MINIREVIEW
Microbial degradation of tetrachloromethane: mechanisms and
perspectives for bioremediation
Christian Penny, Stéphane Vuilleumier & Françoise Bringel
Département Micro-organismes, Génomes, Environnement, Université de Strasbourg, UMR 7156 CNRS, Strasbourg, France
Correspondence: Françoise Bringel,
Département Micro-organismes, Génomes,
Environnement, Université de Strasbourg,
UMR 7156 CNRS, 28 rue Goethe, 67083
Strasbourg Cédex, France. Tel.: 133 3 688
518 15; fax: 133 3 688 520 28;
e-mail: [email protected]
Present address: Christian Penny,
Département Environnement et Agrobiotechnologies, Centre de Recherche Public
– Gabriel Lippmann, Belvaux, Luxembourg.
Received 1 December 2009; revised 21 May
2010; accepted 14 June 2010.
Final version published online 2 August 2010.
MICROBIOLOGY ECOLOGY
DOI:10.1111/j.1574-6941.2010.00935.x
Editor: Ian Head
Keywords
tetrachloromethane; carbon tetrachloride;
bioremediation; cometabolism; reductive
dechlorination; electron shuttles.
Abstract
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 electronproviding 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,
FEMS Microbiol Ecol 74 (2010) 257–275
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 1C (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
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
258
C. Penny et al.
aquifers as a poorly bioavailable, dense non-aqueous-phase
liquid (DNAPL), which dissolves only slowly into groundwater (ITRC-In Situ Bioremediation Team, 2002).
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 environmen-
tally significant metabolic processes, such as methanogenesis
and autotrophy, even at very low concentrations (3 and
80 mM, 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
Table 1. Environmental and chemical data on tetrachloromethane (CCl4)
Features
Origin
Natural
Anthropogenic
Environmental data
Toxicity for human health
Drinking water guideline value
Subsurface half-life
Stratospheric lifetime
Atmospheric concentration
Global warming potential (GWP)w
Ozone-depleting potential (ODP)z
Chlorine equivalents contribution to ozone depletion
Physicochemical properties
Molecular weight
Density
Octanol/water partition coefficient (logPow)‰
Water solubility
Boiling point
Henry’s law constant
Oxidation state
Gibbs free energy values (DG1 0 ) and redox potential
Reductive hydrogenolytic dechlorinationz
Tetrachloromethane ! Trichloromethane
Trichloromethane ! Dichloromethane
Dichloromethane ! Chloromethane
Chloromethane ! Methane
Tetrachloromethane ! Methane
Mineralization
Tetrachloromethane ! CO2 (with H2O as an
electron donor and O2 as an electron acceptor)
Facts
References
Marine algae, oceans,
volcanoes, drill wells. Mean
concentrations in volcanic
gases: 2.0 1.0 p.p.b.
Industrial production. Net
production: 173 000 ton
(1990); 148 000 ton (2000);
9500 ton (2007)
Isidorov et al. (1990), Butler et al.
(1999), Gribble (2003)
Classified in group 2B (possibly
carcinogenic; nongenotoxic;
causes hepatic, renal and
neurological damage)
4 mg L1
7000 years (hydrolysis)
34 5 years (photolysis)
100–130 p.p.t.
1400
1.1
9%
IARC (1999), WHO (2004)
153.8 g mol1
1.594 at 20 1C
2.64
800 mg L1 at 20 1C
76.5 1C
29.5 atm L mol1 at 25 1C
14
UNEP website, http://ozone.unep.org/
Data_Reporting/Data_Access/
WHO (2004)
Vogel et al. (1987)
Allen et al. (2009)
Allen et al. (2009)
Allen et al. (2009)
UNEP (2006)
Butler (2000)
WHO (2004)
WHO (2004)
WHO (2004)
WHO (2004)
WHO (2004)
Dolfing & Janssen (1994)
Dolfing & Janssen (1994)
192.6 kJ/584 mV
170.8 kJ/471 mV
157.4 kJ/402 mV
153.2 kJ/380 mV
674 kJ
551 kJ
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.
w
Based on a 100-year time horizon relative to an identical mass of CO2 (GWP = 1.0, Allen et al., 2009).
z
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).
z
Calculated for aqueous 1 M solutions (pH 7.0; 25 1C; 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).
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
FEMS Microbiol Ecol 74 (2010) 257–275
259
Microbial degradation of tetrachloromethane
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 Mg21-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 L1)
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.
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).
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 2. Minimal inhibitory concentrations (MIC) of tetrachloromethane for selected Proteobacteria
Methylobacterium extorquens DM4 (DSM 6343)
Herminiimonas arsenicoxydans ULPAs1 (DSM 17148)
Pseudomonas putida (DSM 291)
Pseudomonas putida (DSM 3602)
Pseudomonas stutzeri KC (DSM 7136)
Characteristic metabolic trait
Phylogenetic affiliation
MIC (mg L1)
Dichloromethane degradation
Arsenic resistance
Degrades many organic pollutants
Degrades many organic pollutants
Tetrachloromethane degradation
Alphaproteobacteria
Betaproteobacteria
Gammaproteobacteria
Gammaproteobacteria
Gammaproteobacteria
400
400
4 800w
600
4 800w
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 1C; CCl4 added using saturated aqueous solutions (800 mg L1) prepared in the corresponding culture media from ultrapure CCl4
(purity 4 99.9%; Fluka).
w
Limit of water solubility at 20 1C.
FEMS Microbiol Ecol 74 (2010) 257–275
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
260
C. Penny et al.
1
Growth substrate
Electron donor
Carbon and energy source
CO / CS
2
Cellular
metabolism
Terminal
electron acceptor
Reducing
equivalents
Biomass
Oxidised
Fe(III) species,
natural organic matter,
redox active cofactors
Electron shuttles
Reduced
Fe(II) species,
natural organic matter,
redox active cofactors
CO / CO / HCOOH / COCl / CS / CSCl
CHCl / CH Cl / CH Cl / CH
Cl C=CCl / Cl C-CCl
Dipicolinic acid
CCl4
Cu(II):PDTC
complex
3
·CCl 3
:CCl2
Hydrolytic or thiolytic substitution
Hydrogenolysis
C-C coupling reactions or
dihaloelimination
Fig. 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.
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; 4 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 G1C 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
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
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 (14) 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 mol1, 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 perchloroethyleneFEMS Microbiol Ecol 74 (2010) 257–275
Corrinoid
Corrinoid
Cobalamin; b- and c-type
cytochromes
0.6
600
40
40–80
Fermentation
FEMS Microbiol Ecol 74 (2010) 257–275
Fermentation
H2, electron donor;
tetrachloroethene, electron
acceptor
Lactate, electron donor;
tetrachloroethene, electron
acceptor
Autotrophic, sulfate reducing
Clostridium sp. TCAIIB
(anaerobic bioreactor)
Dehalobacter restrictus DSM
9455 (PCE-dechlorinating
column)
Desulfitobacterium hafniense
TCE1 (chloroethene-polluted
soil)w
Desulfobacterium
autotrophicum HRM2 (marine
mud)
Escherichia coli K-12 (human
feces)
Geobacter metallireducens
(mud)
Geobacter sulfurreducens
(ditch surface sediment)
Klebsiella pneumoniae L17
(subsurface forest sediment)
Klebsiella pneumoniae TM2
(CCl4-polluted groundwater)
Methanosaeta concilii DSM
3671 (anaerobic sewage
sludge)
Methanosarcina barkeri DSM
1538 (anaerobic sewage
sludge)
Methanosarcina thermophila
DSM 1825 (thermophilic
digester sludge)
Reduced iron
2–40
3.5
8
65
1
5
2.5–8
Iron reducing
Iron reducing
Fermentation
Methanogenic
Methanogenic
Methanogenic
Cobalamin; cytochromes;
coenzyme F430; zinc
porphorinogen
Cobalamin; cytochromes;
coenzyme F430
Cobalamin; cytochromes;
coenzyme F430
ND; enhanced by reduced
iron and AQDS
ND
Reduced iron; AQDS
ND
0.6–1.3
ND
ND
Fermentation or fumarate
respiration
Iron reducing
65
Cobalamin
Clostridium ruminantium TM5
(CCl4-polluted groundwater)
1–1000
Autotrophic, acetogenic
Acetobacterium woodii DSM
1030 (marine mud)
Cofactor(s)
Culture conditions/targeted
metabolism
Pure bacterial strain, culture
enrichment or consortium (origin)
CCl4
concentrations
(mM)
Tetrachloromethane degradation
Table 3. Reports of microbial degradation of tetrachloromethane from the literature
Chloroform; soluble and cellbound material
Chloroform and unknown
products
Chloroform and unknown
products
Chloroform and unknown
products
Unknown products; chloroform
(traces)
Chloroform and unknown
products
CO2; CS2; chloroform; soluble and
cell-bound material
CO; CH4; chloroform
Chloroform; dichloromethane;
soluble and cell-bound material
Chloroform; dichloromethane
ND
Chloroform; dichloromethane
Egli et al. (1988, 1990), Stromeyer
et al. (1992), Hashsham & Freedman
(1999)
CO2; CO; CS2; chloroform;
dichloromethane; acetate;
pyruvate; lactate; isobutyrate;
hydrophobic and cell-bound
material
Unknown products; chloroform
(traces)
Andrews & Novak (2001),
Baeseman & Novak (2001), Koons
et al. (2001), Novak et al. (1998a, b)
Novak et al. (1998a)
Novak et al. (1998a)
C. Penny et al. (unpublished data)
Li et al. (2009)
McCormick et al. (2002),
McCormick & Adriaens (2004)
Maithreepala & Doong (2009)
Criddle et al. (1990b)
Egli et al. (1987, 1988), Stromeyer
et al. (1992)
Gerritse et al. (1999)
Maillard et al. (2003)
C. Penny, C. Gruffaz, T. Nadalig,
H.M. Cauchie, S. Vuilleumier & F.
Bringel (unpublished data)
Gälli & McCarty (1989)
References
Products
Microbial degradation of tetrachloromethane
261
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
ND
ND
65
0.6–60
3–110
0.3–0.5
9–13
Acidogenic
Acidogenic and
methanogenic
Denitrifying
Denitrifying
Denitrifying
Aquifer material
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
Dichloromethane the sole
carbon and energy source
Methanogenic
Methanogenic
Anaerobic enrichment culture
from an anaerobic digester
Aquifer or sediment material
Anaerobic sugar beet refinery
wastewater treatment reactor,
granular sludge
10
2.5
0.13–340
0.006–0.6
19.5
ND
ND
ND; enhanced by
cobalamin
ND
ND
ND
Cytochrome c type;
enhanced by soil organic
matter and reduced iron
ND
Fermentation
Shewanella putrefaciens 200
(oil pipeline)
0.03–15
Added vitamin B12 or iron
oxides
Menaquinone-1; vitamin
K2
Cu(II):pyridine-2,6bis(thiocarboxylate)
complex
Corrinoid
Sporotalea propionica TM1
(CCl4-polluted groundwater)
Anaerobic digester sludge from
a sewage treatment plant
Anaerobic digester sludge from
a baker yeast factory
Anaerobic sewage effluent from
a water pollution control facility
Aquifer or sediment material
0.6–30
Denitrifying
75–150
80
Acetogenic
Cofactor(s)
Corrinoid; coenzyme F430
Lactate or H2 as an electron
donor
Lactate, formate, H2, electron
donors; Fe(III), electron
acceptor
Lactate, electron donor
40–50
Autotrophic, methanogenic
Methanothermobacter
thermautotrophicus DeltaH
(anaerobic sewage sludge)z
Moorella thermoacetica DSM
512 (horse feces)‰
Pseudomonas stutzeri KC
(groundwater aquifer solids)
Shewanella alga BrY (red alga
Jania sp. surface)
Shewanella oneidensis MR-1
(lake sediment)
CCl4
concentrations
(mM)
Culture conditions/targeted
metabolism
Tetrachloromethane degradation
Pure bacterial strain, culture
enrichment or consortium (origin)
Table 3. Continued.
Products
CO2; chloroform;
dichloromethane; chloromethane
CO2; chloroform; cell-bound
material and unknown products
Chloroform and unknown
products
Chloroform and unknown
products
CO2; CO; CS2; CH4; chloroform;
dichloromethane; acetate;
formate; methanol; (iso)butyrate
soluble and cell-bound material
Chloroform and unknown
products
CO2; CS2; chloroform;
dichloromethane; chloromethane
and cell-bound material
Unknown products; chloroform
(traces)
Chloroform; dichloromethane
CO2; chloroform; volatile, soluble
and cell-bound material
CO2; chloroform; soluble and cellbound material
CO; chloroform
CO2; CS2; CSCl2; chloroform;
soluble and cell-bound material
Chloroform; dichloromethane
Chloroform and unknown
products
References
Van Eekert et al. (1998)
Baeseman & Novak (2001)
Hashsham et al. (1995)
Sherwood et al. (1996)
Sherwood et al. (1999)
Bouwer & McCarty (1983b)
Sponza (2001, 2002)
Mun et al. (2008)
Backhus et al. (1997), Collins &
Picardal (1999), Kim & Picardal
(1999), Picardal et al. (1993, 1995)
C. Penny et al. (unpublished data)
Criddle et al. (1990a), Lee et al.
(1999), Lewis & Crawford (1993,
1995), Lewis et al. (2001), Tatara
et al. (1993)
Gerlach et al. (2000), Workman
et al. (1997)
Fu et al. (2009); Petrovskis et al.
(1994); Ward et al. (2004)
Egli et al. (1988)
Egli et al. (1987, 1990)
262
C. Penny et al.
FEMS Microbiol Ecol 74 (2010) 257–275
FEMS Microbiol Ecol 74 (2010) 257–275
11
z
Former name Desulfitobacterium frappieri.
Former name Methanobacterium thermoautotrophicum.
‰
Former name Clostridium thermoaceticum.
ND, not determined.
w
ND
ND
0.6–6.5
2
ND
5
Mixed anaerobic cultures fed
with acetate, butyrate and
propionate
Mixed anaerobic cultures fed
with glucose, acetate or humic
acid
Sulfate reducing
Sulfate reducing, nitrate
reducing, iron reducing,
methanogenic, fermenting or
mixed electron acceptor
ND; enhanced by humic
acids and AQDS
50–60
Mixed anaerobic
ND
ND; enhanced by
cobalamin, riboflavin or
AQDS
ND
ND
0.3–1.3
100
42–65
Sulfate reducing or
fermentative
ND
Methanogenic
Methanogenic
0.5–10
Methanogenic
Carbon mass balance of tetrachloromethane degradation products of 100%.
Waste-activated sludge
Wastewater treatment plant,
anaerobic distillery granular
sludge
Wastewater treatment plant,
anaerobic digester sludge
Wastewater treatment plant,
granular sludges or wet oxidized
effluents
Wastewater treatment plant
of a sugar corporation, anaerobic
biosolids
Wastewater treatment plant,
anaerobic digester sludge
Wastewater treatment plant,
anaerobic digester sludge
Mixed methanogenic consortium
from a stock reactor
Uncontaminated soil from an
industrial site
Chloroform; dichloromethane
and unknown products
Chloroform; dichloromethane;
chloromethane
Chloroform and unknown
products
Chloroform; dichloromethane
and unknown products
CH4; CO2; CO; CS2; chloroform;
dichloromethane; chloromethane;
hydrophobic and cell-bound
material
CO2
Chloroform; dichloromethane;
perchloroethylene and unknown
products
CO2; CH4; chloroform;
dichloromethane; acetate
Chloroform; dichloromethane;
perchloroethylene
Boopathy (2002)
de Best et al. (1998)
Doong et al. (1996, 1997), Doong &
Chang (2000), Doong & Wu (1996)
Cervantes et al. (2004)
de Best et al. (1999)
Bouwer & McCarty (1983a)
Guerrero-Barajas & Field (2005,
2006)
Shan et al. (2010)
Adamson & Parkin (1999)
Microbial degradation of tetrachloromethane
263
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
264
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 1584 mV (Dolfing & Janssen, 1994; Table 1) is higher
than that for the reduction of common electron acceptors
used in microbial metabolism [MnO2, NO
3 , Fe(OH)3,
,
HCO
].
Accordingly,
CCl
was
proposed
to serve as
SO2
4
3
4
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
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
C. Penny et al.
to identify the organisms involved in CCl4 degradation.
Several reasons may have contributed, such as the concentrations of transformed CCl4 (about 35 mM) 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, zerovalent 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
FEMS Microbiol Ecol 74 (2010) 257–275
Microbial degradation of tetrachloromethane
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
14
CCl4, 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
FEMS Microbiol Ecol 74 (2010) 257–275
265
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 (GuerreroBarajas & 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 band 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
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
266
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 nitratereducing 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/innermembrane 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
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
C. Penny et al.
degradation of CCl4 in the environment, because inorganic
shuttles such as ferric oxides and hydroxides, goethite (aFeOOH), hematite (a-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 mM)
increased two- to sixfold upon addition of small amounts of
AQDS (5–50 mM) 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; GuerreroBarajas & 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
FEMS Microbiol Ecol 74 (2010) 257–275
267
Microbial degradation of tetrachloromethane
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 sulfurcontaining 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 4 70 mg day1
mg1 protein, were mediated by PDTC- and cobalaminproducing 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, groundFEMS Microbiol Ecol 74 (2010) 257–275
water 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 mL1, 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 mL1; 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 mg kg1) by bioaugmentation with strain KC as an inoculum (18 900 L at
2 107 CFU mL1) and feeding with acetate (100 mg L1)
and phosphate (10 mg L1) 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).
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
268
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 cobalamindependent 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 ( 4 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
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
C. Penny et al.
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
FEMS Microbiol Ecol 74 (2010) 257–275
269
Microbial degradation of tetrachloromethane
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 mg day1 mg1 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
FEMS Microbiol Ecol 74 (2010) 257–275
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.
Acknowledgements
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-08AM2c-21, http://www.fnr.lu) is also gratefully acknowledged. Thanks are due to Brett Johnson for his valuable
suggestions.
References
Adamson DT & Parkin GF (1999) Biotransformation of mixtures
of chlorinated aliphatic hydrocarbons by an acetate-grown
methanogenic enrichment culture. Water Res 33: 1482–1494.
Adamson DT & Parkin GF (2001) Product distribution during
transformation of multiple contaminants by a high-rate,
tetrachlorethene-dechlorinating enrichment culture.
Biodegradation 12: 337–348.
Allen NDC, Bernath PF, Boone CD, Chipperfield MP, Fu D,
Manney GL, Toon GC & Weisenstein DK (2009) Global
carbon tetrachloride distributions obtained from the
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
270
Atmospheric Chemistry Experiment (ACE). Atmos Chem Phys
9: 13299–13325.
Amann R & Fuchs BM (2008) Single-cell identification in
microbial communities by improved fluorescence in situ
hybridization techniques. Nat Rev Microbiol 6: 339–348.
Andrews EJ & Novak PJ (2001) Influence of ferrous iron and pH
on carbon tetrachloride degradation by Methanosarcina
thermophila. Water Res 35: 2307–2313.
Araki A, Kamigaito N, Sasaki T & Matsushima T (2004)
Mutagenicity of carbon tetrachloride and chloroform in
Salmonella typhimurium TA98, TA100, TA1535, and TA1537,
and Escherichia coli WP2uvrA/pKM101 and WP2/pKM101,
using a gas exposure method. Environ Mol Mutagen 43:
128–133.
Assaf-Anid N, Hayes KF & Vogel TM (1994) Reductive
dechlorination of carbon tetrachloride by cobalamin (II) in the
presence of dithiothreitol: mechanistic study, effect of redox
potential and pH. Environ Sci Technol 28: 246–252.
Aulenta F, Majone M & Tandoi V (2006) Enhanced anaerobic
bioremediation of chlorinated solvents: environmental factors
influencing microbial activity and their relevance under field
conditions. J Chem Technol Biot 81: 1463–1474.
Aulenta F, Canosa A, Leccese M, Papini MP, Majone M & Viottit
P (2007a) Field study of in situ anaerobic bioremediation of a
chlorinated solvent source zone. Ind Eng Chem Res 46:
6812–6819.
Aulenta F, Catervi A, Majone M, Panero S, Reale P & Rossetti S
(2007b) Electron transfer from a solid-state electrode assisted
by methyl viologen sustains efficient microbial reductive
dechlorination of TCE. Environ Sci Technol 41: 2554–2559.
Aulenta F, Canosa A, Roma LD, Reale P, Panero S, Rossetti S &
Majone M (2009) Influence of mediator immobilization on
the electrochemically assisted microbial dechlorination of
trichloroethene (TCE) and cis-dichloroethene (cis-DCE).
J Chem Technol Biot 84: 864–870.
Backhus DA, Picardal FW, Johnson S, Knowles T, Collins R,
Radue A & Kim S (1997) Soil- and surfactant-enhanced
reductive dechlorination of carbon tetrachloride in the
presence of Shewanella putrefaciens 200. J Contam Hydrol 28:
337–361.
Baeseman JL & Novak PJ (2001) Effects of various environmental
conditions on the transformation of chlorinated solvents by
Methanosarcina thermophila cell exudates. Biotechnol Bioeng
75: 634–641.
Bauchop T (1967) Inhibition of rumen methanogenesis by
methane analogues. J Bacteriol 94: 171–175.
Boopathy R (2002) Anaerobic biotransformation of carbon
tetrachloride under various electron acceptor conditions.
Bioresource Technol 84: 69–73.
Bouwer EJ & McCarty PL (1983a) Transformations of 1- and
2-carbon halogenated aliphatic organic compounds under
methanogenic conditions. Appl Environ Microb 45: 1286–1294.
Bouwer EJ & McCarty PL (1983b) Transformations of
halogenated organic compounds under denitrification
conditions. Appl Environ Microb 45: 1295–1299.
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
C. Penny et al.
Butler JH (2000) Better budgets for methyl halides? Nature 403:
260–261.
Butler JH, Battle M, Bender ML, Montzka SA, Clarke AD,
Saltzmank ES, Sucher CM, Severinghaus JP & Elkins JW
(1999) A record of atmospheric halocarbons during the
twentieth century from polar air. Nature 339: 749–755.
Cervantes FJ, Vu-Thi-Thu L, Lettinga G & Field JA (2004)
Quinone-respiration improves dechlorination of carbon
tetrachloride by anaerobic sludge. Appl Microbiol Biot 64:
702–711.
Chiu PC & Reinhard M (1996) Transformation of carbon
tetrachloride by reduced vitamin B12 in aqueous cysteine
solution. Environ Sci Technol 30: 1882–1889.
Collins R & Picardal F (1999) Enhanced anaerobic
transformations of carbon tetrachloride by soil organic matter.
Environ Toxicol Chem 18: 2703–2710.
Criddle CS & McCarty PL (1991) Electrolytic model system for
reductive dehalogenation in aqueous environments. Environ
Sci Technol 25: 973–978.
Criddle CS, DeWitt JT, Grbić-Galić D & McCarty PL (1990a)
Transformation of carbon tetrachloride by Pseudomonas sp.
strain KC under denitrification conditions. Appl Environ
Microb 56: 3240–3246.
Criddle CS, DeWitt JT & McCarty PL (1990b) Reductive
dehalogenation of carbon tetrachloride by Escherichia coli
K-12. Appl Environ Microb 56: 3247–3254.
Curtis GP & Reinhard M (1994) Reductive dehalogenation of
hexachloroethane, carbon tetrachloride, and bromoform by
anthrahydroquinone disulfonate and humic acid. Environ Sci
Technol 28: 2393–2401.
Davis A, Fennemore GG, Peck C, Walker CR, McIlwraith J &
Thomas S (2003) Degradation of carbon tetrachloride in a
reducing groundwater environment: implications for natural
attenuation. Appl Geochem 18: 503–525.
de Best JH (1999) Anaerobic transformation of chlorinated
hydrocarbons in a packed-bed reactor. PhD Thesis,
Rijksuniversiteit Groningen, the Netherlands.
de Best JH, Salminen E, Doddema HJ, Janssen DB & Harder W
(1998) Transformation of carbon tetrachloride under sulfate
reducing conditions. Biodegradation 8: 429–436.
de Best JH, Hunneman P, Doddema HJ, Janssen DB & Harder W
(1999) Transformation of carbon tetrachloride in an anaerobic
packed-bed reactor without addition of another electron
donor. Biodegradation 10: 287–295.
De Smet MJ, Kingma J & Witholt B (1978) The effect of toluene
on the structure and permeability of the outer and cytoplasmic
membranes of Escherichia coli. Biochim Biophys Acta 506:
64–80.
Doherty RE (2000) A history of the production and use of carbon
tetrachloride, tetrachloroethylene, trichloroethylene and 1,1,
1-trichloroethane in the United States: part 1 – historical
background; carbon tetrachloride and tetrachloroethylene.
Environ Forensics 1: 69–81.
Dojka MA, Hugenholtz P, Haack SK & Pace NR (1998) Microbial
diversity in a hydrocarbon- and chlorinated-solvent-
FEMS Microbiol Ecol 74 (2010) 257–275
Microbial degradation of tetrachloromethane
contaminated aquifer undergoing intrinsic bioremediation.
Appl Environ Microb 64: 3869–3877.
Dolfing J (2003) Thermodynamic considerations for
dehalogenation. Dehalogenation – Microbial Processes and
Environmental Applications (Häggblom MM & Bossert ID,
eds), pp. 89–114. Kluwer Academic Publishers, Dordrecht, the
Netherlands.
Dolfing J & Janssen DB (1994) Estimates of Gibbs free energies of
formation of chlorinated aliphatic compounds. Biodegradation
5: 21–28.
Doong RA & Chang SM (2000) Relationship between electron
donor and microorganism on the dechlorination of carbon
tetrachloride by an anaerobic enrichment culture.
Chemosphere 40: 1427–1433.
Doong RA & Chiang H (2005) Transformation of carbon
tetrachloride by thiol reductants in the presence of quinone
compounds. Environ Sci Technol 39: 7460–7468.
Doong RA & Wu SC (1996) Effect of substrate concentration on
the biotransformation of carbon tetrachloride and 1,1,1trichloroethane under anaerobic condition. Water Res 30:
577–586.
Doong RA, Chen TF & Chang WH (1996) Effects of electron
donor and microbial concentration on the enhanced
dechlorination of carbon tetrachloride by anaerobic consortia.
Appl Microbiol Biot 46: 183–186.
Doong RA, Chen TF & Wu Y (1997) Anaerobic dechlorination of
carbon tetrachloride by free-living and attached bacteria under
various electron-donor conditions. Appl Microbiol Biot 47:
317–323.
Doty SL, James CA, Moore AL et al. (2007) Enhanced
phytoremediation of volatile environmental pollutants with
transgenic trees. P Natl Acad Sci USA 104: 16816–16821.
Du Z, Li H & Gu T (2007) A state of the art review on microbial
fuel cells: a promising technology for wastewater treatment
and bioenergy. Biotechnol Adv 25: 464–482.
Duhamel M & Edwards EA (2007) Growth and yields of
dechlorinators, acetogens, and methanogens during reductive
dechlorination of chlorinated ethenes and dihaloelimination
of 1,2-dichloroethane. Environ Sci Technol 41: 2303–2310.
Dumont MG & Murrell JC (2005) Stable isotope probing –
linking microbial identity to function. Nat Rev Microbiol 3:
499–504.
Dybas MJ, Tatara GM & Criddle CS (1995) Localization and
characterization of the carbon tetrachloride transformation
activity of Pseudomonas sp. strain KC. Appl Environ Microb 61:
758–762.
Dybas MJ, Barcelona M, Bezborodnikov S et al. (1998) Pilot-scale
evaluation of bioaugmentation for in-situ remediation of a
carbon tetrachloride-contaminated aquifer. Environ Sci
Technol 32: 3598–3611.
Dybas MJ, Hyndman DW, Heine R, Tiedje J, Linning K, Wiggert
D, Voici T, Zhao X, Dybas L & Criddle CS (2002)
Development, operation, and long-term performance of a fullscale biocurtain utilizing bioaugmentation. Environ Sci
Technol 36: 3635–3644.
FEMS Microbiol Ecol 74 (2010) 257–275
271
Eastmond DA (2008) Evaluating genotoxicity data to identify a
mode of action and its application in estimating cancer risk at
low doses: a case study involving carbon tetrachloride. Environ
Mol Mutagen 49: 132–141.
Egli C, Scholtz R, Cook AM & Leisinger T (1987) Anaerobic
dechlorination of tetrachloromethane and 1,2-dichloroethane
to degradable products by pure cultures of Desulfobacterium
sp. and Methanobacterium sp. FEMS Microbiol Lett 43:
257–261.
Egli C, Tschan T, Scholtz R, Cook AM & Leisinger T (1988)
Transformation of tetrachloromethane to dichloromethane
and carbon dioxide by Acetobacterium woodii. Appl Environ
Microb 54: 2819–2824.
Egli C, Stromeyer S, Cook AM & Leisinger T (1990)
Transformation of tetra- and trichloromethane to CO2 by
anaerobic bacteria is a non-enzymic process. FEMS Microbiol
Lett 68: 207–212.
El Fantroussi S, Naveau H & Agathos SN (1998) Anaerobic
dechlorinating bacteria. Biotechnol Progr 14: 167–188.
Engl C, Jovanovic G, Lloyd LJ, Murray H, Spitaler M, Ying L,
Errington J & Buck M (2009) In vivo localizations of
membrane stress controllers PspA and PspG in Escherichia coli.
Mol Microbiol 73: 382–396.
Environmental Protection Agency (2008) Green Remediation:
Best Management Practices for Excavation and Surface
Restoration. Environmental Protection Agency, EPA 542-F08-012, Washington, DC. Available at http://www.cluin.org/
greenremediation (accessed 10 May 2010).
Ettwig KF, Butler MK, Le Paslier D et al. (2010) Nitrite-driven
anaerobic methane oxidation by oxygenic bacteria. Nature
464: 543–548.
Eyers L, George I, Schuler L, Stenuit B, Agathos SN & El
Fantroussi S (2004) Environmental genomics: exploring the
unmined richness of microbes to degrade xenobiotics. Appl
Microbiol Biot 66: 123–130.
Field JA & Sierra-Alvarez R (2004) Biodegradability of
chlorinated solvents and related chlorinated aliphatic
compounds. Rev Environ Sci Biotechnol 3: 185–254.
Fu QS, Boonchayaanant B, Tang W, Trost BM & Criddle CS
(2009) Simple menaquinones reduce carbon tetrachloride and
iron (III). Biodegradation 20: 109–116.
Futagami T, Goto M & Furukawa K (2008) Biochemical and
genetic bases of dehalorespiration. Chem Rec 8: 1–12.
Gälli R & McCarty PL (1989) Biotransformation of 1,1,1trichloroethane, trichloromethane, and tetrachloromethane by
a Clostridium sp. Appl Environ Microb 55: 837–844.
Gantzer CJ & Wackett LP (1991) Reductive dechlorination
catalyzed by bacterial transition-metal coenzymes. Environ Sci
Technol 25: 715–722.
Garcia JL, Patel BK & Ollivier B (2000) Taxonomic, phylogenetic,
and ecological diversity of methanogenic Archaea. Anaerobe 6:
205–226.
Gerlach R, Cunningham AB & Caccavo F Jr (2000) Dissimilatory
iron-reducing bacteria can influence the reduction of carbon
tetrachloride by iron metal. Environ Sci Technol 34: 2461–2464.
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
272
Gerritse J, Drzyzga O, Kloetstra G, Keijmel M, Wiersum LP,
Hutson R, Collins MD & Gottschal JC (1999) Influence of
different electron donors and acceptors on dehalorespiration
of tetrachloroethene by Desulfitobacterium frappieri TCE1.
Appl Environ Microb 65: 5212–5221.
Gribble GW (2003) The diversity of naturally produced
organohalogens. Chemosphere 52: 289–297.
Guerrero-Barajas C & Field JA (2005) Enhancement of anaerobic
carbon tetrachloride biotransformation in methanogenic
sludge with redox active vitamins. Biodegradation 16: 215–228.
Guerrero-Barajas C & Field JA (2006) Enhanced anaerobic
biotransformation of carbon tetrachloride with precursors of
vitamin B12 biosynthesis. Biodegradation 17: 317–329.
Hartmann T, Mult S, Suter M, Rennenberg H & Herschbach C
(2000) Leaf age-dependent differences in sulphur assimilation
and allocation in poplar (Populus tremula P. alba) leaves.
J Exp Bot 51: 1077–1088.
Hashsham SA & Freedman DL (1999) Enhanced
biotransformation of carbon tetrachloride by Acetobacterium
woodii upon addition of hydroxocobalamin and fructose. Appl
Environ Microb 65: 4537–4542.
Hashsham SA, Scholze R & Freedman DL (1995) Cobalaminenhanced anaerobic biotransformation of carbon
tetrachloride. Environ Sci Technol 29: 2856–2863.
Holliger C & Schraa G (1994) Physiological meaning and
potential for application of reductive dechlorination by
anaerobic bacteria. FEMS Microbiol Rev 15: 297–305.
Holliger C & Schumacher W (1994) Reductive dehalogenation as
a respiratory process. Antonie von Leeuwenhoek 66: 239–246.
IARC (1999) Re-Evaluation of Some Organic Chemicals,
Hydrazine and Hydrogen Peroxide: Summary of Data Reported
and Evaluation, Vol. 71. International Agency for Research on
Cancer, Lyon, France. Available at http://monographs.iarc.fr/
ENG/Monographs/vol71/volume71.pdf (accessed 10 May
2010).
Isidorov VA, Zenkevich IG & Ioffe BV (1990) Volatile organic
compounds in solfataric gases. J Atmos Chem 10: 329–340.
ITRC In Situ Bioremediation Team (2002) A systematic approach
to in-situ bioremediation of carbon tetrachloride in
groundwater. Proceedings of the 2002 Conference on Application
of Waste Remediation Technologies to Agricultural
Contamination of Water Resources, Kansas City, MO.
Jiang H, Chen Y, Jiang P, Zhang C, Smith TJ, Murrell JC & Jing
XH (2010) Methanotrophs: multifunctional bacteria with
promising applications in environmental bioengineering.
Biochem Eng J 49: 277–288.
Kayser MF, Ucurum Z & Vuilleumier S (2002) Dichloromethane
metabolism and C1 utilization genes in Methylobacterium
strains. Microbiology 148: 1915–1922.
Kim S & Picardal FW (1999) Enhanced anaerobic
biotransformation of carbon tetrachloride in the presence of
reduced iron oxides. Environ Toxicol Chem 18: 2142–2150.
Knox RC & Canter LW (1996) Prioritization of ground water
contaminants and sources. Water Air Soil Poll 88: 205–226.
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
C. Penny et al.
Kobayashi H, Uematsu K, Hirayama H & Horikoshi K (2000)
Novel toluene elimination system in a toluene-tolerant
microorganism. J Bacteriol 182: 6451–6455.
Koons BW, Baeseman JL & Novak PJ (2001) Investigation of cell
exudates active in carbon tetrachloride and chloroform
degradation. Biotechnol Bioeng 74: 12–17.
Kriegman-King MR & Reinhard M (1992) Transformation of
carbon tetrachloride in the presence of sulfide, biotite, and
vermiculite. Environ Sci Technol 26: 2198–2206.
Krone UE, Laufer K, Thauer RK & Hogenkamp HP (1989)
Coenzyme F430 as a possible catalyst for the reductive
dehalogenation of chlorinated C1 hydrocarbons in
methanogenic bacteria. Biochemistry 28: 10061–10065.
Lacerda CMR & Reardon KF (2009) Environmental proteomics:
applications of proteome profiling in environmental
microbiology and biotechnology. Brief Funct Genomic
Proteomic 8: 75–87.
Leach LH & Lewis TA (2006) Identification and characterization
of Pseudomonas membrane transporters necessary for
utilization of the siderophore pyridine-2,6-bis(thiocarboxylic
acid) (PDTC). Microbiology 152: 3157–3166.
Lee CH, Lewis TA, Paszczynski A & Crawford RL (1999)
Identification of an extracellular agent of carbon tetrachloride
dehalogenation from Pseudomonas stutzeri strain KC as
pyridine-2,6-bis(thiocarboxylate). Biochem Bioph Res Co 261:
562–566.
Lewis TA & Crawford RL (1993) Physiological factors affecting
carbon tetrachloride dehalogenation by the denitrifying
bacterium Pseudomonas sp. strain KC. Appl Environ Microb 59:
1635–1641.
Lewis TA & Crawford RL (1995) Transformation of carbon
tetrachloride via sulfur and oxygen substitution by
Pseudomonas sp. strain KC. J Bacteriol 177: 2204–2208.
Lewis TA, Cortese MS, Sebat JL, Green TL, Lee CH & Crawford
RL (2000) A Pseudomonas stutzeri gene cluster encoding the
biosynthesis of the CCl4-dechlorination agent pyridine-2,
6-bis(thiocarboxylic acid). Environ Microbiol 2: 407–416.
Lewis TA, Paszczynski A, Gordon-Wylie SW, Jeedigunta S, Lee
CH & Crawford RL (2001) Carbon tetrachloride
dechlorination by the bacterial transition metal chelator
pyridine-2,6-bis(thiocarboxylic acid). Environ Sci Technol 35:
552–559.
Lewis TA, Leach L, Morales S, Austin PR, Hartwell HJ, Kaplan B,
Forker C & Meyer JM (2004) Physiological and molecular
genetic evaluation of the dechlorination agent, pyridine-2,
6-bis(monothiocarboxylic acid) (PDTC) as a secondary
siderophore of Pseudomonas. Environ Microbiol 6: 159–169.
Li XM, Zhou SG, Li FB, Wu CY, Zhuang L, Xu W & Liu L (2009)
Fe(III) oxide reduction and carbon tetrachloride
dechlorination by a newly isolated Klebsiella pneumoniae
strain L17. J Appl Microbiol 106: 130–139.
Lovley DR (2008) The microbe electric: conversion of organic
matter to electricity. Curr Opin Biotech 19: 564–571.
Lovley DR, Holmes DE & Nevin KP (2004) Dissimilatory Fe(III)
and Mn(IV) reduction. Adv Microb Physiol 49: 219–286.
FEMS Microbiol Ecol 74 (2010) 257–275
Microbial degradation of tetrachloromethane
Luijten M, Roelofsen W, Langenhoff AAM, Schraa G & Stams
AJM (2004) Hydrogen threshold concentrations in pure
cultures of halorespiring bacteria and at a site polluted with
chlorinated ethenes. Environ Microbiol 6: 646–650.
Mack EE, Vidumsky JE & Thomson MM (2001) In-situ
biodegradation of carbon tetrachloride in a stratified aquifer
system. Sixth International In Situ and On Site Bioremediation
Symposium, San Diego, CA, pp. 81–87.
Mägli A, Messmer M & Leisinger T (1998) Metabolism of
dichloromethane by the strict anaerobe Dehalobacterium
formicoaceticum. Appl Environ Microb 64: 646–650.
Maillard J, Schumacher W, Vazquez F, Regeard C, Hagen WR &
Holliger C (2003) Characterization of the corrinoid
iron–sulfur protein tetrachloroethene reductive dehalogenase
of Dehalobacter restrictus. Appl Environ Microb 69: 4628–4638.
Maithreepala RA & Doong R (2008) Effect of biogenic iron
species and copper ions on the reduction of carbon
tetrachloride under iron-reducing conditions. Chemosphere
70: 1405–1413.
Maithreepala RA & Doong R (2009) Transformation of carbon
tetrachloride by biogenic iron species in the presence of
Geobacter sulfurreducens and electron shuttles. J Hazard Mater
164: 337–344.
Martens JH, Barg H, Warren MJ & Jahn D (2002) Microbial
production of vitamin B12. Appl Microbiol Biot 58: 275–285.
Mazumder TK, Nishio N, Fukuzaki S & Nagai S (1987)
Production of extracellular vitamin B-12 compounds from
methanol by Methanosarcina barkeri. Appl Microbiol Biot 26:
511–516.
McCormick ML & Adriaens P (2004) Carbon tetrachloride
transformation on the surface of nanoscale biogenic magnetite
particles. Environ Sci Technol 38: 1045–1053.
McCormick ML, Bouwer EJ & Adriaens P (2002) Carbon
tetrachloride transformation in a model iron-reducing culture:
relative kinetics of biotic and abiotic reactions. Environ Sci
Technol 36: 403–410.
McGregor D & Lang M (1996) Carbon tetrachloride: genetic
effects and other modes of action. Mutat Res 366: 181–195.
Messmer M, Wohlfarth G & Diekert G (1993) Methyl chloride
metabolism of the strictly anaerobic, methyl chloride-utilizing
homoacetogen strain MC. Arch Microbiol 160: 383–387.
Mun CH, Ng WJ & He J (2008) Evaluation of biodegradation
potential of carbon tetrachloride and chlorophenols under
acidogenic condition. J Environ Eng 134: 177–183.
Munsch P, Geoffroy VA, Alatossava T & Meyer JM (2000)
Application of siderotyping for characterization of
Pseudomonas tolaasii and ‘Pseudomonas reactans’ isolates
associated with brown blotch disease of cultivated
mushrooms. Appl Environ Microb 66: 4834–4841.
Musarrat J & Hashsham SA (2003) Customized cDNA
microarray for expression profiling of environmentally
important genes of Pseudomonas stutzeri strain KC. Teratogen
Carcin Mut 23: 283–294.
Nicolaou SA, Gaida SM & Papoutsakis ET (2010) A comparative
view of metabolite and substrate stress and tolerance in
FEMS Microbiol Ecol 74 (2010) 257–275
273
microbial bioprocessing: from biofuels and chemicals, to
biocatalysis and bioremediation. Metab Eng 12: 307–331.
Novak PJ, Daniels L & Parkin GF (1998a) Enhanced
dechlorination of carbon tetrachloride and chloroform in the
presence of elemental iron and Methanosarcina barkeri,
Methanosarcina thermophila, or Methanosaeta concillii.
Environ Sci Technol 32: 1438–1443.
Novak PJ, Daniels L & Parkin GF (1998b) Rapid dechlorination
of carbon tetrachloride and chloroform by extracellular agents
in cultures of Methanosarcina thermophila. Environ Sci Technol
32: 3132–3136.
Pandey J, Chauhan A & Jain RK (2009) Integrative approaches for
assessing the ecological sustainability of in situ
bioremediation. FEMS Microbiol Rev 33: 324–375.
Pecher K, Haderlein SB & Schwarzenbach RP (2002) Reduction
of polyhalogenated methanes by surface-bound Fe(II) in
aqueous suspensions of iron oxides. Environ Sci Technol 36:
1734–1741.
Petrovskis EA, Vogel TM & Adriaens P (1994) Effects of electron
acceptors and donors on transformation of
tetrachloromethane by Shewanella putrefaciens MR-1. FEMS
Microbiol Lett 121: 357–363.
Pfiffner SM, Phelps TJ & Palumbo AV (2000)
Bioaugmentation potential at a carbon tetrachloride
contaminated site. 2nd International Conference on
Remediation of Chlorinated and Recalcitrant Compounds,
Monterey, CA, 2: 389–394.
Phanikumar MS, Hyndman DW, Wiggert DC, Dybas MJ, Witt
ME & Criddle CS (2002) Simulation of microbial transport
and carbon tetrachloride biodegradation in intermittently-fed
aquifer columns. Water Resour Res 38: 4-1–4-13.
Picardal FW, Arnold RG, Couch H, Little AM & Smith ME (1993)
Involvement of cytochromes in the anaerobic
biotransformation of tetrachloromethane by Shewanella
putrefaciens 200. Appl Environ Microb 59: 3763–3770.
Picardal FW, Arnold RG & Huey BB (1995) Effects of electron
donor and acceptor conditions on reductive dehalogenation of
tetrachloromethane by Shewanella putrefaciens 200. Appl
Environ Microb 61: 8–12.
Pierce E, Xie G, Barabote RD et al. (2008) The complete genome
sequence of Moorella thermoacetica (f. Clostridium
thermoaceticum). Environ Microbiol 10: 2550–2573.
Ramos JL, Duque E, Rodriguez-Herva JJ, Godoy P, Haidour A,
Reyes F & Fernandez-Barrero A (1997) Mechanisms for
solvent tolerance in bacteria. J Biol Chem 272: 3887–3890.
Rondon MR, Trzebiatowski JR & Escalante-Semerena JC (1997)
Biochemistry and molecular genetics of cobalamin
biosynthesis. Prog Nucleic Acid Re 56: 347–384.
Rouvière PE & Wolfe RS (1988) Novel biochemistry of
methanogenesis. J Biol Chem 263: 7913–7916.
Scala DJ, Hacherl EL, Cowan R, Young LY & Kosson DS (2006)
Characterization of Fe(III)-reducing enrichment cultures and
isolation of Fe(III)-reducing bacteria from the Savannah River
site, South Carolina. Res Microbiol 157: 772–783.
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
274
Scott D, McKnight DM, Blunt-Harris EL, Kolesar SE & Lovley DR
(1998) Quinone moieties act as electron acceptors in the
reduction of humic substances by humics-reducing
microorganisms. Environ Sci Technol 32: 2984–2989.
Schink B (2006) Microbially driven redox reactions in anoxic
environments: pathways, energetics, and biochemical
consequences. Eng Life Sci 6: 228–233.
Schwarzenbach RP, Escher BI, Fenner K, Hofstetter TB, Johnson
CA, von Gunten U & Wehrli B (2006) The challenge of
micropollutants in aquatic systems. Science 313: 1072–1077.
Semprini L, Hopkins GD, McCarty PL & Roberts PV (1992)
In-situ transformation of carbon tetrachloride and other
halogenated compounds resulting from biostimulation under
anoxic conditions. Environ Sci Technol 26: 2454–2461.
Shan H, Kurtz HD Jr & Freedman DL (2010) Evaluation of
strategies for anaerobic bioremediation of high concentrations
of halomethanes. Water Res 44: 1317–1328.
Sherwood JL, Petersen JN, Skeen RS & Valentine NB (1996)
Effects of nitrate and acetate availability on chloroform
production during carbon tetrachloride destruction.
Biotechnol Bioeng 51: 551–557.
Sherwood JL, Petersen JN & Skeen RS (1999) Biotransformation
of carbon tetrachloride by various acetate- and nitrate-limited
denitrifying consortia. Biotechnol Bioeng 64: 342–348.
Sikkema J, de Bont JAM & Poolman B (1995) Mechanisms of
membrane toxicity of hydrocarbons. Microbiol Rev 59:
201–222.
Simon C & Daniel R (2009) Achievements and new knowledge
unraveled by metagenomic approaches. Appl Microbiol Biot 85:
265–276.
Smith CJ & Osborn AM (2009) Advantages and limitations of
quantitative PCR (Q-PCR)-based approaches in microbial
ecology. FEMS Microbiol Ecol 67: 6–20.
Sponza DT (2001) Performance of upflow anaerobic sludge
blanket (UASB) reactor treating wastewaters containing
carbon tetrachloride. World J Microb Biot 17: 839–847.
Sponza DT (2002) Simultaneous granulation, biomass
retainment and carbon tetrachloride (CT) removal in an
upflow anaerobic sludge blanket (UASB) reactor. Process
Biochem 37: 1091–1101.
Stams AJM, de Bok FAM, Plugge CM, van Eekert MHA, Dolfing J
& Schraa G (2006) Exocellular electron transfer in anaerobic
microbial communities. Environ Microbiol 8: 371–382.
Steele HL & Streit WR (2005) Metagenomics: advances in ecology
and biotechnology. FEMS Microbiol Lett 247: 105–111.
Stromeyer SA, Stumpf K, Cook AM & Leisinger T (1992)
Anaerobic degradation of tetrachloromethane by
Acetobacterium woodii: separation of dechlorinative activities
in cell extracts and roles for vitamin B12 and other factors.
Biodegradation 3: 113–123.
Strous M, Pelletier E, Mangenot S et al. (2006) Deciphering the
evolution and metabolism of an anammox bacterium from a
community genome. Nature 440: 790–794.
Studer A, McAnulla C, Büchele R, Leisinger T & Vuilleumier S
(2002) Chloromethane induced genes define a third C1
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works
C. Penny et al.
utilization pathway in Methylobacterium chloromethanicum
CM4. J Bacteriol 184: 3476–3484.
Suresh B & Ravishankar GA (2004) Phytoremediation – a novel
and promising approach for environmental clean-up. Crit Rev
Biotechnol 24: 97–124.
Tatara GM, Dybas MJ & Criddle CS (1993) Effects of medium
and trace metals on kinetics of carbon tetrachloride
transformation by Pseudomonas sp. strain KC. Appl Environ
Microb 59: 2126–2131.
Toraya T, Honda S & Fukui S (1979) Fermentation of 1,2propanediol with 1,2-ethanediol by some genera of
Enterobacteriaceae, involving coenzyme B12-dependent diol
dehydratase. J Bacteriol 139: 39–47.
UNEP (2006) Production and Consumption of Ozone Depleting
Substances Under the Montreal Protocol: 1986–2004. United
Nations Environment Programme, Nairobi, Kenya. Available
at http://ozone.unep.org.
Van der Zee FP & Cervantes FJ (2009) Impact and application of
electron shuttles on the redox (bio)transformation of
contaminants: a review. Biotechnol Adv 27: 256–277.
Van Eekert MHA, Schröder TJ, Stams AJM, Schraa G & Field JA
(1998) Degradation and fate of carbon tetrachloride in
unadapted methanogenic granular sludge. Appl Environ
Microb 64: 2350–2356.
Vogel TM, Criddle CS & McCarty PL (1987) Transformations of
halogenated aliphatic compounds. Environ Sci Technol 21:
722–736.
Vuilleumier S, Chistoserdova L, Lee MC et al. (2009)
Methylobacterium genome sequences: a reference blueprint to
investigate microbial metabolism of C1 compounds from
natural and industrial sources. PLoS One 4: e5584.
Wang X, Gordon MP & Strand SE (2002) Mechanism of aerobic
transformation of carbon tetrachloride by poplar cells.
Biodegradation 13: 297–305.
Wang X, Dossett MP, Gordon MP & Strand SE (2004) Fate of
carbon tetrachloride during phytoremediation with poplar
under controlled field conditions. Environ Sci Technol 38:
5744–5749.
Ward M, Fu Q, Rhoads K, Yeung C, Spormann A & Criddle CS
(2004) A derivative of the menaquinone precursor 1,4dihydroxy-2-naphthoate is involved in the reductive
transformation of carbon tetrachloride by aerobically
grown Shewanella oneidensis MR-1. Appl Microbiol Biot 63:
571–577.
Watanabe K, Manefield M, Lee M & Kouzuma A (2009)
Electron shuttles in biotechnology. Curr Opin Biotech 20:
633–641.
WHO (2004) Carbon Tetrachloride in Drinking-Water. World
Health Organization, Geneva, Switzerland, WHO/SDE/WSH/
03.04/82. Available at http://www.who.int/entity/water_
sanitation_health/dwq/chemicals/carbontetrachloride.pdf
(accessed 17 May 2010).
Wilmes P, Simmons SL, Denef VJ & Banfield JF (2009) The
dynamic genetic repertoire of microbial communities. FEMS
Microbiol Rev 33: 109–132.
FEMS Microbiol Ecol 74 (2010) 257–275
Microbial degradation of tetrachloromethane
Witt ME, Dybas MJ, Worden RM & Criddle CS (1999) Motilityenhanced bioremediation of carbon tetrachloride-contaminated
aquifer sediments. Environ Sci Technol 33: 2958–2964.
Workman DJ, Woods SL, Gorby YA, Fredrickson JK & Truex MJ
(1997) Microbial reduction of vitamin B12 by Shewanella alga
strain BrY with subsequent transformation of carbon
tetrachloride. Environ Sci Technol 31: 2292–2297.
Xu JP (2006) Microbial ecology in the age of genomics and
metagenomics: concepts, tools, and recent advances. Mol Ecol
15: 1713–1731.
FEMS Microbiol Ecol 74 (2010) 257–275
275
Zhao T, Zhang L, Chen H & Zhao Y (2009) Co-inhibition of
methanogens for methane mitigation in biodegradable wastes.
J Environ Sci 21: 827–833.
Zhou J, Palumbo AV & Strong JM (1999) Phylogenetic
characterization of a mixed microbial community capable of
degrading carbon tetrachloride. Appl Biochem Biotech 80:
243–253.
Zou S, Stensel HD & Ferguson JF (2000) Carbon tetrachloride
degradation: effect of microbial growth substrate and
vitamin B12 content. Environ Sci Technol 34: 1751–1757.
c 2010 Federation of European Microbiological Societies
Journal compilation Published by Blackwell Publishing Ltd. No claim to original French government works