Applied Microbiology and Biotechnology

Applied Microbiology and Biotechnology
© Springer-Verlag 2005
Biodegradation of xenobiotics by anaerobic
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
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.
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
D. oleovorans (P): mineralizes C12–C20 n-alkane
G. spp. (P): oxidizes benzene in Fe(II)-reducing
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
Thauera-related (S/P): denitrifying bacteria
completely mineralize ethylbenzene
Ball et al. 1996; Rabus and
Widdel 1995
D. acetonicum- and Desulfosarcina variabilisrelated: mineralizes o- and m-xylene
Harms et al. 1999; Hess et al.
1997; Rabus and Widdel 1995
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
Pure aerobes isolated; slow under anaerobic
conditions, no pure anaerobes isolated
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
Bacteria namea, source of isolationb, chemical
Desulfovibrio spp. (N): most extensively studied
genera transforming TNT
McCormick et al. 1976
Desulfovibrio spp. (S): uses RDX and HMX as
sole N source
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
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
Dehalobacter restrictus (S): transforms PCE to
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
Dehalococcoides sp. (A): able to grow with VC
and transform VC into ethene
He et al. 2003
Desulfitobacterium dehalogenans (S):
dehalogenates flanking Cl of OH-PCBs
Wiegel et al. 1999
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
Dehalococcoides sp. (S): uses dioxins as the sole Bunge et al. 2003
Bacteria namea, source of isolationb, chemical
electron acceptor
C. sp. (N): degrades DDT as the sole C source.
Degrades other chlorinated pesticides
Aerobacter aerogenes, K. pneumoniae, N.
vulgaris (S): DDT-degrading
Ruppe et al. 2003, 2004;
Sethunathan 1973; Williams
Dehalospirilum multivorans: preferentially
dechlorinates technical toxaphene
Flavobacterium sp. (S): attacks P-insecticides
including diazino and parathion
Sethunathan 1973
K. pneumoniae (D): uses three chlorinated striazines as the sole N source
Ernst and Rehm 1995;
C. bifermentans (D): utilizes Dinoseb as a sole C
Hammill and Crawford 1996
via cometabolism
Strain RZLAS (D): the only pure anaerobe using
LAS as the sole S source
Denger and Cook 1999
Pelobacter propionicus & A. sp. (D): LAEOs
fermented to CH4 and CO2
Wagener and Schink 1988
Unable to isolate a single bacterium using
cationic surfactant as the sole C source
Madsen et al. 2001
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.
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).
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).
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
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
NO2– injection stimulated BTEX
Barker et al. 1987
Fuel spill site
NO2– enhanced m- and p-xylene
Hutchins et al. 1991
2–5 ring PAHs degraded under
SO42–, NO3–-reducing conditions
Rockne and Strand
1998; Rothermich et al.
Stimulated by humic substances;
Finneran and Lovley
field evidence in groundwater using 2001; Kolhatkar et al.
Full-scale reactor
TNT to mineralizable and
nonaromatic products by
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
30 weeks
PCE, TCE, CT Shallow aquifer
Nitrate and acetate injection
Semprini et al. 1992
Type and scale
Performance results
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,
ethene and electron donor
McCarty and Wilson
Major et al. 1991
Nutrients, enrichment culture
injection converted TCE to ethene
Pseudomonas stutzeri KC removed
Dybas et al. 1998
6-l Batch
11% and 23% of total Cl/biphenyl
was reduced after 13 weeks
488-m Long
PCP disappeared: O2-rich > O2-poor
Pignatello et al. 1985
anaerobic > sorption
Pilot digester
>97.5% removal; 95% converted to Chen and Berthouex
UASB, biofilm
and CSTR
95% Removal with C source
provided, dechlorinated at the othen m-, but not p- chlorines;
reactor optimization studies
Dechlorinated under methanogenic
Vargas et al. 2001
In situ anaerobic
HCH converted to MCB and
benzene which was mineralized in
an aerobic treatment plant
Langenhoff 2003
Field studies
Nonpersistent, degraded under Plimited conditions
Ternan et al. 1998
500-l Fermentor;
35 yard3
(=26.8 m3)
Large-scale aerobic systems using
Pseudomonas sp. and recombinant
Escherichia coli; no large-scale
anaerobic processes
Newcombe and
Crowley 1999; Strong
et al. 2000
2,600-l Static
Undetectable by 15 days after
Robert et al. 1993
addition of C and acclimated culture
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.
Batch to
Partially (APEO) to well (AE)
Ferguson and
Brownawell 2003;
Huber et al. 2000;
Ellis et al. 2000
Pagano et al. 1995
Hendriksen et al. 1992;
Ribarova et al. 2002;
Woods et al. 1989
Type and scale
Performance results
Strongly adsorbed, toxic, scarce
anaerobic biodegradation data
Madsen et al. 2001
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
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
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)
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