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Polychlorinated biphenyl-degrading microbial communities in soils and sediments
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Polychlorinated biphenyl-degrading microbial communities in soils and sediments Wolf-Rainer Abraham*¶, Balbina Nogales†, Peter N Golyshin‡, Dietmar H Pieper*¥ and Kenneth N Timmis*§¤ Recent advances in the degradation of polychlorinated biphenyls (PCBs) have focussed on the use of experimental enrichment cultures to obtain PCB-degrading communities, and the use of culture-independent approaches to characterize natural and experimental PCB-degrading communities and to identify the key members in this process. PCB-degrading communities can be surprisingly diverse. Novel types of composite bacteria–mineral biofilm communities have been described. Community metabolism of PCBs may lead to the formation of protoanemonin, a dead-end product in some instances but, in others, a seemingly productive intermediate. Analysis of isotope fractionation and preferred enantiomer degradation has provided new information on degradation of PCBs in anaerobic settings. The first defined community capable of dehalorespiration of PCBs has been described, and important community members identified. Here, we provide an overview of the current knowledge of aerobic and anaerobic degradation of PCBs in microbial consortia and in the environment, including novel approaches to determine in situ PCB degradation. Addresses *Division of Microbiology, Gesellschaft für Biotechnologische Forschung (German Research Centre for Biotechnology), Mascheroder Weg 1, 38124 Braunschweig, Germany † Area de Microbiologia, Universitat de les Illes Balears, 07071 Palma de Mallorca, Spain; e-mail: [email protected] ‡ Department of Microbiology, Technical University of Braunschweig, Spielmannstrasse 7, 38106 Braunschweig, Germany; e-mail: [email protected] § Department of Biological Sciences, University of Essex, Colchester, United Kingdom; e-mail: [email protected] ¶ e-mail: [email protected] ¥ e-mail: [email protected] ¤ e-mail: [email protected] Current Opinion in Microbiology 2002, 5:246–253 1369-5274/02/$ — see front matter © 2002 Elsevier Science Ltd. All rights reserved. Published online 13 May 2002 Abbreviation PCB polychlorinated biphenyl
Introduction Polychlorinated biphenyls (PCBs) are a class of compound in which the aromatic biphenyl carbon skeleton carries between one and 10 chlorine atoms. There are, thus, 209 possible congeners, though typical industrial preparations obtained by random chlorination of biphenyl contain 20–60 PCB congeners. Because they are thermally and chemically highly stable, flame- and oxidation-resistant, have low vapour pressure, are superhydrophobic and have excellent dielectric properties, PCBs were widely used in industry for a multitude of applications, including the manufacture of flame retardants, oil condensers, dielectrics,
plasticizers, heat exchangers and hydraulic fluids. PCBs were first produced in 1000 ton quantities in the early 1930s but, as more uses were found, their production increased exponentially to a peak of 200 000 tons in 1975. However, the realization that the very properties of PCBs that made them industrially useful also made them into toxic, persistent environmental pollutants that bioaccumulate in the food chain led to the interdiction of their production in the mid-1980s. In humans, certain congeners are teratogenic, immunogenic and/or carcinogenic, probably via oxygenated metabolites, and act as environmental oestrogens (so-called endocrine disrupters), which are suspected to be one cause of decreasing fertility in industrialized nations [1]. About 1.5 million tons of PCBs were produced worldwide between the 1930s and the mid-1980s and a substantial fraction of this has entered or will ultimately enter the environment. PCBs are now widely distributed over the Earth and found even in remote parts of the world, such as Antarctica and Northern Greenland. One study reported that the Canadian Arctic presently shows little evidence of reduced PCB loadings and concluded that lifetimes of PCBs in the Arctic measure in decades [2••]. The high chemical stability, superhydrophobicity and toxicity of PCBs make them some of the most serious and persistent environmental pollutants. It is somewhat surprising, therefore, that many microbes and enrichment cultures have been obtained that are able to metabolize and utilize PCBs as carbon and/or energy sources under aerobic or anaerobic conditions. Rates of metabolism attained in the laboratory can be high, which suggests that bioremediation may ultimately be a solution for many PCB-contaminated sites. Nevertheless, in situ removal rates are generally exceptionally slow, so the challenge is to determine why this is so, and to find ways and means to realize the full potential of PCB degraders. Clearly, the multitude of congeners, their poor bioavailability and poor mass transfer, and their toxicity are important factors contributing to poor degradation rates in natural settings, but there are almost certainly others. The key to determining these other factors will be to identify the main microbes involved in the process, and to elucidate how they deal with the problems posed by this class of compounds and how their activities may be enhanced. PCBs are degraded aerobically and anaerobically. As a general rule, highly chlorinated congeners (which are highly stable and highly hydrophobic) are good substrates for anaerobic degradation, possibly via chlororespiration (the initial use of the PCB as an electron acceptor), but are poor
Polychlorinated biphenyl-degrading microbial communities in soils and sediments Abraham et al.
substrates for aerobic degradation. Dechlorination is a progressive process that converts higher-chlorinated congeners to lower-chlorinated forms, or more hydrophobic congeners to less hydrophobic forms. Lower-chlorinated congeners are, in turn, poor substrates for anaerobic dechlorination, but are good substrates for aerobic degradation, in which they act primarily as electron donors. Thus, in recent years, much research has focussed on development of methods to unequivocally demonstrate ongoing reductive dehalogenation in anaerobic environments, on approaches to stimulate such activities, and on characterization of the microbial communities responsible for anaerobic and aerobic biotransformations of PCBs, with the goals of developing biocatalysts for reductive dehalogenation of PCBs and coupled anaerobic–aerobic bioremediation processes. Here, we provide an overview of the current knowledge of aerobic and anaerobic degradation of PCBs in microbial consortia and in the environment. Because the determination of the degradation rates of PCBs in the environment still represents a challenge, novel approaches, such as enantiomer and isotopic ratio measurements, to determine in situ PCB degradation are also included. One focus of the review is on microbial interaction and co-metabolic PCB degradation with carbon sharing in microbial communities, probably the most common situation in the environment.
Enantiomer and isotopic fractionation during biodegradation of PCBs Of the 209 possible PCB congeners, 78 are asymmetrically substituted about their long axis and are, thus, axially chiral in their non-planar conformations. Nineteen of these congeners (atropisomers) have three or four ortho chlorines, which restrict rotation along the central C–C axis; thus, different enantiomers of these congeners exist. Given that only biological processes can affect the enantiomeric composition of chiral compounds, monitoring of the fate of chiral PCBs, and assessment of changes in their composition, can be a powerful approach to detect and investigate PCBbiotransforming activities. The enantiomeric ratios of PCB atropisomers in sediment samples have been determined to monitor the extent of microbial dechlorination [3•]. Eight non-racemic PCB atropisomers were detected, confirming previous reports that indicated reductive dechlorination of PCBs at these sites. Interestingly, the enantiomeric ratios of one PCB were reversed in two different sites, which implies the involvement of different PCB-biotransforming processes with different enantiomer preferences. Many biochemical reactions result in isotopic fractionation of carbon, because enzymes often preferentially transform molecules containing the lighter 12C isotope [4,5]. Thus, during bio-transformation processes, reaction products are enriched in lighter isotopes and contain few heavy isotopes, relative to the original substrate, whereas non-transformed residual substrate is enriched in heavy isotopes and contain few lighter isotopes. However, no significant carbon isotope fractionation was observed during the reductive
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dechlorination of 2,3,4,5-tetrachlorobiphenyl by an anaerobic enrichment culture [6]. The absence of fractionation in this case may be exploited for monitoring biological dechlorination, as an intrinsic trend of decreasing 13C abundance with increasing chloride content is characteristic of commercial preparations of PCBs. Thus, microbial reductive dechlorination should produce lower-chlorinated congeners with lower 13C content than those of native PCBs with similar chlorination levels. Besides carbon fractionation, it has become evident that isotopic fractionation of other elements, such as hydrogen, can testify to metabolic processes [7,8]. A promising approach to detect microbial transformation processes is determination of the isotope ratios of chlorine (37Cl:35Cl) in different PCB congeners in commercial preparations and in environmental samples [9••]. The isotopic ratios of chlorine in commercial PCB mixtures are similar and span a narrow range of 37Cl:35Cl ratios. In contrast, chlorine of PCBs from environmental samples showed a significantly larger range of isotopic ratios; those from one site, the New Bedford Harbor superfund site, showed a pronounced trend of 37Cl enrichment with depth.
Anaerobic dechlorination of PCBs in microcosms and sediments Up-to-date evidence demonstrates that microbial PCB dechlorination is widespread in many anaerobic sediments. Various dechlorination patterns in environmental and laboratory samples have been described. Typically, meta and/or para chlorines are removed to generate primarily orthosubstituted chlorobiphenyls, but ortho dechlorination of several PCB congeners has also been reported [10•]. Different microorganisms seem to be responsible for the different dechlorination activities, and various environmental factors, such as available electron donors and electron acceptors, influence the extent and rate of dechlorination, and the dechlorination route. These effects are due to the specificity of the dehalogenating bacteria that become active, as well as the complex interactions between non-dehalogenating and dehalogenating members of the community. One promising strategy to enhance PCB dechlorination by microbial communities is to exploit the microbes’ ability to adapt to easily degraded halogenated substrates that can be used in dehalorespiration processes. This strategy assumes that the communities are enriched in micro-organisms possessing enzymes or cofactors that have relaxed specificity for halogenated substrates and that perform fortuitous dehalogenation of a variety of halogenated compounds. In one study, it was found that certain PCB congeners could effectively activate or ‘prime’ dechlorination of PCBs, with different congeners ‘priming’ different dechlorination patterns [11]. Certain bromobiphenyls were subsequently observed to be very effective in stimulating dechlorination [12]. The ability of various monocyclic halogenated compounds, including halobenzoates, to prime PCB dechlorination has also been studied and various
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bromoanalogues found to be effective in priming PCB dehalogenation, suggesting that the communities are enriched on the basis of their ability to dehalogenate monocyclic aromatics effectively to dechlorinated PCBs [13]. Again, different haloaromatics were found to prime different dehalogenating processes. Halobenzoates have a significant advantage as primers because, once dehalogenated, they are easily mineralized, whereas halogenated biphenyls are only dehalogenated, at best, to biphenyl, which is recalcitrant under anaerobic conditions. Chlorobenzoates, the reported metabolites of aerobic degradation of PCBs, were not effective as primers at sites studied by DeWeerd and Bedard [13], but seem to prime dehalogenation if material from other locations [14] was used. Thus, soluble chlorobenzoates produced by aerobic processes and transported to anaerobic zones may influence the rate of anaerobic dehalogenation. Organisms mediating the reductive dechlorination of PCBs have been difficult to identify by traditional isolation techniques. Culture-independent community fingerprinting methods have recently been applied to cultures exhibiting reductive dehalogenation. Hou and Dutta [15] characterized microbial communities mediating para- or meta-dechlorination of PCBs, and found sequences related to members of the genus Clostridium to be abundant. Though this did not constitute proof that such organisms actually catalyze dehalogenation, it was recently shown that certain Clostridium species harbor enzymes that catalyze dechlorination of perchloroethylene (PCE) [16]. Although pure cultures of organisms mediating reductive dehalogenation of PCBs have not been obtained thus far, two consortia (one ortho-dechlorinating [17•] and the other meta- or para- dechlorinating with doubly flanked chlorines [18•]) have been thoroughly analyzed. In the case of the highly enriched ortho-dechlorinating culture, the growth of a single microorganism, designated 0-17 and exhibiting 16S rDNA sequence similarity to a group that includes Dehalococcoides ethenogenes, was shown to be dependent on the presence and dechlorination of 2,3,5,6-tetrachlorobiphenyl. In this context, it should be noted that the only organisms described so far to be able to completely dehalogenate trichloroethylene (TCE) [19] or chlorobenzene [20•], and conserve the energy thereby liberated by dehalorespiration for growth, also belong to this group of organisms, a deepbranching phylogenetic group dominated by uncultured organisms and most closely related to the green non-sulphur bacteria. In the culture of microorganisms that dechlorinate PCB congeners with doubly flanked chlorines, a strain designated DF-1 with high sequence similarity to 0-17 was identified as being responsible for PCB dechlorination [21], testifying to the importance of microorganisms from this group for dehalorespiration of PCBs.
Aerobic degradation of PCBs by natural and experimental microbial communities Perhaps surprisingly, despite the chemical stability and toxicity of PCBs, many microorganisms have been isolated
that are able to degrade or transform at least the lowerchlorinated PCB congeners under aerobic conditions in the laboratory. Most of the organisms obtained are either Gram-negative bacteria belonging to a few genera within the γ- and β-subclasses of proteobacteria (such as Burkholderia, Pseudomonas and Sphingomonas) or Grampositives, often Rhodococcus species [22–26]. Each isolate exhibits a particular activity spectrum with regard to the type and extent of PCB congeners metabolized, with some strains having a narrow spectrum and others, notably Burkholderia sp. LB400 [27] and Rhodococcus globerulus P6, being able to transform a broad range of congeners. However, only a few of the large number of bacteria isolated that degrade PCBs aerobically can actually mineralize the substrate (degrade both of the biphenyl rings). The majority degrades only the least chlorinated ring, and releases the second ring as chlorobenzoate [25,28]. If this is also true of as yet uncultivated microorganisms, bacteria capable of mineralizing both aromatic rings of chlorobiphenyls are, for some reason, rare in Nature. Thus, mineralization of chlorobiphenyls appears to necessitate communities of chlorobiphenyl-transforming and chlorobenzoate-degrading organisms [29]. An elegant system to analyze the interactions of such partners has been described by Nielsen et al. [30•]. It consists of a model consortium comprising Burkholderia sp. LB400 (which is capable of metabolizing 3-chlorobiphenyl) and Pseudomonas sp. B13(FR1) (which is capable of mineralizing the chlorobenzoate produced by LB400). When fed citrate, which can be used by both partners as a carbon source, the two strains formed separated biofilm microcolonies. In contrast, when fed with 3-chlorobiphenyl at low concentrations typical of the natural situation, they formed mixed microcolonies in which the close physical association of the two partners ensured efficient transfer of chlorobenzoate from LB400 to B13. A major surprise in PCB catabolism was the finding that released chlorobenzoate may be subjected to non-productive metabolism to the dead-end product protoanemonin (an antibacterial compound that kills PCB degraders) by benzoate-degrading (as opposed to chlorobenzoate-degrading) bacteria in natural communities [31]. This finding hinted at the role of protoanemonin as a possible contributing factor to the extremely slow rates of PCB removal in the environment. However, even though the protoanemonin formed from PCBs (and other chloroaromatics) may be a dead-end toxic metabolite in some situations, latest findings suggest that, despite its toxicity, protoanemonin may be a productive intermediate in chloroaromatic degradation in certain bacteria [32]. Clearly, the diversity of bacterial metabolic pathways is considerably more than currently appreciated, and protoanemonin formation may not be a problem for some bacteria and bacterial communities. Studies of the diversity of microbial communities revealed that most bacteria from the environment are currently not culturable. Assessments of the biodiversities of a broad range of habitats revealed that the culturability
Polychlorinated biphenyl-degrading microbial communities in soils and sediments Abraham et al.
of strains does not follow a statistical distribution, but is related to phylogeny. Hence, for a number of phyla, the diversity is only known from SSU rRNA gene sequences, although not a single member has yet been cultivated. Therefore, the genetic information from these uncultured strains is not accessible for conventional biodegradation studies, and it may harbour novel degradation pathways and novel specificities for PCB congeners. To overcome this difficulty, attempts to clone large DNA fragments from microbial communities and consortia (the so called ‘meta-genome’) are currently under way, and they may provide us with new insights into the degradation of PCBs in the environment. Given that the productive metabolism of a single monochlorinated PCB congener generally requires at least two partners, it is evident that degradation of the multiplicity of substrates in a natural PCB mixture will involve a complex microbial network characterized by a diversity of substrate specificities. The diversity in substrate profiles of the initial enzyme of the pathway, biphenyl dioxygenase, is well-documented [33] and reflected in the co-metabolizing activities of PCB degraders [34,35]. For example, the enzyme of Burkholderia sp. LB400 [36•] has been shown to transform various ortho-substituted biphenyls, such as 2,2′-dichloro-, 2,2′-difluoro-, 2,2′-dibromo-, 2,2′-dinitro- and 2,2′-dihydroxybiphenyl, with concomitant dehalogenation, denitration or dehydroxylation, and to transform dibenzofuran and dibenzodioxin by angular dioxygenation. However, bottlenecks have been observed at various steps in PCB pathways [37]. Recent work has focussed on the 2-hydroxy-6-oxo-6-phenylhexa-2,4-dienoate (HOPDA) hydrolase from LB400, and the accumulation of HOPDAs and chloroacetophenones in the metabolism of certain PCB congeners can be explained by the substrate specificity of this enzyme [38•]. Significant diversity in the substrate specificities of different HOPDA hydrolases has, however, been found [39], emphasizing that different pathway bottlenecks for different PCB congeners operate in different organisms, and that only diverse communities can be expected to able to deal with such complex pollutant mixtures. Of course, enzymes of PCB pathways may not only transform PCBs and their metabolites, but also other related compounds, such as monocyclic aromatics (see, for example, [40]), and vice versa. This substrate overlap means that other pollutants in a site may act as co-substrates that can influence the composition and activity of biphenylmetabolizing communities. Moreover, it was shown that biphenyl can be mineralized by mosaic pathways — in Pseudomonas putida CE2010, tod (toluene) and cmt (cumate) pathways complement one other to provide the ability to mineralize biphenyl, in the absence of an authentic biphenyl pathway [41•]. It is possible that such mosaic routes, both within individual strains and within communities, are important for the metabolism of complex pollutant mixtures.
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From where are PCB-degrading genes derived? This is still an open question. PCBs have been in the environment for about 70 years, and it is unlikely that the PCB-metabolizing genes have developed de novo. Many theories concerning the origin of the initial dioxygenases have been proposed, including those connecting these genes with the degradation of lignin and its derivatives. However, none of these theories have been fully proved and a connection with the degradation of terpenes is an alternative possibility. Such a connection is supported by reports, from some research groups, of an enhanced degradation of PCBs after the addition of terpenes to environmental samples. Given the central importance of microbial communities and microbial interactions in PCB degradation, and the inability to study these using traditional cultivation approaches, a recent focus has been on the use of cultureindependent methods to describe the composition of microbial communities in PCB-polluted environments, their adaptation mechanisms, the identification of the effective pollutant metabolizers, their population dynamics and activity under different environmental conditions, as well as the interactions among populations involved in the degradation of the pollutant or intermediates produced. Analysis of 16S ribosomal RNA (16S rRNA) gene sequences retrieved from polluted sites is currently the method of choice. Lloyd-Jones and Lau [42] analyzed the bacterial diversity in a surface sample from a sandy landfill soil in Canada containing an average concentration of 67 mg kg–1 of Aroclor 1260, and found high bacterial diversity with cloned sequence representatives of several bacterial divisions, such as low guanine + cytosine (G+C) Gram-positives, Cytophagales and the α-, β- and γ-subclasses of the proteobacteria. Sequences related to known PCB degraders were not detected, but the study was limited to only 20 clones. A more extensive study examined the composition of the bacterial community in surface samples of a sandy soil in East Germany [43,44], in which the PCBs (congener patterns similar to Aroclor 1242) were the predominant source of carbon in the site (concentrations as high as 24 g kg–1, with an average of 10–150 mg kg–1). The study mainly focussed on the identification of the metabolically active bacteria through analysis of 16S rRNA sequences amplified from total RNA by reverse transcriptase (RT)-PCR and fluorescence in situ hybridization (FISH). Analysis of 404 cloned sequences revealed an extremely high diversity with representatives of 10 major phylogenetic divisions, including two putative new bacterial groups. The community contained an abundance of β-proteobacteria, mostly Burkholderia spp. and Variovorax spp., with significant abundance of sequence types related to Sphingomonas spp., to acidiphilic bacteria from the α-subclass of proteobacteria commonly found in non-polluted acidic soils, and to the Holophaga-Acidobacterium phylum. Many Burkholderia, Variovorax and Sphingomonas sequences were related to isolates known to degrade pollutants such as PCBs.
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In contrast to the many studies on anaerobic PCBdechlorinating enrichment cultures, there have been only a few reports on experimental aerobic PCB-degrading microbial communities. Luensdorf et al. [45••,46] described a simple experimental system to study de novo development of PCB-degrading microbial communities. This involves a hydrophobic support material floating on a water column in a flask containing PCB-contaminated soil. The soil provides both the inoculum and carbon source for the development of a PCB-supported biofilm community on the substratum. The resulting biofilm was studied morphologically, chemically and microbiologically through 16S rRNA gene fingerprinting and clone library analysis. Interestingly, both the diversity and nature of the biofilm community were similar to that of the inoculum soil, suggesting that the model is a valid representation of a soil PCB biofilm community. The most fascinating finding, however, was that classical biofilms did not develop on the support material, but developed instead on microbe–mineral composite biofilms consisting of bacteria and clay minerals, in which the clay leaflets were arranged around the bacteria to form hutch-like structures. As bacteria lacking clay leaflets were rarely observed to be attached to the support, it seems that the clay minerals performed an essential role in biofilm formation, probably acting as a nutrient shuttle, mediating transfer of PCBs in a palatable form to the bacteria. This suggests that clay minerals may fulfil a similar function in soil and, thus, that the movement and dispersion of clay particles in PCB-contaminated soil may be an important regulating factor in PCB degradation.
Biosurfactants Cells basically exist and function in an aqueous milieu and, hence, the uptake into and excretion from cells of materials, like substrates and wastes, mostly involves their transport through aqueous media. The low water solubility of hydrophobic substances, and their tendency to bind strongly to hydrophobic matrices, may be a crucial factor limiting their rate of passage to and from the cell. PCBs are superhydrophobic, have extremely poor solubility in water and are poorly bioavailable. Microbes that utilize hydrophobic substrates, such as petroleum oil, often produce surfaceactive substances (biosurfactants or bioemulsifiers, or amphipathic compounds with very diverse structures that reduce interfacial tension between aqueous and nonaqueous fluid phases), increase mixing of the two phases, increase the surface area of hydrophobic substrates and their rate of transfer into or through aqueous media and, hence, increase bioavailability of the compound. A number of studies on the influence of surfactants on the biodegradation of hydrophobic organic substrates have been made, mostly in simple experimental systems. Compounds here included: chemically produced surfactants, such as ethoxylated alcohols, sulphates, sulphonates, Tritons, Brij 35 and sodium dodecyl sulphate [47,48]; enzymatically produced cyclodextrins [49,50]; humic substances [51,52]; and biosurfactants, such as carvone,
glucose lipid, rhamnolipids, lipopeptides (for example, lichenysin) and alasan, produced by different microorganisms [53–56]. In most cases, an increase in degradation rates was observed when surfactants or biosurfactants were added. However, two studies involving more complex experimental systems reported that surfactants decreased degradation rates (despite an increase in substrate solubilization owing, perhaps, to desorption of degraders from matrix or substrate surfaces and changes provoked in community composition), with a decrease in degrader populations [57,58•]. In another case, the presence of soil significantly reduced the stimulatory activity of exogenously added surfactant [59]. An elegant concept, based on recombinant bacteria capable of degrading both PCBs and chemical surfactants, was described by Lajoie et al. [60,61]. In this case, exogenous addition of the recombinant strain(s) and cognate surfactant resulted in the simultaneous increase in pollutant bioavailability and degrader populations. Chemical surfactants have the advantage of low price, but are often toxic for biological systems and, hence, are pollutants in their own right. Biosurfactants generally exhibit higher interfacial tension reduction activities than chemical surfactants, and are non- (or less) toxic and readily biodegradable, properties important not only for the natural turnover of hydrophobic substrates by microbes but also for the bioremediation of polluted sites [62•,63•,64]. Moreover, even though biosurfactants are more expensive to produce industrially (ex situ) than are chemical surfactants [65], their in situ production by microbes precisely where they have greatest effect can circumvent cost issues. However, no rules that specify optimal combinations of ‘surfactant–microorganism–pollutant–environment’ have thus far been developed, and empirical approaches are still the order of the day.
Conclusions Most of our knowledge about the degradation of PCBs comes from laboratory experimental systems involving pure cultures of bacteria in laboratory media containing a single PCB congener. However, given that the majority of microbes in environmental samples cannot be cultured at present in laboratory media, there is no certainty that those which have been studied so far are typical of those involved in PCB degradation in the environment. Moreover, it has become increasingly evident that metabolism of substrates in natural environments occurs not via the linear pathways that are familiar to us from pure culture studies, but via complex metabolic networks that involve: yet to be discovered reactions, metabolites and routes; multiple community members exchanging metabolites and regulating carbon flow according to the availability of other substrates and nutrients; prevailing physico-chemical conditions; and community needs. Thus, despite the importance of reductionist approaches for revealing underlying mechanisms, they may not reflect what happens in the environment. Recent efforts have therefore focussed on getting to grips with diverse but functionally integrated
Polychlorinated biphenyl-degrading microbial communities in soils and sediments Abraham et al.
complex microbial communities that metabolize complex PCB mixtures in natural heterogeneous matrices like soil and sediments or near-to-nature experimental systems. Such studies have revealed a number of surprises, including high diversity in PCB-metabolizing microbial communities, the exploitation of PCB-bearing clay minerals to form composite bacterial biofilms, and the ability of PCB degraders to form biofilms directly on PCB droplets. The intimate association of bacteria with the hydrophobic substrate obviously diminishes the problem of bioavailability at the local level. There are clearly more surprises in store. Gaining an understanding of PCB degradation will require: elucidation of the nature of the interacting surfaces (PCBs in the environment do not have the surfaces of the pure chemical, and degraders in a biofilm on a PCB droplet will not have the surfaces of a laboratory planktonic culture), particularly the role of biosurfactants in conditioning these surfaces, and the mechanisms of transfer and uptake of PCBs; the physical and metabolic interactions of the community members primarily involved in metabolism of the PCB carbon skeleton and chlorines; community interactions with and activity modulation by the prevailing physicochemical parameters of the environment; and the role of grazers and phages in regulating community composition and activities. Nevertheless, given the prevalence and diversity of PCB degraders that can be isolated from the environment, we can, perhaps naively, be optimistic about the potential for bioremediation, once we understand the ground rules of engagement between microbial communities and the target compounds.
Acknowledgements Work in the Gesellschaft für Biotechnologische Forschung laboratory has been supported by grants from the Bundesministerium für Bilding und Forschung (Project Number 0319433C, 0318896C, Strategy fund ‘Soil functions’) and the European Union. K Timmis gratefully acknowledges generous support from the Fonds der Chemischen Industrie.
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21. Wu Q, Watts J, Sowers K, May H: Identification of a bacterium that specifically catalyzes the reductive dechlorination of polychlorinated biphenyls with doubly flanked chlorines. Appl Environ Microbiol 2002, 68:807-812.
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22. Williams WA, Lobos JH, Cheetham WE: A phylogenetic analysis of aerobic polychlorinated biphenyl-degrading bacteria. Int J Syst Bacteriol 1997, 47:207-210. 23. Master ER, Mohn WW: Psychrotolerant bacteria isolated from Arctic soil that degrade polychlorinated biphenyls at low temperatures. Appl Environ Microbiol 1998, 64:4823-4829. 24. Kim S, Picardal FW: A novel bacterium that utilizes monochlorobiphenyls and 4-chlorobenzoate as growth substrates. FEMS Microbiol Lett 2001, 185:225-229. 25. Kim S, Picardal F: Microbial growth on dichlorobiphenyls chlorinated on both rings as a sole carbon and energy source. Appl Environ Microbiol 2001, 67:1953-1955. 26. Fedi S, Carnevali M, Fava F, Andracchio A, Zappoli S, Zannoni D: Polychlorinated biphenyl degradation activities and hybridisation analyses of fifteen aerobic strains isolated from a PCB-contaminated site. Res Microbiol 2001, 152:583-592. 27.
Fain MG, Haddock JD: Phenotypic and phylogenetic characterization of Burkholderia (Pseudomonas) sp. strain LB400. Curr Microbiol 2001, 42:269-275.
28. Hrywna Y, Tsoi TV, Maltseva OV, III JFQ, Tiedje JM: Construction and characterization of two recombinant bacteria that grow on orthoand para-substituted chlorobiphenyls. Appl Environ Microbiol 1999, 65:2163-2169. 29. Sàágua MC, Vieira G, Paveia H, Anselmo A: Isolation and preliminary characterization of Bacillus sp. MCS, a gram-positive 4-chlorobiphenyl degrading bacterium. Int Biodeter Biodegr 1998, 42:39-43. 30. Nielsen A, Tolker-Nielsen T, Barken K, Molin S: Role of • commensal relationships on the spatial structure of a surface-attached microbial consortium. Environ Microbiol 2000, 2:59-68. The authors of this paper studied the interactions between a chlorobiphenyltransforming microorganism and a chlorobenzoate-mineralizing microorganism using fluorescent in situ rRNA hybridization and scanning confocal laser microscopy and provide evidence that commensal relationships play a role in determining the spatial distribution of the organisms in microbial communities. 31. Blasco R, Mallavarapu M, Wittich RM, Timmis KN, Pieper DH: Evidence that formation of protoanemonin from metabolites of 4-chlorobiphenyl degradation negatively affects the survival of 4-chlorobiphenyl-cometabolizing microorganisms. Appl Environ Microbiol 1997, 63:427-434. 32. Pelz O, Tesar M, Wittich RM, Moore ERB, Timmis KN, Abraham WR: Towards elucidation of microbial community metabolic pathways: unravelling the network of carbon sharing in a pollutant degrading bacterial consortium by immunocapture and isotopic ratio mass spectrometry. Environ Microbiol 1999, 1:167-174. 33. Furukawa K: Biochemical and genetic bases of microbial degradation of polychlorinated biphenyls. J Gen Appl Microbiol 2000, 46:283-296. 34. Fedi S, Carnevali M, Fava F, Andracchio A, Zappoli S, Zannoni D: Polychlorinated biphenyl degradation activities and hybridization analyses of fifteen aerobic strains isolated from a PCB-contaminated site. Res Microbiol 2001, 152:583-592. 35. Arnett CM, Parales JV, Haddock JD: Influence of chlorine substituents on rates of oxidation of chlorinated biphenyls by the biphenyl dioxygenase of Burkholderia sp. strain LB400. Appl Environ Microbiol 2000, 66:2928-2933. 36. Seeger M, Camara B, Hofer B: Dehalogenation, denitration, • dehydroxylation, and angular attack on substituted biphenyls and related compounds by a biphenyl dioxygenase. J Bacteriol 2001, 183:3548-3555. The authors of this paper showed that the biphenyl dioxygenase of Burkholderia sp. LB400 attacks various 2,2′-substituted biphenyls preferentially at C-2 and C-3. This unusual attack results in unstable intermediates, which spontaneously re-aromatize with concomitant elimination of the substituent. The results indicate that the oxidation of phenols by dioxygenases is actually the result of a di- rather than a monooxygenation. Furthermore, it was proven that certain biphenyl dioxygenases can perform an angular attack at dibenzofuran and dibenzo-p-dioxin. 37.
Furukawa K: Engineering dioxygenases for efficient degradation of environmental pollutants. Curr Opin Biotechnol 2000, 11:244-249.
38. Seah S, Labbe G, Nerdinger S, Johnson M, Snieckus V, Eltis L: • Identification of a serine hydrolase as a key determinant in the microbial degradation of polychlorinated biphenyls. J Biol Chem 2000, 275:15701-15708. The authors studied in detail a hydrolase involved in the degradation of PCBs. Studies on the specificity provides explanations for the occurrence of particular dead-end metabolites that have been observed in the transformation of specific PCB congeners. 39. Seah S, Labbe G, Kaschabek S, Reifenrath F, Reineke W, Eltis L: Comparative specificities of two evolutionarily divergent hydrolases involved in microbial degradation of polychlorinated biphenyls. J Bacteriol 2001, 183:1511-1516. 40. Suenaga H, Goto M, Furukawa K: Emergence of multifunctional oxygenase activities by random priming recombination. J Biol Chem 2001, 276:22500-22506. 41. Ohta Y, Maeda M, Kudo T: Pseudomonas putida CE2010 can • degrade biphenyl by a mosaic pathway encoded by the tod operon and cmtE, which are identical to those of P. putida F1 except for a single base difference in the operator-promoter region of the cmt operon. Microbiology 2001, 147:31-41. Report on an organism able to grow on biphenyl without having an authentic biphenyl metabolic pathway. The pathway mainly responsible for this phenotype is identical to the tod pathway of P. putida F1. The limiting hydrolytic activity of the hydrolase of the tod pathway is overcome by recruitment of an enzyme of the cumate (cmt) pathway. A single base difference in the cmt promoter–operator region compared to P. putida F1 resulted in low constitutive expression and allowed growth. 42. Lloyd-Jones G, Lau PCK: A molecular view of microbial diversity in a dynamic landfill in Québec. FEMS Microbiol Lett 1998, 162:219-226. 43. Nogales B, Moore ERB, Abraham WR, Timmis KN: Identification of the metabolically active members of a bacterial community in a polychlorinated biphenyl-polluted moorland soil. Environ Microbiol 1999, 1:199-212. 44. Nogales B, Moore ERB, Llobet-Brossa E, Rossello-Mora R, Amann R, Timmis KN: Combined use of 16S ribosomal DNA and 16S rRNA to study the bacterial community of polychlorinated biphenyl-polluted soil. Appl Environ Microbiol 2001, 67:1874-1884. 45. Lünsdorf H, Erb RW, Abraham WR, Timmis KN: Bacterial •• construction of clay ‘hutches’: recruitment of clay mineral particles by a PCB-based microbial consortium. Environ Microbiol 2000, 2:161-168. The authors of this paper demonstrated that bacteria recruit and organize PCB-containing clay minerals around themselves into hutch-like structures to form a composite bacteria–mineral biofilm, suggesting that clay minerals may serve as important hydrophobic nutrient shuttles in Nature. 46. Lünsdorf H, Strömpl C, Moore ERB, Osborn AM, Abraham WR, Timmis KN: Approach to analyse interactions of microorganisms, hydrophobic substrates and colloids leading to formation of composite biofilms, and to study initial events in microbiogeological processes. Meth Enzymol 2001, 33:317-331. 47.
Fava F, Di Gioia D: Effects of Triton X-100 and Quillaya Saponin on the ex situ bioremediation of a chronically polychlorobiphenylcontaminated soil. Appl Microbiol Biotechnol 1998, 50:623-630.
48. Billingsley KA, Backus SM, Ward OP: Effect of surfactant solubilization on polychlorinated biphenyl congeners by Pseudomonas LB400. Appl Microbiol Biotechnol 1999, 52:255-260. 49. Wang J-M, Marlowe EM, Miller-Maier RM, Brusseau ML: Cyclodextrin-enhanced biodegradation of phenantrene. Environ Sci Technol 1998, 32:1907-1912. 50. Bardi L, Mattei A, Steffan S, Marzona M: Hydrocarbon degradation by a soil microbial population with beta-cyclodextrin as surfactant to enhance bioavailability. Enzyme Microb Tech 2000, 27:709-713. 51. Conte P, Zena A, Pilidis G, Piccolo A: Increased retention of polycyclic aromatic hydrocarbons in soils induced by soil treatment with humic substances. Environ Pollut 2001, 112:27-31. 52. Fava F, Piccolo A: Effects of humic substances on the bioavailability and aerobic biodegradation of polychlorinated biphenyls in a model soil. Biotechnol Bioeng 2002, 77:204-11. 53. Zhang Y, Maier WJ, Miller RM: Effect of rhamnolipids on the dissolution, bioavailability and biodegradation of phenanthrene. Environ Sci Technol 1997, 31:2211-2217.
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54. Barkay T, Navon-Venezia S, Ron EZ, Rosenberg E: Enhancement of solubilization and biodegradation of polyaromatic hydrocarbons by the bioemulsifier alasan. Appl Environ Microbiol 1999, 65:2697-2702. 55. Singer AC, Gillbert ES, Luepromchai E, Crowly DE: Bioremediation of polychlorinated biphenyl-contaminated soil using carvone and surfactant-grown bacteria. Appl Microbiol Biotechnol 2000, 54:838-843. 56. Fava F, Di Gioia D: Soya lecithin effects on the aerobic biodegradation of polychlorinated biphenyls in an artificially contaminated soil. Biotechnol Bioeng 2001, 72:177-184. 57.
Stelmack PL, Gray MR, Pickard MA: Bacterial adhesion to soil contaminants in the presence of surfactants. Appl Environ Microbiol 1999, 65:163-168.
58. Colores GM, Macur RE, Ward DM, Inskeep WP: Molecular analysis • of surfactant-driven microbial population shifts in hydrocarboncontaminated soil. Appl Environ Microbiol 2000, 66:2959-2964. The authors of this paper observed the following rearrangement of population structure of a microbial community: Alcaligenes spp. and Pseudomonas spp. displaced hydrocarbon-degrading rhodococci and Nocardia upon increase of surfactant concentration. 59. Golyshin PN, Fredrickson HL, Giuliano L, Rothmel R, Timmis KN, Yakimov MM: Effect of novel biosurfactants on biodegradation of polychlorinated biphenyls by pure and mixed bacterial cultures. Microbiologica 1999, 22:257-267.
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60. Lajoie CA, Chen SY, Oh KC, Strom PF: Development and use of field application vectors to express nonadaptive foreign genes in competitive environments. Appl Environ Microbiol 1992, 58:655-663. 61. Lajoie CA, Layton AC, Easter JP, Menn FM, Sayler CS: Degradation of nonionic surfactants and polychlorinated biphenyls by recombinant field application vectors. J Ind Microbiol Biotechnol 1997, 19:252-262. 62. Banat IM, Makkar RS, Cameotra SS: Potential commercial • applications of microbial surfactants. Appl Microbiol Biotechnol 2000, 53:495-508. This paper provides a good general review on the environmental applications of biosurfactants. 63. Ron EZ, Rosenberg E: Natural roles of biosurfactants. Environ • Microbiol 2001, 3:229-236. This is the most recent review on applications of biosurfactants and bioemulsifiers. 64. Wu, QW, Marshall WD: Approaches to the remediation of a polychlorinated biphenyl (PCB) contaminated soil – a laboratory study. J Environ Monitor 2001, 3:281-287. 65. Kim SH, Lim EJ, Lee SO, Lee JD, Lee TH: Purification and characterization of biosurfactants from Nocardia sp. L-417. Biotechnol Appl Biochem 2000, 31:249–253.
Polychlorinated biphenyl-degrading microbial communities in soils and sediments Wolf-Rainer Abraham*¶, Balbina Nogales†, Peter N Golyshin‡, Dietmar H Pieper*¥ and Kenneth N Timmis*§¤ Recent advances in the degradation of polychlorinated biphenyls (PCBs) have focussed on the use of experimental enrichment cultures to obtain PCB-degrading communities, and the use of culture-independent approaches to characterize natural and experimental PCB-degrading communities and to identify the key members in this process. PCB-degrading communities can be surprisingly diverse. Novel types of composite bacteria–mineral biofilm communities have been described. Community metabolism of PCBs may lead to the formation of protoanemonin, a dead-end product in some instances but, in others, a seemingly productive intermediate. Analysis of isotope fractionation and preferred enantiomer degradation has provided new information on degradation of PCBs in anaerobic settings. The first defined community capable of dehalorespiration of PCBs has been described, and important community members identified. Here, we provide an overview of the current knowledge of aerobic and anaerobic degradation of PCBs in microbial consortia and in the environment, including novel approaches to determine in situ PCB degradation. Addresses *Division of Microbiology, Gesellschaft für Biotechnologische Forschung (German Research Centre for Biotechnology), Mascheroder Weg 1, 38124 Braunschweig, Germany † Area de Microbiologia, Universitat de les Illes Balears, 07071 Palma de Mallorca, Spain; e-mail: [email protected] ‡ Department of Microbiology, Technical University of Braunschweig, Spielmannstrasse 7, 38106 Braunschweig, Germany; e-mail: [email protected] § Department of Biological Sciences, University of Essex, Colchester, United Kingdom; e-mail: [email protected] ¶ e-mail: [email protected] ¥ e-mail: [email protected] ¤ e-mail: [email protected] Current Opinion in Microbiology 2002, 5:246–253 1369-5274/02/$ — see front matter © 2002 Elsevier Science Ltd. All rights reserved. Published online 13 May 2002 Abbreviation PCB polychlorinated biphenyl
Introduction Polychlorinated biphenyls (PCBs) are a class of compound in which the aromatic biphenyl carbon skeleton carries between one and 10 chlorine atoms. There are, thus, 209 possible congeners, though typical industrial preparations obtained by random chlorination of biphenyl contain 20–60 PCB congeners. Because they are thermally and chemically highly stable, flame- and oxidation-resistant, have low vapour pressure, are superhydrophobic and have excellent dielectric properties, PCBs were widely used in industry for a multitude of applications, including the manufacture of flame retardants, oil condensers, dielectrics,
plasticizers, heat exchangers and hydraulic fluids. PCBs were first produced in 1000 ton quantities in the early 1930s but, as more uses were found, their production increased exponentially to a peak of 200 000 tons in 1975. However, the realization that the very properties of PCBs that made them industrially useful also made them into toxic, persistent environmental pollutants that bioaccumulate in the food chain led to the interdiction of their production in the mid-1980s. In humans, certain congeners are teratogenic, immunogenic and/or carcinogenic, probably via oxygenated metabolites, and act as environmental oestrogens (so-called endocrine disrupters), which are suspected to be one cause of decreasing fertility in industrialized nations [1]. About 1.5 million tons of PCBs were produced worldwide between the 1930s and the mid-1980s and a substantial fraction of this has entered or will ultimately enter the environment. PCBs are now widely distributed over the Earth and found even in remote parts of the world, such as Antarctica and Northern Greenland. One study reported that the Canadian Arctic presently shows little evidence of reduced PCB loadings and concluded that lifetimes of PCBs in the Arctic measure in decades [2••]. The high chemical stability, superhydrophobicity and toxicity of PCBs make them some of the most serious and persistent environmental pollutants. It is somewhat surprising, therefore, that many microbes and enrichment cultures have been obtained that are able to metabolize and utilize PCBs as carbon and/or energy sources under aerobic or anaerobic conditions. Rates of metabolism attained in the laboratory can be high, which suggests that bioremediation may ultimately be a solution for many PCB-contaminated sites. Nevertheless, in situ removal rates are generally exceptionally slow, so the challenge is to determine why this is so, and to find ways and means to realize the full potential of PCB degraders. Clearly, the multitude of congeners, their poor bioavailability and poor mass transfer, and their toxicity are important factors contributing to poor degradation rates in natural settings, but there are almost certainly others. The key to determining these other factors will be to identify the main microbes involved in the process, and to elucidate how they deal with the problems posed by this class of compounds and how their activities may be enhanced. PCBs are degraded aerobically and anaerobically. As a general rule, highly chlorinated congeners (which are highly stable and highly hydrophobic) are good substrates for anaerobic degradation, possibly via chlororespiration (the initial use of the PCB as an electron acceptor), but are poor
Polychlorinated biphenyl-degrading microbial communities in soils and sediments Abraham et al.
substrates for aerobic degradation. Dechlorination is a progressive process that converts higher-chlorinated congeners to lower-chlorinated forms, or more hydrophobic congeners to less hydrophobic forms. Lower-chlorinated congeners are, in turn, poor substrates for anaerobic dechlorination, but are good substrates for aerobic degradation, in which they act primarily as electron donors. Thus, in recent years, much research has focussed on development of methods to unequivocally demonstrate ongoing reductive dehalogenation in anaerobic environments, on approaches to stimulate such activities, and on characterization of the microbial communities responsible for anaerobic and aerobic biotransformations of PCBs, with the goals of developing biocatalysts for reductive dehalogenation of PCBs and coupled anaerobic–aerobic bioremediation processes. Here, we provide an overview of the current knowledge of aerobic and anaerobic degradation of PCBs in microbial consortia and in the environment. Because the determination of the degradation rates of PCBs in the environment still represents a challenge, novel approaches, such as enantiomer and isotopic ratio measurements, to determine in situ PCB degradation are also included. One focus of the review is on microbial interaction and co-metabolic PCB degradation with carbon sharing in microbial communities, probably the most common situation in the environment.
Enantiomer and isotopic fractionation during biodegradation of PCBs Of the 209 possible PCB congeners, 78 are asymmetrically substituted about their long axis and are, thus, axially chiral in their non-planar conformations. Nineteen of these congeners (atropisomers) have three or four ortho chlorines, which restrict rotation along the central C–C axis; thus, different enantiomers of these congeners exist. Given that only biological processes can affect the enantiomeric composition of chiral compounds, monitoring of the fate of chiral PCBs, and assessment of changes in their composition, can be a powerful approach to detect and investigate PCBbiotransforming activities. The enantiomeric ratios of PCB atropisomers in sediment samples have been determined to monitor the extent of microbial dechlorination [3•]. Eight non-racemic PCB atropisomers were detected, confirming previous reports that indicated reductive dechlorination of PCBs at these sites. Interestingly, the enantiomeric ratios of one PCB were reversed in two different sites, which implies the involvement of different PCB-biotransforming processes with different enantiomer preferences. Many biochemical reactions result in isotopic fractionation of carbon, because enzymes often preferentially transform molecules containing the lighter 12C isotope [4,5]. Thus, during bio-transformation processes, reaction products are enriched in lighter isotopes and contain few heavy isotopes, relative to the original substrate, whereas non-transformed residual substrate is enriched in heavy isotopes and contain few lighter isotopes. However, no significant carbon isotope fractionation was observed during the reductive
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dechlorination of 2,3,4,5-tetrachlorobiphenyl by an anaerobic enrichment culture [6]. The absence of fractionation in this case may be exploited for monitoring biological dechlorination, as an intrinsic trend of decreasing 13C abundance with increasing chloride content is characteristic of commercial preparations of PCBs. Thus, microbial reductive dechlorination should produce lower-chlorinated congeners with lower 13C content than those of native PCBs with similar chlorination levels. Besides carbon fractionation, it has become evident that isotopic fractionation of other elements, such as hydrogen, can testify to metabolic processes [7,8]. A promising approach to detect microbial transformation processes is determination of the isotope ratios of chlorine (37Cl:35Cl) in different PCB congeners in commercial preparations and in environmental samples [9••]. The isotopic ratios of chlorine in commercial PCB mixtures are similar and span a narrow range of 37Cl:35Cl ratios. In contrast, chlorine of PCBs from environmental samples showed a significantly larger range of isotopic ratios; those from one site, the New Bedford Harbor superfund site, showed a pronounced trend of 37Cl enrichment with depth.
Anaerobic dechlorination of PCBs in microcosms and sediments Up-to-date evidence demonstrates that microbial PCB dechlorination is widespread in many anaerobic sediments. Various dechlorination patterns in environmental and laboratory samples have been described. Typically, meta and/or para chlorines are removed to generate primarily orthosubstituted chlorobiphenyls, but ortho dechlorination of several PCB congeners has also been reported [10•]. Different microorganisms seem to be responsible for the different dechlorination activities, and various environmental factors, such as available electron donors and electron acceptors, influence the extent and rate of dechlorination, and the dechlorination route. These effects are due to the specificity of the dehalogenating bacteria that become active, as well as the complex interactions between non-dehalogenating and dehalogenating members of the community. One promising strategy to enhance PCB dechlorination by microbial communities is to exploit the microbes’ ability to adapt to easily degraded halogenated substrates that can be used in dehalorespiration processes. This strategy assumes that the communities are enriched in micro-organisms possessing enzymes or cofactors that have relaxed specificity for halogenated substrates and that perform fortuitous dehalogenation of a variety of halogenated compounds. In one study, it was found that certain PCB congeners could effectively activate or ‘prime’ dechlorination of PCBs, with different congeners ‘priming’ different dechlorination patterns [11]. Certain bromobiphenyls were subsequently observed to be very effective in stimulating dechlorination [12]. The ability of various monocyclic halogenated compounds, including halobenzoates, to prime PCB dechlorination has also been studied and various
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bromoanalogues found to be effective in priming PCB dehalogenation, suggesting that the communities are enriched on the basis of their ability to dehalogenate monocyclic aromatics effectively to dechlorinated PCBs [13]. Again, different haloaromatics were found to prime different dehalogenating processes. Halobenzoates have a significant advantage as primers because, once dehalogenated, they are easily mineralized, whereas halogenated biphenyls are only dehalogenated, at best, to biphenyl, which is recalcitrant under anaerobic conditions. Chlorobenzoates, the reported metabolites of aerobic degradation of PCBs, were not effective as primers at sites studied by DeWeerd and Bedard [13], but seem to prime dehalogenation if material from other locations [14] was used. Thus, soluble chlorobenzoates produced by aerobic processes and transported to anaerobic zones may influence the rate of anaerobic dehalogenation. Organisms mediating the reductive dechlorination of PCBs have been difficult to identify by traditional isolation techniques. Culture-independent community fingerprinting methods have recently been applied to cultures exhibiting reductive dehalogenation. Hou and Dutta [15] characterized microbial communities mediating para- or meta-dechlorination of PCBs, and found sequences related to members of the genus Clostridium to be abundant. Though this did not constitute proof that such organisms actually catalyze dehalogenation, it was recently shown that certain Clostridium species harbor enzymes that catalyze dechlorination of perchloroethylene (PCE) [16]. Although pure cultures of organisms mediating reductive dehalogenation of PCBs have not been obtained thus far, two consortia (one ortho-dechlorinating [17•] and the other meta- or para- dechlorinating with doubly flanked chlorines [18•]) have been thoroughly analyzed. In the case of the highly enriched ortho-dechlorinating culture, the growth of a single microorganism, designated 0-17 and exhibiting 16S rDNA sequence similarity to a group that includes Dehalococcoides ethenogenes, was shown to be dependent on the presence and dechlorination of 2,3,5,6-tetrachlorobiphenyl. In this context, it should be noted that the only organisms described so far to be able to completely dehalogenate trichloroethylene (TCE) [19] or chlorobenzene [20•], and conserve the energy thereby liberated by dehalorespiration for growth, also belong to this group of organisms, a deepbranching phylogenetic group dominated by uncultured organisms and most closely related to the green non-sulphur bacteria. In the culture of microorganisms that dechlorinate PCB congeners with doubly flanked chlorines, a strain designated DF-1 with high sequence similarity to 0-17 was identified as being responsible for PCB dechlorination [21], testifying to the importance of microorganisms from this group for dehalorespiration of PCBs.
Aerobic degradation of PCBs by natural and experimental microbial communities Perhaps surprisingly, despite the chemical stability and toxicity of PCBs, many microorganisms have been isolated
that are able to degrade or transform at least the lowerchlorinated PCB congeners under aerobic conditions in the laboratory. Most of the organisms obtained are either Gram-negative bacteria belonging to a few genera within the γ- and β-subclasses of proteobacteria (such as Burkholderia, Pseudomonas and Sphingomonas) or Grampositives, often Rhodococcus species [22–26]. Each isolate exhibits a particular activity spectrum with regard to the type and extent of PCB congeners metabolized, with some strains having a narrow spectrum and others, notably Burkholderia sp. LB400 [27] and Rhodococcus globerulus P6, being able to transform a broad range of congeners. However, only a few of the large number of bacteria isolated that degrade PCBs aerobically can actually mineralize the substrate (degrade both of the biphenyl rings). The majority degrades only the least chlorinated ring, and releases the second ring as chlorobenzoate [25,28]. If this is also true of as yet uncultivated microorganisms, bacteria capable of mineralizing both aromatic rings of chlorobiphenyls are, for some reason, rare in Nature. Thus, mineralization of chlorobiphenyls appears to necessitate communities of chlorobiphenyl-transforming and chlorobenzoate-degrading organisms [29]. An elegant system to analyze the interactions of such partners has been described by Nielsen et al. [30•]. It consists of a model consortium comprising Burkholderia sp. LB400 (which is capable of metabolizing 3-chlorobiphenyl) and Pseudomonas sp. B13(FR1) (which is capable of mineralizing the chlorobenzoate produced by LB400). When fed citrate, which can be used by both partners as a carbon source, the two strains formed separated biofilm microcolonies. In contrast, when fed with 3-chlorobiphenyl at low concentrations typical of the natural situation, they formed mixed microcolonies in which the close physical association of the two partners ensured efficient transfer of chlorobenzoate from LB400 to B13. A major surprise in PCB catabolism was the finding that released chlorobenzoate may be subjected to non-productive metabolism to the dead-end product protoanemonin (an antibacterial compound that kills PCB degraders) by benzoate-degrading (as opposed to chlorobenzoate-degrading) bacteria in natural communities [31]. This finding hinted at the role of protoanemonin as a possible contributing factor to the extremely slow rates of PCB removal in the environment. However, even though the protoanemonin formed from PCBs (and other chloroaromatics) may be a dead-end toxic metabolite in some situations, latest findings suggest that, despite its toxicity, protoanemonin may be a productive intermediate in chloroaromatic degradation in certain bacteria [32]. Clearly, the diversity of bacterial metabolic pathways is considerably more than currently appreciated, and protoanemonin formation may not be a problem for some bacteria and bacterial communities. Studies of the diversity of microbial communities revealed that most bacteria from the environment are currently not culturable. Assessments of the biodiversities of a broad range of habitats revealed that the culturability
Polychlorinated biphenyl-degrading microbial communities in soils and sediments Abraham et al.
of strains does not follow a statistical distribution, but is related to phylogeny. Hence, for a number of phyla, the diversity is only known from SSU rRNA gene sequences, although not a single member has yet been cultivated. Therefore, the genetic information from these uncultured strains is not accessible for conventional biodegradation studies, and it may harbour novel degradation pathways and novel specificities for PCB congeners. To overcome this difficulty, attempts to clone large DNA fragments from microbial communities and consortia (the so called ‘meta-genome’) are currently under way, and they may provide us with new insights into the degradation of PCBs in the environment. Given that the productive metabolism of a single monochlorinated PCB congener generally requires at least two partners, it is evident that degradation of the multiplicity of substrates in a natural PCB mixture will involve a complex microbial network characterized by a diversity of substrate specificities. The diversity in substrate profiles of the initial enzyme of the pathway, biphenyl dioxygenase, is well-documented [33] and reflected in the co-metabolizing activities of PCB degraders [34,35]. For example, the enzyme of Burkholderia sp. LB400 [36•] has been shown to transform various ortho-substituted biphenyls, such as 2,2′-dichloro-, 2,2′-difluoro-, 2,2′-dibromo-, 2,2′-dinitro- and 2,2′-dihydroxybiphenyl, with concomitant dehalogenation, denitration or dehydroxylation, and to transform dibenzofuran and dibenzodioxin by angular dioxygenation. However, bottlenecks have been observed at various steps in PCB pathways [37]. Recent work has focussed on the 2-hydroxy-6-oxo-6-phenylhexa-2,4-dienoate (HOPDA) hydrolase from LB400, and the accumulation of HOPDAs and chloroacetophenones in the metabolism of certain PCB congeners can be explained by the substrate specificity of this enzyme [38•]. Significant diversity in the substrate specificities of different HOPDA hydrolases has, however, been found [39], emphasizing that different pathway bottlenecks for different PCB congeners operate in different organisms, and that only diverse communities can be expected to able to deal with such complex pollutant mixtures. Of course, enzymes of PCB pathways may not only transform PCBs and their metabolites, but also other related compounds, such as monocyclic aromatics (see, for example, [40]), and vice versa. This substrate overlap means that other pollutants in a site may act as co-substrates that can influence the composition and activity of biphenylmetabolizing communities. Moreover, it was shown that biphenyl can be mineralized by mosaic pathways — in Pseudomonas putida CE2010, tod (toluene) and cmt (cumate) pathways complement one other to provide the ability to mineralize biphenyl, in the absence of an authentic biphenyl pathway [41•]. It is possible that such mosaic routes, both within individual strains and within communities, are important for the metabolism of complex pollutant mixtures.
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From where are PCB-degrading genes derived? This is still an open question. PCBs have been in the environment for about 70 years, and it is unlikely that the PCB-metabolizing genes have developed de novo. Many theories concerning the origin of the initial dioxygenases have been proposed, including those connecting these genes with the degradation of lignin and its derivatives. However, none of these theories have been fully proved and a connection with the degradation of terpenes is an alternative possibility. Such a connection is supported by reports, from some research groups, of an enhanced degradation of PCBs after the addition of terpenes to environmental samples. Given the central importance of microbial communities and microbial interactions in PCB degradation, and the inability to study these using traditional cultivation approaches, a recent focus has been on the use of cultureindependent methods to describe the composition of microbial communities in PCB-polluted environments, their adaptation mechanisms, the identification of the effective pollutant metabolizers, their population dynamics and activity under different environmental conditions, as well as the interactions among populations involved in the degradation of the pollutant or intermediates produced. Analysis of 16S ribosomal RNA (16S rRNA) gene sequences retrieved from polluted sites is currently the method of choice. Lloyd-Jones and Lau [42] analyzed the bacterial diversity in a surface sample from a sandy landfill soil in Canada containing an average concentration of 67 mg kg–1 of Aroclor 1260, and found high bacterial diversity with cloned sequence representatives of several bacterial divisions, such as low guanine + cytosine (G+C) Gram-positives, Cytophagales and the α-, β- and γ-subclasses of the proteobacteria. Sequences related to known PCB degraders were not detected, but the study was limited to only 20 clones. A more extensive study examined the composition of the bacterial community in surface samples of a sandy soil in East Germany [43,44], in which the PCBs (congener patterns similar to Aroclor 1242) were the predominant source of carbon in the site (concentrations as high as 24 g kg–1, with an average of 10–150 mg kg–1). The study mainly focussed on the identification of the metabolically active bacteria through analysis of 16S rRNA sequences amplified from total RNA by reverse transcriptase (RT)-PCR and fluorescence in situ hybridization (FISH). Analysis of 404 cloned sequences revealed an extremely high diversity with representatives of 10 major phylogenetic divisions, including two putative new bacterial groups. The community contained an abundance of β-proteobacteria, mostly Burkholderia spp. and Variovorax spp., with significant abundance of sequence types related to Sphingomonas spp., to acidiphilic bacteria from the α-subclass of proteobacteria commonly found in non-polluted acidic soils, and to the Holophaga-Acidobacterium phylum. Many Burkholderia, Variovorax and Sphingomonas sequences were related to isolates known to degrade pollutants such as PCBs.
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In contrast to the many studies on anaerobic PCBdechlorinating enrichment cultures, there have been only a few reports on experimental aerobic PCB-degrading microbial communities. Luensdorf et al. [45••,46] described a simple experimental system to study de novo development of PCB-degrading microbial communities. This involves a hydrophobic support material floating on a water column in a flask containing PCB-contaminated soil. The soil provides both the inoculum and carbon source for the development of a PCB-supported biofilm community on the substratum. The resulting biofilm was studied morphologically, chemically and microbiologically through 16S rRNA gene fingerprinting and clone library analysis. Interestingly, both the diversity and nature of the biofilm community were similar to that of the inoculum soil, suggesting that the model is a valid representation of a soil PCB biofilm community. The most fascinating finding, however, was that classical biofilms did not develop on the support material, but developed instead on microbe–mineral composite biofilms consisting of bacteria and clay minerals, in which the clay leaflets were arranged around the bacteria to form hutch-like structures. As bacteria lacking clay leaflets were rarely observed to be attached to the support, it seems that the clay minerals performed an essential role in biofilm formation, probably acting as a nutrient shuttle, mediating transfer of PCBs in a palatable form to the bacteria. This suggests that clay minerals may fulfil a similar function in soil and, thus, that the movement and dispersion of clay particles in PCB-contaminated soil may be an important regulating factor in PCB degradation.
Biosurfactants Cells basically exist and function in an aqueous milieu and, hence, the uptake into and excretion from cells of materials, like substrates and wastes, mostly involves their transport through aqueous media. The low water solubility of hydrophobic substances, and their tendency to bind strongly to hydrophobic matrices, may be a crucial factor limiting their rate of passage to and from the cell. PCBs are superhydrophobic, have extremely poor solubility in water and are poorly bioavailable. Microbes that utilize hydrophobic substrates, such as petroleum oil, often produce surfaceactive substances (biosurfactants or bioemulsifiers, or amphipathic compounds with very diverse structures that reduce interfacial tension between aqueous and nonaqueous fluid phases), increase mixing of the two phases, increase the surface area of hydrophobic substrates and their rate of transfer into or through aqueous media and, hence, increase bioavailability of the compound. A number of studies on the influence of surfactants on the biodegradation of hydrophobic organic substrates have been made, mostly in simple experimental systems. Compounds here included: chemically produced surfactants, such as ethoxylated alcohols, sulphates, sulphonates, Tritons, Brij 35 and sodium dodecyl sulphate [47,48]; enzymatically produced cyclodextrins [49,50]; humic substances [51,52]; and biosurfactants, such as carvone,
glucose lipid, rhamnolipids, lipopeptides (for example, lichenysin) and alasan, produced by different microorganisms [53–56]. In most cases, an increase in degradation rates was observed when surfactants or biosurfactants were added. However, two studies involving more complex experimental systems reported that surfactants decreased degradation rates (despite an increase in substrate solubilization owing, perhaps, to desorption of degraders from matrix or substrate surfaces and changes provoked in community composition), with a decrease in degrader populations [57,58•]. In another case, the presence of soil significantly reduced the stimulatory activity of exogenously added surfactant [59]. An elegant concept, based on recombinant bacteria capable of degrading both PCBs and chemical surfactants, was described by Lajoie et al. [60,61]. In this case, exogenous addition of the recombinant strain(s) and cognate surfactant resulted in the simultaneous increase in pollutant bioavailability and degrader populations. Chemical surfactants have the advantage of low price, but are often toxic for biological systems and, hence, are pollutants in their own right. Biosurfactants generally exhibit higher interfacial tension reduction activities than chemical surfactants, and are non- (or less) toxic and readily biodegradable, properties important not only for the natural turnover of hydrophobic substrates by microbes but also for the bioremediation of polluted sites [62•,63•,64]. Moreover, even though biosurfactants are more expensive to produce industrially (ex situ) than are chemical surfactants [65], their in situ production by microbes precisely where they have greatest effect can circumvent cost issues. However, no rules that specify optimal combinations of ‘surfactant–microorganism–pollutant–environment’ have thus far been developed, and empirical approaches are still the order of the day.
Conclusions Most of our knowledge about the degradation of PCBs comes from laboratory experimental systems involving pure cultures of bacteria in laboratory media containing a single PCB congener. However, given that the majority of microbes in environmental samples cannot be cultured at present in laboratory media, there is no certainty that those which have been studied so far are typical of those involved in PCB degradation in the environment. Moreover, it has become increasingly evident that metabolism of substrates in natural environments occurs not via the linear pathways that are familiar to us from pure culture studies, but via complex metabolic networks that involve: yet to be discovered reactions, metabolites and routes; multiple community members exchanging metabolites and regulating carbon flow according to the availability of other substrates and nutrients; prevailing physico-chemical conditions; and community needs. Thus, despite the importance of reductionist approaches for revealing underlying mechanisms, they may not reflect what happens in the environment. Recent efforts have therefore focussed on getting to grips with diverse but functionally integrated
Polychlorinated biphenyl-degrading microbial communities in soils and sediments Abraham et al.
complex microbial communities that metabolize complex PCB mixtures in natural heterogeneous matrices like soil and sediments or near-to-nature experimental systems. Such studies have revealed a number of surprises, including high diversity in PCB-metabolizing microbial communities, the exploitation of PCB-bearing clay minerals to form composite bacterial biofilms, and the ability of PCB degraders to form biofilms directly on PCB droplets. The intimate association of bacteria with the hydrophobic substrate obviously diminishes the problem of bioavailability at the local level. There are clearly more surprises in store. Gaining an understanding of PCB degradation will require: elucidation of the nature of the interacting surfaces (PCBs in the environment do not have the surfaces of the pure chemical, and degraders in a biofilm on a PCB droplet will not have the surfaces of a laboratory planktonic culture), particularly the role of biosurfactants in conditioning these surfaces, and the mechanisms of transfer and uptake of PCBs; the physical and metabolic interactions of the community members primarily involved in metabolism of the PCB carbon skeleton and chlorines; community interactions with and activity modulation by the prevailing physicochemical parameters of the environment; and the role of grazers and phages in regulating community composition and activities. Nevertheless, given the prevalence and diversity of PCB degraders that can be isolated from the environment, we can, perhaps naively, be optimistic about the potential for bioremediation, once we understand the ground rules of engagement between microbial communities and the target compounds.
Acknowledgements Work in the Gesellschaft für Biotechnologische Forschung laboratory has been supported by grants from the Bundesministerium für Bilding und Forschung (Project Number 0319433C, 0318896C, Strategy fund ‘Soil functions’) and the European Union. K Timmis gratefully acknowledges generous support from the Fonds der Chemischen Industrie.
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