bioremediation towards a credible technology

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Microbiology (1998), 144, 5 9 9 4 0 8 Printed in Great Britain

Bioremediation: owards a credible technologyARTICLE Ian M. Head

Tel: +44 191 222 7024. Fax: +44 191 222 5431. e-mail: [email protected]

Fossil Fuels and Environmental Geochemistry (Postgraduate Institute), Drummond Building,University of Newcastle, Newcastle upon Tyne NE1 7RU, UK

Keywords: bioremediation, bioavailability

ContextBioremediation is the technological process wherebybiological systems are harnessed to effect the clean-up ofenvironmental pollutants. Currently, microbial systemsare most widely employed in bioremediation pro-grammes, generally in the treatment of soils and waterscontaminated with organic pollutants. Micro-organismshave a huge metabolic repertoire that enables them todegrade a panoply of organic pollutants and in manycases the complex biochemistry and molecular biologyof the catabolic pathways involved have been unravelled(e.g. Gibson, 1984;Frantz et al., 1987;Evans & Fuchs,1988;Burlage et al., 1989;Abramowicz, 1990;Assinder8c Williams, 1990;Chaudhry & Chapalamadugu, 1991;Cerniglia, 1992; Knackmuss, 1996). Despite valuablebasic knowledge on the mechanisms of pollutant bio-degradation, bioremediation has yet to be accepted as aroutine treatment technology and the environmentalindustry is wary of applying bioremediation for thetreatment of contaminated sites.Bioremediation has the potential to treat contaminantson site with relatively little disturbance to the con-taminated matrix. Furthermore, micro-organisms offerthe possibility that organic pollutants can be completelymineralized to inorganic materials (e.g. CO,, H,O, C1-,NO,), making bioremediation an attractive treatmentstrategy. In contrast, removal of contaminated materialto landfill sites, or extraction of contaminants usingphysical processes such as soil washing, do not destroythe contaminants present, but simply concentrate thecontaminated material in a different location. Why then,if bioremediation offers benefits over other technologies,has it not been more widely adopted to treat en-vironmental contamination ? One reason is that physicaltreatments are rapid and their outcome is generallypredictable in the short term; they are also relativelyinexpensive. Bioremediation too, can be relativelycheap, but methods to confirm efficacy on a field scalehave either been unavailable or simply not appliedroutinely. The unpredictability of bioremediation,meanwhile, stems from a lack of understanding of thebehaviour of microbial populations in natural environ-

ments and how physical, biological and chemical factorsinteract to control their activity against environmentalpollutants. With this realization, has come a differentphilosophy towards bioremediation ; he focus of bio-remediation research has shifted from isolation andconstruction of 'superbugs ' to determining the factorsthat limit pollutant transformations and mineralizationin natural environments. It is now clear that access ofmicro-organisms to pollutants in situ is a critical factorin determining the success of bioremediation. Conse-quently, methods are being developed to enable pre-dictions regarding the feasibility of bioremediationbased on pollutant bioavailability and biodegradation.With knowledge of the causes of unsuccessful bio-remediation, methods can be formulated to overcomethese limitations. Assessing the feasibility of bioremedi-ation within a predictive framework and confirming itsefficacy in the field is the focus of this review.Assessing bioremediation potential:problem of bioavailabilityIt is relatively simple to demonstrate biodegradation ofspecific compounds in a contaminated environmentalsample. Spiking a sample of contaminated soil with apollutant chemical of interest and monitoring its lossand the appearance of degradation end products incomparison with sterilized control samples often de-monstrates rapid biodegradation. Conversion of 14C-labelled compounds to 14C0, is undoubtedly the bestevidence that a microbial community has the ability tomineralize an organic compound, but is insufficient todemonstrate convincingly that bioremediation will besuccessful. For example, added naphthalene and phen-anthrene were rapidly degraded in a polycyclic aromatichydrocarbon (PAH)-contaminated soil, but levels ofindigenous naphthalene and phenanthrene were notreduced appreciably (Ericksonet al., 1993).Comparableobservations have been made in other soils and withdifferent chemicals (Steinberg et al., 1987; Doelman etal., 1990; Weissenfels et al., 1992; Beurskens et al.,1993). Reduced bioavailability results from the inter-action of pollutants with both organic and inorganic

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components of the soil matrix. Access of micro-organ-isms or their enzymes to the pollutant molecules isconstrained and with increasing contact time ( ageing)the proportion of the pollutant that becomes biologicallyunavailable increases. A number of mechanisms areinvolved: diffusion limitation due to sequestration of thepollutant in micropores (


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soils containing aged pollutants and inoculated withspecific pollutant-degrading pure cultures. Recovery ofatrazine with 1 :1 methanol :water for example, wassimilar to the degree of bacterial mineralization of theherbicide, whereas n-butanol extraction efficiency mir-rored the level of phenanthrene mineralization. Thispreliminary study was limited in the scope of pollutantsand soil tested and it is likely that empirical deter-mination of the appropriate extractant for any particularsoil/pollutant system will be required. Nonetheless,because the methodology is straightforward it couldprove an attractive means to assess bioavailability.Mode//ingapproaches. The above methods may be of usein the empirical determination of bioavailability, but donot by themselves allow the likely success of bio-remediation to be determined. Potentially more valuablein this respect is the development of mathematicalmodels that can explain biodegradation behaviour inrelation to both biological and abiotic factors such aspollutant sorption and diffusion.Rigorous kinetic and thermodynamic analyses haverecently been applied to the problem of desorption ofpollutants from soil (Cornelissen et al., 1997) and theeffect of desorption on biodegradation rates (Bosma etal., 1997). The desorption of chlorobenzenes, poly-chlorinated biphenyls (PCBs) and PAHs was studied inlaboratory-contaminated and field-aged sediments (Cor-nelissen et al., 1997).The pollutants were found to occurin three 'compartments '.The pollutants desorbed fromone of the compartments rapidly, from a second slowlyand from a third very slowly. From 8 to 54% of thepollutant was present in the slowly desorbing fraction,with the exact value depending on the particularchemical and sediment. Up to 47%, though moretypically 0.5-16 Yo , of some contaminants was found tobe present in the very slowly desorbing compartment(Cornelissen et al., 1997). The latter can equate tosignificant quantities of pollutant in a contaminated soiland its occurrence in this form has considerable impli-cations for the time that may be required to effectsatisfactory bioremediation. The very slow desorptionphase could only be measured in reasonable time scalesif desorption was carried out at elevated temperatures,and values obtained for the fraction of very slowlydesorbed material at 20 "C and 60 "C were generally ingood agreement. This approach offers the possibilitythat not only the amount of pollutant in differentiallydesorbing pools can be determined, but also the timetaken for complete desorption to occur under fieldconditions can be estimated. This may be invaluable indetermining the length of time required for bioremedi-ation if it is limited by mass transfer of pollutants froma sorbed phase to solution phase, as is generally the casewith aged pollutant residues (Bosma et al., 1997;Carmichael et al., 1997).The biodegradation of a contaminant insitu is a functionof both the catabolic activity of the micro-organismspresent and transport of the contaminant to microbialcells with the ability to degrade the contaminant. It is the

latter that is probably the most important factor indetermining the successof a bioremediation programme.By considering both of these factors, a parameter termedthe bioavailability number (Bn) as been derived (Bosmaet al., 1997).The Bn is equal to a first-order exchangeconstant k (k ontrols the rate of desorption/diffusion ofa sorbed contaminant) , divided by the product of themaximal transformation rate (qmax) and the reciprocalof the half-saturation constant for contaminant trans-formation (K;') (equation 1).This is essentially the ratioof the exchange constant and the first-order rateconstant applicable to Michaelis-Menten kinetics(equation 2 ) at substrate concentrations much less thanK , (equation 3) .Bn = k/qmaXK;' (1)4 = qmax [Sl/(Km+ [Sl) ( 2 )When [S] is much lower than K , this approximates to(3 ) with qmaX/Kmas a first-order rate constant.4 = (qmax/Km) [SI (3 )The Bn is hence a measure of the importance of masstransfer relative to the intrinsic catabolic activity of themicrobial cell or population. Thus, for values of Bn 1,the reverse is true and the rate of transformation is notlimited by desorption and diffusion. In other words, ifmass transfer (k) s large relative to biodegradation(qmax/Km) bioavailability is not likely to limit bio-degradation. The value of k can be calculated fordifferent modes of mass transfer (e.g. linear diffusion vsradial diffusion vs dissolution) using experimentallydetermined values for distribution coefficients anddissolution rate constants (Rijnaarts et al., 1990;Bosmaet al., 1997).Values of qmaxand K , can be determinedby measurement of initial biotransformation rates atdifferent added substrate concentrations in soil slurries.The parameters can then be determined graphically (e.g.using a Lineweaver-Burk-type plot).This approach was used recently to investigate theimportance of mass transport and hence bioavailabilityon a-hexachlorocyclohexane (a-HCH) degradation insoil slurries (Bosma e t al., 1997).The exchange constantwas calculated from a-HCH desorption data using afirst-order model (Rijnaarts et al., 1990) and values forKm and qmaxwere determined using Lineweaver-Burkplots (Bachmann et al., 1988). The Bn for this systemwas calculated at 00164-030 under different mixingregimens that affected mass transfer. Both values werehowever less than one and hence biodegradation waslikely to be sorption limited. When the biodegradationprocess was modelled using an expression that takesaccount of both mass transfer and biodegradation, thefit with experimental data was extremely good (Bosmaet al., 1997).This relatively simple method may thereforeprove useful in predicting the applicability of bio-remediation if it can be shown to be valid over a widerange of conditions.The same authors also developed an expression relating

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the exchange coefficient and the quantity of pollutanttransformed to satisfy cell maintenance requirements tothe threshold concentration below which no furtherbiodegradation would be observed. High and low valuesfor substrate fluxes required to satisfy maintenanceenergy were estimated at pg s-l for copiotrophsand pg s-l for oligotrophic bacteria, using pub-lished data on the maintenance energy requirements fora range of bacteria (e.g. Chesboro et al., 1979; Bouillotet al., 1990; Tros et al., 1996a, b). Exchange coefficientsrepresenting a range of mass transfer conditions fromthe diffusion of hydrophilic compounds in water to thediffusion of hydrophobic compounds in soils were alsocalculated. With this information, threshold pollutantconcentrations expected under different mass transferconditions and with bacterial populations with variousmaintenance energy requirements were determined(Bosma et al., 1997). The range from low maintenanceenergy systems with high mass transfer to high main-tenance energy systems with low mass transfer exhibitedthreshold concentrations that varied from pg 1-' togreater than lo' pg l-l, respectively. It was apparentfrom this analysis that under certain conditions residualconcentrations of pollutant, even in microbiologicallyactive, environments, can be very high and this isconsistent with the levels of residual pollutant oftenmeasured in the field.The fundamental importance of bioavailability in deter-mining the success of bioremediation is becoming morewidely appreciated (e.g. Tabak et al., 1994;Bosma et al.,1997).The development of methods that can potentiallypredict the extent of bioavailability based on relativelysimple measurements and calculations is likely to have asignificant impact on decisions to bioremediate. Cer-tainly, more research is required to determine if pre-dictions made from kinetic and thermodynamic data arewidely applicable and to extend their utility to complexmixtures. As we increase our understanding of themechanisms involved, it is likely that predictions willbecome more reliable and the efficacy, and hence thereputation, of bioremediation will undoubtedly beimproved. Continued isolation of particular organismswith the ability to catabolize specific pollutant chemicalsmay yet reveal much regarding the biochemistry andgenetics of biological transformations of novel xeno-biotics and may even prove useful in 'end-of-pipe'processes for the treatment of well-defined waste-streams. However, improvement of the efficacy ofbioremediation of contaminated matrices in the en-vironment will require a redistribution of research effort,away from such studies and towards a greater under-standing of the limitations on microbial transformationsin situ.Can bioavailability limitation be overcome?The prospects for predictable evaluation of bioremedi-ation efficacy are improving, but it is not clear that thisknowledge will allow development of more effectivemeans to treat organic pollutants biologically in con-taminated soils and sediments.

A number of strategies for improving mass transfer havebeen suggested. Diffusion and desorption of pollutantsare temperature-dependent (Cornelissen et al., 1997)and this is the basis of physico-chemical treatments suchas thermal desorption and steam-stripping. In thecontext of bioremediation it has been suggested that theuse of composting systems, where elevated temperaturesare maintained, should improve mass transfer and hencebioremediation rates (Pignatello & Xing, 1996); apiddegradation of chlorophenols in contaminated soil bycomposting has been demonstrated accordingly (Laine& Jarrgensen, 1996; Jaspers et al., 1997). Physicalgrinding or mixing to disaggregate soils has beensuggested as a means of alleviating mass transferlimitation, and there is some evidence that this iseffective (e.g. Rijnaarts et al., 1990).The cost involved inlarge-scale application of physical disruption to soils tothe degree required to improve mass transfer substan-tially may, however, be prohibitive. The most widelyinvestigated method for improving bioavailability ofsorbed pollutants is treatment with surfactants. Sur-factant-enhanced biodegradation of sorbed and poorlysoluble chemicals is probably due to decreasing thedistribution coefficient (K,) for the pollutant in a soil-water system, i.e. increasing the equilibrium concen-trat ion in the aqueous phase relative to the sorbed phase.This would require the surfactant to be provided at aconcentration greater than the critical micelle concen-tration (CMC; Alexander, 1994; Deitsch & Smith,1995), unfortunately requiring the use of substantialamounts of surfactant that can be both costly and toxicto micro-organisms. There is however evidence thatmineralization rates can be increased even at surfactantconcentrations well below the CMC (Aronstein et al.,1991). Recent evidence suggests that although highsurfactant concentrations do lower K,, concentrationsbelow the CMC increase the mass transfer exchangecoefficient (equivalent to k in equation 1; Deitsch &Smith, 1996). In essence, this means that in abioticsystems the equilibrium concentration in the aqueousphase does not change, but the time for equilibrium tobe reached is reduced. Consequently, enhanced rates ofmineralization can be observed at surfactant concen-trations below the CMC despite no increase in theaqueous equilibrium concentration being noted inabiotic control systems.In a number of cases however, surfactant addition hashad no effect or been detrimental (see Liu et al., 1995 fora summary) due to toxicity of the surfactant at highconcentrations (e.g. Aronstein et al., 1991) or retar-dation of surface attachment of bacterial cells to thehydrophobic contaminant (e.g. Efroymson & Alex-ander, 1991).Furthermore, the identity of the surfactantand its concentration are critical. One study comparedthe effects of seven surfactants (anionic and non-ionic)on desorption of phenanthrene, of which five had noeffect on desorption of the PAH from sterile soil(Aronstein et al., 1991). The effect of the remaining twonon-ionic surfactants on the mineralization of phen-anthrene and biphenyl was quite different in mineral and

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17a(H)21P(H)-hopane

organic-rich soils. It is interesting to note tha t one of thesurfactants, Alfonic 810-60, did not increase the ap-parent aqueous concentration of phenanthrene at equi-librium in sterile slurries of the mineral soil. Despite this,mineralization was markedly enhanced, supporting theview that it is the mass transfer coefficient and not K ,that is affected by surfactant addition.On the basis of pollutant- and soil-specific effects(Aronstein et al., 1991; Providenti et al., 1995), it seemslikely that a universally applicable surfactant treatmentto promote the biodegradation of poorly bioavailablecontaminants will be elusive. Consequently, surfactant-enhanced bioremediation is likely to rely on ad hocrather than proprietary solutions. A need for case bycase evaluation has considerable cost implications forbioremediation and the formulation of generic treatmentpractices would reduce the unit cost. The developmentof reliable, widely applicable practices will be essential ifbioremediation is to compete succesfully with engin-eering solutions. In order to achieve this, it is importantto understand processes rather than rely on empiricalobservations, and the causes of reduced bioavailabilityare now being rigorously investigated.Field demonstration of the efficacy ofbioremediationOne of the greatest challenges faced by advocates ofbioremediation is proof that a chosen treatment iseffective under field conditions. Concentration of con-taminant compounds and residue toxicity are the mostcommon criteria specified by legislative bodies to definean end point for successful bioremediation. Chemicalanalyses (e.g. GC-MS) and toxicological assays (e.g.Microtox) can be applied to accurately identify andquantify organic contaminants or assess residual toxicityfollowing treatment. Nevertheless, a major factorhindering the acquisition of statistically valid proof ofbioremediation efficacy in the field is the heterogeneousdistribution of pollutants. For example, simple gravi-metric analysis of oil residues from the Exxon Valdez oilspill was inadequate for the determination of hydro-carbon bioremediation. The variability in oil loading onbeach sediments, determined from replicate samples,was so great (SEM as high as 224YO;Bragg et al., 1992)that valid conclusions could not be reached. Even aftersieving a contaminated soil from a derelict coal ex-traction plant, the variation in GC-MS analyses ofcertain PAHs in replicate samples can be greater than100% (Mundell, 1994). To overcome this problem anumber of strategies have been devised that either aloneor in combination may provide the body of evidencerequired to demonstrate that bioremediation has beeneffective.Poorly biodegradable components of complexmixturesThe overall biodegradation of contaminants that com-prise a complex mixture of compounds with differentsusceptibilities to biodegradation can be assessed by

*.O r3 7.5xC0r3 7.0nUE= 6.5

Fig, 1. Changes in the ratio of total GC-detectablehydrocarbons (TGCDHC) to hopane during the bioremediationof the Exxon Valdez oil spill. The solid black line representsdata for fertilized beach sediments and the dashed l ine data foruntreated beach material. The grey areas represent the 95%confidence intervals. Inset, structure o f 17a(H)Zlj?(H)-hopane, apoorly biodegradable component of crude oil that has beenused to index the degradation of more labile components ofcrude oil. Redrawn from Bragge t a / . (1992).

measuring the ratio of the degradable components to apoorly degradable component in the mixture. Crude oilis an excellent example of such a contaminant mixtureand this approach has been used successfully to confirmbiodegradation of weathered oil spilled from the ExxonValdez (Bragg e t al., 1994). Crude oil biodegradationstudies have commonly used the ratio of the linearalkane n-heptadecane (n-C,,) to the more resistantbranched alkane pristane (or alternatively n-octa-decane: phytane ratios) to evaluate the extent of biodeg-radation (e.g. Fayad et al., 1992; Pritchard & Costa,1991). A decrease in the ratio of the n-alkane to thebranched alkane is taken as evidence of crude oilbiodegradation. However studies of beach sedimentsfrom Prince William Sound, Alaska, demonstrated thateven the branched alkanes were readily degraded,resulting in anomalously high n-C,, :pristane ratios inhighly degraded oils (Bragg et al., 1992). For this reason,17a(H)21j?(H)-hopane, minor component of the NorthSlope crude spilled at Prince William Sound, was used toindex the biodegradation of the more labile componentsof the oil (Bragget al., 1994). 17a(H)21p(H)-hopanes apentacyclic triterpane (Fig. 1) and was chosen as aconserved internal marker because is known to bepoorly biodegradable. Hopanoids are membrane lipidsfound in many bacterial taxa (Rohmer et al., 1984) butthey exhibit stereochemical differences to those oc-curring in crude oil and can be distinguished readily.Comparisons of fertilizer-treated beach plots with un-treated plots using the ratio of either the total GC-detectable hydrocarbons, total PAH or total resolvablehydrocarbons to the hopane concentration providedconvincing evidence that bioremediation had beeneffective (Fig. 1;Bragg et al., 1994).

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A,, , ,Succinyl product9

COOH

Succinyl product6OOH

Fumaryl productCH3Xylene

In3 coon................................................................................ * ................................................................Fig. 2. Succinyl and fumaryl products from the anaerobiccatabolism of substituted benzenes. These compounds havebeen used as indicators of in situ biodegradation of BTEX inground water.

Whilst assessment of crude oil biodegradation has beenthe main application of this approach to date, it can beapplied to other complex mixtures. PCBs, for example,are normally found as a mixture of congeners withsimilar physico-chemical properties but different bio-logical fates and are therefore amenable to this ap-proach. 3,4-3',4'-tetrachlorobiphenyl is more readilybiodegradable than 2,3,6-3',4'-pentachlorobiphenyl andthe ratio of the two congeners has been used to monitorbioremediation of PCBs in river sediments (Harkness etal., 1993). There is potential for 2,3,6-3',4'-pentachloro-biphenyl to be transformed by 3-, 3'- and 4'-dehalo-genation reactions, compromising its use as a conservedtracer. This is because anaerobic dechlorination hasbeen shown to result in preferential dehalogenation atmeta- and para-positions in PCB molecules (Bedard &May, 1996). However, independent studies have demon-strated that the meta and para chlorines of the 2,3,6-3',4'-substituted congener are not susceptible to re-ductive dehalogenation in anoxic sediments (Bedard etal., 1996).

Analysis of degradation productsAnalysis of metabolic products or intermediates ofcontaminant metabolism in situ is not universallyapplicable as many intermediates may be short-lived oronly accumulate to very low levels. Some features ofmetabolic products applicable in the diagnosis ofsuccessful bioremediation are : an unequivocal bio-chemical relationship with the parent compound ; no

exogenous sources of contamination ; and biologicaland chemical stability under in situ conditions (Beller etal., 1995). Degradation products of trichloroethylene(TCE) metabolism meet some of these criteria in thatbiological degradation produces cis-1,2-dichloroethyl-ene whereas abiotic TCE transformation results in theformation of 1,l-dichloroethylene (Kastner, 1991).Anaerobic biodegradation of the alkyl substituted com-ponents of BTEX (benzene, toluene, ethylbenzene,-ylenes) is also known to produce characteristic by-products under denitrifying and sulphate reducingconditions (Beller et al., 1992; Evans e t al., 1992) andthis has been exploited recently under field conditions(Beller et al., 1995). Benzylsuccinate, benzylfumarateand corresponding homologues accumulate during an-aerobic degradation of BTEX in laboratory incubations(Fig. 2).These compounds could be detected in samplesfrom an anaerobic aquifer contaminated with gasoline,implying that anaerobic transformations of the BTEXhad occurred in the aquifer (Beller et al., 1995).Furthermore, in an aquifer spiked with BTEX-con-taminated water (containing bromide as a conservativetracer) it was possible to demonstrate that the ac-cumulation of the metabolic products followed thedisappearance of the corresponding contaminant hydro-carbons. When BTEX became undetectable, the con-centration of the metabolic products began to decrease.This approach is not fail-safe in determining bio-degradation at field sites, as work on pentachlorophenoldegradation has illustrated. Both photolytic (Steiert &Crawford, 1986) and biodegradative reactions (Apaja-lahti & Salkinoja-Salonen, 1987) produce the sameintermediates. It may thus be impossible to distinguishthese reaction mechanisms under field conditions, andunequivocally associate disappearance of pentachloro-phenol with a biological process, simply by detection ofputative metabolic products.Geochemical indicatorsof bioremediationExcept in fermentation and disproportion reactions, theoxidation of organic compounds is coupled to thereduction of exogenous electron acceptors. In soil andsediment environments these are usually O,, NO,, Fe3+,Mn4+, SO:- or CO,. Depletion of these species oraccumulation of their reduced products in contaminatedrelative to uncontaminated material may be indicativeof organic pollutant metabolism (Borden et al., 1995).This can be achieved most convincingly if the concen-trations are compared with levels of a conservativetracer (e.g. chloride or bromide) that has similarphysico-chemical properties to the analyte of interest,but is not susceptible to biological reduction. Indirectevidence of this nature is far from definitive unlessreduction in contaminant concentrations can be relatedto changes in oxidant concentrations and a mass balanceconstructed. In the field, these objectives are oftendifficult to meet and geochemical analysis of redox-sensitive species at best provides supplementary evidencefor biodegradation.

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BioremediationStable isotopesA rapid method that is used to monitor the progress ofbioremediation is measurement of CO, production.While this gives an indication of increased rates ofbreakdown of organic matter it does not necessarilyfollow that the CO, originates from the target pollutant.The advantage of using stable isotope analysis is that itis possible to determine the source of CO, evolvedduring a remediation process. This is feasible becausethere is very little isotopic fractionation associated withthe aerobic mineralization of organic matter by micro-organisms (Chapelle et al., 1988; Jackson et al., 1996).Whilst potentially a very useful technique to rapidlyassess the biodegradation of contaminants, it is onlyapplicable if the 613C of the endogenous organic matterand the contaminant are measurably different.A good example of the potential of this approach hasbeen published recently (Jackson et al., 1996). The 613Cratio of crude oils is typically -29% to -32%,. Incontrast, organic carbon in soils and sediments domi-nated by C-4 plants exhibits a 613C signature of between-14.4ym and -17.7%. Thus it has been possible toclearly distinguish the rate of CO, production fromhydrocarbon degradation and from mineralization ofendogenous organic matter in oil-contaminated saltmarsh sediments dominated by Spartina (Jackson et al.,1996). In addition, kinetic constants calculated using the613C data for CO, were statistically no different fromthose calculated from data on alkane degradation. Thathydrocarbon degradation had occurred was furtherindicated by reductions in the ratio of labile hydro-carbons to hopane.The carbon isotope ratio of CO, can be readily andinexpensively determined. This method of assessing thebiodegradability of organic pollutants under field con-ditions therefore has considerable promise for adoptionin bioremediation studies (e.g. Aggarwal & Hinchee,1991; Landmeyer et al., 1996; Aggarwal et al., 1997). Itsapplication, however, will be limited to situations wherethe 613C of the contaminant and endogenous organicmatter are known and distinct.A mle for molecular biology in assessingbioremediation?The ability to isolate and enumerate bacteria fromcontaminated sites capable of degrading a particularpollutant is one line of evidence often used to supportthe feasibility of bioremediation. This is particularlytrue if an increase in the population of degradativebacteria occurs following implementation of a treatmentto stimulate biodegradation. Culture-based methodsunderestimate both qualitative and quantitativemeasures of microbial populations by orders of mag-nitude. This has led to the development and applicationof nucleic-acid-based techniques to study the ecologyand diversity of micro-organisms in nature. Since a morecomplete understanding of microbial ecology will un-doubtedly be required to gain maximum benefits from

bioremediation, it is not surprising that molecularbiological methods are now being employed to studybioremediation. This can be viewed as analogous to theisolation of bacteria with appropriate catabolic proper-ties from a polluted site, as a means of implying thatcompetent degradative populations are present at thesite, and thus that bioremediation potential exists.Molecular biological methods have already been usedsuccessfully to evaluate the bioremediation potential ofa contaminated site (Fleming et al., 1993). The ex-pression of a catabolic gene ( n a h A ) involved in theinitial oxidation of naphthalene to 1,2-dihydroxy-naphthalene was investigated at a manufactured gasplant site contaminated with PAHs. RNA was extractedfrom the contaminated soil and mRNA for the n a h Agene quantified using an RNase protection assay (Belin,1996). The extracted RNA was hybridized in solutionwith a radioactively labelled antisense n a h A RNA probetranscribed from a cloned n a h A gene. Treatment of thehybridized RNA preparation with RNase preferentiallydegrades single-stranded RNA leaving probeRNA:mRNA hybrids intact. The amount ofradiolabelled probe RNA: mRNA hybrid is quantifiedin relation to known amounts of target RNA analysedusing the assay. Comparable samples of the contami-nated soil were used to determine the frequency of n a h Agenes in a cultured fraction of the bacterial populationby a colony hybridization protocol. Residual naph-thalene concentrations and the rate of radiolabellednaphthalene catabolism were also measured. Compari-son of the data revealed a good correlation betweenn a h A gene expression, as determined by quantificationof n a h A mRNA, and the other three parameters. Theability to monitor expression of specific catabolic genesmay be of great importance in determining the feasibilityof bioremediation in situ, particularly since samples canbe taken from the site and processed with minimaldisturbance due to storage and transport or alteration ofconditions in the sample, as occurs with culture basedtechniques.More recently, a similar approach has been adopted toexamine the expression of lignin peroxidases (Lamar etal., 1995) and manganese-dependent peroxidases (Boganet al., 1996) in soil inoculated with Phanerochaetechrysosporium. These enzymes are believed to beimportant in the degradation of organic pollutants bywhite-rot fungi. In these studies, reverse transcriptasecompetitive PCR was used to quantify different levels ofthe gene transcripts and transcription of the fungal geneswas shown to be related to the degradation of con-taminant PAHs (Bogan et al., 1996). This was in turnrelated to the activity of manganese peroxidase inextracts of the soil.Although these examples demonstrate the feasibility ofusing molecular methods to monitor bioremediation,routine practical application is perhaps a little way off.A number of factors will determine how widely adoptedthese approaches will be for routine evaluation andmonitoring of bioremediation programmes (Brockman,

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1995).To date, molecular biological studies of catabolicgenes in contaminated samples have focused on wellcharacterized metabolic traits. For example, the oc-currence of meta-cleavage dioxygenase genes involvedin the catabolism of substituted benzenes, naphthaleneand biphenyl have been examined in a number ofsamples. These have naturally targeted genes for whichmost sequence data are available, for instance xyZE-like(e.g. Joshi & Walia, 1996) and 6phC genes (Erb &Wagner-Dobler, 1993). However, meta-cleavage genes(approx. 30 sequences in the public databases) exhibitconsiderable diversity (e.g. Kukor & Olsen, 1996; Dalyet al., 1997) and no single set of PCR primers or geneprobe will allow detection of all variants of knownmeta-cleavage genes, let alone those that remain to bediscovered. This is a particular problem with analysis ofanaerobic pathways in pollutant degradation, for whichthere is growing evidence of importance, but whichremain largely uncharacterized at the molecular level.Notwithstanding methodological limitations, regulatorsand practitioners of bioremediation must be convincedof the benefits of molecular biological techniques. Forthese approaches to be widely embraced, they mustprovide information of practical benefit in assessing thesuccess or progress of bioremediation that cannot beaccessed by other means. Molecular biological analysesmay also be more costly than conventional chemicalanalyses. Nonetheless, if a sufficiently large market formolecular-biological environmental diagnostics exists,it is not unreasonable to assume that expertise in medicaldiagnostics could be adapted to produce cheap, reliable,quantitative molecular assays for biodegradative micro-organisms or catabolic genes. A further hurdle that mustbe overcome is the acceptance of molecular data , whichis often not quantitative and unfamiliar to those notdirectly involved in research using these methods. It istherefore important that researchers communicate thebenefits and limitations of the techniques available anddo not attempt to apply them where tangible benefitsover other approaches cannot reasonably be claimed.Currently, molecular biological techniques are excellentresearch tools that increase our understanding of thedistribution and expression of important catabolic genesand microbial population dynamics during biotreatment(e.g. Massol-Deya et al., 1997). They are not yetsufficiently robust for routine monitoring applications,and as stand-alone assays their use will always belimited. Ultimately. if the techniques can be developedto provide rapid reliable information on the occurrence,quantity and activity of important biodegradative mi-crobial populations that is complementary to chemicaland physical data, it will only be a matter of time beforethey become more widely used.

Concluding remarksDegradation of organic pollutants by micro-organismshas been studied for many decades. The fruit of thislabour is a remarkable understanding of the biochemical

pathways and molecular genetics involved in the ca-tabolism of a relatively small number of intensivelystudied pollutants by a relatively small group of micro-organisms. It may have been assumed from this know-ledge that the catabolic diversity of micro-organismscould be harnessed to solve a wide range of pollutionproblems. On this basis, initial predictions for thebenefits of bioremediation were overstated with theconsequence that it was never likely to live up toexpectations.It is now clear that it is not our knowledge of pollutantcatabolism that limits the successof bioremediation, butrather a restricted understanding of the interplay be-tween the biotic and abiotic factors that determine theoutcome of any particular remedial strategy. Treatingbioremediation as a natural bioengineering process thattakes account of these interactions, coupled with rig-orous quantitative assessment of the outcome of re-medial treatments is required to ensure bioremediationis successful and to raise the credibility of the tech-nology.ReferencesAbramowiu, D. A. (1990). Aerobic and anaerobic biodegradationof PCBs: a review. Crit Rev Biotechnol 10,241-249.Aggarwal, P. K. & Hinchee, R. E. (1991). Monitoring in situbiodegradation of hydrocarbons using stable carbon isotopes.Environ Sci Technol 25, 1179-1180.Aggarwal, P. K., Fuller, M. E., Gurgas, M. M., Manning, 1. F. &Dillon, M. A. (1997). Use of stable isotope analyses for m onitoringthe pathways and rates of intrinsic and enhanced in situbiodegradation. Environ Sci Technol 31, 590-596.Alexander, M. (1994). Biodegradation and Bioremediation. SanDiego :Academic Press.Apajalahti, 1. H. A. & Salkinoja-Salonen, M. 5. (1987). Dechlori-nation and para-hydroxylation of polychlorinated phenols byRhodococcus chlorophenolicus. J Bacteriol 169, 675-681.Aronstein, B. N., Calvillo, Y. M. & Alexander, M. (1991). Effect ofsurfactants at low concentrations on the desorption and bio-degradation of sorbed aromatic compounds in soil. Environ SciTechnol 25, 1728-1731.Assinder, S. J. & Williams, P. A. (1990). The TOL plasmids:determinants of the catabolism of toluene and the xylenes. AdvMicrob Physiol31, 2-69.Bachmann, A., de Bruin, W., Jumelet, 1. C., Rijnaarts, H. H. N. &Zehnder, A. J. B. (1988). Aerobic mineralization of alpha-hexa-chlorocycloh exane in con taminated so il. AppZ Environ MicrobiolBedard, D. L. & May, R. 1. (1996). Characterization of thepolychlorinated biphenyls in the sediments of Woods Pond -evidence fo r microbial dechlorination of Aroclor 1260. EnvironSci Technol 30,237-245.Bedard, D. L., Bunnell, S. C. & Smullen, L. A. (1996). Stimulation ofpara-dechlorination of polychlorinated biphenyls that have per-sisted in Housatonic River sediments for decades. Environ SciTechnol 30,687-694 .Belin, D. (1996). The RNase protection assay. In Basic DNA an dRN A Protocols, pp. 131-136. E dited by A. J . Harwood. Totow a:Humana Press.Beller, H. R., Reinhard, M. & Grbic-Galic, D. (1992). Metabolic by-

54,548-554.

606


9/10

Bioremedia tionproducts of toluene degradation by sulfate-reducing enrichmentcultures. Appl Environ Micro biol5 8, 3192-3195.Beller, H. R Ding, W.-H. & Reinhard, M. (1995). Byproducts ofanaerobic alkylbenzene metabolism useful as indicators of in situbioremediation. Environ Sci Techno1 29, 2864-2870.Beurskens, J. E. M., Dekker, C. G. C Jonkhoff, 1 & Bontstra, L.(1993). Microbial dechlorination of hexachlorobenzene in asedimentation area of the Rhine river. Biogeochemistry 19,61-81.Bhandari, A., Novak, J. T. & Berry, D. F. (1996). Binding of 4-monochlorophenol to soil. Environ Sci Technol 30, 2305-231 1.Bogan, B. W., Schoenike, B Lamar, R. T. & Cullen, D. (1996).Manganese peroxidase mRNA and enzyme activity levels duringbioremediation of polycyclic aromatic hydrocarbon-contami-nated soil with Phanerochaete chrysosporium. Appl EnvironMic robi ol62 , 2381-2386.Borden, R. C Gomez, C. A. & Becker, M. T. (1995). Geochemicalindicators of intrinsic bioremediation. Ground W ate r 33, 180-189.Bosma, T. N. P. Middeldrop, P. J. M., Schraa, G. & Zehnder,A. J. B. (1997). Mass transfer limitation of biotransformation :quantifying bioavailability. Environ Sci Technol 31, 248-252.Bouillot, P. Canales, A., Pareilleux, A., Huyard, A. & Goma, G. J.(1990). Membrane bioreactors for the evaluation of maintenancephenomena in waste-water treatment. J Ferment Bioeng 69,Bragg, J. R. Prince, R. C Wilkinson, 1. B. & Atlas, R. M. (1992).Bioremediation for shoreline cleanup following the 1989 Alaskanoil spill. Houston : Exxon Company.Bragg, J. R Prince, R. C. Harner, E. J. & Atlas, R. M. (1994).Effectiveness of bioremediation for the Exxon Valdez oil spill.Nature 368,413-418.Brockman, F. J. (1995). Nucleic acid-based methods for moni-toring the performance of in situ bioremediation. Mol Ecol 4,Burlage, R. S. Hooper, 5. W. & Sayler, 6. 5. (1989). The TOL(pWW0) catabolic plasmid. Appl Environ Microbiol 5 5 , 1323-1328.Carmichael, L. M., Christman, R. F. & Pfaender, F. K. (1997).Desorption and mineralization kinetics of phenanthrene andchrysene in contaminated soils. Environ Sci Technol 31,126-132.Carroll, K. M., Harkness, M. R Bracco, A. A. & Balcarcel, R. R.(1994). Application of a permeant polymer diffusional model tothe desorption of polychlorinated-biphenyls from Hudson Riversediments. Environ Sci Technol 28, 253-258.Cerniglia, C. E. (1 992). Biodegradation of polycyclic aromatichydrocarbons. Biodegradation 3,351-368.Chapelle, F. H., Morris, 1. T McMahon, P. B. & Zelibor, J. L., Jr(1988). Bacterial metabolism and the 613C composition of groundwater, Floridian aquifer system, South Carolina. Geology 16,Chaudhry, R. G. & Chapalamadugu, S. (1991). Biodegradation ofhalogenated organic compounds. Microbiol Rev 55,59-79.Chesboro, W., Evans, T. & Eifett, R. (1979). Very slow growth ofEscherichia co li. J Bacteriol 139, 625 -638 .Cornelissen, G. , van Noott, P. C. M., Parsons, J. R. & Covers,H. A. J. (1997). Temperature dependence of slow adsorptionand desorption kinetics of organic compounds in sediments.Environ Sci Technol 31,454460.Daly, K., Dixon, A. C Swannell, R. P. J Lepo, J. E. & Head, 1. M.(1997). Diversity among aromatic hydrocarbon-degrading bac-teria and their meta-cleavage genes. J Appl Microbio l83 ,42 1429.

178-183.

567-578.

117-121.

Deitsch, J. J. & Smith, J. A. (1995). Effect of Triton-X-100 on therate of trichloroethene desorption from soil to water. Environ SciTechno l 29, 1069-1080.Doelman, P. L., Haanstra, H., Loonen, H. & Vos, A. (1990).Decomposition of alpha-hexachlorocyclohexane nd beta-hexa-chlorocyclohexane in soil under field conditions in a temperateclimate. Soil Biol Biochem 22, 629-639.Efroymson, R. A. 81Alexander, M. (1991). Biodegradation by anArthrobacter species, of hydrocarbons partitioned into an organicsolvent. Appl Environ Microbiol57, 1441-1447.Erb, R. W. & Wagner-Dobbler, 1. (1993). Detection of poly-chlorinated biphenyl degradation genes in polluted sediments bydirect DNA extraction and polymerase chain reaction. ApplEnviron Microbiol59,4065-4073.Erickson, D. C Loehr, R. C. & Neuhauser, E. F. (1993). PAH lossduring bioremediation of manufactured gas plant site soils. W a t e rRes 27,911-919.Evans, P. J Ling, W., Goldschmidt, B Ritter, E. R. &Young, L Y.(1992). Metabolites formed during anaerobic transformation oftoluene and ortho-xylene and their proposed relationship to theinitial steps of toluene mineralization. Appl Environ MicrobiolEvans, W. C. & Fuchs, G. (1988). Anaerobic degradation ofaromatic compounds. Annu R ev M icrobio l42 ,289-317.Fayad, N. M., Edora, R. L El-Mubarak, A. H. & Polancos, A. B.(1992). Effectiveness of a bioremediation product in degrading theoil spilled in the 1991 Arabian Gulf war. Bull Environ ContamToxicol49 ,787-796.Fleming, S. T Sanseverino, J. & Sayler, G. 5. (1993). Quantitativerelationship between naphthalene catabolic gene-frequency andexpression in predicting PAH degradation in soils at town gasmanufacturing sites. Environ Sci Technol 27, 1068-1074.Frantz, B. Aldrich, T. & Chakrabarty, A. M. (1987). Microbialdegradation of synthetic recalcitrant compounds. Biotechnol AdvGibson, D. T. (1984). Microbial Degradation of Organic Com-pounds. New York: Marcel Dekker.Harkness, M. R., McDermott, J. B Abramowicz, D. A. & 16 otherauthors (1993). In situ stimulation of aerobic PCB biodegradationin Hudson River sediments. Science 259, 503-507.Heitzer, A., Webb, 0. F Thonnard, J. E. & Sayler, G. S. (1992).Specificand quantitative assessment of naphthalene and salicylatebioavailability by using a bioluminescent catabolic reporterbacterium. Appl Environ Microbiol 58, 1839-1846.Heitzer, A., Malachowsky, K., Thonnard, 1. E Bienkowski, P. R.White, D. C. & Sayler, G. 5. (1994). Optical biosensor forenvironmental on-line monitoring of naphthalene and salicylatebioavailability with an immobilized bioluminescent catabolicreporter bacterium. Appl Environ Microbiol60, 1487-1494.Hunter, M. A., Kan, A. T. &Tomson, M. 6. (1996). Development ofa surrogate sediment to study the mechanisms responsible foradsorption/desorption hysteresis. Environ Sci Technol 30,2278 -2285.Jackson, A. W., Pardue, J. H. & Araujo, R. A. (1996). Monitoring ofcrude oil mineralization in salt marshes: use of stable carbonisotope ratios. Environ Sci Technol 30, 1139-1144.Jaspers, C. Ewbank, G., McCarthy, A. 1. &Penninckx, M. 1. (1997).Cattle manure compost for bioremediation of pentachlorophenolin contaminated soil and wood material. T he metabolic fate ofthe xenobiotic. In Proceedings of the Znternational Symposiu m o nEnvironmental Biotechnology, pp. 1-4. Edited by H. Verachtert& W. Verstraete. Antwerp :Technologisch Institut Antwerpen.

58,496-501.

5 , 85-99.

607


10/10

I . M . H E A D

Joshi, B. & Walia, 5. (1996). PCR amplification of catechol 2,3-dioxygenase gene sequences from naturally occurring hydro-carbon-degrading bacteria isolated from petroleum hydrocarboncontaminated groundwater, FEMS Microbiol Ecol 19,s-15.Kan, A. T Fu, G. & Tomson, M. B. (1994). Adsorption/desorptionhysteresis in organic pollutant and soil/sediment interactions.Environ Sci Technol 28, 859-867.Klstner, M. (1991). Reductive dechlorination of tri- and tetra-chloroethylenes depends on transition from aerobic t o anaerobicconditions. Appl Environ MicrobiolS7, 2039-2046.Kelsey, J. W., Kottler, B. D. & Alexander, M. (1997). Selectivechemical extractants to predict bioavailability of soil-agedorganic chemicals. Environ Sci Technol 31,214-217.Knackmuss, H.-J. (1996). Basic knowledge and perspectives ofbioelimination of xenobiotic compounds. J Biotechnol 51, 287-295.Knaebel, D. B Federle, T. W., McAvoy, D. C. & Vestal, J. R.(1994). Effect of mineral and organic soil constituents onmicrobial mineralization of organic compounds in a natural soil.Appl Environ Microbiol60,4500-4508.Kukor, J. 1. & Olsen, R. H. (1996). Catechol 2,3-dioxygenasesfunctional in oxygen-limited (hypoxic) environments. Appl En-viron Microbiol62, 1728-1740.kine, M. M. & Jsrgensen, K. S. (1996). Straw compost andbioremediated soil as iriocula for bioremediation of chlorophenol-contaminated soil. Appl Environ Microbiol62, 1507-1513.Lamar, R. T., Schoenike, B., Van den Wyemelenberg, A., Stewart,P Dietrich, D. M. & Cullen, D. (1995). Quantitation of fungalmRNAs in complex substrates by reverse transcription PCR andits application to Phanerochaete chrysosporium-colonized soil.Appl Environ Microbiol61,2122-2126.Landmeyer,J. E Vroblesky, D. A. & Chapelle, F. H. (1996). Stablecarbon isotope evidenceof biodegradation zonation in a shallowjet-fuel contaminated aquifer. Environ Sci Technol 30,1120-1128.Liu, Z Jacobson, A. M. & Luthy, R. G. (1995). Biodegradation ofnaphthalene in aqueous non-ionic surfactant systems. ApplEnviron Microbiol61, 145-151.Massol-Deya, A., Weller, R Rios-Hernandez, L. Zhou, J.-Z.,Hickey, R. F. & Tiedje, J. M. (1997). Succession and convergence ofbiofilm communities in fixed-film reactors treating aromatichydrocarbons in groundwater. Appl Environ Microbiol 63,270-276.Mundell, K. M. (1994). The biodegradation potential of Phanero-chaete chrysosporium and Penicillium thomii in reducing PAH

concentration in soil from a disused coal extraction plant. MScthesis, University of Newcastle.Pignatello, J. J. &Xing, B. (1996). Mechanisms of slow sorption oforganic chemicals to natural particles. Environ Sci Technol 30,1-11.Pritchard, P. H. & Costa, C. F. (1991). EPAs Alaska oil-spillbioremediation project. Environ Sci Technol 25, 372-379.Providenti, M. A., Fleming, C. A., Lee, H. & Trevors, J. T. (1995).Effect of rhamnolipid biosurfactants or rhamnolipid-producingPseudomonas aeruginosa on phenanthrene mineralization in soilslurries. FEMS Microbiol Ecol 17, 15-26.Rijnaarts, H. H. M., Bachmann, A,, Jumelet, J. C. & Zehnder,A. J. B. (1990). Effect of desorption and intraparticle mass transferon the aerobic biomineralization of a-hexachlorocyclohexane in acontaminated calcareous soil. Environ Sci Technol 24,1349-1354.Rohmer, M., Bouvier-Nave, P. & Ourisson, G. (1984). Distributionof hopanoid triterpenes in prokaryotes. J Gen Microbiol 130,Steiert, J. G. & Crawford, R. L. (1986). Catabolism of penta-chlorophenol by a Flavobacterium sp. Biochem Biophys ResCommun 141,825-830.Steinberg, S M., Pignatello, J. J. & Sawhney, B. L. (1987). Per-sistence of 1,2-dibromoethane in soils - entrapment in intra-particle micropores. Environ Sci Technol 21, 1201-1208.Tabak, H. H Gao, C Lai, L., Yan, X., Pfanstiel, S., Kim, 1. S. &Govind, R. (1994). Determination of bioavailability and bio-degradation kinetics of phenol and alkylphenols in soil. AmChem SOC ymp Series S54,Sl-77.Tros, M. E Schraa, G.& Zehnder, A. J. B. (1996a). Transformationof low concentrations of 3-chlorobenzoate by Pseudomonas sp.strain B13 : kinetics and residual concentrations. Appl EnvironMicrobiol62, 437-442.Tros, M. E Bosma, T. N., Schraa, G. & Zehnder, A. J. B. (1996b).Measurement of minimum substrate concentrations (Smin)in arecycling fermenter and its prediction from the kinetic parametersof Pseudomonas sp. strain B13 from batch and chemostat cultures.Appl Environ Microbiol61, 145-151.Weissenfels, W. D Klewer, H. J. & Langhoff, J. (1992). Adsorp-tion of polycyclic aromatic hydrocarbons (PAHs) by soil particles- nfluence on biodegradability and biotoxicity Appl MicrobiolBiotechnol36,689-696.WU, S. & Gschwend, P. M. (1986). Sorption kinetics of hydro-phobic organic compounds to natural sediments and soils.Environ Sci Technol 20,717-725.

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