Microbial degradation of tetrachloromethane: mechanisms and perspectives for bioremediation


  • Editor: Ian Head

  • Present address: Christian Penny, Département Environnement et Agro-biotechnologies, Centre de Recherche Public – Gabriel Lippmann, Belvaux, Luxembourg.

Correspondence: Françoise Bringel, Département Micro-organismes, Génomes, Environnement, Université de Strasbourg, UMR 7156 CNRS, 28 rue Goethe, 67083 Strasbourg Cédex, France. Tel.: +33 3 688 518 15; fax: +33 3 688 520 28; e-mail: francoise.bringel@unistra.fr


Toxic man-made compounds released into the environment represent potential nutrients for bacteria, and microorganisms growing with such compounds as carbon and energy sources can be used to clean up polluted sites. However, in some instances, microorganisms contribute to contaminant degradation without any apparent benefit for themselves. Such cometabolism plays an important part in bioremediation, but is often difficult to control. Microbial degradation of tetrachloromethane (carbon tetrachloride, CCl4), a toxic ozone-depleting organic solvent mainly of anthropogenic origin, is only known to occur by cometabolic reduction under anoxic conditions. Yet no microbial system capable of using CCl4 as the sole carbon source has been described. Microbial growth based on CCl4 as a terminal electron acceptor has not been reported, although corresponding degradation pathways would yield sufficient energy. Known modes for the biodegradation of CCl4 involve several microbial metabolites, mainly metal-bound coenzymes and siderophores, which are produced by facultative or strictly anaerobic bacteria and methanogenic Archaea. Recent reports have demonstrated that CCl4 dechlorination rates are enhanced by redox-active organic compounds such as humic acids and quinones, which act as shuttles between electron-providing microorganisms and CCl4 as a strong electron acceptor. The key factors underlying dechlorination of CCl4, the practical aspects and specific requirements for microorganism-associated degradation of CCl4 at contaminated sites and perspectives for future developments are discussed.

Tetrachloromethane in the living environment: recalcitrance, toxicity and transformation

Tetrachloromethane (carbon tetrachloride, CCl4) is a volatile chlorinated solvent with biocidal properties, which has been used widely over decades as an industrial degreasing agent, as a pesticide, for dry cleaning and in fire extinguishers (Doherty, 2000). It is toxic and predicted to be carcinogenic, with deleterious effects on stratospheric ozone (Table 1). As a consequence, commercial production and use of CCl4 has been progressively restricted. Its use as a pesticide and grain fumigant was banned in 1986 (ITRC-In Situ Bioremediation Team, 2002). The Montreal protocol on substances that deplete the ozone layer (1987) and its four amendments (London, 1990; Copenhagen, 1992; Montreal, 1997; Beijing, 1999) have implemented a complete phase-out of the use of CCl4, by 1996 for developed countries and by 2010 for developing countries [United Nations Environment Programme (UNEP), 2006]. Currently, CCl4 is still produced, but only as an intermediate in the production of other chemical compounds. Prolonged large-scale use of CCl4 has led to substantial soil and subsurface aquifer contamination and CCl4 is at the top of the priority list of hazardous groundwater contaminants (Knox & Canter, 1996). With an estimated half-life for abiotic hydrolysis of 7000 years in water at 20 °C (Vogel et al., 1987), CCl4 is highly persistent in the environment compared with other halogenated aliphatic compounds. In the case of dichloromethane, for example, published estimates range from 1.5 to 704 years (Vogel et al., 1987). Moreover, the low water solubility of CCl4 (Table 1) leads to its accumulation in subsurface aquifers as a poorly bioavailable, dense non-aqueous-phase liquid (DNAPL), which dissolves only slowly into groundwater (ITRC-In Situ Bioremediation Team, 2002).

Table 1.   Environmental and chemical data on tetrachloromethane (CCl4)
  • *

    Amounts produced minus amounts degraded or used in the manufacture of other chemicals according to the Montreal Protocol (1987; UNEP, 2006); data from 192 countries.

  • Based on a 100-year time horizon relative to an identical mass of CO2 (GWP=1.0, Allen et al., 2009).

  • Ozone impact ratio of a chemical compared with that of an identical mass of CFC-11 (trichlorofluoromethane; ODP=1.0; UNEP, 2006).

  • §

    Substances with a log(POW) between 1.5 and 3 have high biocidal toxicity (Ramos et al., 1997).

  • Calculated for aqueous 1 M solutions (pH 7.0; 25°C; 1 atm; Dolfing & Janssen, 1994). An energy difference of ∼70 kJ allows for the formation of one ATP under physiological conditions (El Fantroussi et al., 1998).

 NaturalMarine algae, oceans, volcanoes, drill wells. Mean concentrations in volcanic gases: 2.0 ± 1.0 p.p.b.Isidorov et al. (1990), Butler et al. (1999), Gribble (2003)
 Anthropogenic*Industrial production. Net production: +73 000 ton (1990); +48 000 ton (2000); −9500 ton (2007)UNEP website, http://ozone.unep.org/Data_Reporting/Data_Access/
Environmental data
 Toxicity for human healthClassified in group 2B (possibly carcinogenic; nongenotoxic; causes hepatic, renal and neurological damage)IARC (1999), WHO (2004)
 Drinking water guideline value4 μg L−1WHO (2004)
 Subsurface half-life7000 years (hydrolysis)Vogel et al. (1987)
 Stratospheric lifetime34 ± 5 years (photolysis)Allen et al. (2009)
 Atmospheric concentration100–130 p.p.t.Allen et al. (2009)
 Global warming potential (GWP)1400Allen et al. (2009)
 Ozone-depleting potential (ODP)1.1UNEP (2006)
 Chlorine equivalents contribution to ozone depletion9%Butler (2000)
Physicochemical properties
 Molecular weight153.8 g mol−1WHO (2004)
 Density1.594 at 20°CWHO (2004)
 Octanol/water partition coefficient (logPow)§2.64WHO (2004)
 Water solubility800 mg L−1 at 20°CWHO (2004)
 Boiling point76.5°CWHO (2004)
 Henry's law constant29.5 atm L mol−1 at 25°CDolfing & Janssen (1994)
 Oxidation state+4 
Gibbs free energy values (ΔG°′) and redox potentialDolfing & Janssen (1994)
 Reductive hydrogenolytic dechlorination
  Tetrachloromethane→Trichloromethane−192.6 kJ/584 mV 
  Trichloromethane→Dichloromethane−170.8 kJ/471 mV 
  Dichloromethane→Chloromethane−157.4 kJ/402 mV 
  Chloromethane→Methane−153.2 kJ/380 mV 
  Tetrachloromethane→Methane−674 kJ 
  Tetrachloromethane→CO2 (with H2O as an   electron donor and O2 as an electron acceptor)−551 kJ 

The toxicity of CCl4 to living organisms is well documented (IARC, 1999; WHO, 2004; Eastmond, 2008), and this also applies to microorganisms. Exposure of bacteria to CCl4 was shown to cause inhibition of a variety of environmentally significant metabolic processes, such as methanogenesis and autotrophy, even at very low concentrations (3 and 80 μM, respectively; Bauchop, 1967; Egli et al., 1988). As with other chlorinated methanes, CCl4 may exert a biostatic effect on methanogenic Archaea due to its structural similarity to other C1 compounds, which is likely to affect methane formation through competitive inhibition of enzymatic reactions or interaction with key cofactors of the pathway (Zhao et al., 2009). As a lipophilic compound with a high octanol/water partition coefficient (Table 1), CCl4 may also cause damage to cellular membranes. As reviewed by Sikkema et al. (1995), cytotoxic organic solvents disturb membrane permeability, thereby disrupting critical functions, for example by dissipation of the membrane potential and through loss of valuable cellular components. For example, cytoplasmic enzymes involved in Escherichia coli central metabolism were released from Mg2+-depleted cells treated with toluene due to structural alterations of the cytoplasmic membrane (de Smet et al., 1978).

Strategies described for microorganisms that tolerate organic solvents involve mechanisms that prevent intracellular exposure to the toxicants, such as membrane adaptation, for example through alterations of phospholipid fatty acid and headgroup composition, to ensure homeostasis of membrane fluidity. Sequestration mediated by membrane vesicles (Kobayashi et al., 2000), active extrusion with energy-driven efflux pumps (reviewed by Nicolaou et al., 2010) and membrane proton-motive force maintenance upon solvent-damaged inner membrane involving phage shock protein synthesis (Engl et al., 2009) may also afford cell protection against the toxic effects of halogenated solvents. In the specific case of CCl4, modifications in the saturated phospholipid content were observed in the aerobic methylotroph Methylobacterium extorquens DM4 (Vuilleumier et al., 2009) exposed to very low (0.13 mM, 20 mg L−1) concentrations of CCl4 (C. Penny, F. Bringel, C. Gruffaz, T. Nadalig, H. Heipieper & S. Vuilleumier, unpublished data). However, it is striking that both CCl4-degrading and -non-degrading bacteria were equally insensitive to the deleterious effects of CCl4 at concentrations near or exceeding its water solubility (Table 2). Clearly, many aspects of the bacterial tolerance to CCl4, as for other halogenated compounds, have yet to be investigated.

Table 2.   Minimal inhibitory concentrations (MIC) of tetrachloromethane for selected Proteobacteria
 Characteristic metabolic traitPhylogenetic affiliationMIC (mg L−1)*
  • *

    Tested for aerobic liquid cultures in 5 mL Difco nutrient broth (strains DM4 and ULPAs1) or CAA medium (Pseudomonas strains; Munsch et al., 2000) in 17-mL Hungate tubes sealed with Viton rubber stoppers (Glasgerätebau Ochs); incubation in a Microtron rotary shaker (Infors, Switzerland) at 100 r.p.m. and 30°C; CCl4 added using saturated aqueous solutions (800 mg L−1) prepared in the corresponding culture media from ultrapure CCl4 (purity >99.9%; Fluka).

  • Limit of water solubility at 20°C.

Methylobacterium extorquens DM4 (DSM 6343)Dichloromethane degradationAlphaproteobacteria400
Herminiimonas arsenicoxydans ULPAs1 (DSM 17148)Arsenic resistanceBetaproteobacteria400
Pseudomonas putida (DSM 291)Degrades many organic pollutantsGammaproteobacteria>800
Pseudomonas putida (DSM 3602)Degrades many organic pollutantsGammaproteobacteria600
Pseudomonas stutzeri KC (DSM 7136)Tetrachloromethane degradationGammaproteobacteria>800

Degradation or transformation of CCl4 is the other major source of toxicity of the compound, as some dechlorination pathways generate toxic intermediates and products (Fig. 1; more details in Tetrachloromethane-degrading bacteria: why not better? and Cometabolism galore: a large panel of low-molecular-weight molecules enhances CCl4 degradation). This mainly seems to be due to intracellular CCl4 transformation by nonspecific reactions, leading to the formation of reactive radicals that, by promoting nonspecific oxidation, can detrimentally affect and inactivate key cellular components, including proteins, DNA and lipids (McGregor & Lang, 1996). This was most clearly shown in investigations involving the Ames test, in which exposure to gaseous CCl4 was shown to have mutagenic effects on Salmonella typhimurium and E. coli tester strains (Araki et al., 2004).

Figure 1.

 Overview of the possible microorganism-mediated transformations of tetrachloromethane. The oval on the left symbolises a bacterium. Reduced electron shuttle compounds have been demonstrated to catalyze the reductive dechlorination of CCl4. The reduced form of these compounds can be regenerated from the oxidized form by diverse types of microbial metabolism. Roles for CCl4 as a carbon source for growth (1) or as a terminal electron acceptor (2) and CCl4-specific dehalogenases (3) have not yet been described.

This paper presents an overview of the prokaryotic organisms mediating CCl4 dechlorination, describes a large panel of reactions and catalysts as well as the thermodynamic and kinetic aspects of this dechlorination, and discusses the physicochemical conditions necessary for microorganism-mediated CCl4 degradation. Perspectives for research to discover new, more efficient bacterial strains and to apply bacterial metabolism for the treatment of sites contaminated with tetrachloromethane are then proposed.

Tetrachloromethane-degrading bacteria: why not better?

The first experiments on microorganisms capable of degrading tetrachloromethane were reported in the early 1980s (Bouwer & McCarty, 1983a, b), almost a century after industrial CCl4 production started at the end of the 19th century in Germany and in England (Doherty, 2000), and almost 150 years after the chemical synthesis of CCl4 was first reported by Regnault in 1839. Since then, bacterial consortia and isolated strains able to degrade CCl4 have been obtained from a large number of sites, not all of which were contaminated with this compound (Table 3).

Table 3.   Reports of microbial degradation of tetrachloromethane from the literature
Pure bacterial strain, culture
enrichment or consortium (origin)
Culture conditions/targeted
Tetrachloromethane degradationReferences
  • *

    Carbon mass balance of tetrachloromethane degradation products of 100%.

  • Former name Desulfitobacterium frappieri.

  • Former name Methanobacterium thermoautotrophicum.

  • §

    Former name Clostridium thermoaceticum.

  • ND, not determined.

 Acetobacterium woodii DSM 1030 (marine mud)Autotrophic, acetogenic1–1000CobalaminCO2; CO; CS2; chloroform; dichloromethane; acetate; pyruvate; lactate; isobutyrate; hydrophobic and cell-bound material*Egli et al. (1988, 1990), Stromeyer et al. (1992), Hashsham & Freedman (1999)
 Clostridium ruminantium TM5 (CCl4-polluted groundwater)Fermentation65NDUnknown products; chloroform (traces)C. Penny, C. Gruffaz, T. Nadalig, H.M. Cauchie, S. Vuilleumier & F. Bringel (unpublished data)
 Clostridium sp. TCAIIB (anaerobic bioreactor)Fermentation0.6NDChloroform; dichloromethaneGälli & McCarty (1989)
 Dehalobacter restrictus DSM 9455 (PCE-dechlorinating column)H2, electron donor; tetrachloroethene, electron acceptor600CorrinoidNDMaillard et al. (2003)
 Desulfitobacterium hafniense TCE1 (chloroethene-polluted soil)Lactate, electron donor; tetrachloroethene, electron acceptor40CorrinoidChloroform; dichloromethaneGerritse et al. (1999)
 Desulfobacterium autotrophicum HRM2 (marine mud)Autotrophic, sulfate reducing40–80Cobalamin; b- and c-type cytochromesChloroform; dichloromethane; soluble and cell-bound materialEgli et al. (1987, 1988), Stromeyer et al. (1992)
 Escherichia coli K-12 (human feces)Fermentation or fumarate respiration0.6–1.3NDCO2; CS2; chloroform; soluble and cell-bound material*Criddle et al. (1990b)
 Geobacter metallireducens (mud)Iron reducing2–40Reduced ironCO; CH4; chloroformMcCormick et al. (2002), McCormick & Adriaens (2004)
 Geobacter sulfurreducens (ditch surface sediment)Iron reducing3.5Reduced iron; AQDSChloroform and unknown productsMaithreepala & Doong (2009)
 Klebsiella pneumoniae L17 (subsurface forest sediment)Iron reducing8ND; enhanced by reduced iron and AQDSChloroform and unknown productsLi et al. (2009)
 Klebsiella pneumoniae TM2 (CCl4-polluted groundwater)Fermentation65NDUnknown products; chloroform (traces)C. Penny et al. (unpublished data)
 Methanosaeta concilii DSM 3671 (anaerobic sewage sludge)Methanogenic1Cobalamin; cytochromes; coenzyme F430Chloroform and unknown productsNovak et al. (1998a)
 Methanosarcina barkeri DSM 1538 (anaerobic sewage sludge)Methanogenic5Cobalamin; cytochromes; coenzyme F430Chloroform and unknown productsNovak et al. (1998a)
 Methanosarcina thermophila DSM 1825 (thermophilic digester sludge)Methanogenic2.5–8Cobalamin; cytochromes; coenzyme F430; zinc porphorinogenChloroform; soluble and cell-bound material*Andrews & Novak (2001), Baeseman & Novak (2001), Koons et al. (2001), Novak et al. (1998a, b)
 Methanothermobacter thermautotrophicus DeltaH (anaerobic sewage sludge)Autotrophic, methanogenic40–50Corrinoid; coenzyme F430Chloroform and unknown productsEgli et al. (1987, 1990)
 Moorella thermoacetica DSM 512 (horse feces)§Acetogenic80CorrinoidChloroform; dichloromethaneEgli et al. (1988)
 Pseudomonas stutzeri KC (groundwater aquifer solids)Denitrifying0.6–30Cu(II):pyridine-2,6-bis(thiocarboxylate) complexCO2; CS2; CSCl2; chloroform; soluble and cell-bound material*Criddle et al. (1990a), Lee et al. (1999), Lewis & Crawford (1993, 1995), Lewis et al. (2001), Tatara et al. (1993)
 Shewanella alga BrY (red alga Jania sp. surface)Lactate or H2 as an electron donor75–150Added vitamin B12 or iron oxidesCO; chloroform*Gerlach et al. (2000), Workman et al. (1997)
 Shewanella oneidensis MR-1 (lake sediment)Lactate, formate, H2, electron donors; Fe(III), electron acceptor0.03–15Menaquinone-1; vitamin K2CO2; chloroform; soluble and cell-bound material*Fu et al. (2009); Petrovskis et al. (1994); Ward et al. (2004)
 Shewanella putrefaciens 200 (oil pipeline)Lactate, electron donor19.5Cytochrome c type; enhanced by soil organic matter and reduced ironCO2; chloroform; volatile, soluble and cell-bound material*Backhus et al. (1997), Collins & Picardal (1999), Kim & Picardal (1999), Picardal et al. (1993, 1995)
 Sporotalea propionica TM1 (CCl4-polluted groundwater)Fermentation65NDUnknown products; chloroform (traces)C. Penny et al. (unpublished data)
 Anaerobic digester sludge from a sewage treatment plantAcidogenic0.6–60NDChloroform; dichloromethaneMun et al. (2008)
 Anaerobic digester sludge from a baker yeast factoryAcidogenic and methanogenic3–110NDCO2; chloroform; dichloromethane; chloromethaneSponza (2001, 2002)
 Anaerobic sewage effluent from a water pollution control facilityDenitrifying0.3–0.5NDCO2; chloroform; cell-bound material and unknown productsBouwer & McCarty (1983b)
 Aquifer or sediment materialDenitrifying9–13NDChloroform and unknown productsSherwood et al. (1999)
 Aquifer materialDenitrifying0.006–0.6NDChloroform and unknown productsSherwood et al. (1996)
 Anaerobic enrichment culture from an anaerobic digesterDichloromethane the sole carbon and energy source0.13–340ND; enhanced by cobalaminCO2; CO; CS2; CH4; chloroform; dichloromethane; acetate; formate; methanol; (iso)butyrate soluble and cell-bound material*Hashsham et al. (1995)
 Aquifer or sediment materialMethanogenic2.5NDChloroform and unknown productsBaeseman & Novak (2001)
 Anaerobic sugar beet refinery wastewater treatment reactor, granular sludgeMethanogenic10NDCO2; CS2; chloroform; dichloromethane; chloromethane and cell-bound material*Van Eekert et al. (1998)
 Mixed methanogenic consortium from a stock reactorMethanogenic0.5–10NDChloroform; dichloromethane and unknown productsAdamson & Parkin (1999)
 Uncontaminated soil from an industrial siteSulfate reducing or fermentative42–65NDCH4; CO2; CO; CS2; chloroform; dichloromethane; chloromethane; hydrophobic and cell-bound materialShan et al. (2010)
 Waste-activated sludgeMethanogenic0.3–1.3NDCO2Bouwer & McCarty (1983a)
 Wastewater treatment plant, anaerobic distillery granular sludgeMethanogenic100ND; enhanced by cobalamin, riboflavin or AQDSChloroform; dichloromethane; perchloroethylene and unknown productsGuerrero-Barajas & Field (2005, 2006)
 Wastewater treatment plant, anaerobic digester sludgeMixed anaerobic50–60NDCO2; CH4; chloroform; dichloromethane; acetatede Best et al. (1999)
 Wastewater treatment plant, granular sludges or wet oxidized effluentsMixed anaerobic cultures fed with acetate, butyrate and propionate5ND; enhanced by humic acids and AQDSChloroform; dichloromethane; perchloroethyleneCervantes et al. (2004)
 Wastewater treatment plant of a sugar corporation, anaerobic biosolidsMixed anaerobic cultures fed with glucose, acetate or humic acid0.6–6.5NDChloroform and unknown productsDoong et al. (1996, 1997), Doong & Chang (2000), Doong & Wu (1996)
 Wastewater treatment plant, anaerobic digester sludgeSulfate reducing11NDChloroform; dichloromethane and unknown productsde Best et al. (1998)
 Wastewater treatment plant, anaerobic digester sludgeSulfate reducing, nitrate reducing, iron reducing, methanogenic, fermenting or mixed electron acceptor2NDChloroform; dichloromethane; chloromethaneBoopathy (2002)

Consortia and strains capable of CCl4 degradation

A major common characteristic of CCl4-degrading bacteria is their ability to grow under anoxic conditions. So far, microbial CCl4 degradation has only been observed under reducing conditions (Table 1), in keeping with the oxidized nature of the carbon in the molecule. The range of culture conditions under which CCl4 degradation has been reported is remarkable, and includes sulfate-reducing, nitrate-reducing, iron-reducing, fermentative and methanogenic conditions (Table 3). Enrichment cultures or consortia capable of CCl4 degradation have been reported (Table 3; >20 cases), but few have been taxonomically characterized, or the consortium member responsible for dehalogenation identified (four cases). For example, Zhou et al. (1999) identified a high G+C Gram-positive bacterium related to Rhodococcus, which represented 70% of a dechlorinating consortium enriched from CCl4-contaminated water in the presence of toluene. However, whether this strain was indeed responsible for CCl4 degradation was not demonstrated. In other investigations, acetogenic anaerobic bacteria were proposed to afford efficient reductive CCl4 removal, in consortia composed mainly of methanogens, sulfate reducers and acetogens isolated from digester sludge of wastewater treatment plants (de Best et al., 1999; Mun et al., 2008). Nevertheless, as detailed in Table 3, the phylogenetic diversity of CCl4-degrading strains is broad: 12 facultative or strict anaerobic bacterial lineages and three methanogenic archaeal lineages were shown to mediate CCl4 degradation.

Bacterial CCl4 degradation: a thermodynamic enigma

No organism capable of using CCl4 as a carbon or an energy source has been isolated and no specific tetrachloromethane dehalogenase is known. This may seem puzzling, given that mineralization of CCl4 to carbon dioxide (CO2) and its reductive transformation to methane are highly exergonic processes (Table 1). However, the favorable energetics of CCl4 transformation under aerobic conditions are somewhat misleading. In terms of its formal oxidation number, the carbon of CCl4 is at the same level (+4) as CO2, so that mineralization of CCl4 to CO2 represents a hydrolytic process without a change in the redox state, and without the release of electrons as a potential energy source. The major contribution to the energetics of this transformation is due to the formation of chloride ions [approximately −131 kJ mol−1, e.g. Dolfing (2003)]. This, however, may not immediately yield metabolically useful energy or carbon for biomass production – the latter would require the reduction of CO2 to carbon at the oxidation state of formaldehyde, HCHO. One possibility for a bacterium to harvest energy from dehalogenation of CCl4 to CO2 for its metabolism would be to exploit a transmembrane gradient generated by dehalogenation. This strategy is used by perchloroethylene-dehalorespiring bacteria: protons generated by the hydrogenase required to deliver electrons for dehalogenation flow back into the cell along their concentration gradient by way of an energy-yielding membrane-bound ATPase (Futagami et al., 2008). In principle, exploitation of a chloride gradient for energy generation could also be envisaged. This has not been observed, possibly because chloride ions diffuse at little or no energy cost across cellular membranes following their concentration gradient (e.g. 100 times more easily than sodium ions). Most likely, the buildup of a transmembrane chloride gradient for subsequent exploitation for energy production would require very profound adjustments as well as evolutionary adaptations of cellular metabolism.

Inspiration from dehalorespiration

The transformation of CCl4 in four successive dehalogenation reactions may be an energetically favorable process overall, but some steps will be energetically more favorable than others, as a function of environmental conditions and redox potential in particular (e.g. Dolfing, 2003); degradation of perchloroethylene by reductive dehalogenation is a well-studied example. It is most favorable by dehalorespiration (Holliger & Schumacher, 1994) under highly anaerobic conditions for the two initial steps to 1,2-dichloroethylene (1,2-DCE), but becomes energetically more favorable under aerobic conditions, with 1,2-DCE serving as a source of energy and possibly also as a carbon source. In addition, other metabolic strategies such as sulfate reduction, iron (III) reduction or even methanogenesis are often energetically competitive with reductive dehalogenation of perchloroethylene under the physicochemical conditions under which this process takes place, setting significant selective constraints for the survival and development of perchloroethylene-degrading organisms in the environment (e.g. Luijten et al., 2004; Aulenta et al., 2007a). Similar constraints will most likely apply to microorganisms involved in CCl4 degradation: in theory, CCl4 indeed represents a favorable electron acceptor in energy-yielding dehalorespiration processes. The redox potential for the CCl4/chloroform couple of +584 mV (Dolfing & Janssen, 1994; Table 1) is higher than that for the reduction of common electron acceptors used in microbial metabolism [MnO2, NO3, Fe(OH)3, SO42−, HCO3]. Accordingly, CCl4 was proposed to serve as an electron acceptor for growth in a benchmark study of a CCl4-degrading mixed community composed of methanogenic Archaea, sulfate-reducing and acetogenic bacteria (de Best et al., 1999). However, given that acetogenic bacteria are capable of autotrophic growth under anoxic conditions (Pierce et al., 2008), the possibility that in this case CO2 acted as an electron acceptor in this consortium was not completely ruled out. In any event, it is intriguing that this most promising work was not pursued further in an attempt to identify the organisms involved in CCl4 degradation. Several reasons may have contributed, such as the concentrations of transformed CCl4 (about 35 μM) too low to characterize the specific biomass buildup, toxicity of CCl4 metabolites or the existence of a functional consortium of several strains, each of which play an essential role in CCl4 degradation.

Which electron donors for bacterial metabolism with CCl4 as an electron acceptor?

The electron donors used to reduce highly chlorinated electron acceptors such as CCl4 are quite varied (Field & Sierra-Alvarez, 2004), and may often involve extracellular electron transfer between different bacteria in dechlorinating consortia (Stams et al., 2006). For example, the associations between hydrogen-producing acetogenic bacteria and electron-consuming methanogens may also support the degradation of halogenated compounds (e.g. Dojka et al., 1998; Duhamel & Edwards, 2007), and such electron-sharing associations are likely to be involved in some of the CCl4-dechlorinating strains and consortia described in Table 3. Addition of hydrogen gas to methanogenic cultures was shown to enhance CCl4 degradation (Novak et al., 1998a). Nevertheless, the presence of electron acceptors other than halogenated compounds determines the levels of redox potential and hydrogen concentration at which reductive dehalogenation metabolism will occur (e.g. Luijten et al., 2004; Aulenta et al., 2007a). Thiol compounds, biogenic iron species and other reducing agents potentially present in the environment may provide alternative reducing equivalents for reductive dechlorination (e.g. H2S, Na2S, zero-valent iron, pyrite, magnetite, goethite; Assaf-Anid et al., 1994; Chiu & Reinhard, 1996; Doong & Chiang, 2005).

Degradation of CCl4: activating an inert molecule

What are the properties and reactivity of the CCl4 molecule that determine the isolation of bacteria capable of degrading it? CCl4 is inert because it has no carbon–hydrogen bonds, and is a tetrahedral symmetric molecule. Its carbon atom does not have an electrophilic nature even though each of its four carbon–chlorine bonds is highly polarized. In other words, two-electron substitutive reactions on CCl4 can be essentially ruled out, and a radical reaction is needed for the cleavage of a carbon–chlorine bond. Thus, unlike the inertness of CCl4 itself, the reaction of the compound by a radicalar mechanism will initially yield two highly reactive entities: a chlorine atom and a halogenated carbon radical. These reactive species are highly toxic, as any biological molecule in close proximity is easily oxidized. For all subsequent dehalogenation reactions on the original CCl4 molecule, it may be very difficult for an organism to control the harvest of carbon and energy from CCl4 for cellular metabolism or biomass production. This situation is quite different from that of lesser chlorinated halogenated methanes chloromethane and dichloromethane, which represent carbon and energy sources for microbial growth under both aerobic and anaerobic conditions (e.g. Messmer et al., 1993; Mägli et al., 1998; Kayser et al., 2002; Studer et al., 2002). For both these compounds and unlike for CCl4, the resulting transformation products are nonchlorinated central metabolic intermediates of microbial methylotrophic metabolism.

CCl4 carbon: can it be assimilated?

Whether CCl4 be used for biomass formation for growth has not been demonstrated. In experiments using radiolabelled 14CCl4, 14C was incorporated into acetate and several other products (pyruvate, lactate, ethanol, isobutyrate) by cultures of Acetobacterium woodii and Moorella thermoacetica (Egli et al., 1988; Hashsham & Freedman, 1999; Adamson & Parkin, 2001). This suggests that the cellular incorporation of carbon monoxide (CO) and CO2 derived from the degradation of CCl4 occurred via the reductive acetyl–CoA pathway (the Wood–Ljungdahl pathway; Fig. 1) and that CCl4-derived carbon may be assimilated under certain conditions (Fig. 1), provided that adequate electron donors are available.

Cometabolism galore: a large panel of low-molecular-weight molecules enhances CCl4 degradation

Many bacteria capable of CCl4 degradation synthesize copious amounts of redox-active low-molecular-weight compounds, which act primarily as cofactors in central metabolic enzymatic electron transfer reactions. These compounds, which include organometallic compounds such as cobalt-containing corrinoids, iron-bound porphyrins (e.g. cytochromes), a nickel-containing factor F430, as well as key cofactors such as riboflavin or menaquinone, enhance the reductive cometabolic dehalogenation of CCl4. For instance, during acetyl-CoA synthesis, methanogenesis, dehalorespiration, fermentation pathways and DNA synthesis (Martens et al., 2002), corrinoid cofactors are produced in a wide variety of taxonomically diverse phyla of CCl4-degrading strains (Table 3; the acetogenic bacteria A. woodii and M. thermoacetica; the enteric bacteria E. coli and Klebsiella pneumoniae; the dehalorespiring bacteria Desulfitobacterium hafniense and Dehalobacter restrictus; and the methanogenic Archaea Methanosarcina barkeri, Methanosarcina thermophila, Methanosaeta concilii and Methanothermobacter thermautotrophicus). The degradation of CCl4 does not always take place inside cells. Cometabolic dechlorination of CCl4 recruits a large panel of low-molecular-weight molecules that can act in an extracellular process, even after cell death.

Use of enzymatic cofactors in CCl4 degradation

Membrane-bound c-type cytochromes, the related hematin (Gantzer & Wackett, 1991; Picardal et al., 1993; Curtis & Reinhard, 1994), riboflavin (vitamin B2; Guerrero-Barajas & Field, 2005), menaquinone (vitamin K2 and analogues; Fu et al., 2009), the reduced form of cobalamin (vitamin B12) and cobamides with diverse ligands to the corrinoid ring (Rondon et al., 1997) catalyze the degradation of CCl4 directly or in conjunction with other factors acting as electron shuttles (Table 3). Different compounds display variable efficiencies in CCl4 degradation. The coenzyme F430 of methyl coenzyme M reductase (Rouvière & Wolfe, 1988; Novak et al., 1998a, b; Baeseman & Novak, 2001; Koons et al., 2001), involved in a late step of methanogenesis, catalyzed the degradation of CCl4 (2.2 mM) at a molar ratio of 0.02 for F430 to CCl4 (Krone et al., 1989). A similar molar ratio of 0.04 for vitamin B12 to CCl4 enabled the reductive degradation of CCl4 (100 nM) (Assaf-Anid et al., 1994). Compared with cyano-, hydroxy- and methylcobalamin, adenosylcobalamin was 10-fold less effective in the dechlorination of CCl4 in an anaerobic enrichment culture (Hashsham et al., 1995). However, compared with riboflavin, cobalamin compounds added to a methanogenic sludge consortium were three times more effective and yielded less potentially toxic chloroform as an end product (Guerrero-Barajas & Field, 2005).

In keeping with a fortuitous, catalytic role of such cofactors in CCl4 degradation, it appears that the more a microorganism is able to produce such catalytic cofactors, the greater its potential to degrade CCl4 and other chlorinated compounds. For example, supplementation of growth medium with porphobilinogen, a de novo vitamin B12 biosynthesis precursor of the corrin ring of cobalamin, enhanced CCl4 biodegradation in methanogenic cultures fed with methanol (Guerrero-Barajas & Field, 2006). Methanogens in particular produce more than one catalytic factor in CCl4 degradation: corrinoids, factor F430, one or more zinc porphyrins and b- and c-type cytochromes (Krone et al., 1989; Baeseman & Novak, 2001). Increased CCl4 degradation was concomitant with increased basal cobalamin production and factor F430 levels in the methanogen M. barkeri (Mazumder et al., 1987; Van Eekert et al., 1998). In cultures of autotrophically grown A. woodii, increased dechlorination rates correlated with a higher content of corrinoid-bound methyltransferases of the acetyl-CoA pathway, compared with growth under heterotrophic conditions (Egli et al., 1988). Anaerobic mixed cultures fed with 1,2-propanediol, a compound whose fermentation to propionaldehyde requires a vitamin B12-dependent diol dehydratase, displayed higher CCl4 degradation kinetics compared with cultures growing with substrates that do not require corrinoid cofactors for their utilization (propionaldehyde, dextrose, acetate) (Toraya et al., 1979; Zou et al., 2000).

Even microorganisms without intrinsic CCl4 degradation activity may facilitate this process, possibly by mediating cofactor regeneration or through the delivery of reducing equivalents. Efficient CCl4 degradation in the nondechlorinating strain Shewanella alga BrY was observed when vitamin B12 was added (Workman et al., 1997). Another explanation is that CCl4 transformation may also take place outside living cells.

Extracellular transformation of CCl4: excreted microbial metal chelators and electron shuttles

The best-known excreted metal chelator involved in CCl4 degradation is pyridine-2,6-bis(thiocarboxylate), or PDTC, a transition metal-chelating molecule identified as a secondary siderophore of Pseudomonas stutzeri KC, a nitrate-reducing bacterium isolated from an aquifer at Seal Beach (CA; Criddle et al., 1990a; Tatara et al., 1993; Lee et al., 1999; Lewis et al., 2001, 2004). Pseudomonas stutzeri KC was found to catalyze extracellular PDTC-dependent CCl4 dehalogenation (Lee et al., 1999; Lewis et al., 2001). Unlike other biomolecules known to mediate reductive CCl4 dechlorination, PDTC is not regenerated by electron addition after CCl4 degradation, but is a true reactant converted to dipicolinic acid in the dehalogenation process (Lewis et al., 2001). Copper, but not iron, nickel and cobalt complexes of PDTC enable dechlorination of CCl4 to CO2, formate and nonvolatile products (Dybas et al., 1995; Lewis & Crawford, 1995; Lewis et al., 2001, 2004). PDTC-dependent CCl4 transformation relies on the pdt gene cluster for biosynthesis of PDTC and on the bipartite outer-membrane/inner-membrane transport system for iron acquisition from Fe(III):PDTC (Lewis et al., 2000; Leach & Lewis, 2006). cDNA microarrays were designed to track the expression of pdt genes to monitor in situ dechlorination activity in CCl4-contaminated environments (Musarrat & Hashsham, 2003). This study demonstrated the iron-independent expression of the pdt operon and its relevance in monitoring CCl4-degrading bacterial subpopulations using DNA microarray technology.

In addition to the low-molecular-weight organometallic compounds of directly biotic origin just discussed, other types of chemicals present in the environment were also repeatedly shown to contribute to CCl4 degradation by acting as electron shuttles (Van der Zee & Cervantes, 2009). Both inorganic metal complexes and organic matter are capable of transferring reducing equivalents produced by microorganisms to halogenated compounds including CCl4 (Watanabe et al., 2009). Such processes may enhance the degradation of CCl4 in the environment, because inorganic shuttles such as ferric oxides and hydroxides, goethite (α-FeOOH), hematite (α-Fe2O3) and magnetite (Fe3O4) represent predominant electron acceptor species in many aquifer sediments (McCormick & Adriaens, 2004). The reduction of surface-bound iron particles generates reactive biogenic Fe(II) species capable of dechlorinating polyhalogenated aliphatic compounds under anoxic conditions (Pecher et al., 2002). The chemical transformation of CCl4 via iron oxidation was enhanced in the presence of dissimilatory iron-reducing bacteria (DIRB), commonly found in soil and groundwater ecosystems (Lovley et al., 2004; Scala et al., 2006) (see Fig. 1). The list of DIRB reported to be involved in CCl4 degradation (Table 3) includes strains of Geobacter sulfurreducens, Geobacter metallireducens, K. pneumoniae, Shewanella putrefaciens and S. alga (Picardal et al., 1995; Gerlach et al., 2000; Maithreepala & Doong, 2008, 2009; Li et al., 2009).

Organic electron shuttle compounds such as natural organic matter, humic acids and quinones are also known to enhance abiotic CCl4 degradation (Maithreepala & Doong, 2008, 2009; Li et al., 2009). Bacteria may provide the required electrons. Soils and aquifers contain large amounts of humic and fulvic acids, and quinones are considered to be the dominant electron acceptors within humic substances (Scott et al., 1998; Collins & Picardal, 1999), supporting dechlorination reactions through one or multiple electron-accepting sites (Curtis & Reinhard, 1994; reviewed by Van der Zee & Cervantes, 2009). For example, the reduced form of the electron shuttle and humic analogue anthraquinone disulfonate (AQDS), anthrahydroquinone disulfonate (AHQDS), catalyzes CCl4 degradation directly or indirectly by transferring electrons to other biomolecules or inorganic compounds able to degrade it reductively (Schink, 2006). AQDS can act as an intermediate electron shuttle, or it can be oxidized, and the electrons generated in the process are delivered directly to CCl4. Oxidized AQDS is returned to its reduced form (AHQDS) by electrons provided by microorganisms (Backhus et al., 1997; Collins & Picardal, 1999) or by abiotic reductants such as thiols (Doong & Chiang, 2005). The degradation of CCl4 (at an initial concentration of 100 μM) increased two- to sixfold upon addition of small amounts of AQDS (5–50 μM) to a Geobacter-dominated consortium (Cervantes et al., 2004). Continuous redox cycling implies that only substoichiometric concentrations of electron shuttles are required for rapid and efficient CCl4 degradation (Hashsham et al., 1995; Cervantes et al., 2004; Guerrero-Barajas & Field, 2005).

CCl4 degradation products and rates

In terms of the toxicity of the final degradation products, the reactions mediated by the Cu(II):PDTC complex and by corrinoid compounds result in the cleanest known transformation of CCl4, to CO2 (10–70% carbon recovery as CO2 from CCl4) and nonvolatile soluble compounds (e.g. acetate, pyruvate; 20–50%) or cell-bound material (4–10% carbon recovery), without accumulation of chloroform (Egli et al., 1988; Criddle et al., 1990a; Hashsham et al., 1995; Lewis et al., 2001). In contrast, the reductive hydrogenolysis of CCl4 successively leads to the formation of chloroform, dichloromethane, chloromethane and finally of methane (Vogel et al., 1987; de Best, 1999;Table 1). In many cases, trichloromethyl and dichlorocarbene radicals are generated in the initial reactions (de Best, 1999). Coupling of two trichloromethyl radicals or one dichlorocarbene radical with a molecule of CCl4 results in hexachloroethane formation, which, under reducing conditions, is readily transformed to perchloroethylene (Cervantes et al., 2004; Guerrero-Barajas & Field, 2006). Similarly, hydrolytic mechanisms yield formate (HCOOH) and CO2, but CO and phosgene (COCl2) were also observed. In the presence of sulfur-containing nucleophiles (e.g. H2S), the product distribution through thiolytic dechlorination generally contains less chloroform, but significant amounts of carbon disulfide (CS2; Kriegman-King & Reinhard, 1992; Hashsham et al., 1995).

In terms of the reaction rates, the highest dechlorination rates and efficiencies, ranging from 0.2 to>70 μg day−1 mg−1 protein, were mediated by PDTC- and cobalamin-producing microbial cultures, including methanogenic Archaea (Krone et al., 1989; Criddle et al., 1990a; Van Eekert et al., 1998; Gerritse et al., 1999; Hashsham & Freedman, 1999; Boopathy, 2002). Such rates are low compared with those observed for the dehalogenation of compounds used as electron acceptors or growth substrates (e.g. two to four orders of magnitude for perchloroethylene; Holliger & Schraa, 1994; Holliger & Schumacher, 1994). Nevertheless, if only cometabolic processes are considered, the rates observed for CCl4 are often higher than those for trichloroethane (by about one order of magnitude) or perchloroethylene (two orders of magnitude) (Gälli & McCarty, 1989; Adamson & Parkin, 1999).

Because the degradation of CCl4 in the environment may be quite rapid under favorable conditions that involve the close interplay across the biotic–abiotic divide of environmental molecules and microorganisms, both the diversity and the mechanisms of these interactions need to be investigated.

Bacteria-mediated remediation of CCl4-contaminated sites: approaches, achievements and perspectives

Remediation strategies for CCl4 usually involve physical and chemical approaches, for example soil excavation, groundwater stripping or venting. These approaches are often associated with a high cost, variable efficiency and hazardous ecological consequences, including the mobilization of this volatile, ozone-depleting compound into the atmosphere (Schwarzenbach et al., 2006; Environmental Protection Agency, 2008). The literature on in situ bioremediation of CCl4 investigating bioaugmentation, biostimulation and natural attenuation approaches is reviewed in this paper. We will then discuss emerging techniques such as phytoremediation (e.g. Suresh & Ravishankar, 2004) and microbial fuel cells (Lovley, 2008).

Bioaugmentation, the more the better

In most pilot-scale studies, bioaugmentation trials have featured the addition of the CCl4-degrading bacterium P. stutzeri KC to CCl4-contaminated aquifers (Dybas et al., 1998, 2002; Pfiffner et al., 2000). Colonization of a test ecosystem by strain KC, inoculated with 1500 L of a cell suspension at 2 × 107 CFU mL−1, was supported by the addition of acetate and phosphate, adjusted to a slightly alkaline pH to favor the development of strain KC. This was optimized by stimulating the chemotactic motility of the strain in a nitrate gradient (Witt et al., 1999), and by evaluating different feeding strategies (Dybas et al., 1998; Phanikumar et al., 2002). Efficient CCl4 removal was also obtained when acetate, nitrate and phosphate were added in a 100 : 10 : 1 C : N : P ratio to a microcosm composed of CCl4-contaminated sediments and groundwater, and inoculated with strain KC (107 cells mL−1; Pfiffner et al., 2000). In this particular study, pH control was unnecessary, and iron or copper had no detectable inhibitory effect on CCl4 removal. A similar approach was then tested on a larger scale by the same authors over 4 years (Dybas et al., 2002). The treatment of 18 000 m3 of contaminated groundwater (sediment concentrations of 23±17 μg kg−1) by bioaugmentation with strain KC as an inoculum (18 900 L at 2 × 107 CFU mL−1) and feeding with acetate (100 mg L−1) and phosphate (10 mg L−1) removed 96% of the CCl4, with chloroform below the detection limit in the treated groundwater. In contrast, inefficient CCl4 removal and chloroform generation were observed when strain KC was absent or not adequately stimulated, indicating that the indigenous bacterial population was not sufficient. To our knowledge, this study is the only published case describing the successful large-scale microorganism-mediated bioremediation of a CCl4-contaminated site using bioaugmentation. Besides strain KC, other strains and consortia may also be used for the bioremediation of CCl4-contaminated sites. Recently, sulfate-reducing and fermentative microbial consortia, in combination with vitamin B12 addition, have been used for the bioaugmentation treatment of halomethane-contaminated soils (Shan et al., 2010).

Biostimulation, the fitter the better

Under simulated environmental conditions in the laboratory, the first attempts to enhance in situ microbial degradation of CCl4 were performed by addition of appropriate electron donors and acceptors (Semprini et al., 1992). Injection into a model shallow subsurface aquifer of acetate as a growth substrate and a potential electron donor, together with electron acceptors nitrate and sulfate, resulted in efficient in situ biodegradation of CCl4 and other halogenated aliphatic hydrocarbons. The sulfate-reducing population, which formed a minor part of the microbial community, was the active player in the conversion of CCl4, whereas the predominant denitrifying microbial population did not participate in CCl4 degradation (Semprini et al., 1992). The major drawback of the approach was that chloroform accumulated to levels up to 30–60% of the initial CCl4 concentration. The addition of catalytic cofactors to increase the anaerobic biotransformation of CCl4, such as commercial vitamin B12, was also proposed (Hashsham et al., 1995; Guerrero-Barajas & Field, 2006). A more cost-effective alternative along the same lines might be to enhance biological in situ bacterial production of cobalamin by the addition of compounds that stimulate cobalamin-dependent metabolic pathways in endogenous microbial populations, such as the fermentation substrate 1,2-propanediol or the vitamin B12 precursor porphobilinogen, in the presence of the methyltrophic methanogen growth substrate methanol (Guerrero-Barajas & Field, 2006). These results are very promising; biostimulation of microbial activity through the addition of appropriate nutrients or key molecules and control of physicochemical conditions on-site may be the most reliable approach for the biological treatment of CCl4 contamination.

Natural attenuation, the power of doing nothing

The first long-term evaluation of the natural attenuation of sites polluted with CCl4 was an 11-year-long monitoring of a DNAPL plume comprising a mixture of CCl4 (>90%), toluene and petroleum oil (Davis et al., 2003). In this strongly reducing groundwater environment, the disappearance of the DNAPL and the concomitant increase of chloroform, dichloromethane and inorganic chloride clearly indicated that the degradation of CCl4 was ongoing. The detection of CS2 most likely came from abiotic processes of reductive dechlorination by iron sulfide species (Davis et al., 2003). In another study of a contaminated chemical manufacturing site, the efficient natural attenuation of CCl4 was limited by in situ bioavailability of carbon and electrons, and by unfavorable physicochemical conditions such as a low pH and a high redox potential (Mack et al., 2001). However, complete CCl4 removal was apparent in a zone of lower redox potentials and increased pH, apparently as a result of the addition of lime-associated organic material in an earlier decontamination treatment, which stimulated microbial dechlorination. This suggests that successful in situ CCl4 biodegradation strongly depends on environmental conditions such as a low oxygen content, low redox potentials and the presence of one or more electron donor species, and that the stimulation of endogenous microbial activity may be the key factor for efficient CCl4 degradation, both by supporting the growth of CCl4-degrading microorganisms and by altering in situ physicochemical conditions favorably.

Future perspectives of microbial biodegradation of CCl4

Considering the ubiquity of microbial CCl4 degraders, the diversity of their energy metabolism and the large range of biomolecules capable of catalyzing the transformation of CCl4, virtually every ecosystem has CCl4 dechlorination potential in store. Further studies should explore the diversity of microbial CCl4 degradation pathways in the environment and evaluate the conditions required for the most efficient degradation. These studies would be facilitated by recent developments in microbial community analysis such as real-time PCR (e.g. Smith & Osborn, 2009), FISH (e.g. Amann & Fuchs, 2008), genotyping techniques (Pandey et al., 2009), metagenomics (Xu, 2006; Simon & Daniel, 2009; Wilmes et al., 2009) and metaproteomics (e.g. Eyers et al., 2004; Steele & Streit, 2005; Lacerda & Reardon, 2009). Stable isotope probing will help to better assess the still unknown fate of CCl4 carbon in the environment. Because this technique has sufficient sensitivity to detect the low levels of CCl4 degraded by microorganisms, it should aid the identification of microorganisms that incorporate carbon from CCl4 into their biomass (Dumont & Murrell, 2005).

Bioremediation strategies based on the potential of the endogenous microbial communities need to focus on the control and definition of on-site redox conditions. Strictly anoxic conditions appear to be most favorable for efficient CCl4 removal. However, such conditions are rarely guaranteed in soils and subsurface aquifers, and may depend on the activity of oxygen-consuming microorganisms. This is particularly important for methanogenic Archaea, which are not only sensitive to low CCl4 concentrations (Bauchop, 1967) but also to exposure to oxygen in trace amounts (Garcia et al., 2000). Thus, the potential of methanogens for CCl4 degradation in bioremediation might be somewhat restricted.

In contrast, under aerobic conditions, bacteria able to degrade CCl4 aerobically have not been reported so far, despite the fact that CCl4 degradation is thermodynamically favorable under these conditions (Table 1). The toxicity associated with oxidative, often oxygenase-mediated transformation of halogenated methanes [see e.g. Jiang et al. (2010) for a recent review] may explain this observation. Higher plants, such as poplar trees, may remove CCl4 from subsurface aquifers by an aerobic mechanism involving cytochrome P450 homologues (Wang et al., 2002). The cytochrome P450 content in these trees appeared to be rate limiting in the process, because transgenic poplar trees overexpressing mammalian cytochrome P450 were more efficient in degrading CCl4 (Doty et al., 2007). The panel of observed dechlorination products and the observed specific degradation rates of 0.1–5 μg day−1 mg−1 protein were similar to those of microbial CCl4 degradation under anoxic conditions (Hartmann et al., 2000; Wang et al., 2002). Although microbial endophytes that could potentially enhance CCl4 degradation or plant-associated CCl4-degrading rhizosphere bacteria were not detected in such experiments (e.g. Wang et al., 2004), the development of close interactions between plant-associated CCl4-degrading microorganisms and a plant partner, combining anaerobic and aerobic or microaerophilic habitats for efficient CCl4 degradation, should be investigated more systematically.

Another potential new avenue being explored for bioremediation is the use of microbial fuel cells (Lovley, 2008) to provide a steady supply of electrons required to stimulate the microbial degradation of CCl4. Microbial fuel cells have already been used in several applications, including electricity generation, wastewater treatment and biohydrogen production (Du et al., 2007). Fine adjustment of working reduction potentials, the combined use of electrodes and bacteria, may help to overcome limitations in exploitable reducing equivalents that often represent a major barrier to the efficient degradation of highly chlorinated compounds (Aulenta et al., 2006). An electrochemical cell for electrolytic reductive dechlorination of CCl4 in aqueous environments was designed and developed almost 20 years ago (Criddle & McCarty, 1991). More recently, the potential of microbial fuel cells was exploited for the reductive dechlorination of trichloroethene in a bioelectrochemically assisted reductive dechlorination process (termed BEARD), which involved the use of selected microbial strains for the transformation of the halogenated compound and methyl viologen as the redox mediator (Aulenta et al., 2007b, 2009). Microorganisms thereby acquired electrons at the cathode via a soluble or an electrode surface-bound methyl viologen intermediate, and subsequently enhanced the reductive dechlorination of trichloroethene into ethene. Whether this approach may be transposed to CCl4 remains to be seen.

To conclude, the physicochemical properties of the CCl4 molecule, the requirements for the temporal sequence of defined redox condition reactions for efficient metabolic exploitation of carbon and energy, the existing competition between CCl4 dehalogenation and other more efficient metabolic strategies and the toxicity of possible reaction intermediates all combine to make the productive use of CCl4 transformation by microorganisms a very challenging prospect. The very significant costs in developing a multistep pathway for CCl4 degradation, involving initial energy investment, efficient carbon funnelling and energy harvest, in the context of a possibly erratic, low-level supply of this compound in the environment (Table 1), make the evolution of such metabolism, and certainly its long-term success quite unlikely. Still, in the light of the recent discoveries of novel anaerobic metabolic pathways, such as Anammox (Strous et al., 2006) and anaerobic methane oxidation (Ettwig et al., 2010), it is not implausible that organisms capable of growing with CCl4 both as a carbon source and as an electron acceptor may exist. A metabolic pathway may exist in which methane generated by the reduction of CCl4 would then be oxidized anaerobically to CO2, to yield the required reducing equivalents for CCl4 reduction to methane. A bacterial strain possessing such a pathway may have the capacity to grow with CCl4 as a carbon and energy source. The increasing sophistication and power of both culture-dependent and culture-independent approaches could be used to investigate microbial functions at the global level of ecosystems (e.g. Xu, 2006; Simon & Daniel, 2009; Wilmes et al., 2009) and will certainly contribute towards improvements in bioremediation strategies for sites contaminated by CCl4 in the near future.


This work was supported by funding from a Programme Pluri-Formations and the Contrat-Plan Etat-Région to REALISE, the Network of Laboratories in Engineering and Science for the Environment in the Alsace Région (France) and by the EC2CO Program of French Institut National des Sciences de l'Univers. Support from the National Research Fund of Luxembourg for PhD and researcher mobility grants to C.P. (Grants No. Ext-BFR-05-085 and FNR-08-AM2c-21, http://www.fnr.lu) is also gratefully acknowledged. Thanks are due to Brett Johnson for his valuable suggestions.