microbiological methods for the cleanup of soil and ground water contaminated with halogenated...

FEMS Microbiology Reviews 63 (1989) 277-300 277 Published by Elsevier

FEMSRE 00126

Microbiological methods for the cleanup of soil and ground water contaminated with halogenated organic compounds

Phil ip M o r g a n and R o b e r t J. W a t k i n s o n

Shell Research Ltd., Sittingbourne Research Centre, Sittingbourne, U.K.

Received 6 March 1989 Revision received 5 May 1989

Accepted 16 May 1989

Key words: Biodegradation; Biorestoration; Land; Aquifers; Haloaliphatics; Haloaromatics

1. S U M M A R Y

There is growing interest in the enhancement of microbial degradative activities as a means of bringing about the in situ cleanup of contaminated soils and ground water. The halogenated organic compounds are likely to be prime targets for such biotechnological processes because of their widespread utilisation and the biodegradability of many of the most commonly used compounds. The aim of this review is to consider the potential for microbiological cleanup of haloorganic-contaminated sites. The technologies available involve the provision of suitable environmental conditions to facilitate maximum biodegradation rates either in the subsurface or in on-site bioreactors. Methodologies include the supply of inorganic nutrients, the supply of oxygen gas, the addition of degradative microbial inocula and the introduction of co-metabolic substrates. The potential efficiencies and limitations of the methods are critically discussed from a microbiological

Correspondence to: Dr. P. Morgan, Shell Research Ltd., Sit- tingbourne Research Centre, Sittingbourne, Kent, ME9 8AG, U.K.

viewpoint with respect to substrate degradability and population responses to supplementation.

2. I N T R O D U C T I O N

Halogenated organic compounds are widely used for a variety of applications ranging from solvation and cleaning through utilisation as intermediates in the chemical industry to direct use, for example as pesticides and wood preservatives. As a consequence, quantities of haloorganic compounds may enter the soil environment in a number of ways. It may be necessary to clean a contaminated area, for example if a site is heavily contaminated or if mobile compounds are likely to taint an aquifer used as a source of potable water.

There exists a variety of physicochemical techniques for the cleanup of contaminated soil and ground water [1,2,2a]. These include excavation followed by incineration or chemical treatment, in situ vapour-phase stripping of volatiles and the extraction of contaminated water for treatment with activated charcoal, resins or chemical agents. All of these processes may be effective if correctly applied and operated but they may prove expensive if large-scale excavation or long-term treat-

0168-6445/89/$03.50 © 1989 Federation of European Microbiological Societies

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ment is necessary. Furthermore, certain ap- proaches may not result in total decontamination. As a result there is growing interest in the use of microbial biodegradative activity for the in situ treatment of contaminated soil and ground water, either alone or in collaboration with one of the physicochemical techniques. The aim of microbially-based clean-up methods is to provide opti- mum environmental conditions either in situ or by means of on-site bioreactors so that biodegradation can proceed at the maximum sustainable rate. The technologies, which have been called by a variety of names including bioreclamation, biorestoration and in situ biotreatment, have been most commonly applied at locations contaminated with hydrocarbon mixtures, especially gasoline [1,3], but appear to be applicable with some modifica- tions at sites contaminated with a variety of organic compounds. The aim of this review is to consider critically the technologies from a microbiological viewpoint in order to highlight areas of potential application or particular difficulty as a function of microbial physiology and ecology.

3. MICROBIAL METABOLISM OF HALO- G E N A T E D ORGANIC COMPOUNDS

Prior to any consideration of biological cleanup of contaminated soils and ground water, it is necessary to understand the extent and limits of microbial metabolism of halogenated organic compounds. It is the aim of this section of the review to survey briefly the metabolism of haloorganic compounds in order to provide information that enables a clearer understanding of the potential applications of the technologies. No attempt will be made to discuss exhaustively the multitude of physiological data that have been obtained and detailed reviews of microbial metabolism are available elsewhere [4-6].

3.1. Halobenzenes Chlorobenzenes are widely utilised as solvents

and disinfectants and may be produced as by- products in a number of industries. Their lipo- philic properties and the volatility of the less substituted molecules means that it is often dif-

ficult to obtain halobenzene-degrading isolates. Indeed, even strains capable of metabolising halobenzenes may be inhibited at relatively low substrate concentrations due to solvent-induced damage to cell membranes. Degradation of mono-, di- and trichlorobenzenes can be performed under aerobic conditions by soil, ground water and waste water populations [7-12]. Conversion to the corresponding chlorocatechols followed by ring cleavage and dechlorination has been ascertained as the metabolic pathway in aerobic degradative isolates [9-12]. The more substituted molecules appear to be highly resistant to aerobic microbial attack, although there is evidence for the reductive dehalogenation of hexachlorobenzene under methanogenic conditions [13,14].

The degradation of nitro-substituted chlorobenzenes has received some attention because of the use of certain of these compounds in anti-fungal treatments. Many genera of fungi have been shown to convert such compounds to the corresponding anilines or anilides and thereby bring about their detoxification [4].

3.2. Halobenzoates Halobenzoates may be produced during the

microbial metabolism of a variety of haloorganic compounds and certain chlorobenzoates may be used as herbicides. Aerobic metabolism of chlorobenzoates has been widely studied. Dehalogena- tion may occur either prior to ring cleavage [15,16] or afterwards [17,18]. Fluorobenzoate degradation has also been described [19] and the degradation of halogen-substituted salicylates (hydroxybenzo- ates) has been investigated [20]. Anaerobic degradation of halobenzoates has also been investigated in detail and appears to be a widespread phenomenon. Metabolism of fluoro-, chloro-, bromo- and iodobenzoates has been demonstrated by organisms growing under denitrifying conditions [21,22]. Methanogenic communities have been shown to degrade a variety of compounds, including 2-fluorobenzoate, many di- and trichlo- robenzoates and monosubstituted chloro-, bromo- and iodobenzoates [23-26].

3.3. Halophenols Chlorinated phenols are widely distributed in

the environment owing to their metabolic produc-

tion during the biodegradation of many chloro- organics and their direct use as antifungal agents. Efficient aerobic degradation of chlorophenols substituted with up to four chlorine atoms has been reported [27,28] and the widely used wood preservative pentachlorophenol can also be metabolised [27-32]. Under methanogenic conditions reductive dehalogenation can occur with the ultimate production of phenol, which can be mineralised [13]. The rates of degradation of different chlorophenol isomers may vary significantly under methanogenic conditions and a population adapted to some compounds may not be able to metabolise others [33,34]. Anaerobic degradation of pentachloro- and pentabromophenols has been widely observed [30,32,35,36]. The aerobic and anaerobic degradation of chlorinated dihydroxy- benzenes (catechols and resorcinols) has also been noted [13,37,38].

3.4. Chlorinated pesticides There exist a large number of structurally di-

verse chloroaromatic compounds used as pesticides. The most commonly used compounds can be grouped by their basic chemical structures and the metabolism of individual compounds with similar structures normally follows the same pattern. Since during manufacture and transport and as a result of normal use these compounds may enter the subsurface environment, it is pertinent to consider the metabolic capabilities of microorganisms on some commonly employed compounds. Detailed reviews of pesticide metabolism are available elsewhere [4,39,40].

Chlorophenoxy herbicides are widely used and the degradation of the most commonly used, 2,4-D (2,4-dichlorophenoxyacetic acid), 2,4,5-T (2,4,5- trichlorophenoxyacetic acid) and MCPA (4- chloro-2-methylphenoxyacetic acid), has been widely studied. Degradation can occur aerobically and anaerobically. Under aerobic conditions the initial metabolic step involves ether cleavage to produce glyoxylate and the appropriate chlorophenol molecule, each of which can be mineralised [40-46].

The degradation of D D T (1,1,1-trichloro-2,2- bis(p-chlorophenyl)ethane) has been particularly studied because of its known environmental per-

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sistence. Biological attack can only occur extremely slowly and transformation normally involves an initial single dechlorination step due to reaction between D D T and iron porphyrin molecules. Further degradation can occur and complete mineralisation has been reported [40,47,48] but the processes are extremely slow. Analogues of D D T have been employed as insecticides but these have also been found to be persistent in the environment. One commonly used compound, methoxychlor (1,1-bis( p-methoxyphenyl)-2,2,2-trichloroethane) can be reductively dechlorinated by microorganisms but metabolism is slow [40,49].

There are many pesticides which may be degraded to chloroanilines by microbial activity under aerobic conditions. These include the phenyl- urea herbicides, such as monuron (3-(4-chlorophenyl)-l , l-dimethylurea) and diuron (3-(3,4-dichlorophenyl-l-methoxy-l-methylurea), the phen- ylcarbamate herbicides, such as CIPC (iso-propyl- N-(3,4-dichlorophenyl)carbamate) and the acyl- anilide herbicides, such as propanil (N-(3,4-dichlorophenyl)propionamide [4]. Mineralisation of mono- and dichloroanilines has been observed and it has been noted that co-metabolism of these molecules is the most common mode of attack [50-53]. Indeed, whilst strains capable of utilising monochloroanilines as sole carbon and energy sources have been isolated no pure cultures of dichloroaniline-degrading organisms have been obtained. Evidence for the degradation of more highly substituted anilines has been obtained but photochemical reactions were necessary to initiate attack [54].

3.5. Polychlorinated biphenyls and dibenzodioxins Polychlorinated biphenyls (PCBs) were widely

used for a variety of industrial applications because of their high thermal, electrical and chemical stabilities. However, it became evident that they were highly persistent in the environment and their manufacture was stopped in many countries. Due to the method of manufacture of PCBs, direct reaction of chlorine with biphenyl, commercial PCB products are a highly complex mixture of different PCB molecules (congeners) containing different numbers of chlorine atoms at different substituent positions. Thus it is not uncommon to

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find more than 50 different congeners in a prod- uct. The study of biodegradation is further com- plicated by the fact that PCBs are highly hydrophobic and sorb strongly onto solids and partition into lipids. As a consequence, many workers have studied the degradation of simple chlorinated biphenyls for reasons of experimental simplicity. A number of organisms have been isolated which are capable of growth upon monochlorobiphenyls and dichlorobiphenyls that are substituted upon one ring only. The major metabolic pathway is the fission of the unsubstituted ring and the end prod- uct of metabolism is the corresponding chlorobenzoate [55-57]. Interestingly, no organism has yet been isolated which is capable of metabolising the chlorobenzoate it produces from the chlorobi-

phenyl and it has been suggested that specific and distinct gene sequences are required for the metabolism of the two compounds [58]. This hy- pothesis is supported by the isolation of a population comprising two strains of microorganisms capable of converting 4-chlorobiphenyl to CO 2 [59]. The first strain converted the chlorobiphenyl to 4-chlorobenzoate and this was mineralised by the second strain. In addition to the conversion to chlorobenzoate, metabolism of 4-chlorobiphenyl to 4 '-chloroacetophenone has been reported [60,61].

Experience gained from research into the degradation of commercial PCB products enables the statement of a number of general rules con- cerning the biodegradability of the congeners

CI

el

CI

2,3-dichlorobiphenyl Single ring substituted,

easily degraded

4,4' dichlorobiphenyl

Sites blocked for 3,4 cleavage. Vulnerable to 2,3 cleavage

Cl ~ C I

el

CI

CI ~ / C l

el ~ C l

CI ~ CI Cl - ~ ~ C I

2,3,2',5'-tetrachlorobiphenyl 3,5,3'.5'-tetrachlorobiphenyl 2,6.2 6' tetrachlorobiphenvl Sites blocked for 2,3 cleavage. Lack of adjacent unsubstituted Chlorines in ortho position Vulnerable to 3.4 cleavage sites makes cleavage difficult hinder enzyme attack

Fig. 1. Example polychlorinated biphenyls showing how the position of the substituent atoms may hinder aerobic biodegradation. Modified with permission from [64]: D.L. Bedard et al. (1986) Appl. Environ. Microbiol. 51,761-768.

within a mixture [4,62-65]. PCBs with greater than 5 substituents are highly recalcitrant whereas those with less are more readily degraded. De- gradation is significantly enhanced if one ring is less substituted or, better, unsubstituted. Since cleavage of the aromatic ring by oxygenase enzymes requires the presence of two adjacent unsubstituted carbon atoms, different isomers with the same degree of substitution will be degraded at different rates. Some examples of the way in which substituent position may affect ring cleavage reactions are given in Fig. 1. Thus, degradation of commercial PCB mixtures is characterised by the differential metabolism of components. With notable exceptions [66], individual PCB-degrading isolates can only attack a few congeners within a PCB mixture and in the environment degradation rates are generally extremely slow [67-69]. There is also evidence that at least some PCBs can be slowly dechlorinated under anaerobic conditions [70]. Dioxins and the related diben- zofurans have no industrial uses but may be produced during certain chemical processes and during the combustion of materials containing chlorinated organic compounds. There are a few reports of co-metabolic attack upon chlorinated dioxins but rates are extremely slow and end-products have not been characterised.

3. 6. Halogenated aliphatics and cycloaliphatics Halogenated aliphatic compounds are widely

used in industry as solvents, plasticisers and as feedstocks for the production of a variety of materials. The metabolism of halogenated aliphatics has been studied under both aerobic and anaerobic conditions. It has been noted that halomethanes and some halogenated ethanes and ethenes are relatively resistant to direct attack under aerobic conditions. A variety of halogenated alkanes, alcohols and fatty acids have been shown to be metabolised aerobically by microbial isolates. Carbon chain lengths of up to 10 atoms have been widely shown to be vulnerable to attack [71-76] and there has been a report of metabolism of 1-chlorohexadecane [71]. Compounds substituted with chlorine, bromine or iodine atoms are all degraded and fluoro-substituted compounds, despite their toxicity, can also be attacked

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by some strains [77,78]. Aerobic breakdown of the more recalcitrant halomethanes and haloethenes appears to be limited to cometabolism by methane-oxidising bacteria. The ability of these organisms to degrade short-chain haloaliphatics is due to the fact that the enzyme methane mono- xygenase (MMO) can fortuitously oxidise a variety of simple organic compounds to oxygenated products [79]. Methane-oxidising populations from ground water, sewage treatment plants and sediments have all been shown to metabolise haloaliphatics. Compounds shown to be degraded include trichloroethene, vinyl chloride, dichloro- ethene, dibromoethane, dichloromethane, di- bromomethane and diiodomethane [80-84]. The pathways responsible for degradation of the chlo- roalkenes involve epoxidation of the double bond to produce unstable compounds that sponta- neously break down to yield simpler products. Thus, trichloroethene is cleaved to produce for- mate, glyoxylate, chloride ions and CO 2 [80] or dichloroacetate and glyoxylate [83]. Vinyl chloride cleaves to yield chloroacetaldehyde and glycoal- dehyde [80]. These cleavage products may be metabolised by the methylotrophs themselves or by other organisms in the community. Isolates of methylotrophic Pseudomonas and Hyphomicro- bium spp. have been shown to convert dichloro-, dibromo- and diiodomethane to formaldehyde which can be fully mineralised by these organisms [85,86]. An alternative fortuitous degradation of trichloroethene has been described for microorganisms possessing toluene dioxygenase activity [87-89].

Under denitrifying conditions a microbial population isolated from a sewage treatment plant has been shown to transform a variety of haloaliphatics including carbon tetrachloride, dichloromethane and tribromomethane [90]. Denitrifying metabolism of dichloromethane has also been described for the methylotrophic Pseudomonas sp. described above [85]. However, it is under methanogenic conditions that anaerobic attack upon haloaliphatics is most diverse. Indeed, even the highly chlorinated compounds which are only slowly attacked in the presence of oxygen may be rapidly dehalogenated by methanogenic consortia. Compounds shown to be dehalogenated under

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Table 1

Capacity for biodegradation of selected halogenated aliphatic compounds by anaerobic consortia [99]

Compound Conditions under which degraded

Tribromomethane (bromoform)] Carbon tetrachloride Bromodichloromethane [ Hexachloroethane ]

Dibromochloropropane ) 1,2-dibromoethane 1,1,1 -trichloroethane 1

Trichloromethane (chloroform) } Tetrachloroethane

Methanogenic, sulphate-reducing and denitrifying.

Methanogenic and sulphate-reducing

Methanogenic.

methanogenic conditions include tetra-, tri- and dichloroethanes and ethenes, brominated methanes and ethanes, carbon tetrachloride and chloroform [91-98]. Products may be fully dehalogenated or may be only partially metabolised. In the latter case there is the risk of persistence of metabolites, including vinyl chloride produced from tetrachlo- roethene [96]. A detailed study of anaerobic degradation of halogenated aliphatics has shown that methanogenic consortia have a greater substrate range than sulphate-reducing populations which in turn are more active than denitrifying populations [99]. These data are summarised in Table 1 and it can be seen that it may be necesary to tailor conditions in order to obtain biological cleanup of different haloaliphatic compounds.

Halogenated cycloaliphatics are less commonly used and their metabolism has therefore received less study. One exception is the insecticide 1,2,3,4,5,6-hexachlorocyclohexane (benzene hexachloride; BHC) which has been widely used. There are four isomers of BHC, designated a, fl, 3' and O, differentiated by the spatial arrangement of the chlorine atoms. Only the 3' isomer has insecticidal activity. It has generally been observed that degradation proceeds most rapidly under anaerobic conditions [40] and that degradation of the isomers under both aerobic and anaerobic conditions occurs at different rates, namely 3' > a > O > fl [40,100,101]. This phenomenon may be due to the relative solubilities of the isomers, with the most

soluble being the most degradable, or may be due to the spatial arrangement of the chlorine atoms. [100] Under anaerobic conditions reductive dechlorination occurs followed by aromatisation of the ring to produce chlorobenzenes which can be hydroxylated to chlorophenols and thereafter fully metabolised.

4. FATE OF H A L O G E N A T E D ORGANIC POLLUTANTS IN SOIL A N D G R O U N D - WATER

The microbial metabolism of a variety of halogenated organic compounds has been discussed and it is evident that numerous substances that may exist as environmental pollutants are potentially biodegradable. The potential for biodegradation in the subsurface environment depends upon the presence of degradative organisms, the existence of environmental conditions suitable for metabolism to occur and contact between substrate and degradative organism or extracellular enzyme. The aim of this section is to review the abiotic fates of haloorganic compounds in soil and ground water and to discuss the environmental properties of the subsurface. Once such factors have been considered the next logical step is to endeavour to optimise biodegradative activity by overcoming the limitations. It is this technology that is the basis of microbiological cleanup of contaminated soil and ground water.

4.1. Transport and sorption The migration of halogenated organic materials

in the subsurface environment depends upon the physical properties of the compound and the composition of the environment itself. Compounds that are liquid at ambient temperatures will tend to penetrate rapidly into the soil column, unless infiltration is minimised due to waterlogging or freezing of the soil. Such migration is most marked for the halogenated solvents and their fates on reaching the groundwater table are determined by their densities and aqueous solubilities. The major- ity of solvent compounds are denser than water and will penetrate through the aquifer to accu- mulate as pools of solvent upon impermeable con-

Residual solvent held in soil pores

Surface

Vadose (unsaturated) zone

Water table

!:~:i:!:: . . . . ..:.~.,.

Confining stratum ::::::::::::::::5:::::::::::::::::::: "" ' " " ' ~ " ~ ' " ' Aquifer i:i:~!?~!:~:Wiiik... I o i ~ i :

Pools of Impermeable free solvent layers

Fig. 2. Diagrammatic illustration of migration of halogenated organic solvents through the subsurface.

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fining strata (Fig. 2). Dissolution of the materials may result in their long-distance transport within the ground water and the accumulations of the compounds may act as reservoirs for long-term solubilisation and dispersal.

The migration of solid-phase contaminants, unless carded within a solvent phase, is dependent upon their solubility in water. Many halogenated contaminants are hydrophobic and tend to adsorb strongly onto soil particles. For example, polychlorinated biphenyls are highly hydrophobic and are generally poorly mobile in soil [69]. Sorption of the more highly substituted haloaliphatics may also be significant [98]. Furthermore, a variety of compounds may be incorporated into soil humic material. The hydroxy-group present in halogenated phenolics renders them vulnerable to oxidative coupling reactions with humic components [102]. 2,4-D [103], 2,4-dichlorophenol [104] and MCPA [105], for example, have all been shown to be bound in this way. 3,4-Dichloroaniline is also strongly bound to humus both by physical and chemical processes [51,106,107]. In general, binding to humus results in the reduced bioavaila- bility of contaminants but highly hydrophobic materials may become more mobile following incorporation into humic acid fractions, as has been observed for DDT [108]. In general, evidence sug-

gests that many non-solvent haloorganics are strongly retained in the subsurface [109], although the more hydrophilic compounds may be rapidly transported through soil fissures and macropores. Migration is usually constrained by soil-compound interactions. For example, the migration of 4-chlorophenol was retarded by 28% relative to the rate of water flow in a sandy aquifer containing little organic carbon [110] whereas pentachlorophenol was found not to be dispersed from a contaminated site because of its insolubility [111].

4.2. Environmental factors limiting contaminant elimination

The elimination of compounds from a subsurface environment may occur by physicochemical or biological means. Volatilisation is particularly important for solvent-type halogenated organic compounds and is a function of the vapour pressure of the compound, environmental temperature, wind speed and the degree of vertical penetration. Volatilisation of compounds may assist a biodegradation process by removing mem- brane-damaging components. Photochemical reactions on slowly-infiltrating material may result in the production of more readily degraded oxygenated compounds, as has been observed for 2,4,5-trichloroaniline in lake water [112]. However,

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it is microbiological reactions that normally account for most of contaminant elimination in the soil and ground water environments.

There exists a great diversity of microorganisms in soil and ground water environments and both bacteria and fungi may make major contributions to the degradation of contaminants. The general microbial ecology of the subsurface has been ex- tensively reviewed [e.g. 113-115] and will not be discussed in detail in this report. However, reitera- tion of a few general points may serve to illustrate the requirements for successful biological cleanup. Since the soil surface is the site of entry of water, organic carbon, oxygen and many inorganic nutrients, rapid microbial utilisation of these will generally occur and key substrates may be depleted in the subsoil. Microbial activity therefore tends to be reduced in the subsoil, although subsoils are by no means microbially barren. Below the unsaturated subsoil (vadose zone) the ecology of the ground water aquifer(s) depends upon their nutrient and oxygen status. Active aquifer populations under aerobic, n i t ra te- reducing [116-118], sulphate-reducing [119] and methanogenic [120] conditions have been studied. The overall rate of microbial activity in a soil or aquifer depends upon a complexity of environmental factors, the individual importance of which depends upon population structure and, in the case of biodegradation, the substrate under consideration. The interaction between air and water in soil pore spaces in the unsaturated zone impacts upon a variety of environmental properties. Decreasing water content not only reduces water activity and solute transport but also the migration of unicellular microorganisms since these require continuous water films for translocation. Excessive waterlogging, however, reduces the rate of oxygen transport and, furthermore, the limited solubility of air in water results in a relatively small reservoir of oxygen in aqueous solution. Excessively low or high environmental temperature or pH, while selecting for a specific microbial population, are not generally conducive to rapid biodegradation. Inorganic nutrient supply, particularly that of nitrogen or phosphorus, frequently limits microbial activity, particularly in sites containing large quantities or organic carbon.

As discussed above, microbial metabolic diversity potentially enables attack upon a variety of halogenated materials in soil and ground water but it is evident that physicochemical parameters may severely constrain biodegradation rates in the environment even when metabolically competent organisms are present. It is the aim of biological cleanup methods to optimise environmental conditions so that the indigenous microorganisms or, for certain technologies, introduced degradative strains can metabolise the contaminants at the maximum possible rate and thereby bring about site cleanup.

5. TECHNOLOGIES FOR TH E MICROBIO- LOGICAL DECONTAMINATION OF SOILS AND G R O U N D W A T E R

The methodologies available for site cleanup can be loosely grouped into in situ techniques and extraction-treatment techniques. The former involves provision, as appropriate, of inorganic nutrients, organic growth substrates, oxygen and microbial inocula to contaminated soil or ground water and allowing degradation to proceed in situ. The latter processes are used for treatment of extracted ground water or solutions leached from soil and involve the use of bioreactors on site. The techniques may be used in parallel with each other or with certain of the relatively well established physicochemical cleanup processes, such as ad- sorption of organics onto activated charcoal or air-stripping of volatiles [1-3]. It is the aim of this section of the review to discuss in detail the methodologies and microbiological principles of biotechnological cleanup methods.

An obvious essential prerequisite to any cleanup scheme is the assessment and analysis of the location both from a microbiological viewpoint and to determine the nature and degree of contamination. In addition, any work involving deeper soil regions or ground water requires an understanding of site hydrogeology. Sampling schemes need careful design and some information is available to assist in the process [e.g., 121]. The task of ensuring that representative and reliable samples are taken for microbiological analysis [122] is essen-

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tial, since it is necessary to ascertain how biodegradation rates are affected by the proposed treatment methodologies. Once site analysis has demonstrated the suitability of the location for treatment, preparations for the process may com- mence. Frequently, it may be possible to remove large quantities of contaminant by direct pumping of free solvent or by vapour-phase stripping of volatiles and this will reduce the quantity of material that requires biodegradation. Except at those locations where contamination is limited to the surface soil, and mobility of contaminants or degradation products can be shown not to occur, it is always necessary to control the outflow of leachate a n d / o r ground water from a treatment site in order to prevent contaminant dispersal. Surface soil may be piled upon impermeable liners of clay or plastic if contamination is localised and has not penetrated deeply. If the contamination is deep or the surface cannot be disturbed, direct ground water control is necessary. If the ground water table is shallow, trenches cut across the direction of water flow may prove suitable [123-125] but most commonly, and invariably for deep ground water, control is achieved by pumping. Strategically positioned wells can create a downdraw of the water table and cause all flow to occur towards them. This is a very effective means of preventing migration away from a treatment site [123-125].

5.1. Treatment of surface-contaminated soil Disruptive treatment of surface soil is only

feasible where contamination is limited to the upper 0.5 m of the soil column and where the site is clear for earth movement. The simplest microbiological method for decontamination involves fertilisation of the soil and frequent tilling to ensure good mixing and aeration. Adjustment of site pH may be achieved by the application of lime and irrigation may be used if the soil is dry. Such technology has not been widely applied for site decontamination but is the method employed for the disposal of oily wastes in the ' land farming' process. Experience has proved that land farming results in the safe and efficient biodegradation of oils [3,126] and there is no reason to suggest that microbial degradation of halogenated

organic compounds could not be enhanced in the same way.

Composting is a related process which involves excavation of the surface soil. The soil is mixed with fertiliser and a bulking agent such as straw or wood chippings to enhance oxygen penetration. The soil is piled into low mounds, irrigated as necessary and allowed to incubate. An attempt at field-scale cleanup of chlorophenol-contaminated soil has been reported [127]. The soil was mixed with wood bark, piled into mounds and irrigated, fertiliser solution being used as necessary. Effi- cient degradation was observed, especially during the warmer months, and temperatures were in- creased as a consequence of microbial utilisation of the wood bark. A simple aerobic composting system has also been tested for the decontamination of soil containing hexachlorocyclohexane. En- hanced biodegradation was observed, particularly in summer, provided that aerobiosis was main- tained [128,129].

5.2. Supplementation of subsurface enoironments with inorganic nutrients and oxygen

The provision of nutrients and, when applicable, oxygen to the subsurface in order to enhance biodegradation rates in situ is normally performed in a closed ground water system in order to prevent contaminants being dispersed away from the site. The principles of two types of enhanced in situ biodegradation systems are il- lustrated diagrammatically in Fig. 3 [1-3]. Both involve ground water extraction, supplementation and reintroduction. When contamination is confined to the aquifer, treated water may be rein- jected directly but if supply of nutrient solution to the normally unsaturated zone is required then this must be allowed to percolate through the soil column either from infiltration trenches or from a spray irrigation system.

The necessity for and composition of supple- ments depends upon the composition of the soil, the properties of the indigenous microorganisms and the nature of the contamination. Once opti- mum nutrient supply has been determined it may prove feasible to employ either continuous or batch additions and the latter may offer some economic benefits. However, provision of nutrients to the

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(a)

Surface

Nutrients Compressed air

(b)

Water table

Nutrients

Aquifer

b v . . v v ,

contamination

nation

Wal

Fig. 3. Diagrammatic representation of the principles involved in in situ microbiological cleanup of the subsurface. (a) Illustrates a location where otdy the aquifer is contaminated and a pumping-injection circulation system permits continuous supply of nutrients a n d / o r oxygen. (b) Illustrates a location where nutrient supply to the subsurface is being performed by allowing percolation to occur

for an irrigation sprayfield.

unsaturated subsoil may be rendered more difficult by different zones of permeability, which may result in preferential flow through certain areas. Furthermore, the dispersal of inorganic nutrients from percolation or injection points may be limited by chemical interactions with the soil matrix. Ammonium ions are efficiently bound to clay minerals by cation exchange and phosphate anions react with calcium, iron and aluminium ions to produce insoluble phosphate salts. If this

latter process occurs to any great extent there is the risk that precipitates will block the soil pores to nutrient flow and thereby inhibit the cleanup process. This has indeed been observed under field conditions [130], particularly in iron- and calcium-rich soils. Some proponents of in situ biotreatment claim to enhance phosphate mobility by supplementation with calcium or iron-seques- tering compounds (possibly EDTA or polyphos- phate salts could be used) but information is

proprietary and corroborative evidence is not available. Nitrate moves freely through soils but may be rapidly biologically sequestered.

The necessity for oxygen for biodegradation of halogenated organic compounds varies depending upon the contaminant. If oxygen is necessary, it may be very rapidly consumed by an active population at a cleanup location and oxygen supply may become the key limitation of biodegradation rate. The provision of oxygen in ground water can be achieved by direct sparging of the aquifer [1-3]. However, the limited solubility of both air and pure oxygen in water, typically 8-10 mg 1-1 and 35-50 mg 1-1, respectively, limits the efficiency of this process. The problems inherent in the supply of oxygen in aqueous solution have resulted in the suggestion that oxygen-yielding compounds may be advantageous. Hydrogen peroxide has been the most commonly employed but its use does have disadvantages. Firstly, it may be strongly inhibitory of microbial activity at relatively low concentrations. Even at 200 mg 1-~ some inhibition of biodegradation has been observed [2,131] and at above 500 mg 1-1 very severe inhibition occurs [130]. Secondly, the chemical reactivity of hydrogen peroxide limits its potential use. It reacts rapidly with many soil organics and inorganic salts to produce oxygenated products with a consequent loss of oxygen carrying capacity and the risk of precipitation of insoluble materials resulting in pore blockage [1,2]. Furthermore, breakdown by chemical reactions or soil-bound catalase and related enzymes may result in bubble formation which again may block pore spaces. Some purveyors of microbiological cleanup technologies claim to stabilise hydrogen peroxide solutions in soils by use of proprietary additives [132,133], including phosphate salts, but corroborative evidence is lacking. The degassing may also be eliminated by keeping the peroxide concentration at or below 100 mg 1-1 and thus decomposition will yield 50 mg 1-1 or less of molecular oxygen which should not exceed the solubility of oxygen in water [113,134]. However, in such cases the supply of oxygen will not be improved over the use of pure oxygen gas but may have advantages in terms of cost and ease of use. An alternative to the use of hydrogen peroxide is ozone [135] but

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this is difficult to handle, expensive to produce and likely to be just as prone to oxidation of soil compounds. The provision of oxygen to the unsaturated zone can be achieved by percolation of oxygenated solutions but the carrying capacity is limited. It may be preferable to force-aerate the soil by sinking perforated pipes and either pumping compressed air or pulling a vacuum [1,3,136]. Case histories of enhanced in situ biodegradation using oxygen and/or nutrient supplementation have been described. The principal field experi- ments described to date involve the cleanup of sites contaminated with gasoline which is both readily biodegradable and volatile [1,3]. However, reports demonstrate that the nutrient addition processes can be efficient and that sparging of ground water may be beneficial [137-139]. Re- ports of successful use of hydrogen peroxide and ozone have also been made [131-133,135,139]. There have been fewer attempts at using these methods to decontaminate sites containing halogenated organic compounds but microbiological cleanup of aquifers polluted with dichloromethane and other organic solvents has been reported [1,140,141]. Microbiological cleanup by enhancement of anaerobic degradative activity in situ has received less study. Anaerobiosis may be induced by enhanced microbial activity at a cleanup site and could be ensured by nitrogen- sparging of the groundwater or by flooding of the unsaturated zone. A trial under laboratory conditions using a sandy clay soil containing trichloroethane demonstrated that continuous feed with inorganic nutrients and ethanol resulted in the selection of a methanogenic community capable of complete degradation of the contaminant [142]. Under field conditions, soil flooding has been shown to enhance degradation of a variety of compounds including DDT [143], 2,4-D [144] and the pesticide toxaphene [145]. The use of induced denitrifying conditions to enhance anaerobic degradation of halogenated organics has not been reported. However there is a report of nitrate-supplementation resulting in enhanced degradation of mineral oil contamination in an aquifer [145a].

5.3. Analogue enrichment for co-metabolism An additional approach proposed for the in

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situ biodegradation of haloorganic compounds has been the enhancement of co-metabolism by the supply of the pr imary growth substrate. This approach may be particularly beneficial if the contaminant is only poorly degraded as a primary substrate but more rapidly metabolised by co- metabolism. It is the potential for enhanced degradation of certain haloliphatics by methane- oxidising bacteria [80-84,146] that has received the most attention in view of the ease with which methane supplementation can be achieved, the relative recalcitrance of many haloaliphatics under aerobic conditions and the low cost of methane gas. It has been clearly demonstrated that the passage of methane through soil columns may result in the degradation of halogenated ethenes [147] and limited-scale field trials of the technology have been performed [148]. For the field study, a confined sandy aquifer contaminated with a range of haloaliphatic compounds, principally trichloroethane and trichloroethene, was supplemented with alternate long pulses of methane and air. It was found that no inorganic nutrients were required, but if the supply of gases was simulta-

neous, methane oxidation occurred so rapidly that it resulted in excessive biomass production close to the injection points. This not only reduced permeability but also prevented long distance nutrient dispersal and could be avoided by the alternate additions employed. Enhanced degradation was found with a reduction in trichloroethene concentrations of 20-30% within one treatment season. It was suggested that degradation rate was limited by the relatively low solubilities of methane and air in water and the long adaptat ion period noted prior to the onset of biodegradation.

The use of primary substrates to enhance co- metabolism of haloaromatics has received less study, frequently because the substrates themselves may be environmental contaminants. En- hanced degradation under laboratory conditions has been demonstrated, for example, using aniline to increase 3,4-dichloroaniline degradation [52], biphenyl to enhance PCB degradation [68,149,150] and phenol to enhance 4-chlorophenol degradation [150]. Since co-metabolism of halogenated organic compounds may frequently result in a biodegradation rate superior to that obtained when

Wate

Fig. 4. Diagrammatic representation of the principles involved in a biotreatment system involving a sunace bioreactor. Ground water is withdrawn, supplemented with nutrients and introduced to a suitable bioreactor. The effluent is allowed to reinfiltrate via a trench

system.

the compound is used as sole carbon and energy source the use of analogue enrichment is worthy of further investigation. This approach may be especially beneficial when the contaminant is present at concentrations too low to induce degradative activity.

5. 4. Use of surface bioreactors As an alternative to enhancing degradation in

situ, it may be possible to extract the ground water and treat this in a bioreactor prior to rein- filtration. The technology may be used in isolation or as an extension to an in situ process (Fig. 4). A significant problem that may be encountered with extraction processes is that many haloorganic compounds may strongly sorb to soil and will be difficult to remove. Mobilisation of contaminants by percolation with detergent solutions has been proposed but detergents may cause soil pore blockage due to precipitation and will themselves require biodegradation [2,3].

The removal of haloorganics in bioreactors has received extensive study in respect of the treatment of industrial effluent streams and it is possible to make general comments about contaminant degradability in such systems. A wide variety of compounds are degradable in bioreactors, including halophenols, halobenzoates, monochlorobe- nzene and a number of pesticides [151,152]. Haloaliphatics can be degraded rapidly under methanogenic conditions [92-93,152a] but vapour-phase stripping is the most important removal mechanism for volatile contaminants under most conditions [140,153]. It should be noted that sorption of halogenated organics to sludge may be very significant and such sludge would need careful disposal. Specific reports of site decontamination by use of bioreactor systems are relatively few. A complex mixture of materials in a waste water lagoon was reported to be efficiently degraded once the lagoon was aerated and supplemented with nutrients to produce a type of activated sludge system [154]. However, this paper did not differentiate between biodegradation and loss by volatilisation. Degradation of halomethanes in ground water introduced into a bioreactor was found to be unimportant relative to vapour-phase removal [1,140]. A system has been described using

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a combined biotreatment system involving ground water extraction with a feed into an aerobic bioreactor, the effluent from which was nutrient- and oxygen-supplemented and allowed to reinfiltrate into the soil in order to enhance in situ activity [123]. Similar processes have successfully cleaned haloaliphatic-contaminated aquifers [140,141] but the relative contributions of biodegradation and vapour-phase removal are unclear.

Whilst further studies are evidently required, available evidence suggests that, provided treatment is specifically tailored to the contaminants present at a site, degradation within bioreactors should proceed efficiently. A variety of reactors may be employed under aerobic or anaerobic conditions, as appropriate, including activated sludge plants, trickling filters, rotating disk reactors, lagoons and fluidised bed systems [155,156]. Fur- thermore, the use of sequential aerobic and anaerobic systems may permit optimal degradation of different components within a complex mixture.

5.5. The use of microbial inocula to enhance biological cleanup

At many contaminated locations there may exist a degradative microbial population that has adapted to the introduced compounds. It is the intention of in situ microbiological cleanup to utilise these degradative organisms by optimising environmental conditions for their activity. How- ever, at sites contaminated with highly recalcitrant materials or where local environmental conditions have prevented the development of a degradative population, it may be considered advantageous to inoculate the subsurface with degradative microorganisms. Two strategies for strain provision have been employed, classical enrichment techniques and molecular biology. The latter techniques has been shown to be successful in producing haloaromatic-degrading strains but their use as inocula under field conditions is likely to be re- stricted by political considerations. Before consid- ering strain deployment, it is valuable to discuss the practicalities of inoculation, particularly for subsoil and ground water where addition of cells is not straightforward.

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The first consideration prior to the use of inocula under field conditions is the competitiveness of the introduced strains. Introduced organisms may lose degradative activity in soils if they pref- erentially utilise more readily degradable carbon sources in situ. Furthermore, introduced organisms may be killed off by microbial toxins, by preda- tion, if they are unable to sequester nutrients competitively or if they are unable to utilise excessively high or low substrate concentrations [157]. A second difficulty in the subsurface is the introduction of cells to the contaminated sites. Injec- tion of a cell suspension has most commonly been proposed but, unless the site is highly permeable, the migration of organisms through a particulate matrix may be relatively poor. Unicellular organisms can only move within water films and under non-saturated conditions movement of even the most motile organisms is likely to be limited to a few centimetres per day [158,159]. Under saturated conditions migration is significantly greater, particularly if the water is fast-flowing. Even so, migration of unicellular microbes is re- stricted by filtration effects and sorption to clay minerals [160-162] and such processes will not only limit distribution but may result in pore blockage and a consequent inhibition of the cleanup process. It has been suggested that spores or miniaturised bacterial cells induced by starva- tion (ultramicrobacteria) [163] could overcome some of these difficulties by virtue of their small sizes and potential for resuscitation in situ. Little information is available, although it has been shown that miniaturised bacterial cells could be transported through soil and rock formations and be induced to grow in situ by nutrient supplementation [164,165]. Also of interest as inocula are mycelial microorganisms since the penetration of mycelial filaments through soils may be extremely efficient, even under relatively dry conditions [166]. The potential for the utilisation of one group of fungi, the white-rot organisms, is discussed in section 5.6, below.

The successful use of microbial inocula to enhance cleanup of haloorganic-contaminated soil has been reported. The addition o f Rhodococcus chlorophenolicus to a composting process for cleanup of chlorophenol-contaminated soil en-

hanced biodegradation under laboratory conditions but not in the field [127]. In contrast, pentachlorophenol degradation in topsoil was enhanced under both laboratory and field conditions following inoculation with an Arthrobacter sp. [167] and a Flavobacterium sp. [168]. It was noted that high substrate concentrations proved inhibitory to the degradative strains and that repeated inoculations were needed to maintain biodegradation rates. The degradation of 2,4,5-T was found to be enhanced in soil inoculated with a strain of Pseudo-

monas cepacia and this organism died rap id ly in its absence [169]. The inoculation of a Pseudo-

monas sp. into a soil polluted with chlorobenzenes resulted in the rapid degradation of all of the mono- and dichlorobenzenes present and 1,2,4-trichlorobenzene [170]. The strain was found to survive for approximately 60 days in soil free of chlorobenzenes but the degradative capacity was rapidly lost. Inoculation of PCB-contaminated soil with an Acinetobacter sp. was found to enhance biodegradation of some congeners within the mixture but others remained unmetabolised [68,149]. Addition of a bromacil-degrading Pseudomonas

sp. to contaminated soil under laboratory conditions was found to result in degradation of bromacil from 50 mg kg-1 to below the detection limit within one week [170a]. The use of inocula in bioreactor systems used for site decontamination has also been considered [171] but the need for inoculation is variable and ensuring strain competitiveness is difficult.

The problems with the recalcitrance of certain haloorganic contaminants and the frequent re- quirement for microbial consortia to ensure complete mineralisation [172] has led to the utilisation of genetic techniques in attempts to construct more efficient degradative strains. Natural genetic exchange has been most commonly employed. Pseudomonas B13 has been particularly studied because it can utilise some chlorobenzoates and chlorophenols but cannot use all isomers. By transferring the pWWO plasmid, which codes for a broad substrate range but not chloroorganic degradation, from Pseudomonas mt-2, into B13, it proved possible to construct a strain with a wider chloroaromatic substrate range [173]. A similar experiment utilising B13 and strain WR401, which

was capable of utilising methyl-substituted salicylates, resulted in the construction of a strain capable of degrading chlorosalicylates [174]. It has also proved possible to construct 4-chloro-2-nitrophenol-degrading strains by conjugating the nitrophenol-utilising Pseudomonas N31 with either of the chlorophenol-degrading strains Pseudomonas B13 or Alcaligenes eutrophus JMP134 [175]. The use of an undefined mixture of cells from sites contaminated with 2,4,5-T resulted in the selection of strains capable of utilising this compound as sole carbon and energy source [176]. Conjugation between a DDT-degrading strain of Pseudomonas and strains carrying plasmids for naphthalene metabolism resulted in the production of strains which could more rapidly metabolise DDT and one strain which could degrade kelthane (1,1-bis- (p-chlorophenyl)-2,2,2-trichloroethanol) [177]. It was subsequently demonstrated [178] that the kelthane-degrading strain could be used under laboratory conditions for the cleanup of soil contaminated with this compound.

The utilisation of constructed strains as inocula may indeed permit more effective microbiological cleanup of contaminated soil and ground water. The full potential of the technology has yet to be realised [179,180] and further study of strain stability in situ is required [181].

5. 6. The utilisation of white-rot fungi for haloorganic breakdown

The white-rot fungi are a group of Basidiomy- cetes capable of degrading lignin which is a random polymer of phenolic subunits. In order to cleave lignin, an extracellular 'ligninase' is produced. This enzyme generates hydroxyl and other oxygen radicals from hydrogen peroxide and these are responsible for random cleavage of the lignin molecule to produce soluble intermediates [182]. It has been shown that this random chemical attack mediated by the enzyme can also cleave a variety of otherwise relatively recalcitrant xenobiotic compounds. Indeed, complete mineralisation of the compounds may occur. The most commonly studied fungus in biodegradation studies has been Phanerochaete chrysosporium but available evidence suggests that other white-rot fungi have the same range of degradative capacity. Among the

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compounds shown to be degraded are simple chlorophenols, chlorobenzoates, chloroanilines and pentachlorophenol [182,183], dichloroaniline [184], D D T [185,186], hexachlorocyclohexane [186], 2,3,7,8-tetrachlorodibenzo-p-dioxin [186] and polychlorinated biphenyls [186,187]. There is some evidence to suggest that the degradation of D D T by P. chrysosporium is due to an enzyme system other than ligninase since degradation has been noted in the absence of ligninase expression [187a]. The environmental conditions that result in the induction of ligninase production are not entirely clear but frequently in laboratory culture it is nitrogen-limitation that brings about maximum activity. Whilst further research is necessary to assess the potential of white-rot fungi for biotreatment processes, particularly with respect to enzyme activity and survival in situ, it is evident that these organisms may have uses in the cleanup of contaminated land. The use of P. chrysosporium in a rotating disc bioreactor for the treatment of chloropheno!ic effluents from paper manufacture has been described [188] and a similar system may be of benefit for ground water decontamination. An alternative approach may be to inoculate strains into contaminated soil. In such a situation, mycelial penetration may serve to distribute the organisms [166] and it is possible that extracellular enzyme activity may permit degradation of soil- bound contaminants which would not otherwise be available for metabolism.

6. THE POTENTIAL FOR MICROBIOLOGI- CAL CLEANUP OF C O N T A M I N A T E D SOIL AND G R O U N D W A T E R

Microbiological cleanup is a developing technology founded upon basic principles of microbial ecology and physiology. It is not a universal pan- acea and cannot be applied at many locations for a variety of reasons. It has been the aim of this article to describe the biodegradation capacity of microorganisms from an ecological viewpoint and to illustrate the methods which may be applied to optimise degradation rates under field conditions.

In conclusion, it is necessary to draw together these themes and endeavour to realistically assess

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the advantages and limitations of the technologies. In cannot be overemphasised that every contaminated location is unique and the applicability of any cleanup technology will vary from site to site. The initial consideration for the biotechnolo- gist in site assessment is the degradability of the contaminants. Many compounds are degradable or cometabolisable under either or both aerobic and anaerobic conditions but once the potential for degradability is assessed it is necessary to address the question of sustainable degradation rates and to set a realistic timescale for a proposed treatment [153]. The environmental persistence of such materials as D D T and PCBs is evidence of microbial fallibility, and biological cleanup of sites contaminated with this type of compound is unlikely to be generally feasible unless, at best, an extremely long treatment period is acceptable. Having assessed the degradability of the compound it is necessary to consider its distribution and availability. It may be the case that the compounds are strongly sorbed or extremely insoluble, in which case degradation may be limited by dissolution [189,190]. The subsurface itself may prevent the use of biological treatment methodologies by being of low permeability and therefore unsuitable for the flow of air or aqueous media necessary in the processes [1-3,130].

Having established that biotreatment is a viable option it is necessary to select a suitable technology. For more highly substituted aromatic molecules and for haloalkanes, it may be beneficial to employ anaerobic technologies and for chlo- roethenes enhancement of methane-oxidation to enable co-metabolism of the contaminants may be a useful approach. The selection of in situ enhancement or extraction-biotreatment processes may be made on the basis of determination of contaminant distribution and degradative activity in the subsurface. Also essential for in situ enhancement are determinations of the potential impacts on permeability of the use of inorganic nutrients or hydrogen peroxide. The selection and optimisation of a microbial cleanup technology requires careful microbiological and hydrogeologi- cal investigations but if applied correctly field evidence suggests that treatment processes for topsoil [127-129] and for subsoils and ground

water [1-3,131-133,135,137-139] may be successful.

What are the potential advantages and disadvantages of in situ biological cleanup? The first advantage is that, unless a land farming or corn- posting procedure is chosen, there is minimal site disruption beyond the sinking of bore holes or the positioning of infiltration sites. This may render the technology highly suitable for built-up areas and for industrial sites. Secondly the process should be safe, provided that ground water outflow is controlled, since normally only indigenous microbes and non-persistent, non-toxic nutrient additions are made. The use of inocula, if required, should pose no hazards if the strains have been isolated from the environment. The use of genetically modified strains is potentially more politically contentious. Thirdly, for simple molecules the process should be relatively rapid and efficient since the production of poorly degradable intermediates is unlikely. Indeed, complete mineralisation is possible. Potential disadvantages include the persistence of poorly degradable compounds. Although the exciting developments in our knowledge of the degradative capabilities of white-rot fungi [182-188] may illustrate some potential for the biodegradation of otherwise recalcitrant molecules, the technology is relatively untested and biotreatment at such sites is unlikely to be a viable option in the immediate future. A further disadvantage is the inability of the technologies to function at poorly permeable locations but, similarly, few physicochemical treatment methods are suitable at such sites.

It is evident that in situ microbiological decontamination is a technology of great potential and that the microbiologist has a major role to play in the development and application of methods. Our understanding of subsurface microbiology is clearly incomplete and further research should also address such aspects as inoculum survival and dispersal, fungal activity, anaerobic degradation, nutrient supply and the effects of pollutant availability. Calculations suggest that microbiological cleanup may be a cost-effective method of site decontamination [191] but successful application will depend upon the existence of a sound founda- tion of microbiological data.

REFERENCES

[1] Thomas, J.M., Lee, M.D., Bedient, P.B., Borden, R.C., Canter, L.W. and Ward, C.H. (1987) Leaking Under- ground Storage Tanks: Remediation with Emphasis on in situ Biorestoration. U.S. Environmental Protection Agency Report No. EPA/600/2-87/008. Robert S. Kerr Environmental Research Laboratory, Ada, OK.

[21 Lee, M.D., Wilson, J.T. and Ward, C.H. (1987) In situ restoration techniques for aquifers contaminated with hazardous wastes. J. Hazard. Mater. 14, 71-82.

[2] (a) Bouwer, E., Mercer, J., Kavanaugh, M. and Digiano, F. (1988) Coping with ground water contamination. J. Water Pollut. Control Fed. 60, 1415-1427.

[3] Morgan, P. and Watkinson, R.J. (1989) Hydrocarbon degradation in soils and methods for soil biotreatment. CRC Crit. Rev. Biotechnol. 8, 305-333.

[4] Rochkind-Dubinsky, M.L., Sayler, G.S. and Blackburn, J.W. (1987) Microbiological Decomposition of Chlo- rinated Aromatic Compounds. Marcel Dekker, New York.

[5] Lal, R. (1984) Insecticide Microbiology. Spinger-Verlag, Berlin.

[6] Gibson, D.T. (1984) Microbial Degradation of Organic Compounds. Marcel Dekker, New York.

[7] Bouwer, E.J. and McCarty, P.L. (1982) Removal of trace chlorinated organics by activated carbon and fixed film bacteria. Environ. Sci. Technol. 16, 836-843.

[8] Bouwer, E.J. (1984) Biotransformation of organic micro- pollutants in the subsurface. Proc. NWWA/API Conf. Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Restoration, Nov. 1984, pp. 66-80. N W W A , Worthington, OH.

[9] Reineke, W. and Knackmuss, H.-J. (1984) Microbial metabolism of haloaromatics: isolation and properties of a chlorobenzene-degrading bacterium. Appl. Environ. Microbiol. 47, 395-402.

[10] Spain, J.C. and Nishino, S.F. (1987) Degradation of 1,4-dichlorobenzene by a Pseudornonas sp. Appl. En- viron. Microbiol. 53, 1010-1019.

[11] de Bont, J.A.M., Vorage, M.J.A.W., Hartmans, S. and van den Tweel, W.J.J. (1986) Microbial degradation of 1,3-dichlorobenzene. Appl. Environ. Microbiol. 52, 677-680.

[12] Haigler, B.E., Nishino, S.F. and Spain, J.C. (1988) De- gradation of 1,2-dichlorobenzene by a Pseudomonas sp. Appl. Environ. Microbiol. 54, 294-301.

[13] Tiedje, J.M., Boyd, S.A. and Fathepure, B.Z. (1987) Anaerobic degradation of chlorinated aromatic hydrocarbons. Dev. Ind. Microbiol. 27, 117-127.

[14] Fathepure, B.Z., Tiedje, J.M. and Boyd, S.A. (1988) Reductive dechlorination of hexachlorobenzene to tri- and dichlorobenzenes in anaerobic sewage sludge. Appl. Environ. Microbiol. 54, 327-330.

[15] Reineke, W. (1984) Microbial metabolism of halogenated aromatic compounds, in Microbial Degradation of Organic Compounds (Gibson, D.T., ed.), pp. 319-360, Marcel Dekker, New York.

293

[16] Klages, U. and Lingens, F. (1979) Degradation of 4-chlo- robenzoic acid by a Nocardia species. FEMS Microbiol. Lett. 6, 201-203.

[17] Dorn, E., Hellwig, M., Reineke, W. and Knackmuss, H.-J. (1974) Isolation and characterization of a 3-chlorobenzoate-degrading pseudomonad. Arch. Microbiol. 99, 61-70.

[18] Hartmann, J., Reineke, W. and Knackmuss, H.-J. (1979) Metabolism of 3-chloro-, 4-chloro- and 3,5-dichlorobenzoate by a pseudomonad. Appl. Environ. Microbiol. 37, 421-428.

[19] Milne, G.W.A., Goldman, P. and Holzman, J.L. (1967) The metabolism of 2-fluorobenzoic acid, II. Studies with 1802. J. Biol. Chem. 246, 11232-11238.

[20] Crawford, R.L., Olsen, P.E. and Frick, T.D. (1979) Catabolism of 5-chlorosalicylate by a Bacillus isolated from the Mississippi river. Appl. Environ. Microbiol. 38, 379-384.

[21] Schennen, U., Braun, K. and Knackmuss, H.-J. (1985) Anaerobic degradation of 2-fluorobenzoate by benzoate-degrading, dentrifying bacteria. J. Bacteriol. 161, 321-325.

[22] van den Tweel, W.J.J., Kok, J.B. and de Bont, J.A.M. (1987). Reductive dechlorination of 2,4-dichlorobenzoate to 4-chlorobenzoate and hydrolytic dehalogenation of 4-chloro-, 4-bromo-, and 4-iodobenzoate by Alcaligenes dentrificans NTB-1. Appl. Environ. Microbiol. 53, 810-815.

[23] Horowitz, A., Suflita, J.M. and Tiedje, J.M. (1983) Re- ductive dehalogenation of halobenzoates by anaerobic lake sediment microorganisms. Appl. Environ. Micro- biol. 45, 1459-1465.

[24] Suflita, J.M., Horowitz, A., Shelton, D.R. and Tiedje, J.M. (1982) Dehalogenation: a novel pathway for the anaerobic biodegradation of haloaromatic compounds. Science 218, 1115-1117.

[25] de Weerd, K.A., Suflita, J.M., Linkfield, T., Tiedje, J.M. and Pritchard, P.H. (1986) The relationship between reductive dehalogenation and other substituent removal reactions catalyzed by anaerobes. FEMS Microbiol. Ecol. 38, 331-339.

[26] Ide, A., Niki, Y., Sakamoto, F., Watanabe, I. and Watanabe, H. (1972) Decomposition of pentachlorophenol in paddy soil. Agric. Biol. Chem. 36, 1937-1944.

[27] Steiert, J.G. and Crawford, R.L. (1985) Microbial degradation of chlorinated phenols. Trends Biotechnol. 3, 300-305.

[28] Steiert, J.G., Pignatello, J.J. and Crawford, R.L. (1987) Degradation of chlorinated phenols by a pentachlorophenol-degrading bacterium. Appl. Environ. Microbiol. 53, 907-910.

[29] Pignatello, J.J., Martinson, M.M., Steiert, J.G., Carlson, R.E. and Crawford, R.L. (1983) Biodegradation and photolysis of pentachlorophenol in artificial freshwater streams. Appl. Environ. Microbiol. 46, 1028-1031.

[30] Suflita, J.M. and Miller, G.D. (1985) The microbial metabolism of chlorophenolic compounds in ground water aquifers. Environ. Toxicol. Chem. 4, 751-758.

294

[31] Schmidt, E., Hellwig, E. and Knackmuss, H.-J. (1983) Degradation of chlorophenols by a defined mixed microbial community. Appl. Environ. Microbiol. 46, 1038- 1044.

[32] Krumme, M.L. and Boyd, S.A. (1988) Reductive dechlorination of chlorinated phenols in anaerobic upflow bioreactors. Water Res. 22, 171-177.

[33] Boyd, S.A. and Shelton, D.R. (1984) Anaerobic biodegradation of chlorophenols in fresh and acclimated sludge. Appl. Environ. Microbiol. 47, 272-277.

[34] Sahm, H., Brunner, M. and Schoberth, S.M. (1986) Anaerobic degradation of halogenated aromatic compounds. Microb. Ecol. 12, 147-153.

[35] Mikesell, M.D. and Boyd, S.A. (1986) Complete reductive dechlorination and mineralisation of pentachlorophenol by anaerobic microorganisms. Appl. Environ. Microbiol. 52, 861-865.

[36] Guthrie, M.A., Kirsch, E.J., Wukasch, R.F. and Grady, C.P.L. (1984) Pentachlorophenol biodegradation - - II anaerobic. Water Res. 18, 451-461.

[37] Schwien, U., Schmidt, E., Knackmuss, H.-J. and Re- ineke, W. (1988) Degradation of chlorosubstituted aromatic compounds by Pseudomanas sp. strain B13: fate of 3,5-dichlorocatechol. Arch. Microbiol. 150, 78-84.

[38] Fathepure, B.Z., Tiedje, T.M. and Boyd, S.A. (1987) Reductive dechlorination of 4-chlororesorcinol by anaerobic microorganisms. Environ. Toxicol. Chem. 6, 929-934.

[39] Hill, I.R. and Wright, S.J.L. (1978) Pesticide Microbi- ology. Academic Press, London.

[40] Lal, R. and Saxena, D.M. (1982) Accumulation, metabolism, and effects of organochlorine insecticides on microorganisms. Microbiol. Rev. 46, 95-127.

[41] Evans, W.C., Smith, B.S.W., Fernley, H.N. and Davies, J.I. (1971) Bacterial metabolism of 2,4-dichloropheno- xyacetate. Biochem. J. 122, 543-551.

[42] Sinton, G.L., Fan, L.T., Erickson, L.E. and Lee, S.M. (1986) Biodegradation of 2,4-D and related compounds. Enzyme Microb. Technol. 8, 395-403.

[43] Kilbane, J.J., Chatterjee, D.K., Karns, J.S., Kellogg, S.T. and Chakrabarty, A.M. (1982) Biodegradation of 2,4,5- trichlorophenoxyacetic acid by a pure culture of Pseudo- monas cepacia. Appl. Environ. Microbiol. 44, 72-78.

[44] Rosenberg, A. and Alexander, M. (1980) Microbial metabolism of 2,4,5-trichlorophenoxyacetic acid in soil, soil suspensions, and axenic culture. J. Agric. Food Chem. 28, 297-302.

[45] McCall, P.J., Vrona, S.A. and Kelley, S.S. (1981) Fate of uniformly carbon-14 ring labelled 2,4,5-trichlorophenoxyacetic acid and 2,4-dichlorophenoxyacetic acid. J. Agric. Food Chem. 29, 100-107.

[46] Mikesell, M.D. and Boyd, S.A. (1985) Reductive dechlorination of the pesticides 2,4-D, 2,4,5-T and pentachlorophenol in anaerobic sludges. J. Environ. Qual. 14, 337-340.

[47] Barik, S. (1984) Metabolism of insecticides by microorganisms, in Insecticide Microbiology (Lal, R., ed.), pp. 87-130, Springer-Verlag, Berlin.

[48] Focht, D.D. (1972) Microbial degradation of DDT metabolites to carbon dioxide, water, and chloride. Bull. Environ. Contam. Toxicol. 7, 52-56.

[49] Fogel, S., Lancione, R.L. and Sewall, A.E. (1982) En- hanced biodegradation of methoxychlor in soil under sequential environmental conditions. Appl. Environ. Mi- crobiol. 44, 113-120.

[50] You, I.S. and Bartha, R. (1982) Metabolism of 3,4-dichloroaniline by Pseudomonas putida. J. Agric. Food Chem. 30, 274-277.

[51] Saxena, A. and Bartha, R. (1983) Microbial mineralization of humic acid-3,4-dichloroaniline complexes. Soil Biol. Biochem. 15, 59-62.

[52] You, I.S. and Bartha, R. (1982) Stimulation of 3,4-dichloroaniline mineralization by aniline. Appl. Environ. Microbiol. 44, 678-681.

[53] Zeyer, J. and Kearney, P.C. (1982) Microbial metabolism of propranil and 3,4-dichloroaniline. Pest. Biochem. Physiol. 17, 215-223.

[54] Hwang, H.-M., Hodson, R.E. and Lee, R.F. (1985) Pho- tochemical and microbial degradation of 2,4,5-trichloroaniline in a freshwater lake. Appl. Environ. Microbiol. 50, 1177-1180.

[55] Furukawa, R., Tomizuka, N. and Kamibayashi, A. (1979) Effect of chlorine substitution on the bacterial metabolism of various polychlorinated biphenyls. Appl. En- viron. Microbiol. 38, 301-310.

[56] Masse, R., Messier, F., Prloquin, L., Ayotte, C. and Sylvestre, M. (1984) Microbial biodegradation of 4-chlorobiphenyl, a model compound of chlorinated biphenyls. Appl. Environ. Microbiol. 47, 947-951.

[57] Yagi, O. and Sudo, R. (1980) Degradation of polychlorinated biphenyls by microorganisisms. J. Water Pollut. Control Fed. 52, 1035-1043.

[58] Furukawa, K. and Chakrabarty, A.M. (1982) Involve- ment of plasmids in total degradation of chlorinated biphenyls. Appl. Environ. Microbiol. 44, 619-626.

[59] Sylvestre, M., Masse, R., Ayotte, C., Messier, F. and Fauteaux, J. (1985) Total biodegradation of 4-chlorobiphenyl (4CB) by a two-membered bacterial culture. Appl. Microbiol. Biotechnol. 21, 192-195.

[60] Barton, M.R. and Crawford, R.L. (1988) Novel biotrans- formations of 4-chlorobiphenyl by a Pseudomonas sp. Appl. Environ. Microbiol. 54, 594-595.

[61] Bedard, D.L., Haberl, M.L., May, R.J. and Brennan, M.J. (1987) Evidence for novel mechanisms of polychlorinated biphenyl metabolism in Alcaligenes eutrophus H850. Appl. Environ. Microbiol. 53, 1102-1112.

[62] Furukawa, K. and Matsumura, F. (1976) Microbial metabolism of polychlorinated biphenyls. Studies on the relative degradability of polychlorinated biphenyl components by Alcaligenes sp. J. Agric. Food Chem. 24, 251-256.

[63] Safe, S.H. (1984) Microbial degradation of polychlorinated biphenyls, in Microbial Degradation of Organic Compounds (Gibson, D.T., ed.), pp. 361-369, Marcel Dekker, New York.

[64] Bedard, D.L., Unterman, R., Bopp, L.H., Brennan, M.J.,

Haberl, M.L. and Johnson, C. (1986) Rapid assay for screening and characterizing microorganisms for the ability to degrade polychlorinated biphenyls. Appl. Environ. Microbiol. 51, 761-768.

[65] Dmochewitz, S. and Ballschmiter, K. (1988) Microbial transformation of technical mixtures of polychlorinated biphenyls (PCB) by the fungus Aspergillus niger. Chem- osphere 17, 111-121.

[66] Bopp, L.H. (1986) Degradation of highly chlorinated PCBs by Pseudomonas strain LB400. J. Ind. Microbiol. 1, 23-29.

[67] Fairbanks, B.C., O'Connor, G.A. and Smith, S.E. (1987) Mineralization and volatilization of polychlorinated biphenyls in sludge-amended soils. J. Environ. Qual. 16, 18-25.

[68] Focht, D.D. and Brunner, W. (1985) Kinetics of biphenyl and polychlorinated biphenyl metabolism in soil. Appl. Environ. Microbiol. 50, 1058-1063.

[69] Moein, G.J., Smith, A.J. and Stewart, P.L. (1976) Fol- low-up study of the distribution and fate of polychlorinated biphenyls and benzenes in soil and ground water samples after an accidental spill of transformer fluid. Proc. 3rd Nat. Conf. Control Hazard Mater. Spills, pp. 368-372, Information Transfer Inc., Rockville, Mary- land.

[70] Brown, J.F., Bedard, D.L., Brennan, M.J., Carnahan, J.C., Feng, H. and Wagner, R.E. (1987) Polychlorinated biphenyl degradation in aquatic sediments. Science 236, 709-712.

[71] Omori, T., Kimura, T. and Kodama, T. (1987) Bacterial cometabolic degradation of chlorinated paraffins. Appl. Microbiol. Biotechnol. 25, 553-557.

[72] Yokota, T., Fuse, H., Omori, T. and Minoda, Y. (1986) Microbial dehalogenation of haloalkanes mediated by oxygenase or halidohydrolase. Agric. Biol. Chem. 50, 453-460.

[73] Janssen, D.B., Jager, D. and Witholt, B. (1987) Degrada- tion of n-haloalkanes and a, o~-dibaloalkanes by wild- type and mutants of Acinetobacter sp. strain G J70. Appl. Environ. Microbiol. 53, 561-566.

[74] Omori, T. and Alexander, M. (1978) Bacterial dehalogenation of halogenated alkanes and fatty acids. Appl. Environ. Microbiol. 35, 867-871.

[75] Scholtz, R., Schmuckle, A., Cook, A.M. and Leisinger, T. (1987) Degradation of eighteen 1-monohaloalkanes by Arthrobacter sp. strain HA1. J. Gen. Microbioi. 133, 267-274.

[76] Janssen, D.B., Scheper, A., Dijkhuizen, L. and Witholt, B. (1985) Degradation of halogenated aliphatic compounds by Xanthobacter autotrophicus G J10. Appl. En- viron. Microbiol. 49, 673-677.

[77] Goldman, P. (1965) The enzymatic cleavage of the carbon-fluorine bond in fluoroacetate. J. Biol. Chem. 240, 3434-3438.

[78] Kawasaki, H., Tone, N. and Tonomura, K. (1981) Purifi- cation and properties of haloacetate halidohydrolase specified by plasmid from Moraxella sp. strain B. Agric. Biol. Chem. 45, 29-34.

295

[79] Dalton, H. and Stirling, D.I. (1982) Co-metabolism. Phil. Trans. R. Soc. Lond. B 297, 481-491.

[80] Fogel, M.M. Taddeo, A.R. and Fogel, S. (1986) Biode- gradation of chlorinated ethenes by a methane-utilizing mixed culture. Appl. Environ. Microbiol. 51,720-724.

[81] Pignatello, J.J. (1987) Microbial degradation of 1,2-dibromoethane in shallow aquifer materials. J. Environ. Qual. 16, 307-312.

[82] Fogel, S., Findlay, M., Moore, A. and Leahy, M. (1987) Biodegradation of chlorinated chemicals in groundwater by methane oxidising bacteria. Proc. NWWA/API Conf. Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Restoration, Nov. 1987, pp. 187-205, NWWA, Dublin, OH.

[83] Little, C.D., Palumbo, A.V., Herbes, S.E., Lidstrom, M.E., Tyndall, R.L. and Gilmer, P.J. (1988) Trichloro- ethylene biodegradation by a methane-oxidizing bacterium. Appl. Environ. Microbiol. 54, 951-956.

[84] Kiecka, G.M. (1982) Fate and effects of methylene chloride in activated sludge. Appl. Environ. Microbiol. 44, 701-707.

[85] Brunner, W., Staub, D. and Leisinger, T. (1980) Bacterial degradation of dichloromethane. Appl. Environ. Micro- biol. 40, 950-958.

[86] Stucki, G., G~illi, R., Ebersold, H.-R. and Leisinger, T. (1981) Dehalogenation of dichloromethane by cell ex- tracts of Hyphomicrobium DM2. Arch. Microbiol. 130, 366-371.

[87] Nelson, M.J.K., Montgomery, S.O., Mahaffey, W.R. and Pritchard, P.H. (1987) Biodegradation of trichloroethylene and the involvement of an aromatic biodegradative pathway. Appl. Environ. Microbiol. 53, 949-954.

[88] Nelson, M.J.K., Montgomery, S.O. and Pritchard, P.H. (1988) Trichloroethylene metabolism by microorganisms that degrade aromatic compounds. Appl. Environ. Mi- crobiol. 54, 604-606.

[89] Wackett, L.P. and Gibson, D.T. (1988) Degradation of trichloroethylene by toluene dioxygenase in whole-cell studies with Pseudomonas putida F1. Appl. Environ. Microbiol. 54, 1703-1708.

[90] Bouwer, E.J. and McCarty, P.L. (1983) Transformations of halogenated organic compounds under dentitrification conditions. Appl. Environ. Microbiol. 45, 1295-1299.

[91] Bouwer, E.J. and McCarty, P.L. (1983) Transformations of 1- and 2-carbon halogenated aliphatic organic compounds under methanogenic conditions. Appl. Environ. Microbiol. 45, 1286-1294.

[92] Vogel, T.M. and McCarty, P.L. (1987) Abiotic and biotic t ransformations of 1,1,1-trichloroethylene under methanogenic conditions. Environ. Sci. Technol. 21, 1208-1213.

[93] Fathepure, B.Z. and Boyd, S.A. (1988) Reductive dechlorination of perchloroethylene and the role of methano-. gens. FEMS Microbiol. Lett. 49, 149-156.

[94] Barrio-Lage, G.A., Parsons, F.Z., Nassar, R.S. and Lorenzo, P.A. (1987) Biotransformation of trichloroethene in a variety of subsurface materials. Environ. Toxicol. Chem. 6, 571-578.

296

[95] Bouwer, E.J. (1984) Biotransformation of organic micro- pollutants in the subsurface. Proc. NWWA/API Conf. Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Restoration, Nov. 1984, pp. 66-80, NWWA, Worthington, OH.

[96] Wood, P.R., Lang, R.F. and Payan, I.L. (1985) Anaerobic transformation, transport, and removal of volatile chlorinated organics in ground water, in Ground Water Quality (Ward, C.H., Giger, W. and McCarty, P.L., eds.), pp. 493-511, Wiley, New York.

[97] Kleopfer, R.D., Easley, D.M., Haas, B.B., Deihl, T.G., Jackson, D.E. and Wurrey, C.J. (1985) Anaerobic degradation of trichloroethylene in soil. Environ. Sci. Tech- nol. 19, 277-280.

[98] Schwarzenbach, R.P. and Giger, W. (1985) Behaviour and fate of halogenated hydrocarbons in ground water, in Ground Water Quality (Ward, C.H., Giger, W. and McCarty, P.L., eds.), pp. 446-471, Wiley, New York.

[99] Bouwer, E.J. and Wright, J.P. (1988) Transformations of trace halogenated aliphatics in anoxic biofilm columns. J. Contam. Hydrol. 2, 155-169.

[100] Bachman, A., de Bruit, W., Jumelet, J.C., Rijnaarts, H.H.N. and Zehnder, A.J.B. (1988) Aerobic biominerali- zation of alpha-hexachlorocyclohexane in contaminated soil. Appl. Environ. Microbiol. 54, 548-554.

[101] MacRae, I.C., Raghu, C.K. and Castro, T.F. (1967) Persistence and biodegradation of four common isomers of benzene hexachloride in submerged soils. J. Agric. Food Chem. 15, 911-914.

[102] Bollag, J.-M. and Loll, M.J. (1983) Incorporation of xenobiotics into soil humus. Experientia 39, 1221-1231.

[103] Wolf, D.C. and Martin, J.P. (1976) Decomposition of fungal mycelia and humic-type polymers containing carbon-14 from ring and side-chain labelled 2,4-D and chlorpropham. Soil Sci. Soc. Am. Proc. 40, 700-704.

[104] Bollag, J.-M., Liu, S.-Y. and Minard, R.D. (1980) Cross- coupling of phenolic humus constituents and 2,4-dichlorophenol. Soil Sci. Soc. Am. J. 44, 52-56.

[105] Bollag, J.-M., Sj~blad, R.D. and Minard, R.D. (1977) Polymerization of phenolic intermediates of pesticides by a fungal enzyme. Experientia 33, 1564-1566.

[106] Hsu, T.-S. and Bartha, R. (1974) Interaction of pesticide-derived chloroaniline residues with soil organic matter. Soil Sci. 116, 444-452.

[107] Hsu, T-.S. and Bartha, R. (1976) Hydrolyzable and non- hydrolyzable 3,4-dichloroaniline complexes and their re- spective rates of biodegradation. J. Agric. Food Chem. 24, 118-122.

[108] Ballard, T.M. (1971) Role of humic carrier substances in DDT movement through forest soil. Soil Sci. Soc. Am. Proc. 35, 145-147.

[109] Mortland, M.M. (1985) Interaction between organic molecules and mineral surfaces, in Ground Water Qual- ity (Ward, C.H., Giger, W. and McCarty, P.L., eds.), pp. 370-386, Wiley, New York.

[110] Sutton, P.A. and Barker, J.F. (1985) Migration and at- tenuation of selected organics in a sandy aquifer - - a natural gradient experiment. Ground Water 23, 10-16.

[111] Goerlitz, D.F., Troutman, D.E., Godsey, E.M. and Franks, B.J. (1985) Migration of wood-preserving chemicals in contaminated groundwater in a sand aquifer at Pensacola, Florida, Environ. Sci. Technol. 19, 955 961.

[112] Hwang, H.-M., Hodson, R.E. and Lee, R.F. (1985) Pho- tochemical and microbial degradation of 2,4,5-trichloroaniline in a freshwater lake. Appl. Environ. Microbiol. 50, 1177-1180.

[113] Alexander, M. (1977) Introduction to Soil Microbiology, 2nd edition. Wiley, New York.

[114] Nedwell, D.B. and Gray, T.R.G. (1987) Soils and sediments as matrices for microbial growth, in Ecology of Microbial Communities (Fletcher, M., Gray, T.R.G. and Jones, J.G., eds.), pp. 22-54, Cambridge University Press, Cambridge.

[115] Lynch, J.M. (1983) Soil Biotechnology. Microbiological Factors in Crop Productivity. Blackwell, Oxford.

[116] Hirsch, P. and Rades-Rohkohl, E. (1983) Microbial diversity in a groundwater aquifer in northern Germany, Dev. Ind. Microbiol. 24, 183-200.

[117] Ventullo, R.M. and Larson, R.J. (1985) Metabolic diversity and activity of heterotrophic bacteria in ground water. Environ. Toxicol. Chem. 4, 759-771.

[118] Ladd, T.I., Ventullo, R.M., Wallis, P.M. and Costerton, J.W. (1982) Heterotrophic activity and biodegradation of labile and refractory compounds by ground water and stream microbial populations. Appl. Environ. Microbiol. 44, 321-329.

[119] Kuhn, E.P., Colberg, P.J., Schnoor, J.L., Wanner, O., Zehnder, A.J.B. and Schwartzenbach, R.P. (1985) Micro- bial transformations of substituted benzenes during infiltration of river water to groundwater: laboratory column studies. Environ. Sci. Technol. 19, 961-968.

[120] Wilson, B.H., Smith, G.B. and Rees, J.F. (1986) Bio- transformations of selected alkylbenzenes and halogenated aliphatic hydrocarbons in methanogenic aquifer material: a microcosm study. Environ. Sci. Technol. 20, 997-1002.

[121] British Standards Institution (1988) Code of Practice for the Identification of Potentially Contaminated Land and its Investigation. Publication No. DD175:1988. BSI, Milton Keynes.

[122] Ghiorse, W.C. and Balkwill, D.L. (1985) Microbiological characterization of subsurface environments, in Ground Water Quality (Ward, C.H., Giger, W. and McCarty, P.L., eds.), pp. 387-401, Wiley, New York.

[123] Jhaveri, V., Mazzacca, A.J. and Snyder, H. (1986) Method and apparatus for treating hydrocarbon and halogenated hydrocarbon contaminated ground water. European Patent 0,066,898.

[124] CONCAWE (1981) Revised Inland Oil Spill Clean-up Manual. Report No. 7/81. CONCAWE, The Hague.

[125] Atwood, D.F. and Gorehck, S.M. (1985) Optimal hydra- uric containment of contaminated ground water. Proc. 5th Nat. Syrup. Expo. Aquifer Rest. and Ground Water Monit., May 1985, pp. 328-344, NWWA, Worthington, OH.

[126] CONCAWE (1980) Sludge Farming: a Technique for the

Disposal of Oily Refinery Wastes. Report No. 3/80. CONCAWE, The Hague.

[127] Valo, R. and Salkinoja-Salonen, M. (1986) Bioreclama- tion of chlorophenol-contaminated soil by composting. Appl. Microbiol. Biotechnol. 25, 68-75.

[128] Doelman, P., Haanstra, L. and Vos, A. (1988) Microbial degradation by the autochthonous soil population of alpha and beta HCH under anaerobic field conditions in temperate regions. Chemosphere 17, 481-487.

[129] Doelman, P., Haanstra, L. and Vos, A. (1988) Microbial sanitation of soil with alpha and beta HCH under aerobic glasshouse conditions. Chemosphere 17, 489-492.

[130] Lee, M.D., Jamison, V.W. and Raymond, R.L. (1987) Applicability of in-situ bioreclamation as a remedial action alternative. Proc. NWWA/API Conf. Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Restoration, Nov. 1987, NWWA, Dublin, OH.

[131] van den Berg, R , Socz6, E.R., Verheul, J.H.A.M. and Eikelboom, D.H. (1987) In Situ Biorestoration of a Sub- soil, Contaminated with Oil. Report No. RIVM 728518003, National Institute of Public Health and En- vironmental Protection, Bilthoven, The Netherlands.

[132] Raymond, R.L., Brown, R.A., Norris, R.D. and O'Neill, E.T. (1986) Stimulation of biooxidation processes in subterranean formations. US Patent 4,588,506.

[133] Brown, R.A., Norris, R.D. and Raymond, R.L. (1984) Oxygen transport in contaminated aquifers with hydrogen peroxide. Proc. NWWA/API Conf. Petroleum Hy- drocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Restoration, Nov. 1984, pp. 441-450, NWWA, Worthington, OH.

[134] Yaniga, P.M. and Smith, W. (1984) Aquifer restoration via accelerated in situ biodegradation of organic contaminants. Proc. NWWA/API Conf. Petroleum Hydro- carbons and Organic Chemicals in Ground Water: Pre- vention, Detection and Restoration, Nov. 1984, pp. 451-470, NWWA, Worthington, OH.

[135] Nagel, G., Kiihn, W., Werner, P. and Sontheimer, H. (1982) Grundwassersanierung durch infiltration von ozontem wasser. GWF-Wasser/Abwasser 123, 399-407.

[136] Kuhlheimer, P.D. and Sunderland, G.L. (1986) Biotrans- formation of petroleum hydrocarbons in deep unsaturated sediments. Proc. NWWA/API Conf. Petro- leum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Retoration, Nov. 1986, pp. 445-461, NWWA, Worthington, OH.

[137] Jamison, V.W., Raymond, R.L. and Hudson, J.O. (1975) Biodegradation of high-octane gasoline in groundwater. Dev. Ind. Microbiol. 16, 305-312.

[138] Jamison, V.W., Raymond, R.L. and Hudson, J.O. (1976) Biodegradation of high octane gasoline, in Proceedings of the Third International Biodegradation Symposium (Sharpley, J.M. and Kaplan, A.M., eds.), pp. 187-196, Applied Science, London.

[139] Brown, R.A., Norris, R.D. and Brubaker, G.R. (1985) Aquifer restoration with enhanced bioreclamation. Pol- hit. Eng. 17, 25-28.

297

[140] Flathman, P.E., McCloskey, M.J., Vondrick, J.J. and Pimlett, D.M. (1985) In situ physical/biological treatment of methylene chloride (dichloromethane) contaminated ground water. Proc. 5th Nat. Syrup. Expo. Aquifer Restoration and Ground Water Monitoring, pp. 571-597, NWWA, Worthington, OH.

[141] Quince, J.R., Ohneck, R.J. and Vondrick, J.J. (1985) Response to an environmental incident affecting ground water. Proc. 5th Nat. Syrup. Expo. Aquifer Restoration and Ground Water Monitoring, pp. 598-608, NWWA, Worthington, OH.

[142] Boyer, J.D., Ahlert, R.C. and Kosson, D.S. (1987) De- gradation of 1,1,1-trichloroethane in bench-scale bioreactors. Hazard. Waste Hazard. Mater. 4, 241-260.

[143] Castro, T.F. and Yoshida, T. (1974) Effect of organic matter on the biodegradation of some organochlorine insecticides in submerged soils. Soil Sci. Plant Nutr. 20, 363-370.

[144] Yoshida, T. and Castro, T.F. (1975) Degradation of 2,4-D, 2,4,5-T and pichloram in two Philippine soils. Soil Sci. Plant Nutr. 21, 397-404.

[145] Mirsatari, S.G., McChesney, M.M., Craigmill, A.C., Winterlin, W.L. and Seiber, J.N. (1987) Anaerobic microbial dechlorination: an approach to on-site treatment of toxaphene-contaminated soil. J. Environ. Sci. Health B 22, 663-690.

[145] (a) Gottfreund, J., Schmitt, G. and Schweisfurth, R. (1985) Erfahrungen fiber den abbau von kohlenwassers- toffen im untergrund mittels nitratdosierung. Forum St~idte-Hygeine 36, 184-187.

[146] Fliermans, C.B., Phelps, T.J., Ringelberg, D., Mikell, A.T. and White, D.C. (1988) Mineralization of trichloroethylene by heterotrophic enrichment cultures. Appl. En- viron. Microbiol. 54, 1709-1714.

[147] Wilson, J.T. and Wilson, B.H. (1985) Biotransformation of trichloroethylene in soil. Appl. Environ. Microbiol. 49, 242-243.

[148] Semprini, L, Roberts, P.V., Hopkins, G.D. and Mackay, D.M. (1987) A Field Evaluation of in-situ Biodegrada- tion for Aquifer Restoration. U.S. Environmental Protec- tion Agency report No. EPA/600/2-87/096. Robert S. Kerr Environmental Research Laboratory, Ada, OK.

[149] Brunner, W., Sutherland, F. and Focht, D.D. (1985) Enhanced biodegradation of polychlorinated biphenyls in soils by analog enrichment and bacterial inoculation. J. Environ. Qual. 14, 324-328.

[150] Focht, D.D. (1987) Analog enrichment decontamination process. U.S. Patent 4,664,805.

[151] Knackmuss, H.-J. (1983) Xenobiotic degradation in industrial sewage: haloaromatics as target substrates. Bio- chem. Soc. Symp. 48, 173-190.

[152] SchOnborn, W. (1986) Biotechnology Volume 8: Micro- bial Degradation. VCH Verlagsgesellschaft, Weinheim.

[152] (a) Boyer, J.D., Ahlert, R.C. and Kosson, D.S. (1988) Pilot plant demonstration of in-situ biodegradation of 1,1,1-trichloroethane. J. Water Pollut. Control Fed. 60, 1843-1849.

[153] Sims, R.C., Doucette, W.J., McLean, J.E., Grenney, W.J.

298

and Dupont, R.R. (1988) Treatment Potential for 56 EPA Listed Hazardous Chemicals in Soil. US Environ- mental Protection Agency report no. EPA/600/S6- 88/001, Robert S. Kerr Environmental Research Laboratory, Ada, OK.

[154] Sloan, R. (1987) Bioremediation demonstrated at hazardous waste site. Oil Gas J. 85, 61-66.

[155] Lee, M.D. and Ward, C.H. (1985) Biological methods for the restoration of contaminated aquifers. Environ. Toxi- col. Chem. 4, 743-750.

[156] Westmeier, F. and Rehm, H.J. (1985) Biodegradation of 4-chlorophenol by entrapped Alicaligenes sp. A7-2. Appl. Microbiol. Biotechnol. 22, 301-305.

[157] Goldstein, R.M., Mallory, L.M. and Alexander, M. (1985) Reasons for possible failure of inoculation to enhance biodegradation. Appl. Environ. Microbiol. 50, 977-983.

[158] Scher, F.M., Kloepper, J.W. and Singleton, C.A. (1985) Chemotaxis of fluorescent Pseudomonas spp. to soybean exudates in vitro and in soil. Can. J. Microbiol. 31, 570-574.

[159] Wong, P.T.W. and Griffin, D.M. (1976) Bacterial movement at high matric potentials - - I. In artificial and natural soils. Soil Biol. Biochem. 8, 215-218.

[160] McCoy, E.L. and Hagedorn, C. (1979) Quantitatively tracing bacterial transport in saturated soil systems. Water, Air, Soil Pollut. 11,467-479.

[161] Madsen, E.L. and Alexander, M. (1982) Transport of Rhizobium and Pseudomonas through soil. Soil Sci. Soc. Am. J. 46, 557-560.

[162] Gerba, C.P. and Bitton, G. (1985) Microbial pollutants: their survival and transport pattern to groundwater, in Ground Water Quality (Ward, C.H., Giger, W. and McCarty, P.L., eds.), pp. 65-88, Wiley, New York.

[163] Morgan, P. and Dow, C.S. (1986) Bacterial adaptations for growth in low nutrient environments, in Microbes in Extreme Environments (Herbert, R.A. and Codd, G.A., eds.), pp. 187-214, Academic Press, London.

[164] MacLeod, F.A., Lappin-Scott, H.M. and Costerton, J.W. (1988) Plugging of a model rock system by using starved bacteria. Appl. Environ. Microbiol. 54, 1365-1372.

[165] Lappin-Sc~tt, H.M., Cusack, F. and Costerton, J.W. (1988) Nutrient resuscitation and growth of starved cells in sandstone cores: a novel approach to enhanced oil recovery. Appl. Environ. Microbiol. 54, 1373-1382.

[166] Rayner, A.D.M., Boddy, L. and Dowson, C.G. (1987) Genetic interactions and developmental versatility during establishment of decomposer basidiomycetes in wood and tree litter, in Ecology of Microbial Communities (Fletcher, M., Gray, T.R.G. and Jones, J.G., eds.), pp. 83-123, Cambridge University Press, Cambridge.

[167] Edgehill, R.U. and Finn, R.K. (1983) Microbial treatment of soil to remove pentachlorophenol. Appl. En- viron. Microbiol. 45, 1122-1125.

[168] Crawford, R.L. and Mohn, W.W. (1985) Microbiological removal of pentachlorophenol from soil using a Flavobacterium. Enzyme Microb. Technol. 7, 617-620.

[169] Chatterjee, D.K., Kilbane, J.J. and Chakrabarty, A.M.

(1982) Biodegradation of 2,4,5-trichlorophenoxyacetic acid in soil by a pure culture of Pseudomonas cepacia. Appl. Environ. Microbiol. 44, 514-516.

[170] van der Meer, J.R., Roelofson, W., Schraa, G. and Zehnder, A.J.B. (1987) Degradation of low concentrations of dichlorobenzenes and 1,2,4-trichlorobenzene by Pseudomonas sp. strain P51 in non-sterile soil columns. FEMS Microbiol. Ecol. 45, 333-341.

[170] (a) Chaudhry, G.R. and Cortez, L. (1988) Degradation of bromacil by a Pseudomonas sp. Appl. Environ. Micro- biol. 54, 2203-2207.

[171] Finn, R.K. (1983) Use of specialized microbial strains in the treatment of industrial waste and in soil decontamination. Experientia 39, 1231-1236.

[172] Slater, J.H. and Lovatt, D. (1984) Biodegradation and the significance of microbial communities, in Microbial Degradation of Organic Compounds (Gibson, D.T., ed.), pp. 439-485, Marcel Dekker, New York.

[173] Reineke, W. and Knackmuss, H.-J. (1979) Construction of haloaromatics utilising bacteria. Nature 177, 385-386.

[174] Rubio, M.A., Engesser, K.-H. and Knackmuss, H.-J. (1986) Microbial metabolism of chlorosalicylates: accelerated evolution by natural genetic exchange. Arch. Microbiol. 145, 116-122.

[175] Bruhn, C., Bayly, R.C. and Knackmuss, H.-J. (1988) The in vivo construction of 4-chloro-2-nitrophenol assimila- tory bacteria. Arch. Microbiol. 150, 171-177.

[176] Kellogg, S.T., Chatterjee, D.K. and Chakrabarty, A.M. (1981) Plasmid-assisted molecular breeding: new technique for enhanced biodegradation of persistent toxic chemicals. Science 214, 1133-1135.

[177] Golovleva, L.A., Pertsova, R.N., Boronin, A.M., Grischenkov, V.G., Baskunov, B.P. and Polyakova, A.V. (1982) Degradation of polychloroaromatic insecticides by Pseudomonas aeruginosa containing biodegradation plasmids. Microbiology 51, 772-777.

[178] Golovleva, L.A., Pertsova, R.N., Boronin, A.M., Travkin, V.M. and Kozlovsky, S.A. (1988) Kelthane degradation by genetically engineered Pseudomonas aeruginosa BS827 in a soil ecosystem. Appl. Environ. Microbiol. 54, 1587-1590.

[179] Jain, R.K. and Sayler, G.S. (1987) Problems and potential for in situ treatment of environmental pollutants by engineered microorganisms. Microbiol. Sci. 4, 59-63.

[180] Pemberton, J.M. and Wynn, E.C. (1984) Genetic engineering and biological detoxification/degradation of insecticides, in Insecticide Microbiology (Lal, R., ed.), pp. 139-168, Springer-Verlag, Berlin.

[181] Jain, R.K., Sayler, G.S., Wilson, J.T., Houston, L. and Pacia, D. (1987) Maintenance and stability of introduced genotypes in groundwater aquifer material. Appl. En- viron. Microbiol. 53, 996-1002.

[182] Bumpus, J.A. and Aust, S.D. (1987) Biodegradation of environmental pollutants by the white rot fungus Phanerochaete chrysosporium: involvement of the lignin degrading system. BioEssays 6, 166-170.

[183] Arjmand, M. and Sandermann, H. (1985) Mineralization

of chloroaniline/lignin conjugates and of free chloroanilines by the white rot fungus Phanerochaete chrysosporium. J. Agric. Food Chem. 33, 1055-1060.

[184] Hallinger, S., Zeigler, W., Walln~fer, P.R. and En- gelhardt, G. (1988) Verhalten von 3,4-dichloroanilin in wachsenden pilzkulturen. Chemosphere 17, 543-550.

[185] Bumpus, J.A. and Aust, S.D. (1987) Biodegradation of DDT [1,1,1-trichloro-2,2-bis(4-chlorophenyl)ethane] by the white rot fungus Phanerochaete chrysosporium. Appl. Environ. Microbiol. 53, 2001-2008.

[186] Bumpus, J.A., Tien, M., Wright, D. and Aust, S.D. (1985) Oxidation of persistent environmental pollutants by a white rot fungus. Science 228, 1434-1436.

[187] Eaton, D.C. (1985) Mineralization of polychlorinated biphenyls by Phanerochaete chrysosporium: a ligninolytic fungus. Enzyme Microb. Technol. 7, 194-196.

[187] (a) K~Shler, A., Jiiger, A., Willershausen, H. and Graf, H. (1988) Extracellular ligninase of Phanerochaete chrysosporium Burdsall has no role in the degradation of DDT. Appl. Microbiol. Biotechnol. 29, 618-620.

299

[188] Chang, H.-M., Joyce, T.W., Kirk, T.K. and Huynh, V.-B. (1985) Process of degrading chloro-organics by white-rot fungi. U.S. Patent 4,554,075.

[189] Stucki, G. and Alexander, M. (1987) Role of dissolution rate and solubility in biodegradation of aromatic compounds. Appl. Environ. Microbiol. 53, 292-297.

[190] Thomas, J.M., Yordy, J.R., Amador, J.A. and Alexander, M. (1986) Rates of dissolution and biodegradation of water-insoluble organic compounds. Appl. Environ. Microbiol. 52, 290-296.

[191] Hinchee, R.E., Downey, D.C. and Coleman, E.J. (1987) Enhanced bioreclamation, soil venting and ground-water extraction: a cost-effectiveness and feasibility compari- son. Proc. NWWA/API Conf. Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Restoration, Nov. 1987, pp. 147-164, NWWA, Dublin, OH.

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