolic and regulatoryinteractions even within a single organism (Pazos et al., 2003; Trigoet al., 2009). Pazos et al. (2003) considered the biodegradationprocess as a single interconnecting network (metabolic cooperation),with metabolic activities and substrates and intermediate compoundsflowing freely in the environment and less boundaries existingbetween bacterial species. The ‘network theory,’ thus forms a basisfor studying the functional properties and mechanisms involved inthe organization of biological systems and predicting their responsesto environmental (both internal and external) variations (Feist andPalsson, 2008).

Formalization and categorization of many biodegradation re-actions and pathways have been done in the University of MinnesotaBiocatalysis/Biodegradation Database (UM-BBD) (Ellis et al., 2006).With information on about 900 compounds, 600 enzymes, 1000reactions and 350microbial entries, the UM-BBD is useful for applyingthe system biology approaches. The most likely metabolic pathwayfor any given compound is predictable using the ‘reaction rules’ forparticular functional groups (Ellis et al., 2008). Pazos et al. (2005)developed a database, ‘Metarouter’, based on the informationavailable at the UM-BBD. Using the Metarouter, Gomez et al. (2007)showed the existence of a correlation between the frequency of 149chemical triads (chemotypes) common in organo-chemical com-pounds and the global capacity of microorganisms to metabolizethem. These authors developed a predictive tool (http://www.pdg.cnb.uam.es/BDPSERVER) which can provide the biodegradativeoutcome of the compounds as biodegradable or recalcitrant, depend-ing on the type of environmental fate defined. Trigo et al. (2009)suggested that (i) the central metabolism of the global biodegradationnetworks involves transferases, isomerases, hydrolases and ligases,(ii) linear pathways converging on particular intermediates form afunnel topology, (iii) the novel reactions exist in the exterior part ofthe network, and (iv) the possible pathway between compoundsand the central metabolism can be arrived at by considering all therequired enzymes in a given organism and intermediate compounds.

Biodegradation in the natural environment is beyond the ‘com-plete system’ of a single cell, where the ‘system’ is extremely complexinvolving multiple biotic and abiotic components. Nevertheless,there exists the coordination of microbial communities to mediateand transfer substrates and products between species and commu-nities (Abraham et al., 2002). The application of molecular siteassessment (Fleming et al., 1998; Sayler et al., 1995) and molecularecological techniques for community profiling (Malik et al., 2008), soilmetagenomics using isotope distribution analysis (Villas-Boas andBruheim, 2007), and functional genomics and proteomics (Zhao andPoh, 2008) can help in identifying the partners and the patterns ofresponses to external stimuli within the network and the ‘systemcomplexities’ of contaminated sites. Stable isotope probing (SIP)analyses, either DNA-SIP (Winderl et al., 2010) or RNA-SIP (Bombachet al., 2010), provide opportunities to link microbial diversity withfunction and identify those culturable as well as yet-to-be culturedorganisms which are involved in biodegradation in the field (Cupples,2011). Likewise, the high-throughput approaches such as DNAmicroarrays, metagenomics, metatranscriptomics, metaproteomics,metabolomics, and whole cell-based biosensors are useful to char-acterize the contaminated sites, identify new degradative activities

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and monitor bioremediation efficiency (Desai et al., 2010; Kakirdeet al., 2010; Stenuit et al., 2008).

4. In situ and ex situ bioremediation

Bioremediation approaches are generally classified as in situ orex situ. In situ bioremediation involves treating the polluted materialat the site while ex situ involves the removal of the polluted materialto be treated elsewhere (Aggarwal et al., 1990). In situ bioremediationcan be described as the process whereby organic pollutants arebiologically degraded under natural conditions to either carbondioxide and water or an attenuated transformation product. It is alow-cost, low maintenance, environment-friendly and sustainableapproach for the cleanup of polluted sites. With the need for ex-cavation of the contaminated samples for treatment, the cost of ex situbioremediation approaches can be high, relative to in situ methods.In addition, the rate of biodegradation and the consistency of theprocess outcome differ between the in situ- and ex situ bioremediationmethods. While the methods of both in situ and ex situ remediationdepend essentially on microbial metabolism, the in situ bioremedia-tion methods are preferred to those of ex situ for ecological resto-ration of contaminated soil and water environments (Jorgensen,2007). Three different types of in situ bioremediation process are(i) bioattenuation which depends on the natural process of degrada-tion, (ii) biostimulation where intentional stimulation of degradationof chemicals is achieved by addition of water, nutrient, electrondonors or acceptors, and (iii) bioaugmentation where the microbialmembers with proven capabilities of degrading or transformingthe chemical pollutants are added (Madsen, 1991). The suitability ofa particular bioremediation technology is determined by severalfactors, such as site conditions, indigenous population of microor-ganism, and the type, quantity and toxicity of pollutant chemicalspecies present.

4.1. Bioattenuation: the natural way

During bioattenuation (natural attenuation), the pollutants aretransformed to less harmful forms or immobilized. Such transforma-tion and immobilization processes are largely due to biodegradationby microorganisms (Smets and Pritchard, 2003), and to some extentby the reactions with naturally-occurring chemicals and sorption onthe geologic media. The natural attenuation processes are contami-nant-specific, accepted as methods for treating fuel components(e.g., BTEX) (Atteia and Guillot, 2007), but not for many other classes.The time required for natural attenuation varies considerably withsite conditions. Many polluted sites may not require an aggressiveapproach to remediation, and bioattenuation is efficient and cost-effective (Davis et al., 1994; Mulligan and Yong, 2004). In fact, avariety of bioremediation techniques have been successfully em-ployed at over 400 cleanup sites throughout the USA, at costs whichare approximately 80–90% lower than other cleanup technologies,based on the physical and chemical principles. With minimal sitedisturbance, the post-cleanup costs are also substantially reduced.Consequently, the global demand for bioremediation along withphytoremediation technologies is valued to be about US $1.5 billionper annum (Singh et al., 2009). Industrial and environmentalbiotechnologies also prefer newer paths, resulting in processes with‘clean technologies’, with maximum production and fewer residues.Bioattenuation alone becomes inadequate and protracted in manycases since many soils are oligotrophic in nature or lack appropriatemicroorganisms.

4.2. Biostimulation: importance of correct nutrient ratios

The acceleration of microbial turnover of chemical pollutants gen-erally depends on the supply of carbon, nutrients such as N and P,

temperature, available oxygen, soil pH, redox potential, and the typeand concentration of organic pollutant itself (Carberry and Wik,2001). To stimulate microbial degradation, nutrients in the form offertilizers (water soluble (e.g., KNO3, NaNO3, NH3NO3, K2HPO4 andMgNH4PO4), slow release (e.g., customblen, IBDU, max-bac), and oleo-philic (e.g., Inipol EAP22, F1, MM80, S200)) are added (Nikolopoulouand Kalogerakis, 2008). As a thumb rule for oil spill remediation,around 1–5% N by weight of oil with a ratio of N:P between 5 and 10:1is applied (Swannell et al., 1996). These additions may be insufficientor inaccurate for polluted sites with different types of pollutants.Formulation of nutrient-treatment strategies and maintenance of con-trol on the degradation rates and the outcomes of degradation needto be tailored to specific site/pollutant combinations. Limitations ofnutrients such as nitrogen and phosphorus onmicrobial decompositionof organic matter and the possible ecological implications of theseeffects for carbon flow through natural ecosystems are well known(Sterner and Elser, 2002). Wolicka et al. (2009) optimized the C:N:Pratio (at the level of 100:9:2, 100:10:1 or 250:10:3) before commencingin situ remediation of BTEX.

The ‘ecological stoichiometry’ is concerned with the supplies ofnutrients, and their elemental stoichiometry relative to the nutritionaldemands of the cell's innate physiology. It also exemplifies the effectsof resource (nutrient) supply rates and supply ratios on the structureand function of microbial communities (Smith, 2002). Smith et al.(1998) applied the resource-ratio theory to hydrocarbon degradationand demonstrated that the changes in nitrogen and phosphorussupply ratios not only altered the biodegradation rates of hydrocar-bons (hexadecane and phenanthrene) but also the microbial com-munity composition significantly. In addition, the changes in absolutenutrient supply levels, at constant supply ratio, were found to altertotal hydrocarbon degrader biomass, with altered rates of hydrocar-bon degradation. The ‘resource-ratio approach’ to gain information onthe ecophysiological status of pollutant-degrading microorganismshas many practical implications. Basically, it provides the theoreticalframework for optimizing nutrient formulation and application inbiostimulation approaches.

4.3. Bioaugmentation: when the locals aren't up to the task?

Often, the biological response lags behind, up to weeks or months,in the polluted sites with no exposure history. The ‘soil activation,’ aconcept which is based on the cultivation of biomass from a fractionof a contaminated soil and the subsequent use as an inoculum forbioaugmentation for the same soil was attempted by Otte et al. (1994)for degradation of PCP and PAHs. The soils with microbiota, adaptedby prior exposure to degradation of organic pollutants such ashydrocarbons can be a source of microorganisms for remediating soilsfreshly contaminated with hydrocarbons. Priming with 2% bioreme-diated soil was found to increase biodegradation of PAH constituentsof a fuel oil-treated soil (Lamberts et al., 2008). Similar priming effectof exhaustively bioremediated soils for hydrocarbon degradation wasobserved by Greenwood et al. (2009). Exposure history and adaptivestatus of microbial degraders thus determine the lag period ofdegradation. In addition, ascertaining the history of exposure ofchemical pollutants in the contaminated sites has even becomesignificant in the environmental forensics such as the 1989 ExxonValdez oil spill case (Peters et al., 2005) and for ecological engineeringsuch as the 2010 Gulf of Mexico oil spill case (Mitsch, 2010).

Pre-adaptation of catabolic bacteria to the target environment,prior to inoculation, improves survival, persistence and degradativeactivities, leading to enhanced remediation of the polluted soil(Megharaj et al., 1997). Sphingomonas sp. RW1 which contained amini transposon Tn-5 lacZ was pre-adapted to soil by growing in thesoil extract medium. The pre-adapted bacterium exhibited bettersurvival and efficient degradation of dibenzo-p-dioxin and dibenzo-furan in the polluted soil, compared to the unadapted bacterium,

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grown only in the nutrient-rich medium. Sudden exposure to stressesin soil (oligotrophic conditions that generally exist in soils, starvationor susceptibility/resistance, etc.) determines the physiological re-sponse of bacteria and their subsequent survival and activities.

Pre-exposure and subsequent re-exposure of a chemical pollutantenhances the metabolic potential of microorganisms (Reddy andSethunathan, 1983). The phenomenon of retaining specific metaboliccapacity after pre-exposure over long periods of time is referred toas ‘soil memory.’ The soil memory makes a contribution to thesubsequent natural attenuation. Now, in a typical bioaugmentationapproach, microorganisms are amended to a polluted site to hastendetoxification and/or degradation. There are many reports onbioaugmentation for treatment of soils containing organic pollutants(Brunner et al., 1985). Gilbert and Crowley (1998) found that therepeated application of carvone-induced bacteria enhanced biodeg-radation of PCBs in soil. To improve efficiency of bioaugmentation,microorganisms of different physiological groups and of differentdivisions can also be brought together. Bender and Phillips (2004)suggested the use of microbial mats which occur in nature as stratifiedcommunities of cyanobacteria and bacteria to remediate organiccontaminants by degrading and completely mineralizing the contam-inants. Wolicka et al. (2009) applied aerobic microbial communities,selected from those adapted to utilize one type of BTEX compound, forbioremediation of soil contaminated with BTEX.

A successful strategy for in situ bioremediation can be thecombination, in a single bacterial strain or in a syntrophic bacterialconsortium, of different degrading abilities with genetic traits thatprovide selective advantages in a given environment (Diaz, 2004). Thepresent strain selection procedures dwell on isolating ‘superbugs’with high resilience to environmental stresses, those harboring cata-bolically superior enzymes, and those species that are not humanpathogens (Singer et al., 2005). Most laboratory strains which arecapable of degrading organic pollutants constitute a fraction ofculturable microorganisms, making only small contributions tobioaugmentation (Watanabe, 2001). Paul et al. (2005) also pointedout that only a fraction of total microbial diversity has been harnessedso far while the genetic resource for degradation of recalcitrant andxenobiotic pollutants is vast.

Bioaugmentation efforts are met with failures more often dueto lesser efficiency, competitiveness and adaptability, relative to theindigenous members of natural communities. For example, the wellknown bacteria capable of degrading PCBs in laboratory culturemedia survived poorly in natural soils, and when these strains wereinoculated to remediate PCB-contaminated soils, the resultantwas thefailure of bioaugmentation (Blasco et al., 1995). Further investigationsrevealed that formation of an antibiotic compound, protoanemonin,from 4-chlorocatechol via the classical 3-oxoadipate pathway by thenative microorganisms was the reason for poor survival of theintroduced specialist PCB-degrading strains (Blasco et al., 1995, 1997).Indeed, bioaugmentation itself is undesirable in all the environmen-tally sensitive locations, especially those protected from the intro-duction of exotic flora or fauna. Scott et al. (2010) proposed a newstrategy of using a free enzyme-based product to remediate waterbodies contaminated with atrazine. The ecological or environmentalissues associated with degrading organisms can be circumvented bythis strategy. The soils do have exoenzymes (cell-free enzymes)which include proteases, and the presence of proteases along withother inhibitors may limit the longevity of free enzymes applied forbioremediation. The cell-free approach can only be used for viableand efficient enzymes that are not dependent on diffusible co-factorssuch as NAD (particularly hydrolases), and cannot be applied in caseswhere the enzyme activity (e.g., most oxygenases) is lost when thecells are broken (Scott et al., 2008). Orica Watercare (Australia) hascommercialized for the first time a free-enzyme for phosphotriesterinsecticides under the trade name LandGuardTM which was proven tobe successful and cost effective. Nevertheless, the technical feasibility

of such strategy needs careful evaluation for many contaminants ortheir mixtures. Immobilizing enzymes on suitable carriers will makethem more stable and resistant to changes in pH, temperature andsubstrate concentrations (Gainfreda and Rao, 2004; Kandelbaueret al., 2004). Other limitations for enzymes include: (a) expensiveproduction costs for pure enzymes, (b) reduced activity due tosorption in soils requiring repeated doses, and (c) the issues withdelivery of enzymes, immobilized enzymes in particular, to come incontact with the pollutant in the contaminated site. Selection ofsuitable carrier materials for immobilizing enzymes will not only helpto increase their longevity but also allow their re-use thus makingthem more cost-effective. Further research into cheap nutrientsources for growing microorganisms may lower production costs ofpure enzymes. Also, more research is required into the mechanismsof delivery of enzymes for their in situ application.

4.3.1. Biosurfactant-producing and pollutant-degrading microbialinoculants: dual role

Most of the biosurfactants are anionic or nonionic; the structure isa characteristic of the microorganism producing the surfactant underthe specific growth conditions (Mulligan and Gibbs, 1993; Zhang andMiller, 1995). Relative to a synthetic surfactant (Tween-80), thebiosurfactant (rhamnolipid) was found to enhance the solubility andthe subsequent degradation of phenanthrene by Sphingomonas sp.(Pei et al., 2010). The biosurfactants can be toxic or even utilizedpreferentially by the pollutant-degrading microorganisms. But, theapplication of biosurfactant-producing and pollutant-degrading mi-croorganisms offers dual advantages of a continuous supply ofbiodegradable surfactant and the ability to degrade pollutant(s)(Moran et al., 2000; Rahman et al., 2002). In a recent report, Hua et al.(2010) demonstrated that a salt-tolerant Enterobacter cloacae mutantcould be used as an agent for bioaugmentation of petroleum- and salt-contaminated soil due to increased K+ accumulation inside andexopolysaccharide level outside the cell membrane.

4.3.2. Genetic engineering of microorganismsMicroorganisms respond differently to various kinds of stresses

and gain fitness in the polluted environment. This process can beaccelerated by applying genetic engineering techniques. The recom-binant DNA and other molecular biological techniques have enabled(i) amplification, disruption, and/or modification of the targetedgenes that encode the enzymes in the metabolic pathways, (ii)minimization of pathway bottlenecks, (iii) enhancement of redox andenergy generation, and (iv) recruiting heterologous genes to give newcharacteristics (Liu et al., 2006; Shimizu, 2002; Timmis and Piper,1999). Various genetic approaches have been developed and used tooptimize the enzymes, metabolic pathways and organisms relevantfor biodegradation (Pieper and Reineke, 2000). New information onthe metabolic routes and bottlenecks of degradation is still accumu-lating, requiring the need to reinforce the available molecular toolbox(Stegmann, 2001). Nevertheless, the introduced genes or enzymes,even in a single modified organism, need to be integrated within theregulatory and metabolic network for proper expression (Cases andLorenzo, 2005).

There are some drawbacks with the field release of geneticallyengineered microorganisms (GEMs), which include the decreasedlevels of fitness and the extra energy demands imposed by thepresence of foreign genetic material in the cells (Saylor and Ripp,2000; Singh et al., 2011). More importantly, there remains a great riskof mobile genetic elements entering the environment and beingacquired by undesirable organisms. The biotechnological innovationsformaking GEMs are numerous. According to Pandey et al. (2005), theadvances such as the programmed cell death based on the principleof killer–anti-killer gene(s) after detoxification can help to develop‘suicidal genetically engineered microorganisms’ (S-GEMs) that canlead to safe and efficient bioremediation. Few GEMs have been used

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for field application because of strict regulations for the release ofGEMs into the environment (Ezezika and Singer, 2010). The only GEMapproved for field testing in the USA for bioremediation wasPseudomonas fluorescens HK44, possessing a naphthalene catabolicplasmid (pUTK21), mutagenized by transposon insertion of lux genes(Ripp et al., 2000). The transition of genetically engineered microor-ganisms from the laboratory to the field environments is hampereddue to the lack of information on the population dynamics ofintroduced genetically engineered microorganisms in the field andpoor physiological control of catabolic gene expression in theengineered organisms under nutrient and other stresses (Cases andLorenzo, 2005). The bioengineering and environmental release ofthose engineered microorganisms has to overcome several obstacleswhich include inconsistencies in risk assessment procedures andpublic health concerns before their effective application in the field.Selecting an indigenous bacterium able to grow rapidly andwithstandthe local stressful conditions for genetic engineering to enhance thebiodegradation capabilities will be more advantageous over otherbacterial strains.We hope, in 5 to 10 years from now, research into thefield release of GEMs will help in designing them for alleviation orprevention of any perceived risks and eventually gaining public andregulatory acceptance in bioremediation of contaminated sites.

5. Bioremediation technologies

Bioremediation technologies based on the principles of biostimu-lation and bioaugmentation include bioventing, land farming, biore-actor, and composting. From these technologies which are at differentstages of development in terms of experimentation and acceptance,the choice of technology option can bemade consideringmany factorswhich include the class of organic contaminants and the cost ofoperation. Sebate et al. (2004) proposed a protocol for biotreatabilityassays in two phases, for the successful application of bioremediationtechnology. In the first phase, the type and metabolic activity ofindigenous microorganisms at the polluted site and the presence ofpossible inhibitors are to be assayed to knowwhether bioremediationitself is appropriate. In the second phase, the influences of nutrients,surfactant, and specialized inocula amendment are to be evaluated inmicrocosms to identify the appropriate treatment for the polluted site.Recently, Bento et al. (2005) reiterated the need for a detailed site-specific characterization studies since the soil properties and theindigenous soil microbial population affect the degree of biodegrada-tion. These conclusions were drawn from a comparative study onnatural attenuation, biostimulation and bioaugmentation on degra-dation of total petroleum hydrocarbons (TPHs) in contaminated soilscollected from Long Beach, California, USA and Hong Kong, China.Among the technological options available for bioremediation, some

of them have been employed at the field sites of USA (Table 1).Improvements in reliability, cost efficiency and speed of remediationcan be achieved by the use of various methods ranging from minimalintervention (bioattenuation), through in situ introduction of nutri-ents and/or bacterial inocula, improvements of physicochemicalconditions or development of novel methods (Romantschuk et al.,2000).

The fate of pollutants is largely influenced by the competingprocesses of degradation and sorption which refers to both adsorp-tion, occurring on surfaces (e.g., between a charged compound andclay) and absorption, i.e., the sorption beyond the surface into aseparate portion defined by the surface (e.g., partitioning into organicmatter). The soil sorption of neutral, and hydrophobic compoundsis dependent on soil and sediment organic content; one of the usefulparameters for describing sorption of neutral, and hydrophobiccompounds is the organic carbon partition coefficient (Koc) that iscorrelated to its Kow (Karickhoff et al., 1979). For successfulbioremediation treatment, the pollutants as substrates must beavailable and accessible either to microorganisms or their extracel-lular enzymes for metabolism to occur. Another important limitingfactor is microbial movement. Because of low bioavailability andaccessibility of pollutants, biphasic (‘hockey stick’) kinetics ofbiodegradation, consisting of an initial period of fast degradation,followed by a second, much slower phase, is commonly observed insoils and sediments during bioremediation (Semple et al., 2004).These constraints drive a constant demand for developing innovativetreatment methods.

5.1. Composting and addition of composted material

Traditionally, the practice of composting is intended to reducevolume and water content of vegetable wastes, to destroy pathogens,and to remove odor-producing compounds. This technology is nowapplied for handling polluted soil or sediments by two chief ways:(i) composting of polluted soils for efficient degradation, and (ii)addition of composted materials. Additions of composted materialwere found to improve degradation of two herbicides, benthiocarb(S-4-chlorobenzyl diethylthiocarbamate) and MCPA (4-chloro-2-methylphenoxyacetic acid) in soil (Duah-Yentumi and Kuwatsuka,1980). Van Gestel et al. (2003) reported that the impact of diesel onthe composting process was negligible when soil was spiked withdiesel oil and mixed with biowaste (vegetable, fruit and gardenwaste) at a 1:10 ratio (fresh weight) and composted in a monitoredcomposting bin system. The spent mushroom waste from Pleurotusostreatus was found to degrade and mineralize DDT in soil (Purnomoet al., 2010). On the contrary, Alvey and Crowley (1995) observed thatadditions of compost suppressed soil mineralization of atrazine

Table 1Field use of selected bioremediation technologies in the United States of America.Adapted from US EPA, 2011.

Technology Site Pollutant

In situ bioremediation of soil Dover Air Force Base, Building 719 site, Delaware TCE (Trichloroethylene), TCA (1,1,1-trichloroethane) cis-DCE(cis-dichloroethylene) (In situ cometabolic bioventing)

In situ bioremediation of water Avco Lycoming Superfund site, Williamsport,Pennsylvania

TCE, cis-DCE vinyl chloride (VC), hexavalent chromium, cadmium

In situ bioremediation Ciba-Geigy, Dover Township, New Jersey VOCs (volatile organic compounds)Composting Dubose Oil Products Co., Cantonment, Florida VOCs, PAHs (polycyclic aromatic hydrocarbons) PCP

(pentachlorophenol)Land treatment Burlington, Northern Superfund site, MN PAHs, SVOCs (semi-VOCs) (treatment with lime, cow manure)Land treatment Scott Lumber Company, Alton, MO PAHs, Benzo(a)pyreneIn situ source treatment US DOE, Savannah River, River site, SC TCE, PCE (polychloroethylene) (nitrogen, phosphorus,

methane addition)Ex situ bioremediation of soil and sludge Southeastern Wood Preserving Superfund site,

Canton, MississippiPAHs (slurry phase bioremediation)

Slurry-Phase bioremediation French Limited Superfund site, Crosby, Texas VOCs, SVOCs, PCBs (Polychlorinated biphenyls) PCP

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relative to rates in unamended soils or in soils amendedwith starch orrice hulls, probably due to the high nitrogen content of the compost.

The critical parameters for composting depend on the type ofcontaminants and waste materials to be used for composting. Thecomposting efficiency essentially depends on temperature and soil/waste amendment ratio as the two important operating parametersfor bioremediation (Antizar-Ladislao et al., 2005). According to Baheriand Meysami (2002), the increase in the bulking agents such as peatmoss, pine wood shavings, bran flakes, or a mixture of these agentsfrom 6 to 12% led to an increase of 4–5% in the biodegradation of totalpetroleum hydrocarbons. In another study, the soil amendment withsludge-only or compost-only in a ratio of 1:0.1, 1:0.3, 1:0.5, and 1:1(soil/amendment, wet weight basis) increased the rates, but highermix ratios did not increase the degradation rates of total petroleumhydrocarbons correspondingly (Namkoong et al., 2002). For theoptimum removal of aged PAH during composting, Guerin (2000)recommended to keep moisture and amendment ratio constant.During the composting-bioremediation, not only the contaminant butalso the waste amendment and the operating conditions willdetermine the rate of biodegradation.

Organic pollutants can be degraded during the first phase of rapiddecomposition during composting. Heat which is generated bymicrobial metabolism is trapped in the compost matrix and most ofthe microbial decomposition and biomass formation occur during thethermophilic stage of composting. The mixing of remediated soil withcontaminated soil can increase the effectiveness of compostingbecause the remediated soil with acclimated microorganisms signif-icantly influences pollutant degradation in the composting process(Hwang et al., 2001). The mineralization may be only a small fractionof pollutant degradation, with other prominent fates being partialdegradation to secondary compounds, volatilization, and adsorptionto compost (Buyuksonmez et al., 1999). In the composting matrices,microorganisms can degrade pollutants into innocuous compounds,transform pollutants into less toxic substances and/or aid in lockingup the chemical pollutants within the organic matrix, therebyreducing pollutant bioavailability. Even in the compost remediationstrategy, the bioavailability and biodegradability of pollutants arethe two most important factors which determine the degradationefficiency (Semple et al., 2001). Cai et al. (2007) showed that theefficiency of composting processes differed among the manuallyturned compost, inoculated manually turned compost, continuouslyaerated compost and intermittently aerated compost for bioremediat-ing sewage sludge contaminated with PAHs, with the intermittentlyaerated compost treatment showing higher removal rate of highmolecular weight PAHs. Composting or the use of compostedmaterials can be applied to the bioremediation of polluted soils.However, the nature of waste or soil organic matter that consists ofhumic materials play an important role in binding of the contami-nants such as PAHs and making them accessible to microbes fordegradation. Plaza et al. (2009) reported that composting will inducesignificant modifications to the structural and chemical properties ofthe humic material fraction including loss of aliphatic materials, anincreased polarity and aromatic polycondensation resulting in adecrease in PAH-binding. Recently, Sayara et al. (2010) demonstratedthat stable composts in municipal solid wastes enhanced biodegra-dation of PAH particularly during the initial phase of composting.Humic material which accumulates with an increase in stability of thecompost is known to act like a surfactant and plays an important rolein releasing PAHs sorbed to the soil. PAH degradation mostly occursduring mesophilic stage of composting, while thermophilic stage isinhibitory for biodegradation (Antizar-Ladislao et al., 2004; Haderleinet al., 2006; Sayara et al., 2009).

Similar to any other technology, composting has both advantagesand limitations. Addition of compost to contaminated soil for bio-remediation makes it a sustainable technology since the biodegrad-able organic waste in the compost is being utilized for beneficial

activity. Also, composting improves the soil structure, nutrient statusand microbial activity. During composting the contaminant can dis-appear via different mechanisms such as mineralization by microbialactivity, transformation to products, volatilization, and formationof nonextractable bound residues with organic matter. The fate ofnonextractable bound residues of contaminants in composting isanother area of interest that requires more research into their release,behavior and risk. One of the critical knowledge gaps of composting islack of sufficient knowledge about microorganisms involved duringvarious stages of composting, the thermophilic stage in particular,which is almost like a blackbox. In fact, there are conflicting viewsabout the role of the thermophilic stage of composting in bioreme-diation of contaminants. Added to this complexity is the fate of boundresidues and whether or not they pose a risk in the future. Knowledgeabout (a) the nature and activity of microorganisms involved invarious stages of composting, and (b) the degree of stability of com-post and its humic matter content will greatly assist in betterdesigning of composting as a bioremediation strategy for contami-nated soils.

5.2. Electrobioremediation

Electrobioremediation as a hybrid technology of bioremediationand electrokinetics for the treatment of hydrophobic organiccompounds is becoming popular. It involves passage through pollutedsoil of a direct current between appropriately distributed electrodesand uses microbiological phenomena for pollutant degradation andelectrokinetic phenomena for the acceleration and orientation oftransport of pollutants (or their derivatives) and the pollutant-degrading microorganisms (Chilingar et al., 1997; Li et al., 2010).The electrokinetics is the use of weak electric fields of about 0.2to 2.0 V cm−1 to soil (Saichek and Reddy, 2005) and the basicphenomena which make up electrokinetic remediation are diffusion,electrolysis, electroosmosis, electrophoresis, and electromigration.Since the present electrokinetic approaches mainly aim at pollutantextraction through transport over large distances, the impact of directcurrent on organism–soil interactions and organism–compound isoften neglected (Wick et al., 2007). Shi et al. (2008) showed thatdirect current (X=1 V cm−1; J=10.2 mA cm−2) as typically used forelectrobioremediationmeasures had no negative effect on the activityof a PAH-degrading soil bacterium (Sphingomonas sp. LB126), andthe DC-exposed cells exhibited up to 60% elevated intracellular ATPlevels and yet remained unaffected on all other levels of cellularintegrity and functionality. Information on the direct reduction oroxidation of the pollutant at the electrode and the changes inmicrobial community due to generation of hydrogen or oxygen at theelectrode is limited.

Luo et al. (2005) developed the non-uniform electrokinetic systemwith periodic polarity-reversal to accelerate the movement and in situbiodegradation of phenol in a sandy loam soil. Although reversing thepolarity of an electric field increased the consumption of electricity,a higher and more uniform removal of phenol from the soil wasobserved. The 2-dimensional (2-D) non-uniform electric field en-hanced the in situ bioremediation process by promoting the masstransfer of organics to degrading bacteria. When tested at bench-scalewith a sandy loam soil and 2,4-dichlorophenol (2,4-DCP) atbidirectional and rotational modes, the 2-D non-uniform electricfield stimulated the desorption and the movement of 2,4-DCP. About73.4% of 2,4-DCP was removed at the bidirectional mode and about34.8% was removed at the rotational mode, which also maintainedremediation uniformity in soil, in 15 days (Fan et al., 2007). In anelectrochemical cell packed with an inert support, the application oflow intensity electric current led to the degradation of hexadecane aswell as higher biomass production by Aspergillus niger (Velasco-Alvarez et al., 2011).

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During electrobioremediation, the transport of PAH-degradingbacteria, Sphingomonas sp. L138 and Mycobacterium frederiksbergenseLB501 from the surface into the subsurface occurred due to elec-troosmosis (Wick et al., 2004). Niqui-Arroyo and Ortega-Calvo (2007)integrated biodegradation and electroosmosis for the enhancedremoval of PAHs from the creosote-polluted soils. The residualconcentrations of total biodegradable PAHs, remaining after biore-mediation in soil slurries, were two-fold lower in electrokineticallypretreated soils than in untreated soils. The remediation rate of in situbioremediation will be otherwise very slow due to limited masstransfer of pollutants to the degrading bacteria. Very recently,Maillacheruvu and Chinchoud (2011) demonstrated synergisticremoval of contaminants by the electrokinetically transported aerobicmicrobial consortium.

There are limitations with electrobioremediation technology thatneed to be overcome, and these include: (i) solubility of the pollutantand its desorption from the soil matrix, (ii) the availability of the righttype of microorganisms at the site of contamination, (iii) the ratiobetween target and nontarget ion concentrations, (iv) requirement ofa conducting pore fluid to mobilize pollutants, (v) heterogeneity oranomalies found at sites, such as large quantities of iron or iron oxides,large rocks or gravel, and (vi) toxic electrode effects on microbialmetabolism or dielectric cell membrane breakdown or changes in thephysicochemical surface properties of microbial cells (Sogorka et al.,1998; Velizarov, 1999; Virkutyte et al., 2002).

5.3. Microbe-assisted phytoremediation

Pollutant effects on plant growth are concentration-dependentand different plant species respond differently. Low doses of pollutantcan increase plant weight while high doses can inhibit, a phenomenonreferred to as ‘hormesis’ (Calabrese and Blain, 2009). In general, plantscan promote dissipation of organic pollutants by immobilization,removal, and promotion of microbial degradation. Some organiccompounds are transported across plant membranes, releasedthrough leaves via evapotranspiration (phytovolatilization) orextracted, transported and accumulated in plant tissues (phytoex-traction) or degraded via enzymatic processes (phytodegradation).Some of the non-volatile compounds are sequestered in planta and areless bioavailable (phytostabilization). Several limitations of bioreme-diation such as the inability of degrading microorganisms to competewith indigenous microflora, insufficient microbial activities at sub-surface, poor support of native as well as pollutant-degrading micro-flora by available or limiting nutrients, heterogeneity of bioavailablecontaminants, and toxic or inhibitory compounds in the pollutantmixture requires the union of phytoremediation and other bioreme-diation strategies (Gerhardt et al., 2009).

Plants have several miles of roots per acre, suggesting the potentialof pollutant degradation in the rhizosphere (Boyajian and Carreira,1997). Sugars, organic acids, and larger organic compounds whichconstitute about 10–50% of plant's photosynthate are deposited insoils (Kumar et al., 2006), and the carbon cycling from CO2

assimilation by plants to root exudation to incorporation to microbialbiomass to microbial respiration takes about just 5 h (Ostle et al.,2003). In the rhizosphere which is dependent on morphology, pro-portion of fine roots, water and nutrient conditions, root exudation,and associatedmicrobial communities, theremay be either promotionor competition between the pollutant degraders and other microbialmembers. Ma et al. (2010) suggested from a meta-analysis thatthe activity of PAH decomposers in soil is more likely to be enhancedby root activities than to be inhibited by other microorganisms inthe rhizosphere, despite the variations due to species, habitats,contamination types and doses. The complex aromatic compoundssuch as flavonoids and coumarins which aid microbial colonizationof roots are structurally similar to PCBs, PAHs and PHC, providingopportunities as the analogue-enrichment for stimulating degradative

pathways in microorganisms (Holden and Firestone, 1997). Rhizor-emediation, an integral component of phytoremediation can occurnaturally or can be triggered by introducing specific pollutant-degrading microbes or plant growth promoting microorganisms(Gerhardt et al., 2009). Since the root depth of herbaceous plantsvaries from plant to plant, from soil to soil, and season to season, thepresence of contaminants in soils which is deeper than the root zoneof plants requires excavation, other agronomic practices or selectionof trees with deeper roots. Nevertheless, most of the recalcitrantorganic contaminants are typically found in the top few cm of thesoil. Dendroremediation, which is a type of phytoremediation usingtrees may be useful in attenuating certain pollutants such as 2,4,6-trinitrotoluene and trichloroethylene from soil and groundwater(Susarla et al., 2002).

Plants produce many secondary plant metabolites (SPMEs)which include allelopathic chemicals, root exudates, phytohormones/phytoalexins, phytosiderophores, and phytoanticipins and are derivedfrom isoprenoid, phenylpropanoid, alkaloid or fatty acid/polyketidepathways (Hadacek, 2002). Singer et al. (2004) argued that SPMEsare pollutant analogues within the network of suprametabolism,having implications for predicting the fate of pollutants. Gilbert andCrowley (1998), and Kim et al. (2003) showed that SPMEs such aslimonene, cymene, carvone and pinene enhanced degradation ofPCBs. Pseudomonas putida PCL1444, isolated from the rhizosphere ofLolium multiflorum cv. Barmultra when grown in PAH-polluted soildegraded the PAHs and protected the plant from the pollutant, byefficient utilization of root exudates for growth and high transcriptionof naphthalene catabolic genes (Kupier et al., 2002). Narasimhan et al.(2003) applied the rhizosphere metabolomics-driven approach,which has been referred to profiling of root exudates for identificationof targeted compounds for creating the nutritional bias, to degradePCBs (2Cl-biphenyl, 4Cl-biphenyl and Aroclor 1254 at 53 μM) in therhizosphere of Arabidopsis. The growth of gfp-tagged Pseudomonasputida PML2 was increased due to the exudation of SPMEs such asphenylpropanoids and consequently PCB degradation was enhanced.The rhizosphere metabolomics-driven approach will become an im-portant tool for engineering phytoremediation systems.

The activity and the numbers of the pollutant-degrading endo-phytes are both plant- and contaminant-dependent (Siciliano et al.,2001). Contaminants such as TCE and methyl tert-butyl ether whichare routinely assimilated in the transpiration pathways of plantsmay be degraded effectively by the pollutant-degrading endophytes.Methylobacterium sp. strain BJ001, a phytosymbiotic bacteriumisolated from tissue culture plantlets of Populus deltoides×nigraDN34 was found to transform 2,4,6-trinitrotoluene and mineralizehexahydro-1,3,5-trinitro-1,3,5-triazine and octahydro-1,3,5,7-tetra-nitro-1,3,5-tetrazocine to CO2 (Van Aken et al., 2004). Barac et al.(2004) demonstrated that the engineered endophyte (Burkholderiacepacia strain L.S.2.4 containing the toluene-degrading plasmid,pTOM), when applied to surface-sterilized yellow lupine seeds lednot only to the protection against the phytotoxic effects of toluenebut also decreased emissions from the transpiration stream of itshost. The pollutant-degrading endophytes are relatively free fromthe competition for nutrients and water among the colonizers in therhizosphere. Greater opportunities for employing the endophyte-assisted phytoremediation, either through naturally-occurring orengineered endophytes exist, especially for the mobile pollutants.Phytostimulation of pollutant degradation by microorganisms inthe rhizosphere or inside the plants can offer many economic andenvironmental advantages compared to the conventional strategiesemployed in biostimulation. But, the disadvantages include hydro-phobicity and chemical stability of pollutants that influence thephytostabilization and the rates of degradation by the associatedmicroorganisms (Van Aken et al., 2010), and plant root exudationwhich modifies the structure and activities of pollutant-degradingmicroorganisms (Corgie et al., 2004). Besides, phytoremediation in

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the field is also challenged by many obstacles which include theinability to mitigate plant stress factors and non-availability ofsuitable methods for the assessment of phytoremediation (Gerhardtet al., 2009).

6. Bioremediation monitoring and efficacy testing: relevance

Monitoring and efficacy testing for bioremediation are essential forthe purposes of efficiency and economics. There is a strong need totest the efficacy. The ‘conservative biomarkers’, the internal markerssuch as dimethyl chrysene which are recalcitrant can be used totest the efficacy of bioremediation (Huang et al., 2005). Theconcentration of an individual pollutant can be normalized to theinternal marker and the relative ratio of a specific pollutant tothe internal marker should decrease during the remediation process.Indices such as the carbon preference index, average chain length andvarious n-alkane/acyclic isoprenoid ratios which are used for thechemical fingerprinting in the environmental forensics can be appliedto distinguish the plant- or microbe-derived hydrocarbons from thehydrocarbons of petrogenic or anthropogenic origin. The epicuticularwaxes derived from leaf cuticles are generally abundant with odd-numbered n-alkane peaks in the range of C25–C31, while the even-numbered carbon compounds are abundant in petroleum. The carbonpreference indices indicating the ratio of odd-numbered to even-numbered carbon compounds provide information on the predom-inance of phytogenic or petrogenic hydrocarbon contamination (Jeng,2006).

Several microbiological methods are currently employed for thegeneral soil quality assessment. The monitoring and efficacy testingfor bioremediation require a careful selection from them, besidesusing specific information on abundance of microbial members orgenes, and microbial processes and activities since factors such aswater content, temperature and many others determine the courseof attenuation (Table 2). The global regulatory networks in which setsof operons, scattered on the bacterial genome and representingdisparate functions such as response to nutrient starvation are co-ordinately controlled in microorganisms (Gottesman, 1984). Thesignal transduction and effector proteins which are involved in thenitrogen regulation (Reitzer, 2003) or the involvement of cra, crp andrelA/spoT modulons and the accumulated levels of alarmone guano-sine 3′,5′-bis(diphosphate) and cAMP (Hardiman et al., 2007) can bethe basis of a means to monitor changes in nutrient limitation ofmicrobial response during the bioremediation process. Many func-tional genes namely, nahAc, alkB and xylE which are involved in thedegradation of naphthalenes, n-alkanes and toluene, respectively areknown from the cultured microorganisms. With optimized assays, thefunctional gene abundances which seem to reflect the type as well asthe actual degradation rates can be used to assess the efficacy ofbioremediation (Salminen et al., 2008). Recently, Kao et al. (2010)used the culture-based method, real-time polymerase chain reaction

of genes such as phenol hydroxylase, ring-hydroxylating toluenemonooxygenase, naphthalene dioxygenase, toluene monooxygenase,toluene dioxygenase and biphenyl dioxygenase, and denaturinggradient gel electrophoresis fingerprinting analysis for microbialcommunities to evaluate the effectiveness of bioremediation of apetroleum contaminated site. Compound specific carbon isotope (CSI)analysis has emerged recently as a powerful tool to quantify and/ordistinguish biodegradation from other abiotic processes such assorption, volatilization etc. of contaminants like chlorinated solvents(TCE, PCE, DCE) and aromatic hydrocarbons (benzene, toluene,xylene, ethyl benzene, naphthalene, etc.) and to confirm intrinsicbiodegradation during natural attenuation process in the contami-nated aquifers (Fischer et al., 2007, 2008; Hunkeler et al., 2005;Meckenstock et al., 2004). In stable carbon isotope analysis, the lighterisotope is preferentially utilized by microorganisms leaving behindthe heavier isotope thereby resulting in a distinct fractionationpattern among 12C and 13C.

Toxicity testing should be an integral part of the bioremediationprogram since a reduction in toxicity is a necessary characteristic ofbioremediation process. The toxicity of a pollutant to microorganismsis also regarded as a direct measure of bioavailability (Ronday et al.,1997). Megharaj et al. (2000) suggested that chemical analysis inconjunction with bioassays were necessary for toxicological estima-tions. Despite the importance of toxicological assays, only fewbioremediation studies have attempted to include one or two suchassays. The toxicity of fuel spills followed by bioremediation treat-ment was assessed by Microtox measurements, seed germination andplant growth assays (Leung et al., 1997; Wang and Bartha, 1990).Although the standardized toxicity test system such as Microtoxwhich employs Vibrio fischeri, a bioluminescent marine bacterium,has certain advantages, the ecological relevance of toxicity tests canbe improved by use of ecologically relevant (aquatic or terrestrial)representatives from different trophic levels. The efficacy of biore-mediation (bioaugmentation with Pseudomonas sp. strain ADP orbiostimulation with citrate) of atrazine-contaminated soils was testedby ecotoxicological endpoints such as plant biomass production,earthworm reproduction, microalgae growth, and cladoceran repro-duction (Chelinho et al., 2010). Since no single organism is con-sistently sensitive to all pollutants, it is pertinent to include a batteryof bioassays, by involving members from different trophic levels ofthe food chain. Every bioremediation technology thus requires theuse of experimental controls and performance indicators (Table 3)for both process optimization and implementation of regulatorydecisions.

7. Conclusions

Environmental effects of many chemical substances could not beanticipated earlier, which is exemplified by the ‘hero to villain’ statusof DDT (2,2-bis (p-chlorophenyl)-1,1,1-trichloroethane) (Beard,

Table 2Factors affecting bioremediation of organic pollutants in soil and aquatic systems.Adapted from Michels et al., 2000.

Factor Effect

Water content Transport of pollutants and the degraded products; degradation of pollutantsTemperature Composition of communities (all the interacting organisms living together in a specific habitat)

and velocity of degradation; pollutants persist longer at lower temperaturepH Microorganisms and enzymes exhibit pH-dependent activity maximaRedox potential Concentrations and ratios of electron donors/acceptors determine pathways and efficiency of degradationSolubility, volatility, particle size, sorption, occlusion Bioavailability is determinedOrganic matter Influence degradation and sorption/entrapmentNutrients Growth and reproduction of microorganismsAuxiliary (co-)substrates Enable co-metabolic transformation of contaminantsCo-contaminants Influence bioavailability and enhance/inhibit biodegradationMicrobial communities Pollutant-tolerant members within communities determine the rate of degradation

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2006). Considerable researches on chemical pollutants now providethe necessary body of knowledge to understand their recalcitranceand toxic nature. With this useful information, the policy makers haveto decide whether remediation is necessary and practical (Alcocket al., 2011). Regulations that limit the disposal of chemicals andescalation in the costs of physical and chemical treatments makebioremediation technologies more attractive. Wider usage of biore-mediation technologies desires new governmental regulations whichinclude risk-based criteria in cleanup treatments. Each strategy ofbioremediation process has certain specific advantages and disad-vantages, which need to be considered for each situation. Factorswhich limit the efficiency of microbial degradation of organicpollutants are numerous. Besides the bioavailability of the pollutantitself, low temperature, anaerobic conditions, low levels of nutrientsand co-substrates, the presence of toxic substances, and the physio-logical potential of microorganisms are particularly important inthe polluted sites (Romantschuk et al., 2000). The biological re-sponse to environmental pollutants varies within a microbial guild(Ramakrishnan et al., 2010), and the presence of co-contaminants canelicit variable responses (Ramakrishnan et al., 2011). The choice ofmethods in each technology requires careful consideration. What isnow important is to gain a better understanding on the metaboliccooperation among the microbial communities. The studies on thestructure and functions of microbial communities in the pollutedsites on different spatial and temporal scales and their responses todifferent stimuli using community fingerprinting and environmentalgenomics techniques can show the way. It is impractical to restore allnatural functions of some polluted soils (multifunctional remedia-tion). Hence, the application of the principle of function-directedremediation may be sufficient enough to minimize the risks due topersistence and further spreading of pollutants.

Acknowledgements

BR and KV thank the Government of Australia (Department ofEducation, Employment and Workplace Relations) for the EndeavourResearch Fellowship and Endeavour Executive Award, respectively.

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