Applied Microbiology and Biotechnology
© Springer-Verlag 2005
10.1007/s00253-004-1864-3
Mini-Review
Biodegradation of xenobiotics by anaerobic
bacteria
Chunlong Zhang1 and George N. Bennett2
(1) Department of Environmental Sciences, University of Houston-Clear Lake, Houston,
TX 77058, USA
(2) Department of Biochemistry and Cell Biology, Rice University, 6100 Main St., Houston,
TX 77005, USA
George N. Bennett
Email: gbennett@bioc.rice.edu
Phone: +1-713-3484920
Fax: +1-713-3485154
Received: 28 September 2004 Revised: 29 November 2004 Accepted:
30 November 2004 Published online: 26 January 2005
Abstract Xenobiotic biodegradation under anaerobic conditions such as in groundwater,
sediment, landfill, sludge digesters and bioreactors has gained increasing attention over the last
two decades. This review gives a broad overview of our current understanding of and recent
advances in anaerobic biodegradation of five selected groups of xenobiotic compounds
(petroleum hydrocarbons and fuel additives, nitroaromatic compounds and explosives,
chlorinated aliphatic and aromatic compounds, pesticides, and surfactants). Significant advances
have been made toward the isolation of bacterial cultures, elucidation of biochemical
mechanisms, and laboratory and field scale applications for xenobiotic removal. For certain
highly chlorinated hydrocarbons (e.g., tetrachlorethylene), anaerobic processes cannot be easily
substituted with current aerobic processes. For petroleum hydrocarbons, although aerobic
processes are generally used, anaerobic biodegradation is significant under certain circumstances
(e.g., O2-depleted aquifers, oil spilled in marshes). For persistent compounds including
polychlorinated biphenyls, dioxins, and DDT, anaerobic processes are slow for remedial
application, but can be a significant long-term avenue for natural attenuation. In some cases, a
sequential anaerobic-aerobic strategy is needed for total destruction of xenobiotic compounds.
Several points for future research are also presented in this review.
Introduction
Anaerobic biodegradation of xenobiotic compounds has been a subject of extensive research
during the last two decades. Consequently, our current understanding of the dissipation
mechanisms of xenobiotics in natural anaerobic environments has considerably improved. Many
anaerobe-based bioreactors and remediation systems have been developed to effectively clean-up
contaminated media. The purpose of this review is to summarize recent advances in our
understanding and briefly describe biotechnological applications for the biodegradation of five
major groups of xenobiotic compounds: petroleum hydrocarbons and related fuel additives,
nitroaromatic compounds and explosives, chlorinated aliphatic and aromatic compounds,
pesticides, and surfactants.
The review is not intended to be exhaustive, but focuses on representative anaerobes, their
biochemical mechanisms, and potential biotechnological and environmental implications.
Several excellent reviews have been published on anaerobic biodegradation of xenobiotics, both
in general (Janke and Fritsche 1985; Mogensen et al. 2003a; Schink 2002) or focused on specific
compounds including petroleum hydrocarbons (Chakraborty and Coates 2004; Heider and Fuchs
1997; Prince 1993; Spormann and Widdel 2000; Widdel and Rabus 2001), explosives (Ahmad
and Hughes 2000; Esteve-Nuñez et al. 2001; Gorontzy et al. 1994; Marvin-Sikkema and de Bont
1994; Peres and Agathos 2000), chlorinated compounds (Abramowicz 1990; Bedard 2003; Chen
2004; El Fantroussi et al. 1998; Fetzner 1998; Haggblom et al. 2000; Ohtsubo et al. 2004), and
pesticides (Sethunathan 1973; Williams 1977).
There are several reasons why anaerobic biodegradation of xenobiotics is important to
researchers and practitioners. Aerobic processes require expensive O2 delivery systems,
maintenance is often high due to biofouling in subsurface remedial applications (Baker and
Herson 1994), and there are high energy costs and sludge production when bioreactors are
employed (Jewell 1987; McCarty and Smith 1986). In addition, anaerobic conditions naturally
prevail in most cases for contaminated groundwater, and some xenobiotic compounds [e.g.,
tetrachloroethylene, polychlorinated biphenyls (PCBs), and nitro-substituted aromatics] can be
efficiently transformed or mineralized only by anaerobic bacteria. In some cases, aerobic
degradation does not occur without a prior anaerobic process (Master et al. 2002).
Major groups of anaerobic organisms involved in
xenobiotic biodegradation
Like their aerobic counterparts, anaerobic bacteria able to degrade xenobiotic compounds are
diverse and present in various anaerobic habitats, including sediments, water-laden soils,
gastrointestinal contents, reticulo-ruminal contents, feedlot wastes, sludge digesters, groundwater,
and landfill sites (Williams 1977). Anaerobes use natural organics such as proteins,
carbohydrates, and many others as carbon and energy sources. Many of the so-called xenobiotic
compounds of environmental concern have naturally occurring relatives (Wackett et al. 2002).
For other xenobiotics, repeated exposure has resulted in the adaptation and evolution of
anaerobic bacteria capable of metabolizing these man-made compounds.
Table 1 lists the major groups of anaerobic microorganisms involved in biodegradation of
selected xenobiotic compounds. The pure bacterial cultures given in this table are by no means
exhaustive but are representative of each compound category. In reporting these bacteria with
compound-specific metabolic capability, two classical strategies are commonly employed. Some
researchers have chosen to employ pure cultures of previously isolated anaerobic strains to test
with specific compounds, whereas others have focused on the isolation and identification of new
strains from anaerobic bacterial consortia or enrichment cultures (El Fantroussi et al. 1998).
Without a systematic screening approach, the number of bacterial cultures successfully isolated
is limited since only a small portion of what is present in the actual microbial habitat has been
tested. In other cases, several syntrophic bacterial strains of a bacterial consortium co-exist to
metabolize a specific compound (El Fantroussi et al. 1998; Janke and Fritsche 1985; Williams
1977). Despite these limitations, the diversity of anaerobic microorganisms able to biodegrade
xenobiotic compounds is apparent.
Table 1 Major groups of anaerobic microorganisms involved in xenobiotic biodegradation. PAH
Polycyclic aromatic hydrocarbon, MTBE methyl tert-butyl ether, TNT trinitrotoluene, DNT
dinitrotoluene, RDX hexahydro-1,3,5-trinitro-1,3,5-triazine, HMX octahydro-1,3,5,7-tetranitro1,3,5,7-tetrazocine, PCE tetrachloroethylene, TCE trichloroethene, DCE cis-dichloroethene, VC
vinyl chloride, PCB polychlorinated biphenyls, PCP pentachlorophenol, LAS linear alkylbenzene
sulfonate, LAEOs linear alcohol ethyoxylates
Bacteria namea, source of isolationb, chemical
Compounds
Reference
action
Alkane
Benzene
Toluene
D. oleovorans (P): mineralizes C12–C20 n-alkane
G. spp. (P): oxidizes benzene in Fe(II)-reducing
conditions
Dechloromonas spp. (S): mineralizes benzene
into CO2 in 5 days
G. metallireducens (S): first pure culture (Fe3+
reducing) for toluene oxidation
Azoarcus and Thauera spp. (S/D): facultative
toluene-oxidizing nitrate-reducers
Aeckersberg et al. 1991
Coates et al. 2001; RooneyVarga et al. 1999
Chakraborty and Coates
2004; Lovley et al. 1989
Ethylbenzene
Thauera-related (S/P): denitrifying bacteria
completely mineralize ethylbenzene
Ball et al. 1996; Rabus and
Widdel 1995
Xylene
D. acetonicum- and Desulfosarcina variabilisrelated: mineralizes o- and m-xylene
Harms et al. 1999; Hess et al.
1997; Rabus and Widdel 1995
PAHs
Acidovorax, Bordetella, Pseudomonas,
Sphingomonas, and Variovorax (S): degradation
Eriksson et al. 2003; Rockne
complete for naphthalene and partial for 3–5 ring
et al. 2000
PAHs; P. stutzeri and Vibriop pelagius related
(S): mineralizes 7–20% naphthalene
MTBE
Pure aerobes isolated; slow under anaerobic
conditions, no pure anaerobes isolated
TNT, DNT
Veillonella alkalescens (D): the earliest evidence Esteve-Nuñez et. 2001;
of anaerobic TNT degradation C. spp. and
Hughes et al. 1999;
Finneran and Lovley 2001;
Stocking et al. 2000
Compounds
Bacteria namea, source of isolationb, chemical
action
Desulfovibrio spp. (N): most extensively studied
genera transforming TNT
Reference
McCormick et al. 1976
Desulfovibrio spp. (S): uses RDX and HMX as
sole N source
RDX, HMX
Providencia sp., and M. morganii (S): transforms
into nitroso derivatives
Boopathy et al. 1998; Kitts et
Serratia marcescens (M): RDX ring cleavage
al. 1994; Young et al. 1997;
Zhang and Hughes 2003;
similar to McCormick s pathway
Zhao et al. 2002
C. acetobutylicum (N): transforms RDX into
NHOH and NH2 derivatives
K. pneumoniae (D): degrades RDX into HCHO,
CO2 and N2O
A. woodii, C. formicoaceticum, Methanolobus
tindarius, Methanosarcina sp., Methanosarcina
mazei, Methanosarcina thermphila, Sporomusa
ovata (N): previously known strains transforming
PCE & TCE
El Fantroussi et al. 1998;
Fathepure and Boyd 1988;
Jablonski and Ferry 1992;
Terzenbach and Blaut 1994
Desulfitobacterium sp. (S): transforms PCE to
TCE and trace amount of DCE
PCE, TCE
Dehalobacter restrictus (S): transforms PCE to
ethene
Desulfitobacterium frappieri (S/D): tranforms
PCE & TCE into cis-DCE
Dehalococcoides ethenogenes (N): completes
PCE & TCE degradation into ethene
De Bruin et al. 1992; Gerritse
et al. 1996; Gerritse et al.
1999; Magnuson et al. 2000;
Maymo-Gatell et al. 1997;
Sung et al. 2003
Desulfuromonas michiganensis (S): able to grow
on free-phase PCE
VC
Dehalococcoides sp. (A): able to grow with VC
and transform VC into ethene
He et al. 2003
PCBs
Desulfitobacterium dehalogenans (S):
dehalogenates flanking Cl of OH-PCBs
Wiegel et al. 1999
PCP
Desulfitobacterium frappieri (S/D): 90–99% PCP
removal forming 3-CP
Beaudet et al. 1998; Bouchard
et al. 1996; Shelton and
Desulfitobacterium halogenans (S),
Tiedje 1984; Tartakovsky et
Desulfitobacterium chlororespirans (C),
Desulfomnile tiedje (N): dechlorinates at o- and al. 1999
m- position
Dioxins
Dehalococcoides sp. (S): uses dioxins as the sole Bunge et al. 2003
Compounds
Bacteria namea, source of isolationb, chemical
action
Reference
electron acceptor
C. sp. (N): degrades DDT as the sole C source.
Degrades other chlorinated pesticides
Chlorinated
pesticide
Aerobacter aerogenes, K. pneumoniae, N.
vulgaris (S): DDT-degrading
Ruppe et al. 2003, 2004;
Sethunathan 1973; Williams
1977
Dehalospirilum multivorans: preferentially
dechlorinates technical toxaphene
P-based
pesticide
Flavobacterium sp. (S): attacks P-insecticides
including diazino and parathion
Sethunathan 1973
Carbamate
pesticide
K. pneumoniae (D): uses three chlorinated striazines as the sole N source
Ernst and Rehm 1995;
Dinoseb
pesticide
C. bifermentans (D): utilizes Dinoseb as a sole C
Hammill and Crawford 1996
via cometabolism
Anionic
surfactant
Strain RZLAS (D): the only pure anaerobe using
LAS as the sole S source
Denger and Cook 1999
Nonionic
surfactant
Pelobacter propionicus & A. sp. (D): LAEOs
fermented to CH4 and CO2
Wagener and Schink 1988
Cationic
surfactant
Unable to isolate a single bacterium using
cationic surfactant as the sole C source
Madsen et al. 2001
aBacteria:
A Acetobacterium, C Clostridium, D Desulfobacterium, G Geobacter, K Klebsiella, M
Morganella, N Nocardia, P Pseudomonas
bSource: A Aquifer materials, C compost; D sludge; M manure; N not specified; P petroleum related sites;
S soil or sediment
cChemicals: CP chlorophenol, Dinoseb 2-sec-butyl-4,6-dinitrophenol
Pure cultures summarized in Table 1 have been isolated under strict anaerobic conditions
(sulfate-reducing and methanogenic). Example bacteria in this category include Clostridia,
Desulfobacterium, Desulfovibrio, Methanococcus, Methanosarcina, and most of the newly
isolated dehalogenating bacteria (e.g., Dehalococcoides). For practical purposes, some of the
facultative denitrifying microorganisms are also included in the table such as Flavobacterium
and Klebsiella to illustrate their potential role in these environmental communities. Anaerobic
bacteria isolated from environmental compartments and bioreactors are preferentially illustrated
over anaerobes of pathological origin.
Attention has focused on the isolation of anaerobic bacteria that play a role in the degradation of
two types of compounds due to their widespread environmental problems: the petroleum
hydrocarbons [benzene-toluene-ethylbenzene-xylene (BTEX); polycyclic aromatic hydrocarbons
(PAHs)] and chlorinated compounds including the pesticide DDT [1,1,1-trichloro-2,2-bis(pchlorophenyl)ethane]. In particular, extensive efforts have focused on the latter, partly because
halogenated organic compounds probably cause about half of the environmental problems
attributable to organic pollution in the world today (Tiedje et al. 1993), and partly because
anaerobic biodegradation is a preferred strategy. Following the discovery of the insecticidal
properties of DDT in the late 1930s, its subsequent use and the awareness of its environmental
persistence, more than 300 bacterial strains have been shown to convert DDT into DDD [1,1dichloro-2,2-bis(p-chlorophenyl)ethane] (Cookson 1995) and several novel dechlorinating strains
have been reported (Chacko et al. 1966; Guenzi and Beard 1967; Matsumura and Boush 1971;
Wedemeyer 1966) from the late 1960s to the 1970s. Research on the biodegradation of DDT
declined drastically after it was banned in the 1970s (Quensen et al. 2001) and the focus during
the last 10 years has been directed toward chlorinated aliphatic hydrocarbons due to their
worldwide prevalence. A pure culture of Dehalococcoides ethenogenes was able to completely
dechlorinate tetrachloroethylene (PCE) into innocuous ethene (Magnuson et al. 2000; MaymoGatell et al. 1997; McCarty 1997), and Desulfuromonas michiganensis can even grow on freephase PCE (Sung et al. 2003). Most PCE-dechlorinating bacteria convert PCE into
trichloroethene (TCE) or further into cis-dichloroethene (DCE) (Bagley and Gossett 1990), while
for others the more toxic vinyl chloride (VC) is produced as the end-product. Several recent
efforts have therefore been made to isolate VC-transforming bacteria. Dehalococcoides sp.,
which can grow on VC and transform it into ethene in the presence of lactate and pyruvate as
electron donors (He et al. 2003), is one such isolate.
Anaerobic degradation of the monoaromatic BTEX hydrocarbons was considered to be
negligible prior to the 1980s, partially due to the favorable energetics of aerobic bacteria
(Chakraborty and Coates 2004). These compounds have been shown to serve as carbon and
energy sources for diverse anaerobic bacteria under nitrate-reducing, Fe(III)-reducing, sulfatereducing and methanogenic conditions. Except for p-xylene, isolation of pure bacterial cultures
degrading all other BTEX compounds has been successful (Table 1). Like BTEX, 2- to 4-ring
PAHs are quite readily biodegradable aerobically (Cerniglia 1992), and anaerobic degradation of
PAHs was formerly thought impossible. However, naphthalene biodegradation through
denitrification has been documented (Eriksson et al. 2003; Mihelcic and Luthy 1988), and
phenanthrene biodegradation through similar conditions was also reported (Rockne and Strand
1998). A few PAH-degrading bacterial strains have been successfully isolated but none were
able to produce complete mineralization. As a concurrent contaminant with BTEX and PAHs in
many petroleum-contaminated sites, methyl tert-butyl ether (MTBE) is mainly susceptible to
aerobic degradation; however, anaerobic metabolism of MTBE has been reported (Finneran and
Lovley 2001; Kolhatkar et al. 2002; Somsamak et al. 2001; Stocking et al. 2000).
Anaerobic degradation of halogenated phenol, particularly pentachlorophenol (PCP), has been
the subject of several studies due to its wide use as a wood preservative. Pure cultures able to
dechlorinate PCP into 3-chlorophenol have been isolated; some bacteria preferentially remove Cl
at the ortho and meta positions (Beaudet et al. 1998; Tartakovsky et al. 1999). However, no
single bacterial culture with an ability for complete dechlorination and mineralization has yet
been isolated. For polychlorinated biphenyls (PCBs), although reductive dechlorination has been
observed frequently in many contaminated sediments and aquifers with an array of
microorganisms (Quensen et al. 1988), only recently have pure cultures been characterized (Wu
et al. 2002a, b). A strain was isolated that could dechlorinate hydroxylated PCBs (Wiegel et al.
1999). A pure culture that could use dioxin as the sole electron acceptor was isolated (Bunge et
al. 2003). The isolation of dioxin-degrading bacteria is a good example of how bacteria have
evolved to metabolize toxic xenobiotic compounds.
The biodegradation of nitroaromatic explosives [trinitrotoluene (TNT); dinitrotoluene (DNT)]
has been studied for more than two decades. Clostridium and Desulfovibrio spp. have been
extensively studied for their pathways transforming these compounds into amino- and
hydroxyamino-derivatives under anaerobic conditions. Unlike aerobic mineralization pathways
(e.g., DNT mineralization can be readily demonstrated under aerobic conditions, Zhang et al.
2000a, b), significant mineralization of TNT and DNT under anaerobic conditions has never
been achieved and anaerobic mineralizing bacteria never isolated. On the other hand, for nonaromatic explosives such as RDX (hexahydro-1,3,5-trinitro-1,3,5-triazine) and HMX (octahyrdo1,3,5,7-tetranitro-1,3,5,7-tetrazocine), pure bacterial cultures able to transform both agents have
been isolated (Boopathy et al. 1998; Kitts et al. 1994; Young et al. 1997; Zhao et al. 2002).
With a significant number of pesticides in use, dissimilar chemical structures and limited pure
bacterial isolates, generalizations regarding pesticide-degrading microorganisms are difficult to
make. For instance, in the United States alone, over 125 herbicides, 300 insecticides and 325
fungicides are registered (Cookson 1995). The most extensively studied pesticide has been DDT
due to its persistent nature in the environment. The biodegradability of many other new synthetic
pesticides are of less concern due to the shorter half-life associated with biotic and abiotic
processes. Furthermore, studies on the biodegradation of pesticides appear to be focused mostly
on aerobic bacteria, despite some limited studies on the isolation of anoxic bacterial cultures (e.g.,
Ruppe et al. 2003, 2004).
Synthetic surfactants have created environmental problems due to the use of alkyl benzene
sulfonate (ABS) detergents that were later replaced by linear alkylbenzene sulfonate (LAS) in
the late 1960s. A common misconception is that surfactants are readily removed through aerobic
processes in municipal wastewater treatment plants due to sorption and aerobic biodegradation.
This is also why biodegradability data of surfactants are predominantly aerobic (Swisher 1987).
A significant percentage of surfactants escape aerobic processes and accumulate in anaerobic
sludge digesters. A conservative estimate shows that approximately 20% of surfactants reached
the anaerobic compartment (AISE and CESIO 1999). In addition, renewed interest in surfactant
biodegradation is based on the recent finding that many alkyl phenol polyethoxylates show
toxicity to fish and are suspected of being endocrine disrupters. While the importance of
anaerobic pathways is still in debate, research efforts to isolate anaerobic surfactant degrading
bacteria (Table 1) are limited.
Biochemistry of xenobiotic biodegradation
Hydrocarbons and fuel additives
The anaerobic biochemical pathways of petroleum hydrocarbons and related fuel additives have
been the subjects of many investigations during the last two decades. For hydrocarbons, the
elucidation of anaerobic BTEX (particularly toluene) degradation pathways is probably the most
advanced (Boll et al. 2002). This is not surprising since saturated alkanes are less of a health
concern, although quantitatively they are more important than BTEX (Gieg and Suflita 2002).
Saturated alkanes are more susceptible to aerobic bacterial attack than unsaturated aliphatic
hydrocarbons (i.e., alkene, alkyne). It is also well established that alkanes with long carbon
chains and straight structures are more prone to aerobic biodegradation and the same is likely to
be the case for anaerobes. The most common aerobic pathway for alkane degradation is
oxidation of the terminal methyl group into a carboxylic acid through an alcohol intermediate,
and eventually complete mineralization through -oxidation (Cookson 1995; Leahy and Colwell
1990). Several physiologically and phylogenetically distinct anaerobes have been shown to
degrade alkanes (Aeckersberg et al. 1991; Ehrenreich et al. 2000; Rabus et al. 2001; Rueter et al.
1994). Methane can also be formed from alkanes by anaerobic organisms (Zengler et al. 1999).
Recent data with an n-hexane-utilizing denitrifying isolate pointed to a pathway involving initial
enzymatic attack by fumarate (–OOCCH=CHCOO–) addition in a manner similar to that for
toluene as discussed below (Krieger et al. 2001; Rabus et al. 2001; Wilkes et al. 2002). Another
pathway reported in a sulfate-reducing bacterium, Hxd3 (Aeckersberg et al. 1991), involves
carboxylation followed by removal of a terminal two-carbon unit to reduce the original alkane
length by one carbon as the fatty acid is formed (So et al. 2003). Observations of a carbon
addition reaction internal to the chain were also made in studies of strain SK-01 (So and Young
1999a, b).
Similarly, anaerobic MTBE metabolism is not as well understood as aerobic pathways. In the
presence of oxygen, aerobes attack MTBE with a monooxygenase. The biochemical mechanisms
of the recalcitrant ether bond cleavage have been explained in a review by Fayolle et al. (2001).
With anaerobic bacteria, the cleavage involves methyl transferases and tetrahydrofolate for the
degradation of lignin (a naturally occurring ether compound) and hydroxyl group addition during
fermentation of polyethylene glycols (-O-CH2-CH2OH). Anaerobic degradation of MTBE has
been demonstrated using compound-specific carbon isotope analyses in a groundwater site
(Kolhatkar et al. 2002), and transformation of MTBE has been observed under sulfate-reducing
conditions (Somsamak et al. 2001).
Figure 1 delineates the major enzymes and intermediates involved in anaerobic degradation of
BETX compounds. Variations in pathways exist since various bacteria depend on different
electron acceptors corresponding to differing redox conditions (Chakraborty and Coates 2004).
Complete mineralization has been reported for all BTEX compounds except p-xylene, and
research has elucidated the initial enzymatic reactions shown in Fig. 1. A difference from aerobic
mechanisms, which involve molecular oxygen, is the introduction of oxygen through H2O to
form oxygenated monoaromatic compounds that are susceptible to further ring cleavage. In some
cases, for example in the anaerobic degradation of p-cresol, oxidation of the methyl group via
addition of oxygen derived from water occurs (Bossert et al. 1989; Bossert and Young 1986).
Also shown in Fig. 1 is benzoyl coenzyme A (benzoyl-CoA), a common intermediate for BTEX
compounds. Benzoyl-CoA is formed through the addition of fumarate to the BTEX compounds
through the enzymatic action of benzylsuccinate synthase (BSS) (for toluene) or
methylbenzylsuccinate synthase (for o- and m-xylene) (Biegert et al. 1996). Studies on the
mechanism have demonstrated that these are glycyl radical enzymes (Beller and Spormann 1998;
Krieger et al. 2001; Leuthner et al. 1998). After formation of benzylsuccinate, it is converted to
the CoA derivative benzylsuccinyl-CoA by a CoA transferase and then oxidized to benzoyl-CoA
and succinyl-CoA for further metabolism (Leutwein and Heider 1999). The genes encoding the
benzyl succinate synthase have been isolated (Hermuth et al. 2002) and, in strain EbN1, are near
another operon encoding enzymes required for conversion of benzyl succinate to benzoyl-CoA
(Kube et al. 2004). The enzyme benzylsuccinyl-CoA dehydrogenase is encoded by bbsG in
Thauera aromatica (Leutwein and Heider 2002). Benzoyl-CoA is transformed to 1,5-diene-1carboxyl-CoA through the key enzyme, benzoyl-CoA reductase. After a series of hydration and
dehydrogenation steps, 3 mol acetyl-CoA and 1 mol CO2 is formed from each mole of BTEX
compound (Mogensen et al. 2003a).
Fig. 1 Anaerobic pathways for the biodegradation of petroleum hydrocarbons [benzene-tolueneethylbenzene-xylene (BTEX); adapted from Chakraborty and Coates 2004; Mogensen et al.
2003]. A Fumarate (HOOCCH=CHCOOH), E1 benzylsuccinate synthase (BSS), E2
ethylbenzylsuccinate synthase, E3 ethylbenzene dehydrogenase, E4 ethylbenzylsuccinate
synthase, E5 benzoyl-CoA reductase
The anaerobic biochemical pathways for PAHs have been studied only in the last few years, with
a focus on naphthalene and phenanthrene. Pure cultures of sulfate-reducing (Galushko et al. 1999)
and nitrate-reducing (Rockne et al. 2000) bacteria that degrade naphthalene have been isolated.
Like monoaromatic hydrocarbons, research has focused on the rate-limiting step of the initial
enzymatic attack. In contrast to earlier work that supported phenol as the major intermediate in
the fermentation of naphthalene [D. Grbic-Galic (1990) Microbial degradation of homocyclic
and heterocyclic aromatic hydrocarbons under anaerobic conditions. Unpublished report,
Department of Civil Engineering, Stanford University], recent work by several research groups
has identified 2-naphthoic acid (2-NA) as a common intermediate (Fig. 2) (Zhang et al. 2000a, b).
This acid is formed through carboxylation with the addition of a C1 unit (Zhang and Young 1997)
or fumarate, catalyzed by naphthyl-2-methyl-succinate synthase in the case of a substituted 2methylnaphthalene (Sullivan et al. 2001). The latter is analogous to the benzoyl-CoA pathway of
monoaromatic BTEX degradation. Researchers have identified several intermediates including
two ring-cleaved products (Annweiler et al. 2000, 2002; Meckenstock et al. 2000, Fig. 2).
Fig. 2 Anaerobic pathways for the biodegradation of polycyclic aromatic hydrocarbons (PAHs)
(adapted from Annweiler et al. 2000, 2002). A Fumarate (HOOCCH=CHCOOH), E1 naphthyl-2methyl-succinate synthase
Nitroaromatic compounds and explosives
The metabolic scheme in Fig. 3 illustrates major intermediates and end-products representative
of several anaerobic TNT pathways reported to date (Esteve-Nuñez et al. 2001). TNT has three
highly oxidized NO2 groups at the 2,4,6-positions. Because of their electrophilic nature, these
external NO2 groups are amenable to enzymatic reduction. In the meantime, since -electrons in
the benzene ring are shielded by four functional groups (3NO2 and 1CH3) due to steric hindrance,
the aromatic structure is very stable, preventing enzymatic attack that could lead to ring cleavage.
This unique chemical structure explains, to a large extent, why biotransformation of TNT occurs
rapidly but appreciable mineralization has never been achieved in either aerobic or anaerobic
systems even with more than two decades of intensive research effort (Hawari et al. 2000).
Fig. 3 Anaerobic pathways for the biodegradation of nitroaromatic explosives [trinitrotoluene
(TNT)] (adapted from Esteve-Nuñez et al. 2001). A Bamberger rearrangement, E1 carbon
monoxide dehydrogenase (CODH), E2 nitrite reductase, E3 the combination of enzymes
including hydrogenase, pyruvate-ferredoxin oxidoreductase, or CODH for the first step and
sulfite reductase for the final step of the reaction process (Preuss et al. 1993)
An advantage of anaerobic TNT biotransformation at low redox potential is to minimize
oxidative polymerization and the toxic azoxy compounds that can be readily formed in the
presence of oxygen. Among an array of end-products proposed or identified (Fig. 3), the amino
(NH2) and hydroxylamino (NHOH) derivatives from the reduction of NO2 groups are frequently
reported. Results have also shown the removal of NO2 groups as nitrite (NO2–), and the oxidation
of CH3 into benzoic acids (Esteve-Nuñez and Ramos 1998; Esteve-Nuñez et al. 2000). Boopathy
and Kulpa (1992) even noted the formation of NH4+ from the reductive elimination of NH2 and
proposed a pathway that included toluene as the transformation end-product. The role of
triaminotoluene (TAT), hydroxylamino intermediates, and the resulting compounds from
subsequent hydroxyl addition para to NHOH (through Bamberger rearrangement) are
incompletely known under environmental conditions but have been studied in laboratory
experiments (Hughes et al. 1998; 1999). TAT is considered to be a dead-end product that
precludes further mineralization (Hawari et al. 2000). While hydroxylamino intermediates are
not stable, their transient toxicity could be an issue in remediation systems (Tadros et al. 2000).
The good news, however, is that both compounds are strongly, or even irreversibly, adsorbed to
soils—a mechanism that may hold promise for remediation (Daun et al. 1998; Xue et al. 1995),
and the chemically unstable nature of these compounds reduces long-term toxicity risks (Padda
et al. 2000, 2003). The use of cyclodextrins for desorption of TNT-related compounds has been
studied with various soils; however, the suitability of this practice over the long term is unclear
(Sheremata and Hawari 2000).
The enzymes involved in anaerobic TNT transformation have not been fully characterized,
although several key proteins have been implicated, including ferredoxins, hydrogenases, carbon
monoxide dehydrogenase (CODH), pyruvate-ferredoxin oxidoreductases, and sulfite reductase
(Huang et al. 2000; Preuss et al. 1993). Perhaps more important to revitalize future research
efforts is the search for new microorganisms capable of TNT ring cleavage and mineralization
(Hawari et al. 2000).
Unlike nitroaromatic TNT, the nonaromatic cyclic nitroamines (RDX and HMX) have weak C–
N bonds. Initial enzymatic attack able to change N–NO2 or C–H bonds can readily destabilize
the cyclic structure and cause further molecular fragmentation. RDX is generally recalcitrant
under aerobic conditions, therefore anaerobic metabolism has been the subject of investigation.
Unfortunately, our understanding of RDX biodegradation has been limited since an early
pathway study by McCormick et al. (1981). In several recent studies on the examination of
approximately 24 hypothetical metabolites proposed in McCormick s pathways, only a few were
confirmed, several intermediates were excluded, and many other new metabolites were identified
(Adrian and Chow 2001; Hawari et al. 2000; Zhang and Hughes 2003). The full product analysis
of RDX biodegradation is particularly challenging because it involves gas-phase mineralization
products, unstable nitroso- and hydroxyamino intermediates, as well as small molecules such as
formaldehyde and methanol. At the present time, enzymatic analysis is even more speculative
despite the recent characterization of one enzyme (nitrate oxidoreductase) involved in RDX
biotransformation (Bhushan et al. 2002).
Chlorinated aliphatic and aromatic hydrocarbons
The general features of anaerobic biodegradation of chlorinated compounds has been reviewed
(Haggblom et al. 2000, 2003). The pathways for degradation of chlorinated aliphatic
hydrocarbons (CAHs) such as PCE are well established (Fig. 4). Much remains to be understood
about the biochemical mechanisms, including the enzymes and the associated genes encoding
these metabolic enzymes in bacteria with various dechlorinating activities. A strain that has
activity on PCE and a variety of diverse halogen compounds is Dehalococcoides ethenogenes
195 (Fennell et al. 2004; Maymo-Gatell et al. 1997). Related Dehalococcoides-like organisms
have been studied (Cupples et al. 2004; Maymo-Gatell et al 2001). Aerobic bacteria can grow on
the VC intermediate of PCE degradation (Coleman et al. 2002a, b). Such information is critical
so that complete PCE dechlorination can be achieved and the dechlorination rate can be
maximized by maintaining optimal conditions such as redox, electron donors (normally H2), and
competing electron acceptors (e.g., nitrate, sulfate).
Fig. 4 Anaerobic pathways for the biodegradation of chlorinated aliphatic tetrachloroethylene
(PCE) (adapted from Cookson 1995; Rittmann and McCarty 2001). E1 PCE reductive
dehydrogenase (PCE-RDase), E2 trichloroethene reductive dehydrogenase (TCE-RDase)
PCE is one of the highly chlorinated (more oxidized) CAHs with no known microorganism
capable of aerobic biodegradation. Due to its high electron negative character, PCE can be used
as an electron acceptor (the oxidant) that is susceptible to reduction into the thermodynamically
more stable VC or ethene. Reduction is accomplished either through co-metabolism (fortuitous
modifications by bacteria that use other primary substrates for carbon and energy) or a novel
biochemical mechanism known as dehalorespiration, where PCE is used as electron acceptor and
energy generated from exergonic dehalogenation reactions is used for bacterial growth (Cookson
1995; El Fantroussi et al. 1998). The electrons needed for reductive dehalogenation of PCE are
generated from the oxidation of H2 (as electron donor, Fig. 4), which originates from the
fermentation of other organic compounds (DiStefano et al. 1992). Since dechlorinating bacteria
compete with H2-utlilizing methanogens for H2, and a low H2 concentration is favored for
dechlorinating bacteria, in practice, slow-release fermentation compounds such as fatty acids and
decaying bacterial biomass are preferred (Chen 2004; Rittmann and McCarty 2001).
Several enzymes and electron carriers responsible for PCE and TCE dechlorination have been
characterized. Three of the four known PCE reductive dehalogenases (PCE-RDases)
dechlorinate PCE or TCE to cis-DCE, but the PCE-RDase from D. ethenogenes can use PCE as
sole substrate, converting it into TCE (Magnuson et al. 1998). Five chloroethene RDases have a
subunit molecular mass of 50–65 kDa and contain cobalamin and Fe-S clusters, and four
enzymes are membrane bound (Holliger et al. 1999). TCE-RDase, located on the exterior of the
cytoplasmic membrane, catalyzes the dechlorination of TCE to ethene. The gene encoding this
enzyme, tceA, was cloned and sequenced via an inverse PCR approach (Magnuson et al. 2000).
In studies on PCE respiration in D. multivorans, PCE dehalogenase was found in the cytoplasm
and was not tightly bound to the cell membrane (Neumann et al. 1996).
The ability of anaerobic consortia (Kazumi et al. 1995) and individual organisms (Song et al.
2000, 2001) to act on chlorinated or fluorinated aromatics (Vargas et al. 2000) has been reported.
Little is known about the biochemical mechanisms (particularly enzymes) of the anaerobic
biodegradation of chlorinated aromatics including PCP, PCBs, and dioxins. Various anaerobic
PCP pathways have been proposed, and an illustration of putative pathways is shown in Fig. 5. It
is likely that bacteria may take several paths simultaneously for the removal of five chlorines
leading to the formation of phenol (the rate-limiting steps) and eventually mineralization to CH4
and CO2. It is also apparent that the pathway (i.e., regiospecificity of chlorine removal) is
dominated by the redox potentials and whether the bacteria are acclimated prior to PCP
degradation. As can be seen from Fig. 5, certain bacteria preferentially remove chlorines in the
order of para > ortho > meta (Path A, Fig. 5) (Bryant et al. 1991), whereas in others an ortho >
para > meta order of chlorine removal has been reported (Path B, Fig. 5) (Mikesell and Boyd
1986). While Fig. 5 is overly simplified, a detailed description of anaerobic PCP pathways is
summarized by Nicholson et al. (1992). Preferential chlorine removal has practical ramifications
since some intermediates (e.g., 3,4,5-trichlorophenol) are more toxic than the parent compound,
while others are possible dead-end products.
Fig. 5 Anaerobic pathways for the biodegradation of chlorinated aromatic pentachlorophenol
(PCP) (adapted from Bryant et al. 1991; Mikesell and Boyd 1986). The letters o, m, p denote
dechlorination at the o, m, and p positions
PCBs and dioxins, although dissimilar in chemical structure, share some common features with
regard to their biodegradability. PCBs contain 209 different compounds (congeners) with
between 1 and 10 Cl substitutions on the backbone biphenyl structure. A typical synthetic PCB
mixture contains 60–80 different congeners. Dioxins have 1–8 Cl atoms substituted for H
atoms on dibenzo-p-dioxin, giving a total of 75 possible chlorinated derivatives, the most toxic
of which, i.e., 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), is commonly referred to as
dioxin. For both PCBs and dioxins, the less chlorinated compounds are more amenable to
aerobic biodegradation. Nevertheless, reductive dechlorination is generally faster for the more
highly chlorinated compounds. Anaerobic biodegradation of both PCBs and dioxins has been
reported (Bunge et al. 2003) and can be enhanced by acclimation of bacteria to structurally
similar, or dissimilar yet readily biodegradable, halogenated aromatic compounds, a process
called priming (Deweerd and Bedard 1999; Haggblom and Young 1990, 1995; Wu et al. 1997,
1998). Early studies by Quensen et al. (1988) indicated that PCB dechlorination occurred
primarily from the meta and para positions, yielding less toxic and more readily degraded
products. A sequential anaerobic-aerobic treatment has recently been shown to be successful in
removing PCBs from contaminated soil (Master et al. 2002). The degradation pattern of PCBs is
complex. Extensive meta and moderate ortho dechlorination were noted in a sediment slurry
study (Wu et al. 1998), but a subsequent study using a sediment-free system indicated that
bacteria specifically removed doubly flanked chlorines (i.e., chlorines bound to C that are
flanked on both sides by other Cl–C bonds) while leaving ortho chlorines intact (Wu et al. 2000).
The bacterium DF-1 dechlorinated several polychlorinated benzenes as well as PCB (Wu et al.
2002a).
Like those of PCBs, the dechlorination patterns of dioxins are difficult to generalize due to the
limited data available and the presence of a variety of dioxin congeners. Nevertheless, several
laboratory studies and field analysis of signature compounds have all indicated predominately
the initial lateral dechlorination (i.e., chlorines in the lateral 2,3,7,8 positions relative to the peri
1,4,6,9 positions), producing a characteristic 1,4-pattern of dioxin derivatives (Gaus et al. 2002;
Vargas et al. 2001). This generalization, however, contrasts with recent work by Bunge et al.
(2003) who proposed an initial peri-dechlorination pathway, demonstrating the diversity of
dechlorinating bacteria. Although no ring cleavage has been reported thus far, dechlorination is
of importance because of the reduction and even the elimination of toxicity. While the current
focus is on dechlorination of highly chlorinated aromatic compounds, including PCP, PCBs and
dioxin, there is less awareness within the research community of the fate and effects of the less
chlorinated degradation products with higher aqueous solubility and a lower octanol/water
partitioning coefficient (Mogensen et al. 2003a). In this context, dechlorination could be a
blessing in disguise if it yields compounds that are more readily bioavailable and mobile
(Dolfing and Beurskens 1995; Mogensen et al. 2003a).
Pesticides
The biochemical principles of pesticide biodegradation are no different from those of organic
compounds discussed earlier. Although a wealth of information is available, our current
understanding remains dispersed among a variety of pesticides and detailed biochemical
pathways are still unknown for many pesticides, even those in common use. Nevertheless, the
types of biochemical reactions are limited to a few (Alexander 1981). Under anaerobic
conditions, the enzymatic reactions common to many pesticides include dechlorination,
hydrolysis, nitro reduction, and dealkylation (Williams 1977). A bacterium may be partially
responsible for these metabolic activities, and in some cases the bacteria may have a metabolic
shift from one pathway to another (Barik et al. 1979). To illustrate, Fig. 6 describes the anaerobic
reactions of three structurally distinct pesticides, 2,4-dichlorophenoxyacetic acid (2,4-D),
parathion (o,o-diethyl-o-p-nitropheno phosphorothioate), and atrazine.
Fig. 6 Anaerobic pathways for the biodegradation of three selected pesticides: a 2,4dichlorophenoxyacetic acid (2,4-D), b parathion (o,o-diethyl-o-p-nitropheno phosphorothioate),
and c atrazine (adapted from Crawford et al. 1998; Mikesell and Boyd 1985; Sethunathan 1973;
Wackett et al. 2002). AtzA Atrazine chlorohydrase, AtzB hydroxylatrazine hydrolase, AtzC Nisopropylammelide amidohydrolase
Reductive dechlorination is common to all halogenated pesticides (Fig. 6a, c), including aliphatic
(fumigants), cyclic aliphatic (lindane), aromatic (DDT; PCP, Fig. 5), phenoxyalkanotes (2,4-D),
aniline-based (alachlor), and cyclodiene (aldrin) (Cookson 1995). While lightly halogenated
pesticides are more biodegradable under aerobic conditions, it is commonly believed that highly
halogenated pesticides often biodegrade more rapidly under anaerobic conditions.
Hydrolysis of phosphate esters, catalyzed by esterase, is an important mechanism for
organophosphate pesticides. For example, an esterase hydrolyzed the P–O–C linkage in
parathion subsequent to a nitro reduction, which leads to the formation of p-aminophenol
(Fig. 6b). Various other esterases catalyze degradation of aliphatic and aromatic ester pesticides
(e.g., carbamates; Sethunathan 1973). The degradation of the nitrogen-containing pesticide
atrazine shown in Fig. 6c (partially aerobic processes) involves hydrolytic dechlorination,
dealkylation, and the cleavage of C–N in the cyclic ring, yielding ultimate mineralization to CO2
and NH3. Anaerobic degradation of atrazine by mixed consortium and in wetlands by a cometabolic process has been reported (Ghosh and Philip 2004; Kao et al. 2001; Seybold et al.
2001). Although N-dealkylated intermediates were not confirmed under denitrifying conditions,
evidence of a hydroxyatrazine intermediate and ring cleavage was provided by Crawford et al.
(1998) with the bacterial isolate M91-3. Three enzymes have been characterized in Pseudomonas
sp. ADP, including atrazine chlorohydrase (AtzA), hydroxylatrazine hydrolase (AtzB), and Niso-pylammelide amidohydrolase (AtzC) (Wackett et al. 2002).
Surfactants
For each of the three major surfactants (anionic, nonionic and cationic), current understanding of
the anaerobic biochemical pathways is based on a few limited studies. Even with recent advances
in sensitive analytical instrumentation, such as high resolution GC-MS, LC-MS and tandem MS,
many of the putative pathways are based on a few tentatively identified intermediates. Other
added challenges include cumbersome derivatization procedures, effects of sorption, difficulty in
obtaining pure surfactant homologues, and the requirement for a consortium of bacteria to
completely degrade a surfactant with various moieties. In this context, any detailed discussion on
anaerobic surfactant degradation pathways must be speculative. Described briefly below are
likely bacterial strategies in attacking nonionic and anionic surfactant moieties based on several
recent studies using anaerobic microorganisms.
The nonionic linear alcohol ethyoxylates (LAE) have the common structural formula
CH3(CH2)mO(CH2CH2O)nH (m=7–17, n=1–25). Initial bacterial attack proposed by Steber and
Wierich (1987) included the central scission at the center ether bond linking the alkyl chain with
the ethoxy (EO) chain–a strategy well-known for aerobic bacteria. Wagener and Schink (1988)
suggested that the initial step is a hydroxyl group exchange reaction, followed by a shortening of
the EO chain by stepwise cleavage of acetaldehyde. Huber et al. (2000) recently concluded that
central scission is unlikely and that the first step of microbial attack is cleavage of the terminal
EO unit, releasing acetaldehyde stepwise and shortening the EO until the lipophilic moiety is
reached. Another major nonionic surfactant, nonylphenol ethoxylates (NPEO), has a benzene
ring with an EO chain para to the C9H19 functional group. Although rapid mineralization has
been reported, a recent study by Ferguson and Brownawell (2003) concluded that aromatic ring
mineralization was not a major pathway for NPEO biodegradation.
Anionic linear alkylbenzene sulfonate (LAS) is a mixture of related isomers and homologues
consisting of a para-sulfonated benzene molecule with an alkyl chain attached to any position
except the terminal one. This structural uniqueness requires the alteration of an alkyl chain, a
benzene ring, and a sulfonate linkage for complete mineralization (Mogensen et al. 2003b).
Under aerobic conditions, LAS biodegradation is initiated with an -oxidation of the terminal
methyl group of the alkyl chain to form a carboxylic acid. Further degradation proceeds by a
stepwise shortening of the alkyl chain by -oxidation, leaving a short-chain sulfophenyl
carboxylic acid. The aromatic ring hydrolyzes to form a dihydroxy-benzene structure that is
opened before desulfonation of the formed sulfonated dicarboxylic acid (Madsen et al. 2001).
Such needed information is lacking for LAS biodegradation under various anaerobic conditions.
C12–3LAS was desulfonated under sulphur-limited anoxic conditions (Denger and Cook 1999),
suggesting that LAS may not be entirely persistent. Current data on anaerobic biodegradability
does not allow an accurate survey of anaerobic biodegradation pathways of surfactants.
Practical applications of anaerobic processes in
xenobiotic biodegradation
Conventional anaerobic processes have been used for the treatment of concentrated municipal
and industrial wastewaters for over a century as they enjoy energy savings from methane and
lower sludge production than aerobic activated sludge processes (Jewell 1987; McCarty and
Smith 1986). The rapidly growing knowledge of chemical-specific bacteria and biochemical
pathways suggests that the treatment of xenobiotics, commonly at very low concentration, is
technically feasible and, in many instances, also economically viable. Evidence for xenobiotic
biodegradability under various anaerobic environments has stemmed predominately from labscale studies using serum bottles, microcosms, columns, and small-sized bioreactors. These labscale studies, along with many field and large-scale demonstrations are based mostly on
indigenous bacteria or enriched cultures. Biodegradation tests using pure bacterial cultures or
studies aimed at isolation of pure cultures (Table 1) have been almost exclusively lab-scale,
although there are a few reported uses of pure bacterial inocula in pilot and field tests (e.g.,
Dybas et al. 1998; El Fantroussi et al. 1999). A summary in Table 2 focuses on practical
applications of anaerobic bacterial consortia in field or large-scale studies and, whenever
applicable, documented sources from peer reviewed journals are selected although many studies
of this type are reported frequently in conference proceedings and industrial notices.
Table 2 Selected large and field scale anaerobic processes in xenobiotic degradationa. BTEX
Petroleum hydrocarbons (benzene-toluene-ethylbenzene-xylene), CT carbon tetrachloride, CF
chloroform, UASB upflow anaerobic sludge bioreactor, CSTR continuous stirred tank reactor,
APEO, alkylphenylethoxylate, AE alcohol ethyoxylate, 3-MCP 3-monochlorophenol, HCH
hexachlorocyclobenzene, MCB monochlorobenzene
Compounds Type and scale
Performance results
Reference
Microcosm
44% n-alkanes removed in
12 months
Salminen et al. 2004
8–10 m Aquifer
7.5 tons hydrocarbon removed in
120 days
Batterman 1983
Aquifer
NO2– injection stimulated BTEX
removal
Barker et al. 1987
Fuel spill site
NO2– enhanced m- and p-xylene
removal
Hutchins et al. 1991
PAHs
Microcosm
2–5 ring PAHs degraded under
SO42–, NO3–-reducing conditions
Rockne and Strand
1998; Rothermich et al.
2002
MTBE
Microcosm;
groundwater
Stimulated by humic substances;
Finneran and Lovley
field evidence in groundwater using 2001; Kolhatkar et al.
13
2002
C of MTBE
Full-scale reactor
TNT to mineralizable and
nonaromatic products by
Clostridium
Alkane
BTEX
TNT, DNT,
RDX, HMX
Funk et al. 1995
Providing sucrose and NH4Cl, 98%
5×1.8×2 m Sludge
TNT, DNT and RDX removal in
Lenke et al. 1998
reactor
30 weeks
PCE, TCE, CT Shallow aquifer
Nitrate and acetate injection
Semprini et al. 1992
Compounds
Type and scale
Performance results
Reference
transformed 95–97% CT to CF
Industrial site
Significant ethene and CH4 in a
TCE-contaminated aquifer
4.5-Acre chemical Strong correlation between PCE,
plant
ethene and electron donor
McCarty and Wilson
1992
Major et al. 1991
Aquifer
Nutrients, enrichment culture
injection converted TCE to ethene
Pilot-scale
Pseudomonas stutzeri KC removed
Dybas et al. 1998
CT
6-l Batch
bioreactor
11% and 23% of total Cl/biphenyl
was reduced after 13 weeks
488-m Long
stream
PCP disappeared: O2-rich > O2-poor
Pignatello et al. 1985
anaerobic > sorption
Pilot digester
>97.5% removal; 95% converted to Chen and Berthouex
3-MCP
2001
UASB, biofilm
and CSTR
95% Removal with C source
provided, dechlorinated at the othen m-, but not p- chlorines;
reactor optimization studies
Dioxins
Microcosm
Dechlorinated under methanogenic
Vargas et al. 2001
conditions
Chlorinated
pesticide
In situ anaerobic
bioscreen
HCH converted to MCB and
benzene which was mineralized in
an aerobic treatment plant
Langenhoff 2003
P-Based
pesticide
Field studies
Nonpersistent, degraded under Plimited conditions
Ternan et al. 1998
Carbamate
pesticide
500-l Fermentor;
35 yard3
(=26.8 m3)
bioreactor
Large-scale aerobic systems using
Pseudomonas sp. and recombinant
Escherichia coli; no large-scale
anaerobic processes
Newcombe and
Crowley 1999; Strong
et al. 2000
Dinoseb
pesticide
2,600-l Static
reactors
Undetectable by 15 days after
Robert et al. 1993
addition of C and acclimated culture
Anionic
surfactant
3.5-l Digester;
field data
14–25% LAS12 was transformed in
a CSTR reactor. Field data support
anaerobic biodegradation in
sediments, landfill and soils
Federle and Pastwa
1988; Haggensen et al.
2002; Mogensen et al.
2003b
Nonionic
surfactant
Batch to
microcosm
Partially (APEO) to well (AE)
biodegradable
Ferguson and
Brownawell 2003;
Huber et al. 2000;
PCBs
PCP
Ellis et al. 2000
Pagano et al. 1995
Hendriksen et al. 1992;
Ribarova et al. 2002;
Woods et al. 1989
Compounds
Type and scale
Performance results
Reference
Cationic
surfactant
Batch
Strongly adsorbed, toxic, scarce
anaerobic biodegradation data
Madsen et al. 2001
aLaboratory
microcosm studies were included for certain xenobiotic compounds
Compounds particularly suited to anoxic/anaerobic processes have included highly halogenated
compounds such as carbon tetrachloride (CT), PCE, PCBs and some of the organochlorine
pesticides that persist under aerobic conditions. Nonhalogenated compounds such as
nitroaromatic and aminoaromatic compounds, including herbicides and hazardous energetic
organonitro compounds, persist under aerobic conditions and decompose only under
anoxic/anaerobic conditions (Baker and Herson 1994). Morgan and Watkinson (1989) indicated
that the persistent nature of compounds such as DDT and PCBs is evidence of microbial
fallibility, and therefore biological cleanup of sites contaminated with this type of compound is
unlikely to be generally feasible unless an extremely long treatment period is acceptable. The
debate continues over whether persistent organic pollutants (POP) can be remediated by any
biological means. Studies demonstrated that DDE [1,1-bis(chlorophenyl)-ethylene], a toxic
byproduct of DDT, can be biodegraded into DDMU [1,1-bis(p-chlorophenyl)-2-chloro-ethylene]
under methanogenic and sulfidogenic conditions (Quensen et al. 1998). DDMU has one less Cl
atom and does not bioaccumulate as readily as its parent, and is also subject to dechlorination.
This finding, however, was discounted by others who believe that the rate was insignificant in
the field and the dechlorinating bacteria are often less favorable in competing with other bacteria
(Renner 1998). Despite much success in lab studies, in practice timely remediation of POPs such
as PCBs and DDT still relies heavily on non-biological means such as sediment dredging and
natural capping.
Chlorinated aliphatic hydrocarbons provide perhaps the most successful example of anaerobic
biodegradation in anoxic aquifer environments. Under proper conditions, deliveries of electron
donors and nutrients significantly stimulated the activities of reductive dechlorination in many
field studies (Major et al. 1991; McCarty and Wilson 1992; Semprini et al. 1992). Field success,
however, often entails expensive monitoring of the contaminant plume and the end-products
including methane and ethene. In several cases where indigenous bacteria were unable to
dechlorinate, bioaugmentation with pure dechlorinating bacteria has been shown to be successful
(Dybas et al. 1998; Ellis et al. 2000).
Field experience in remediating hydrocarbons using aerobic bacteria and pathways dates back to
the early 1970s. For instance, Raymond (1974) received a patent on a process designed to
remove hydrocarbon contaminants from groundwater by stimulating indigenous aerobic bacteria
with nutrients and oxygen. Anaerobic processes, however, have received little attention and have
had limited success in the field even with monoaromatic hydrocarbons (BTEX). Recently, field
data have suggested anaerobic biodegradation could be a significant process in contaminated
aquifers depleted of oxygen (Table 2). Field evidence regarding the exclusive role of anaerobes
are sometimes equivocal since groundwater normally considered to be anoxic can sometimes
contain dissolved oxygen (DO) as high as 1 mg/l (Batterman 1983; Hutchins et al. 1991;
Steinbach et al. 2004). Future research and field demonstrations with hydrocarbons, both in
terrestrial and marine environments, are likely to increase. For the terrestrial environment,
research is motivated largely by the clean-up of gasoline spills in leaking underground storage
tanks and the increased recognition of natural attenuation as part of the remedial strategy. For the
marine environment, work is largely driven by oil spills, particularly of crude oil. Prince (1993)
stated that there is room to extend current applications to oiled marshes and other anaerobic
sediments as these are the frequent recipients of spill incidents. Thirty percent of gasoline sold in
the United States contains 11% by volume MTBE and crude oils are composed of more than
75% aliphatic and aromatic hydrocarbons (Stocking et al. 2000). Anaerobic MTBE
biodegradation is still considered to be a rare occurrence, therefore remedial applications for
MTBE and other fuel oxygenates are almost exclusively aerobic processes (Fayolle et al. 2001;
Stocking et al. 2000).
Anaerobic processes for the degradation of explosive compounds have been employed in both in
situ and ex situ reactor systems (Funk et al. 1995; Lenke et al. 1998). Processes such as land
farming, composting and slurry reactors have been very successful in transforming or
detoxifying explosives and, in some cases, result in complete mineralization. Since
mineralization of explosives is very unlikely in anaerobic processes, remediation is often
achieved by two strategies, i.e., transformation into innocuous products or irreversible binding
with soil components. Recently, increasing evidence has pointed toward the use of sequential
anaerobic-aerobic processes to destroy nitroaromatic explosives (Esteve-Nuñez et al. 2001;
Hawari et al. 2000).
The anaerobic biodegradation of pesticides and surfactants has witnessed limited in situ and ex
situ applications relative to their extensive usage and disposal. Most pesticide biodegradation
studies stem from the need to minimize dispersion outside of the agriculture environment, and
remedial applications are limited to some contaminated pesticide manufacturing sites and
accidental spills as shown in Table 2 (Langenhoff 2003; Newcombe and Crowley 1999; Roberts
et al. 1993; Strong et al. 2000; Ternan et al. 1998). There is a paucity of data regarding the
anaerobic biodegradation of surfactants, and surfactants commonly in use are considered as not
persistent in the environment as implied from the extensive aerobic biodegradation database
currently available. Surfactants are in fact the most abundant organic species in domestic sewage
sludge, where concentrations exceeding g/kg levels are frequently observed (Mogensen et al.
2003a). Field monitoring data support evidence of anaerobic biodegradation in sediment below
sewage treatment plant (STP) outfalls, domestic septic systems, landfill sites receiving sludge,
and subsurface soils beneath laundromat wastewater discharge (Federle and Pastwa 1988). One
area of needed research is the anaerobic biodegradation in sludge digesters of municipal STPs.
Such anaerobic digesters are generally not designed for the removal of surfactants, hence
improved designs and optimization of various anaerobic reactor systems has been the subject of
several studies (Haggensen et al. 2002; Mogensen et al. 2003a). Further research is needed with
regard to surfactants of current environmental concern, particularly LAS and NPEO (Ferguson
and Brownawell 2003; Huber et al. 2000).
Conclusions and future prospects
The mounting evidence accumulated during the last two decades supports the argument that
anaerobic biodegradation, once considered to be negligible, could be significant for a variety of
xenobiotic compounds in anaerobic environments such as groundwater, sediment, landfill,
sludge digesters and bioreactor systems. The elucidation of biochemical mechanisms using
isolated bacteria strains, and laboratory feasibility studies using mainly enrichment cultures has
enabled successful large- and field-scale in situ and ex situ remediation applications (Tables 1, 2).
For certain highly chlorinated hydrocarbons (e.g., PCE), anaerobic processes cannot easily be
substituted with current aerobic processes. For petroleum hydrocarbons, although aerobic
processes are generally used, anaerobic biodegradation could become significant, and an
economically viable option under certain circumstances (e.g., oxygen-depleted aquifer, oilspilled marsh). For persistent compounds including PCBs, dioxins, and DDT, anaerobic
processes are slow for remedial applications, but can represent a significant avenue if natural
attenuation is an option. For many xenobiotic compounds, particularly PCBs and explosives,
anaerobic processes could be complementary to aerobic processes for complete contaminant
destruction.
With the increasing appreciation of anaerobic processes, along with recent advances in
biochemical, molecular technology and analytical instrumentation, new strains will continue to
be isolated and novel enzymes and biochemical pathways will be characterized. Further research
will be needed to characterize genes encoding the enzymes that bacteria have evolved to degrade
such xenobiotics. Recombinant strains, although still a debated issue in practice, have been
explored in the case of aerobic microorganisms and show some success in outdoing the
performance of indigenous bacteria (Shimazu et al. 2001; Wackett et al. 2002). Genetically
engineered microorganisms capable of multiple pathways are likely to offer solutions to some of
the most recalcitrant xenobiotic compounds, most likely at contained wastestreams associated
with industrial facilities. An ignored area of research is the characterization of enrichment
cultures. This is particularly important for recalcitrant compounds that require a consortium of
syntrophic bacteria. Elucidating the ecology of these bacterial consortia is critical, but such
information is almost nonexistent. A related approach involving the sequential use of anaerobic
and aerobic bacteria (Esteve-Nuñez et al. 2001; Lenke et al. 1998; Master et al. 2002) may also
allow advances in treatment to be attained.
Other knowledge gaps include the understanding and manipulation of bacterial strategies in
utilizing compounds with various functional moieties. Not only the initial enzymatic attack but
also the complete mineralization potential needs to be characterized. Not discussed in this review
are the optimization of anaerobic processes and the provision of optimal electron donors and
acceptors.
Acknowledgements Research in authors laboratories has been supported by the Welch
Foundation (C-1268) and BC-0022, DSWA, EIH and SERDP. This material is also based on
work supported in part by the United States Army Research Laboratory and the United States
Army Research Office (Grant DOD Army W911NF-04-1-0179)
References
Abramowicz DA (1990) Aerobic and anaerobic biodegradation of PCBs: a review. Crit Rev
Biotechnol 10:241–251
Adrian NR, Chow T (2001) Identification of hydroxylamino-dinitroso-1,3,5-triazine as a
transient intermediate formed during the anaerobic biodegradation of hexahydro-1,3,5-trinitro1,3,5-triazine. Environ Toxicol Chem 20:1874–1877
Aeckersberg F, Bak F, Widdel F (1991) Anaerobic oxidation of saturated hydrocarbons to CO2
by a new type of sulfate-reducing bacterium. Arch Microbiol 156:5–14
Ahmad F, Hughes JB (2000) Anaerobic transformation of TNT by Clostridium. In: Spain JC,
Hughes JB, Knackmuss H-J (eds) Biodegradation of nitroaromatic compounds and explosives,
Lewis, Boca Raton, Fla., pp 185–212
AISE and CESIO (1999) Environmental relevance of anaerobic biodegradability of surfactants.
http://www.aise-net.org/PDF/anaerobicBiopub1.pdf, p 6
Alexander M (1981) Biodegradation of chemicals of environmental concern. Science 211:132–
138
Annweiler E, Materna A, Safinowski S, Kappler A, Richnow HH, Michaelis W, Meckensrock
RU (2000) Anaerobic degradation of 2-methylnaphthalene by a sulfate reducing enrichment
culture. Appl Environ Microbiol 66:5329–5333
Annweiler E, Michaelis W, Meckenstock RU (2002) Identical ring cleavage products during
anaerobic degradation of naphthalene, 2-methylnaphthalene, and tetralin indicate a new
metabolic pathway. Appl Environ Microbiol 68:852–858
Bagley DM, Gossett JM (1990) Tetrachloroethene transformation to trichloroethene and cis-1,2dichloroethene by sulfate-reducing enrichment cultures. Appl Environ Microbiol 56:2511–2516
Baker KH, Herson DS (1994) Bioremediation. McGraw Hill, New York, NY
Ball HA, Johnson HA, Reinhard M, Spormann AM (1996) Initial reactions in anaerobic
ethylbenzene oxidation by a denitrifying bacterium, strain EB1. J Bacteriol 178:5755–5761
Barik S, Wahid PA, Ramakrishna C, Sethunathan N (1979) A change in the degradation pathway
of parathion after repeated applications to flooded soil. J Agric Food Chem 27:1391–1392
Barker JF, Patrick GC, Major DW (1987) Natural attenuation of aromatic hydrocarbons in a
shallow sand aquifer. Ground Water Monit Rev 7:64–71
Batterman G (1983) A large scale experiment on in situ biodegradation of hydrocarbon in the
subsurface. In: Ground water in water resources planning, vol II. Proc Int Symp. IASA
Publication 142. International Association of Hydrological Sciences, London, p 93
Beaudet R, Levesque MJ, Villemur R, Lanthier M, Chenier M, Lepine F, Bisaillon JG (1998)
Anaerobic biodegradation of pentachlorophenol in a contaminated soil inoculated with a
methanogenic consortium or with Desulfitobacterium frappieri strain PCP-1. Appl Microbiol
Biotechnol 50:135–141
Bedard DL (2003) Polychlorinated biphenyls in aquatic sediments: environmental fate and
outlook for biological treatment. In: Bossert ID, Haggblom MM (eds) Dehalogenation. Kluwer,
Norwell, Mass., pp 443–465
Beller HR, Spormann AM (1998) Analysis of the novel benzylsuccinate synthase reaction for
anaerobic toluene activation based on structural studies of the product. J Bacteriol 180:5454–
5457
Bhushan B, Halasz A, Spain J, Thiboutot S, Ampleman G, Hawari J (2002) Biotransformation of
hexahydro-1,3,5-trinitro-1,3,5-triazine catalyzed by a NAD(P)H: nitrate oxidoreductase from
Aspergillus niger. Environ Sci Technol 36:3104–3108
Biegert T, Fuchs G, Heider J (1996) Evidence that anaerobic oxidation of toluene in the
denitrifying bacterium Thauera aromatica is initiated by formation of benzylsuccinate from
toluene and fumarate. Eur J Biochem 238:661–668
Boll M, Fuchs G, Heider J (2002) Anaerobic oxidation of aromatic compounds and
hydrocarbons. Curr Opin Chem Biol 6:604–611
Boopathy R, Gurgas M, Ullian J, Manning JF (1998) Metabolism of explosive compounds by
sulfate-reducing bacteria. Curr Microbiol 37:127–131
Boopathy R, Kulpa CF (1992) Trinitrotoluene as a sole nitrogen source for a sulfate-reducing
bacterium Desulfovibrio sp. (B strain) isolated from an anaerobic digester. Curr Microbiol
25:235–241
Bossert ID, Young LY (1986) Anaerobic oxidation of p-cresol by a denitrifying bacterium. Appl
Environ Microbiol 52:1117–1122
Bossert ID, Whited G, Gibson DT, Young LY (1989) Anaerobic oxidation of p-cresol mediated
by a partially purified methylhydroxylase denitrifying bacterium. J Bacteriol 171:2956–2962
Bouchard B, Beaudet R, Villemur R, McSween G, Lepine F, Bisaillon JG (1996) Isolation and
characterization of Desulfitobacterium frappieri sp. nov., an anaerobic bacterium which
reductively dechlorinates pentachlorophenol to 3-chlorophenol. Int J Syst Bacteriol 46:1010–
1015
Bryant FO, Hale DD, Rogers JE (1991) Regiospecific dechlorination of pentachlorophenol by
dichlorophenol-adapted microorganisms in freshwater, anaerobic sediment slurries. Appl
Environ Microbiol 57:2293–2301
Bunge M, Adrian L, Kraus A, Lorenz WG, Andreesen JR, Gorisch H, Lechner U (2003)
Reductive dehalogenation of chlorinated dioxins by the anaerobic bacterium Dehalococcoides
ethenogenes sp. strain CBDB1. Nature 421:357–360
Cerniglia CE (1992) Biodegradation of polycyclic aromatic hydrocarbons. Biodegradation
3:351–368
Chacko CI, Lockwood JL, Zabik M (1966) Chlorinated hydrocarbon pesticides: degradation by
microbes. Science 154:893–895
Chakraborty R, Coates JD (2004) Anaerobic degradation of monoaromatic hydrocarbons. Appl
Microbiol Biotechnol 64:437–446
Chen G (2004) Reductive dehalogenation of tetrachloroethylene by microorganisms: current
knowledge and application strategies. Appl Microbiol Biotechnol 63:373–377
Chen ST, Berthouex PM (2001) Treating an aged pentachlorophenol- (PCP-) contaminated soil
through three sludge handling processes, anaerobic sludge digestion, post-sludge digestion and
sludge land application. Water Sci Technol 44:149–56
Coates JD, Chakraborty R, Lack JG, O Connor SM, Cole KA, Bender KS, Achenbach LA
(2001) Anaerobic benzene oxidation coupled to nitrate reduction in pure culture by two strains of
Dechloromonas. Nature 411:1039–1043
Coleman NV, Mattes TE, Gossett JM, Spain JC (2002a) Biodegradation of cis-dichloroethene as
the sole carbon source by a beta-proteobacterium. Appl Environ Microbiol 68:2726–2730
Coleman NV, Mattes TE, Gossett JM, Spain JC (2002b) Phylogenetic and kinetic diversity of
aerobic vinyl chloride-assimilating bacteria from contaminated sites. Appl Environ Microbiol
68:6162–6171
Cookson JT Jr (1995) Bioremediation engineering: design and application. McGraw-Hill, New
York, NY
Crawford JJ, Sims GK, Mulvaney RL, Radosevich M (1998) Biodegradation of atrazine under
denitrifying conditions. Appl Microbiol Biotechnol 49:618–623
Cupples AM, Spormann AM, McCarty PL (2004) Comparative evaluation of chloroethene
dechlorination to ethene by Dehalococcoides-like microorganisms. Environ Sci Technol
38:4768–4774
Daun G, Lenke H, Reuss M, Knackmuss H-J (1998) Biological treatment of TNT-contaminated
soil. 1. Anaerobic cometabolic reduction and interaction of TNT and metabolites with soil
components. Environ Sci Technol 32:1956–1963
De Bruin WP, Koterman MJJ, Posthumus MA, Schraa G, Zehnder AJB (1992) Complete
biological reductive transformation of tetrachloroethene to ethane. Appl Environ Microbiol
58:1996–2000
Denger K, Cook AM (1999) Linear alkylbenzenesulphonate (LAS) bioavailable to anaerobic
bacteria as a source of sulphur. J Appl Microbiol 86:165–168
Deweerd KA, Bedard DL (1999) Use of halogenated benzoates and other halogenated aromatic
compounds to stimulate the microbial dechlorination of PCBs. Environ Sci Technol 33:2057–
2063
DiStefano TD, Gossett JM, Zinder SH (1992) Hydrogen as an electron donor for dechlorination
of tetrachloroethene by an anaerobic mixed culture. Appl Environ Microbiol 58:3622–3629
Dolfing J, Beurskens JEM (1995) The microbial logic and environmental significance of
reductive dehalogenation. Adv Microbial Ecol 14:143–206
Dybas MJ, Barcelona M, Bezborodnikov S, Davies S, Forney L, Heuer H, Kawka O, Mayotte T,
Sepulveda-Torres L, Smalla K, Sneathen M, Tiedje J, Voice T, Wiggert DC, Witt ME, Criddle
CS (1998) Pilot-scale evaluation of bioaugmentation for in-situ remediation of a carbon
tetrachloride-contaminated aquifer. Environ Sci Technol 32:3598–3611
Ehrenreich P, Behrends A, Harder J, Widdel F (2000) Anaerobic oxidation of alkanes by newly
isolated denitrifying bacteria. Arch Microbiol 173:58–64
El Fantroussi S, Naveau H, Agathos SN (1998) Anaerobic dechlorinating bacteria. Biotechnol
Prog 14:167–188
ElFantroussi S, Belkacemi M, Top EM, Mahillon J, Vaveau H, Agathos SN (1999)
Bioaugmentation of a soil bioreactor designed for pilot-scale anaerobic bioremediation studies.
Environ Sci Technol 33:2992–3001
Ellis DE, Lutz EJ, Odom JM, Buchanan RJ Jr, Bartlett CL, Lee MD, Harkness MR, Deweerd
KA (2000) Bioaugmentation for accelerated in situ anaerobic bioremediation. Environ Sci
Technol 34:2254–2260
Eriksson M, Sodersten E, Yu Z, Dalhammar G, Mohn WW (2003) Degradation of polycyclic
aromatic hydrocarbons at low temperature under aerobic and nitrate-reducing conditions in
enrichment cultures from northern soils. Appl Environ Microbiol 69:275–84
Ernst C, Rehm HJ (1995) Utilization of chlorinated s-triazines by a new strain of Klebsiella
pneumoniae. Appl Microbiol Biotechnol 42:763–768
Esteve-Nuñez A, Ramos JL (1998) Metabolism of 2,4,6-trinitrotoluene (TNT) by Pseudomonas
sp. JLR11. Environ Sci Technol 32:3802–3808
Esteve-Nuñez A, Luchessi, Phillipps B, Schink B, Ramos JL (2000) Respiration of 2,4,6trinitrotoluene by Pseudomonas sp. strain JLR11. J Bacteriol 182:1352–1355
Esteve-Nuñez A, Caballero A, Ramos JL (2001) Biological degradation of 2,4,6-trinitrotoluene.
Microbiol Mol Biol Rev 65:335–352
Fathepure BZ, Boyd SA (1988) Dependence of tetrachloroethylene dechlorination on
methanogenic substrate consumption by Methanosarcina sp. strain DCM. Appl Environ
Microbiol 54:2976–2980
Fayolle F, Vandecasteele J-P, Monot F (2001) Microbial degradation and fate in the environment
of methyl tert-butyl ether and related fuel oxygenates. Appl Microbiol Biotechnol 56:339–349
Federle TW, Pastwa GM (1988) Biodegradation of surfactants in saturated subsurface sediments:
a field study. Ground Water 26:761–770
Fennell DE, Nijenhuis I, Wilson SF, Zinder SH, Haggblom MM (2004) Dehalococcoides
ethenogenes strain 195 reductively dechlorinates diverse chlorinated aromatic pollutants.
Environ Sci Technol 38:2075–2081
Ferguson PL, Brownawell BJ (2003) Degradation of nonylphenol ethoxylates in estuarine
sediment under aerobic and anaerobic conditions. Environ Toxicol Chem 22:1189–1199
Fetzner S (1998) Bacterial dehalogenation. Appl Microbiol Biotechnol 50:633–657
Finneran KT, Lovley DR (2001) Anaerobic degradation of methyl tert-butyl ether (MTBE) and
tert-butyl alcohol (TBA). Environ Sci Technol 35:1785–1790
Funk SB, Crawford DL, Crawford RL, Mead G, Davis-Hooker W (1995) Full-scale anaerobic
bioremediation of trinitrotoluene contaminated soils. Appl Biochem Biotechnol 51:625–633
Galushko A, Minz D, Schink B, Widdel F (1999) Anaerobic degradation of naphthalene by a
pure culture of a novel type of marine sulphate-reducing bacterium. Environ Microbiol 1:415–
420
Gaus C, Brunskill GJ, Connell DW, Prange J, Muller JF, Papke O, Webber R (2002)
Transformation processes, pathways, and possible sources of distinctive polychlorinated
dibenzo-p-dioxin signatures in sink environments. Environ Sci Technol 36:3452–3549
Gerritse J, Renard V, Gomes TMP, Lawson PA, Collins MD, Gottschal J (1996)
Desulfitobacterium sp. strain, an anaerobic bacterium that can grow by reductive dechlorination
of tetrachloroethene or ortho-chlorinated phenols. Arch Microbiol 165:132–140
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 Microbiol 65:5212–5221
Ghosh PK, Philip L (2004) Atrazine degradation in anaerobic environment by a mixed microbial
consortium. Water Res 38:2276–2283
Gieg LM, Suflita JM (2002) Detection of anaerobic metabolites of saturated and aromatic
hydrocarbons in petroleum-contaminated aquifers. Environ Sci Technol 36:3755–3762
Gorontzy T, Drzyzga O, Kahl MW, Bruns-Nagel D, Breitung J, von Loew E, Blotevogel KH
(1994) Microbial degradation of explosives and related compounds. Crit Rev Microbiol 20:265–
284
Guenzi WD, Beard WE (1967) Anaerobic biodegradation of DDT to DDD in soil. Science
156:1116–1117
Haggblom MM, Young LY (1990) Chlorophenol degradation coupled to sulfate reduction. Appl
Environ Microbiol 56:3255–3260
Haggblom MM, Young LY (1995) Anaerobic degradation of halogenated phenols by sulfatereducing consortia. Appl Environ Microbiol 61:1546–1550
Haggblom MM, Knight VK, Kerkhof LJ (2000) Anaerobic decomposition of halogenated
aromatic compounds. Environ Pollut 107:199–207
Haggblom MM, Ahn YB, Fennell DE, Kerkhof LJ, Rhee SK (2003) Anaerobic dehalogenation
of organohalide contaminants in the marine environment. Adv Appl Microbiol 53:61–84
Haggensen F, Mogensen AS, Angelidaki I, Ahring BK (2002) Anaerobic treatment of sludge:
focusing on reduction of LAS concentration in sludge. Water Sci Technol 46:159–165
Hammill TB, Crawford RL (1996) Degradation of 2-sec-butyl-4,6-dinitrophenol (Dinoseb) by
Clostridium bifermentans KMR-1. Appl Environ Microbiol 62:1842–1846
Harms G, Zengler K, Rabus R, Aeckersberg F, Minz D, Rosselló-Mora R, Widdel F (1999)
Anaerobic oxidation of o-xylene, m-xylene, and homologous alkylbenzenes by new types of
sulfate-reducing bacteria. Appl Environ Microbiol 65:999–1004
Hawari J, Beaudet S, Halasz A, Thiboutots S, Ampleman G (2000) Microbial degradation of
explosives: biotransformation versus mineralization. Appl Microbiol Biotechnol 54:605–618
He J, Ritalahti KM, Yang K-L, Koenigsberg SS, Löffler FE (2003) Detoxification of vinyl
chloride to ethene coupled to an anaerobic bacterium. Nature 424:62–65
Heider J, Fuchs G (1997) Anaerobic metabolism of aromatic compounds. Eur J Biochem
243:577–96
Hendriksen HV, Larsen S, Ahring BK (1992) Influence of a supplemental carbon source on
anaerobic dechlorination of pentachlorophenol in granular sludge. Appl Environ Microbiol
58:365–370
Hermuth K, Leuthner B, Heider J (2002) Operon structure and expression of the genes for
benzylsuccinate synthase in Thauera aromatica strain K172. Arch Microbiol 177:132–138
Hess A, Zarda B, Hahn D, Haner A, Stax D, Hohener P, Zeyer J (1997) In situ analysis of
denitrifying toluene- and m-xylene degrading bacteria in a diesel fuel-contaminated laboratory
aquifer column. Appl Environ Microbiol 65:2136–2141
Holliger C, Wohlfarth G, Diekert G (1999) Reductive dechlorination in the energy metabolism
of anaerobic bacteria. FEM Microbiol Rev 22:383–398
Huang S, Lindahl PA, Wang C, Bennett GN, Rudolph FB, Hughes JB (2000) 2,4,6Trinitrotoluene reduction by carbon monoxide dehydrogenase from Clostridium thermoaceticum.
Appl Environ Microbiol 66:1474–1478
Huber M, Meyer U, Rys P (2000) Biodegradation mechanisms of linear alcohol ethoxylates
under anaerobic conditions. Environ Sci Technol 34:1737–1741
Hutchins SR, Down WC, Wilson JT, Smith GB, Kovacs DA, Fine DD, Douglass RH, Hendrix
DJ (1991) Effect of nitrate addition on bioremediation of fuel-contaminated aquifer: field
demonstration. Ground Water 29:571–580
Hughes JB, Wang C, Yesland K, Richardson A, Bhadra R, Bennett G, Rudolph F (1998)
Bamberger rearrangement during TNT metabolism by Clostridium acetobutylicum. Environ Sci
Technol 32:494–500
Hughes JB, Wang CY, Zhang C (1999) Anaerobic Biotransformation of 2,4-dinitrotoluene and
2,6-dinitrotoluene by Clostridium acetobutylicum: a pathway through dihydroxylamino
intermediates. Environ Sci Technol 33:1065–1070
Jablonski PE, Ferry JG (1992) Reductive dechlorination of trichloroethylene by the CO-reduced
CO dehydrogenase enzyme complex from Methanosarcina thermophila. FEMS Microbiol Lett
96:55–60
Janke D, Fritsche W (1985) Nature and significance of microbial cometabolism of xenobiotics. J
Basic Microbiol 25:603–619
Jewell WJ (1987) Anaerobic sewage treatment, part 6. Environ Sci Technol 21:14–21
Kao CM, Wang JY, Wu MJ (2001) Evaluation of atrazine removal processes in a wetland. Water
Sci Technol 44:539–544
Kazumi J, Haggblom MM, Young LY (1995) Diversity of anaerobic microbial processes in
chlorobenzoate degradation: nitrate, iron, sulfate and carbonate as electron acceptors. Appl
Microbiol Biotechnol 43:929–936
Kitts CL, Cunningham DP, Unkefer PJ (1994) Isolation of three hexahydro-1,3,5-trinitro-1,3,5triazine degrading species of the family Enterobacteriaceae from nitramine explosivecontaminated soil. Appl Environ Microbiol 60:4608–4711
Kolhatkar R, Kuder T, Philp P, Allen J, Wilson JT (2002) Use of compound-specific stable
carbon isotope analyses to demonstrate anaerobic biodegradation of MTBE in groundwater at a
gasoline release site. Environ Sci Technol 36:5139–5146
Krieger J, Roseboom W, Albracht SP, Spormann AM (2001) A stable organic free radical in
anaerobic benzylsuccinate synthase of Azoarcus sp. strain T. J Biol Chem 276:12924–12927
Kube M, Heider J, Amann J, Hufnagel P, Kuhner S, Beck A, Reinhardt R, Rabus R (2004)
Genes involved in the anaerobic degradation of toluene in a denitrifying bacterium, strain EbN1.
Arch Microbiol 181:182–194
Langenhoff A (2003) In-situ bioremediation of pesticides.
(http://www.mep.tno.nl/Informatiebladen_eng/304e.pdf)
Leahy JG, Colwell RR (1990) Microbial degradation of hydrocarbons in the environment.
Microbiol Rev 54:305–315
Lenke H, Warrelmann J, Daun G, Hund K, Sieglen U, Knackmuss H-J (1998) Biological
treatment of TNT-contaminated soil. 2. Biologically induced immobilization of the contaminants
and full-scale application. Environ Sci Technol 32:1964–1971
Leutwein C, Heider J (1999) Anaerobic toluene-catabolic pathway in denitrifying Thauera
aromatica: activation and beta-oxidation of the first intermediate, (R)-(+)-benzylsuccinate.
Microbiology 145:3265–3271
Leutwein C, Heider J (2002) (R)-Benzylsuccinyl-CoA dehydrogenase of Thauera aromatica, an
enzyme of the anaerobic toluene catabolic pathway. Arch Microbiol 178:517–524
Leuthner B, Leutwein C, Schulz H, Horth P, Haehnel W, Schiltz E, Schagger H, Heider J (1998)
Biochemical and genetic characterization of benzylsuccinate synthase from Thauera aromatica:
a new glycyl radical enzyme catalysing the first step in anaerobic toluene metabolism. Mol
Microbiol 28:615–628
Lovley DR, Baedecker MJ, Lonergan DJ, Cozzarelli IM, Phillips EJP, Siegel DI (1989)
Oxidation of aromatic contaminants coupled to microbial iron reduction. Nature 339:297–299
Madsen T, Boyd HB, Nylén D, Pedersen AR, Petersen GI, Simonsen F (2001). Environmental
and health assessment of substances in household detergents and cosmetic detergent products.
Environmental Project No. 615. http://www.mst.dk/udgiv/publications/2001
Magnuson JK, Stern RV, Gossett JM, Zinder SH, Burris DR (1998) Reductive dechlorination of
tetrachloroethene to ethene by a two-component enzyme pathway. Appl Environ Microbiol
64:1270–1275
Magnuson JK, Romine MF, Burris DR, Kingsley MT (2000) Trichloroethene reductive
dehalogenase from Dehalococcoides ethenogenes: Sequence of tceA and substrate range
characterization. Appl Environ Microbiol 66:5141–5147
Major DW, Hodgins WW, Butler BJ (1991) Field and laboratory evidence of in situ
biotransformation of tetrachloroethene to ethene and ethane at a chemical transfer facility in
North Toronto. In: Hinchee RE, Olfenbuttel (eds) On site bioremediation: processes for
xenobiotic and hydrocarbon treatment. Butterworth-Heinemann, Stoneham, Mass. pp 141–171
Marvin-Sikkema FD, de Bont JA (1994) Degradation of nitroaromatic compounds by
microorganisms. Appl Microbiol Biotechnol 42:499–507
Master ER, Lai VW-M, Kuipers B, Cullen WR, Mohn WW (2002) Sequential anaerobic-aerobic
treatment of soil contaminated with weathered Aroclor 1260. Environ Sci Technol 36:100–103
Matsumura F, Boush GM (1971) DDT metabolized by microorganisms from Lake Michigan.
Nature 230:325–326
Maymo-Gatell X, Chien Y, Gossett JM, Zinder SH (1997) Isolation of a bacterium that
reductively dechlorinates tetrachloroethene to ethene. Science 276:1568–1571
Maymo-Gatell X, Nijenhuis I, Zinder SH (2001) Reductive dechlorination of cis-1,2dichloroethene and vinyl chloride by Dehalococcoides ethenogenes . Environ Sci Technol
35:516–521
McCarty PL (1997) Breathing with chlorinated solvents. Science 276:1521–1522
McCarty PL, Smith DP (1986) Anaerobic wastewater treatment. Environ Sci Technol 20:1200–
1206
McCarty PL, Wilson JT (1992) Natural anaerobic treatment of a TCE plume St. Joseph,
Michigan, NPL site. In: United States Environmental Protection Agency (ed) Bioremediation of
hazardous wastes. EPA /600/R-92/126. US EPA, Washington, D.C., pp 57–50
McCormick NG, Cornell JH, Kaplan AM (1981) Biodegradation of hexahydro-1,3,5-trinitro1,3,5-trazine. Appl Environ Microbiol 42:817–823
McCormick NG, Feeherry FE, Levinson HS (1976) Microbial transformation of 2,4,6-TNT and
other nitroaromatic compounds. Appl Environ Microbiol 31:949–958
Meckenstock RU, Annweiler E, Michaelis W, Richnow HH, Schink B (2000) Anaerobic
naphthalene degradation by a sulfate-reducing enrichment culture. Appl Environ Microbiol
66:2743–2747
Mihelcic JR, Luthy RG (1988) Microbial-degradation of acenaphthene and naphthalene under
denitrification conditions in soil-water systems. Appl Environ Microbiol 54:1188–1198
Mikesell MD, Boyd S (1985) Reductive dechlorination of the pesticides 2,4-D and 2,4,5-T, and
pentachlorophenol in anaerobic sludges. J Environ Qual 14:337–340
Mikesell MD, Boyd SA (1986) Complete reductive dechlorination and mineralization of
pentachlorophenol by anaerobic microorganisms. Appl Environ Microbiol 52:861–865
Mogensen AS, Dolfing J, Haagensen F, Ahring BK (2003a) Potential for anaerobic conversion
of xenobiotics. Adv Biochem Eng Biotechnol 82:69–134
Mogensen AS, Haagensen F, Ahring BK (2003b) Anaerobic degradation of linear alkylbenzene
sulfonate. Environ Toxicol Chem 22:706–711
Morgan P, Watkinson RJ (1989) Microbiological methods for the clean up of soil and
groundwater contaminated with halogenated organic compounds. FEMS Microbiol Rev 63:277–
300
Neumann A, Wohlfarth G, Diekert G (1996) Purification and characterization of
tetrachloroethene reductive dehalogenase from Dehalospirillum multivorans. J Biol Chem
271:16515–16519
Newcombe D, Crowley DE (1999) Bioremediation of atrazine-contaminated soil by repeated
applications of atrazine-degrading bacteria. Appl Microbiol Biotechnol 51:877–882
Nicholson DK, Woods SL, Istok JD, Peek DC (1992) Reductive dechlorination of chlorophenols
by a pentachlorophenol-acclimated methanogenic consortium. Appl Environ Microbiol 58:2280–
2286
Ohtsubo Y, Kudo T, Tsuda M, Nagata Y (2004) Strategies for bioremediation of polychlorinated
biphenyls. Appl Microbiol Biotechnol 65:250–258
Padda RS, Wang CY, Hughes JB, Bennett GN (2000) Mutagenicity of trinitrotoluene and its
metabolites formed during anaerobic degradation by Clostridium acetobutylicum ATCC 824.
Environ Toxicol Chem 19:2871–2875
Padda RS, Wang C, Hughes JB, Bennett GN (2003) Mutagenicity of nitroaromatic explosives
during anaerobic transformation by Clostridium acetobutylicum. Environ Toxicol Chem
22:2293–2297
Pagano JJ, Scrudato RJ, Roberts RN, Bemis JC (1995) Reductive dechlorination of PCBcontaminated sediments in an anaerobic bioreactor system. Environ Sci Technol 29:2584–2589
Peres CM, Agathos SN (2000) Biodegradation of nitroaromatic pollutants: from pathways to
remediation. Biotechnol Annu Rev 6:197–220
Pignatello JJ, Johnson LK, Martinson MM, Carlson RE, Crawford RL (1985) Response of the
microflora in outdoor experimental streams to pentachlorophenol: compartmental contributions.
Appl Environ Microbiol 50:127–132
Preuss A, Fimpel J, Diekert G (1993) Anaerobic transformation of 2,4,6-trinitrotoluene (TNT).
Arch Microbiol 159:345–353
Prince RC (1993) Petroleum spill bioremediation in marine environments. Crit Rev Microbiol
19:217–242
Quensen JF III, Tiedje JM, Boyd SA (1988) Reductive dechlorination of polychlorinated
biphenyls by anaerobic microorganisms from sediments. Science 242:752–754
Quensen JF III, Mueller SA, Jain MK, Tiedje JM (1998) Reductive dechlorination of DDE to
DDMU in marine sediment microcosms. Science 280:722–724
Quensen JF III , Tiedje JM, Jain MK, Mueller SA (2001) Factors controlling the rate of DDE
dechlorination to DDMU in Palos Verdes margin sediments under anaerobic conditions. Environ
Sci Technol 35:286–291
Rabus R, Widdel F (1995) Anaerobic degradation of ethylbenzene and other aromatic
hydrocarbons by new denitrifying bacteria. Arch Microbiol 163:96–103
Rabus R, Wilkes H, Behrends A, Armstroff A, Fischer T, Pierik AJ, Widdel F (2001) Anaerobic
initial reaction of n-alkanes in a denitrifying bacterium: evidence for (1-methylpentyl)succicate
as initial product and for involvement of an organic radical in n-hexane metabolism. J Bacteriol
183:1707–1715
Raymond RL (1974) Reclamation of hydrocarbon contaminated ground water. US Patent
3 846 290, 5 November 1974
Renner R (1998) Natural remediation of DDT, PCBs debated. Environ Sci Technol 32:360–
363A
Ribarova I, Topalova J, Ivanov I, Kozuharov D, Dimkov R, Cheng C (2002) Anaerobic
sequencing batch reactor as initiating stage in complete pentachlorophenol biodegradation.
Water Sci Technol 46:565–569
Rittmann BE, McCarty PL (2001) Environmental biotechnology: principles and applications.
McGraw-Hill, New York
Roberts DJ, Kaake RH, Funk SB, Crawford DL, Crawford RL (1993) Anaerobic remediation of
dinoseb from contaminated soil. An on-site demonstration. Appl Biochem Biotechnol 39–
40:781–789
Rockne KJ, Strand SE (1998) Biodegradation of bicyclic and polycyclic aromatic hydrocarbons
in anaerobic enrichments. Environ Sci Technol 32:3962–3967
Rockne KJ, Chee-Sanford JC, Sanford RA, Hedlund BP, Staley JT, Strand SE (2000) Anaerobic
naphthalene degradation by microbial pure culture under nitrate-reducing conditions. Appl
Environ Microbiol 66:1595–1601
Rooney-Varga JN, Anderson RT, Fraga JL, Ringelberg D, Loveley DR (1999) Microbial
communities associated with anaerobic benzene degradation in a petroleum-contaminated
aquifer. Appl Environ Microbiol 65:3056–3063
Rothermich MM, Hayes LA, Lovley DR (2002) Anaerobic, sulfate-dependent degradation of
polycyclic aromatic hydrocarbons in petroleum-contaminated harbor sediment. Environ Sci
Technol 36:4811–4817
Rueter P, Rabus R, Wilkes H, Aeckersberg F, Rainey FA, Jannasch HW, Widdel F (1994)
Anaerobic oxidation of hydrocarbons in crude oil by new types of sulphate-reducing bacteria.
Nature 372:455–458
Ruppe S, Neumann A, Vetter W (2003) Anaerobic transformation of compounds of technical
toxaphene. I. Regiospecific reaction of chlorobornanes with geminal chlorine atoms. Environ
Toxicol Chem 22:2614–2621
Ruppe S, Neumann A, Vetter W (2004) Anaerobic transformation of compounds of technical
toxaphene. II. Fate of compounds lacking geminal chlorine atoms. Environ Toxicol Chem
23:591–598
Salminen JM, Tuomi PM, Suortti A-M, Jørgensen KS (2004) Potential for aerobic and anaerobic
biodegradation of petroleum hydrocarbons in boreal subsurface. Biodegradation 15:29–39
Schink B (2002) Anaerobic digestion: concepts, limits and perspectives. Water Sci Technol
45:1–8
Semprini L, Hopkins GD, McCarty PL, Roberts PV (1992) In situ biotransformation of carbon
tetrachloride and other halogenated compounds resulting from biostimulation under anoxic
conditions. Environ Sci Technol 26:2454–2461
Sethunathan N (1973) Microbial degradation of insecticides in flooded soil and in anaerobic
cultures. Residue Rev 47:143–165
Seybold CA, Mersie W, McNamee C (2001) Anaerobic degradation of atrazine and metolachlor
and metabolite formation in wetland soil and water microcosms. J Environ Qual 30:1271–1277
Shelton DR, Tiedje JM (1984) Isolation and partial characterization of bacteria in an anaerobic
consortium that mineralizes 3-chlorobenzoic acid. Appl Environ Microbiol 48:840–848
Sheremata TW, Hawari J (2000) Cyclodextrins for desorption and solubilization of 2,4,6trinitrotoluene and its metabolites from soil. Environ Sci Technol 36:3462–3468
Shimazu M, Mulchandani A, Chen W (2001) Simultaneous degradation of organophosphorus
pesticides and p-nitrophenol by a genetically engineered Moraxell sp. with surface-expressed
organophosphorus hydrolase. Biotechnol Bioeng 76:318–324
So CM, Young LY (1999a) Initial reactions in anaerobic alkane degradation by a sulfate reducer,
strain AK-01. Appl Environ Microbiol 65:5532–5540
So CM, Young LY (1999b) Isolation and characterization of a sulfate-reducing bacterium that
anaerobically degrades alkanes. Appl Environ Microbiol 65:2969–2976
So CM, Phelps CD, Young LY (2003) Anaerobic transformation of alkanes to fatty acids by a
sulfate-reducing bacterium, strain Hxd3. Appl Environ Microbiol 69:3892–3900
Somsamak P, Cowan RM, Haggblom MM (2001) Anaerobic biotransformation of fuel
oxygenates under sulfate-reducing conditions. FEMS Microbiol Ecol 37:259–264
Song B, Palleroni NJ, Haggblom MM (2000) Isolation and characterization of diverse
halobenzoate-degrading denitrifying bacteria from soils and sediments. Appl Environ Microbiol
66:3446–3453
Song B, Palleroni NJ, Kerkhof LJ, Haggblom MM (2001) Characterization of halobenzoatedegrading, denitrifying Azoarcus and Thauera isolates and description of Thauera
chlorobenzoica sp. nov. Int J Syst Evol Microbiol 51:589–602
Spormann AM, Widdel F (2000) Metabolism of alkylbenzenes, alkanes, and other hydrocarbons
in anaerobic bacteria. Biodegradation 11:85–105
Steber J, Wierich P (1987) The anaerobic degradation of detergent range fatty alcohol
ethoxylates. Studies with 14C-labbelled model surfactant. Water Res 21:661–667
Steinbach A, Seifert R, Annweiler E, Michaelis W (2004) Hydrogen and carbon isotope
fractionation during anaerobic biodegradation of aromatic hydrocarbons-a field study. Environ
Sci Technol 38:609–616
Stocking AJ, Deeb RA, Flores AE, Stringfellow W, Talley J, Brownell R, Kavanaugh MC
(2000) Bioremediation of MTBE: a review from a practical perspective. Biodegradation 11:187–
201
Strong LC, McTavish H, Sadowsky MJ, Wackett LP (2000) Field-scale remediation of atrazinecontaminated soil using recombinant Escherichia coli expressing atrazine chlorohydrolase.
Environ Microbiol 2:91–98
Sullivan ER, Zhang X, Phelps C, Young LY (2001) Anaerobic mineralization of stable-isotopelabeled 2-methylnaphthalene. Appl Environ Microbiol 67:4353–4357
Sung Y, Ritalahti KM, Sanford RA, Urbance JW, Flynn SJ, Tiedje JM, Löffler FE (2003)
Characterization of two tetrachloroethene-reducing, acetate-oxidizing anaerobic bacteria and
their description as Desulfuromonas michiganesis sp. nov. Appl Environ Microbiol 69:2964–
2974
Swisher RD (1987) Surfactant biodegradation, 2nd edn. Dekker, New York, N.Y.
Tadros MG, Crawford A, Mateo-Sullivan A, Zhang C, Hughes JB (2000) Toxic effects of
hydroxylamino intermediates on algae Selenastrum capricornutum. Bull Environ Contam
Toxicol 64:579–585
Tartakovsky B, Levesque M, Dumortier R, Beaudet R, Guiot SR (1999) Biodegradation of
pentachlorophenol in a continuous anaerobic reactor augmented with Desulfitobacterium
frappieri PCP-1. Appl Environ Microbiol 65:4357–4362
Ternan NG, McGrath JW, McMullan G, Quinn JP (1998) Organophosphonates: occurrence,
synthesis and biodegradation by microorganisms. World J Microbiol Biotechnol 14:635–647
Terzenbach DP, Blaut M (1994) Transformation of tetrachloroethylene to trichloroethylene by
homoacetogenic bacteria. FEMS Microbiol Lett 123:213–218
Tiedje JM, Quensen JF III, Chee-Sanford J, Schimel JP, Boyd SA (1993) Microbial reductive
dechlorination of PCBs. Biodegradation 4:231–240
Vargas C, Song B, Camps M, Haggblom MM (2000) Anaerobic degradation of fluorinated
aromatic compounds. Appl Microbiol Biotechnol 53:342–347
Vargas C, Fennell DE, Häggblom MM (2001) Anaerobic reductive dechlorination of chlorinated
dioxins in estuarine sediments. Appl Microbiol Biotechnol 57:786–790
Wackett LP, Sadosky MJ, Martinez B, Shapir N (2002) Biodegradation of atrazine and related striazine compounds: from enzymes to field studies. Appl Microbiol Biotechnol 58:39–45
Wagener S, Schink B (1988) Fermentative degradation of nonionic surfactants and polyethylene
glycol by enrichment cultures and by pure cultures of homoacetogenic and propionate-forming
bacteria. Appl Environ Microbiol 54:561–565
Wedemeyer G (1966) Dechlorination of DDT by Aerobacter aerogenes. Science 152:647
Widdel F, Rabus R (2001) Anaerobic biodegradation of saturated and aromatic hydrocarbons.
Curr Opin Biotechnol 12:259–276
Wiegel J, Zhang X, Wu Q (1999) Anaerobic dehalogenation of hydroxylated polychlorinated
biphenyls by Desulfitobacterium dehalogenans. Appl Environ Microbiol 65:2217–2221
Wilkes H, Rabus R, Fischer T, Armstroff A, Behrends A, Widdel F (2002) Anaerobic
degradation of n-hexane in a denitrifying bacterium: further degradation of the initial
intermediate (1-methylphentyl)succinate via C-skeleton rearrangement. Arch Microbiol
177:235–243
Williams PP (1977) Metabolism of synthetic organic pesticides by anaerobic microorganisms.
Residue Rev 66:63–135
Woods SL, Ferguson JF, Benjamin MM (1989) Characterization of chlorophenol and
chloromethoxybenzene biodegradation during anaerobic treatment. Environ Sci Technol 23:62–
68
Wu Q, Bedard DL, Wiegel J (1997) Temperature determines pattern of anaerobic microbial
dechlorination of Aroclor 1260 primed by 2,3,4,6-tetrachlorobiphenyl in Woods Pond sediments.
Appl Environ Microbiol 63:4818–4825
Wu Q, Sowers KR, May HD (1998) Microbial reductive dechlorination of Aroclor 1260 in
anaerobic slurries of estuarine sediments. Appl Environ Microbiol 64:1052–1058
Wu Q, Sowers KR, May HD (2000) Establishment of a polychlorinated biphenyl-dechlorinating
microbial consortium, specific for doubly flanked chlorines, in a defined, sediment-free medium.
Appl Environ Microbiol 66:49–53
Wu Q, Milliken CE, Meier GP, Watts JE, Sowers KR, May HD (2002a) Dechlorination of
chlorobenzenes by a culture containing bacterium DF-1, a PCB dechlorinating microorganism.
Environ Sci Technol 36:3290–3294
Wu Q, Watts JE, Sowers KR, May HD (2002b) Identification of a bacterium that specifically
catalyzes the reductive dechlorination of polychlorinated biphenyls with doubly flanked
chlorines. Appl Environ Microbiol 68:807–812
Xue SK, Iskandar IK, Selim HM (1995) Adsorption-desorption of 2,4,6-trinitrotoluene and
hexahydro-1,3,5-trinitro-1,3,5-triazine in soils. Soil Sci 160:317–327
Young DM, Unkefer PJ, Ogden KL (1997) Biotransformation of hexahydro-1,3,5-trinitro-1,3,5triazine (RDX) by a prospective consortium and its most effective isolate Serratia marcescens.
Biotechnol Bioeng 53:515–522
Zengler K, Richnow HH, Rosselló-Mora R, Michaelis W, Widdel F (1999) Methane formation
from long-chain alkanes by anaerobic microorganisms. Nature 401:266–269
Zhang C, Hughes JB (2003) Biodegradation pathways of hexahydro-1,3,5-trinitro-1,3,5-triazine
(RDX) by Clostridium acetobutylicum cell-free extract. Chemosphere 50:665–671
Zhang C, Hughes JB, Nishino SF, Spain J (2000a) Slurry-phase biological treatment of 2,4dinitrotoluene and 2,6-dinitrotoluene: Role of bioaugmentation and effects of high dinitrotoluene
concentrations. Environ Sci Technol 34:2810–2816
Zhang X, Young LY (1997) Carboxylation as an initial reaction in the anaerobic metabolism of
naphthalene and phenanthrene by sulfidogenic consortia. Appl Environ Microbiol 63:4759–4764
Zhang X, Sullivan ER, Young LY (2000b) Evidence for aromatic ring reduction in the
biodegradation pathway of carboxylated naphthalene by a sulfate-reducing consortium.
Biodegradation 11:117–124
Zhao J-S, Halasz A, Paquet L, Beaulieu C, Hawari J (2002) Biotransformation of hexahydro1,3,5-trnitro-1,3,5-triazine and its mononitroso derivative hexahydro-1-nitroso-3,5-dinitro-1,3,5triazine by Klebsiella pneumoniae strain SCZ-1 isolated from an anaerobic sludge. Appl Environ
Microbiol 68:5336–5341