United States
Environmental Protection
Agency
Office of
Research and
Development
Office of Solid Waste
and Emergency
Response
EPA/540/4-90/054
January 1991
&EPA Ground Water Issue
Reductive Dehalogenation of Organic
Contaminants in Soils and Ground Water
Judith L. Sims, Joseph M. Suflita, and Hugh H. Russell
The Regional Superfund Ground-Water Forum is a group of
ground-water scientists, representing EPA's Regional Superfund
Offices, organized to exchange up-to-date information related to
ground-water remediation of superfund sites. One of the major
issues of concern to the Forum is the transport and fate of
contaminants in soil and ground water as related to subsurface
remediation. Process which influence the behavior of
contaminants in the subsurface must be considered both in
evaluating the potential for movement as well as in designing
remediation activities at hazardous waste sites. Such factors not
only tend to regulate the mobility of contaminants, but also their
form and stability. Reductive dehalogenation is a process which
may prove to be of paramount importance in dealing with a
particularly persistent class of contaminants often found in soil
and ground water at superfund sites. This paper summarizes
concepts associated with reductive dehalogenation and describes
applications and limitations to its use as a remediation technology.
For further information contact Dr. Hugh Russell, FTS 743-2444
at RSKERL-Ada.
Abstract
Introduction and large scale production of synthetic halogenated
organic chemicals over the last 50 years has resulted in a group
of contaminants which tend to persist in the environment and
resist both biotic and abiotic degradation. The low solubility of
these types of contaminants, along with theirtoxicity and tendency
to accumulate in food chains, make them particularly relevant
targets for remediation activities.
Although the processes involved in dechlorination of many of
these organic compounds are well understood in the fields of
chemistry and microbiology, technological applications of these
processes to environmental remediation are relatively new—
particularly at pilot or field scale. It is well established, however,
thatthere are several mechanisms which result in dehalogenation
of some classes of organic contaminants, often rendering them
less offensive environmentally. These include: stimulation of
metabolic sequences through introduction of electron donor and
acceptor combinations; addition of nutrients to meet the needs
of dehalogenating microorganisms; possible use of engineered
micro-organisms; and use of enzyme systems capable of
catalyzing reductive dehalogenation.
The current state of research and development in the area of
reductive dehalogenation is discussed along with possible
technological applications of relevant processes and mechanisms
to the remediation of soil and ground water contaminated with
chlorinated organics. In addition, an overview of research needs
is suggested which might be of interest for development of in situ
systems to reduce the mass of halogenated organic contaminants
in soil and ground water.
Introduction
Large scale production of synthetic halogenated organic
compounds, which are often resistant to both biotic and abiotic
degradation, has occurred only in the last few decades (Hutzinger
and Verkamp 1981). However, naturally occurring halogenated
organic compounds have existed in marine systems for perhaps
Superfund Technology Support Center for Ground Water
Robert S. Kerr Environmental
Research Laboratory
Ada, OK
-------
millions of years. These compounds, including aliphatic and
aromatic compounds containing chlorine, bromine, or iodine,
are produced by macroalgae and invertebrates. The presence
of these natural compounds, at potentially high concentrations,
may have resulted in populations of bacteria that are effective
dehalogenators (King 1988). Microorganisms exposed to
halogenated compounds in soil and ground water may also have
developed enzymatic capabilities similar to those in marine
environments. Enzyme systems that have evolved to degrade
nonchlorinated compounds may also be specific enough to
degrade those that are chlorinated. (Tiedje and Stevens 1987).
Many halogenated organic compounds are not very soluble and
tend to be highly lipophilic, therefore having the potential to
bioaccumulate in some food chains. These chemical properties,
along with their toxicity and resistance to degradation, present
the potential for adverse health effects and ecosystem
perturbations upon exposure (Rochkind et al. 1986).
Recent research findings indicate that anaerobic processes that
remove halogens fromthese compounds produce dehalogenated
compounds that are generally less toxic, less likely to
bioaccumulate, and more susceptible to further microbial attack,
especially by aerobic microorganisms utilizing oxidative
biodegradative processes. Both aromatic and nonaromatic
organic compounds are subject to these dehalogenation
processes. Technological applications of these processes for
remediation of contaminated soils and ground waters is of a
relatively new concept.
Recent research also has shown that anaerobic dehalogenation
reactions specifically involving reductive processes can effectively
degrade a wide variety of halogenated contaminants in soil and
ground water (Vogel et al. 1987, Kuhn and Suflita 1989a).
Organic compounds generally represent reduced forms of carbon,
making degradation by oxidation energetically favorable.
However, halogenated organic compounds are relatively oxidized
by the presence of halogen substituents, which are highly
electronegative and thus more susceptible to reduction. A
compound with more halogen substituents is therefore more
oxidized and more susceptible to reduction. Thus, with increased
halogenation, reduction becomes more likely than does oxidation
(Vogel etal. 1987).
An organic compound is considered to be reduced if a reaction
leads to an increase in its hydrogen content or a decrease in its
oxygen content; however, many reduction reactions (e.g., the
vicinal reduction process) do not involve changes in the hydrogen
or oxygen content of a compound. Oxidation and reduction
reactions are more precisely defined in terms of electron transfers.
An organic chemical is said to be reduced if it undergoes a net
gain of electrons as the result of a chemical reaction (electron
acceptor), and is said to be oxidized if it undergoes a net loss of
electrons (electron donor). Under environmental conditions,
oxygen commonly acts as the electron acceptor when present.
When oxygen is not present (anoxic conditions), microorganisms
can use organic chemicals or inorganic anions as alternate
electron acceptors under metabolic conditions referred to as
fermentative, denitrifying, sulfate-reducing or methanogenic.
Generally, organic compounds present at a contaminated site
represent potential electron donors to support microbial
metabolism. However, halogenated compounds can act as
electron acceptors, and thus become reduced in the reductive
dehalogenation process. Specifically, dehalogenation by
reduction is the replacement of a halogen such as chloride,
bromide, fluoride, or iodide on an organic molecule by a hydrogen
atom. Vicinal reduction occurs when two halogens are released
while two electrons are incorporated into the compound.
An organic chemical would be expected to be reduced if the
electrode potential of the specific soil or ground-water system, in
which the chemical is present, is less than that of the organic
chemical (Dragun 1988). The electrode potential is described by
the oxidation-reduction (redox) status of the system, refering to
potential for the transfer of electrons to a reducible material. The
electron (e~) participates in chemical reactions in soil and ground
water similar to the hydrogen ion (H+) in that electrons are
donated from a reduced compound to an oxididized. Redox
potential (Eh) is usually reported in volts and is measured using
a reference electrode in combination with a metallic electrode,
such as platinum, which is sensitive and reversible to oxidation-
reduction conditions.
The redox potential of a soil system is complex. The oxidation
state of each soil constituent, such as organic compounds,
humus, iron, manganese, and sulfur, contributes to the measured
redox potential. The contribution of each constituent in a system
varies with such factors as soil water content, oxygen activity,
and pH. Well-oxidized soils have redox potentials of 0.4 to 0.8
V, while extremely reduced soils may have potentials of-0.1 to
-0.5 V (Dragun 1988).
The potential for anaerobic biological processes to reductively
dehalogenate organic compounds may be important in the
bioremediation of soils and aquifers contaminated with these
compounds. These environments often become anaerobic due
to depletion of oxygen by the microbial degradation of more
easily degradable organic matter. When compounds can be
degraded under anaerobic conditions, the cost associated with
the maintenance of an aerobic environment by providing air,
ozone, or hydrogen peroxide would be eliminated (Suflita et al.
1988).
While anaerobic biological mediated reductive dehalogenation
mechanisms were demonstrated as early as 1983 (Allan, 1955),
the utilization of this process as a remedial alternative to reduce
the overall mass of halogenated organic compounds from soil
and ground water is a new concept and still subject to field
demonstrations.
For this reason research is currently underway to better define
the basic mechanisms of reductive dehalogenation reactions.
Such approaches may include: (1) stimulation of desirable
metabolic sequences in soil and ground water through the
intentional introduction of suitable electron donor and acceptor
combinations (Suflita et al. 1988); (2) addition of adequate
nutrients to meet the nutritional requirements of dehalogenating
microorganisms (Palmer et al. 1989); (3) use of engineered
microorganisms with optimum dehalogenating activity (Palmer
et al. 1989); and (4) addition of cell-free enzymes capable of
-------
catalyzing reductive dehalogenation reactions (DeWeerd and
Suflita 1989).
Dehalogenation Mechanisms
Anaerobic reductive dehalogenation is only one of the
mechanisms available to remove halogens from organic
compounds. Other anaerobic dehalogenation processes are
identified in Figure 1 (Kuhn and Suflita 1989a). The reactions are
classified according to the type of compound undergoing
dehalogenation, i.e., aromatic or nonaromatic.
Dehalogenation of Aromatic Compounds
Two mechanisms of dehalogenation for aromatic compounds
under anaerobic conditions have been observed: reduction and
hydrolysis. Reductive mechanisms are recognized as the
predominant pathway for removal of halogens from homocyclic
aromatic rings under anaerobic conditions, while hydrolytic
dehalogenation (including both chemically and enzymatically
mediated reactions) is the preferred mechanism for heterocyclic
aromatic compounds (Suflita et al. 1982; Kuhn and Suflita
1989a). However, Adrian and Suflita (1989) have recently
demonstrated reductive debromination of the herbicide bromacil
under methanogenic conditions. This is the first report of
reductive dehalogenation of a heterocyclic aromatic compound.
Reductive Dehalogenation of Aromatic Compounds
Many classes of halogenated aromatic compounds have been
shown to be degraded by reductive dehalogenation processes
(Table 1). Evidence for the involvement of microorganisms in
aryl or aromatic reductive dehalogenation reactions include: (1)
the specificity of the reductive reaction; (2) characteristic lag
periods required before significant dehalogenation is observed;
(3) the absence of activity in autoclaved controls; and (4) the
isolation of aryl dehalogenating bacteria.
Reductive dehalogenation is rare in well-aerated environments.
Methanogenic conditions, in which the typical redox potential is
-0.3 V, the preferred electron acceptor is carbon dioxide, and the
Anaerobic Dehalogenation Mechanisms
Aromatic Compounds-
2e-+ H+ X
H
1
Reduction
OH X
Hydrolysis
OH
N
"NT
• 'Nonaromatic Compounds. -
Reduction
(Hydrogenolysis)
Hydrolysis
(Substitution)
2e- + H+ X
XV ./
'- -^ >
—X
OH X
>C-OH
Vicinal Reduction
(Dihalo-Elimination)
X X 2e-
-c-c- V
X
4
X
Dehydrohalogenation _ Q _ Q _
H
HX
4
Figure 1. Examples of anaerobic dehalogenation mechanisms for aromatic and nonaromatic pesticides (Kuhn and Suflita, 1989a)
-------
product is methane (Dragun 1988), appear to be optimal forthis
type of biotransformation. Genthneretal. (1989), have recently
investigated dehalogenation of monochlorophenols and
monochlorobenzoates under four anaerobic enrichment
conditions: methanogenic, nitrate-reducing, sulfate-reducing,
and bromoethane sulfonic acid (BESA)-amended. BESA is a
potent inhibitor of methanogenesis and was used to promote
reductive dechlorination as a terminal electron process.
Aquatic sediments used as inocula were collected from a salinity
gradient that included both freshwater and estuarine
environments and varying degrees of exposure to industrial
effluents. Degradation was observed least often in enrichments
with nitrate or sulfate, and most often when amended with 1 mM
BESA. In contrast to 1mM BESA, 10mM BESA prevented or
delayed the degradation of several of the chloroaromatic
compounds, suggesting inhibitionofmethanogenesisorinhibition
of reductive dechlorination by BESA. Other sulfur oxyanions
also have been shown to inhibit anaerobic dehalogenation
reactions where sulfate is present as an inorganic contaminant
(DeWeerd et al. 1986, Gibson and Suflita 1986, Suflita et al.
1988, Kuhn and Suflita 1989b). Additional research is being
conducted in environments where sulfate occurs naturally. King
(1988) showed that sulfate-reducing bacteria did not participate
in dehalogenation of 2,4-dibromophenol (DBP), a naturally
occurring halogenated organic compound in some marine
sediments, but did appear to degrade phenol, a metabolic
product of DBP dehalogenation.
The reductive dehalogenation of chlorinated compounds, as
shown in Table 1, is characterized by their specificity for
compounds within a particular chemical class, for example
benzoates, phenols, or phenoxyacetates (Suflita et al. 1982,
Gibson and Suflita 1986, Suflita and Miller 1985, Kuhn and
Suflita 1989a). Recently, however, research has shown that
cross-acclimation between compound classes can occur. Struijs
and Rogers (1989) demonstrated the reductive dehalogenation
of dichloroanilines by anaerobic microorganisms in pond
sediments acclimated to dehalogenate dichlorophenols. Since
both hydroxyl and amino groups have a tendency to donate
electrons, the authors hypothesized that organisms that were
capable of dechlorinating dichlorophenols, which have been
shown to be relatively non-persistent in the environment, could
possibly dechlorinate the more persistent dichloroanilines. The
monochloroanilines produced by dechlorination of the
dichloroanilines were stable under anaerobic conditions, but
have been shown previously to be readily degraded under
aerobic conditions (Zeyer and Kearney 1982, Zeyer et al. 1985).
The specificity of dehalogenation also is dependent on the
position of halogens on the aromatic ring within a class of
compounds. For example, chlorinated benzoates are generally
more readily dehalogenated at the meta position, followed by
the ortho and para positions (Suflita etal. 1982, Genthneretal.
1989). Hydroxy, alkoxy, and nitrogen-substituted aromatic
compounds generally are dehalogenated faster at ortho and
para halogens (Kuhn and Suflita 1989a, 1989b), however,
Genthner et al. (1989) recently have shown that the order of
degradability of monochlorophenols was meta > ortho > para..
Mikesell and Boyd (1986) have shown that three groups of
acclimated microorganisms can act in concert to completely
Class of Halogenated Aromatic Compounds
Examples of Specific Compounds
Carboxylated Benzenes
Oxygen-Substituted Benzenes
Nitrogen-Substituted Benzenes
Cyano-Substituted Benzenes
Methylene-substituted Benzenes
Chlorinated Benzenes
Polychlorinated biphenyls
Amiben
Dicamba
2,3,6-trichlorobenzoate
Pentachlorophenol
Chlorinated phenoxyacetates (e.g.,
2,4-D, 2,4,5-T
Halogenated diphenyl ether herbicides
(e.g., chloronitrofen)
3,4-Dihalogenated aromatic compounds
(diuron, DCPU, linuron, DCIPC,
propanil)
Pentachloronitrobenzene
2,4,5,6-tetrachloroisophthalonitrile
(TPN)
Benthiocarb
Hexachlorobenzene
Araclors (commercial PCB products)
Table 1. Classes of halogenated aromatic compounds demonstrated to be susceptible to degradation by reductive dehalogenation
processes (Kuhn and Suflita 1989a).
-------
dehalogenate pentachlorophenol (PCP) to form phenol, a
substrate that was labile under the methanogenic conditions of
their experiments. Each type of microorganism, acclimated to
one of three monochlorophenol isomers, transformed PCP by
removal of halogens from the same relative ring positions at
which they dehalogenated the monochlorophenol substrates.
The 2-chlorophenol adapted cells dehalogenated PCP at the
two ortho positions as well as from the para position. Similarly,
4-chlorophenol adapted cells cleaved the para chlorine of PCP
in addition to the two ortho substituents. In contrast, the 3-
chlorophenol adapted cells exclusively dehalogenated the meta
position.
Other studies of PCP degradation have shown accumulation of
tri- and tetrachlorophenol intermediates, which indicates that
higher halogenated phenols tend to be more readily
dehalogenated than their lesser halogenated congeners.
Similarly, dehalogenation of chlorinated anilines shows shorter
lag periods and faster dehalogenation rates with multi-
halogenated compounds compared to di- and monohalogenated
anilines. Dehalogenation of aromatic amines occurs
predominately at the ortho and para positions as has been
demonstrated with the dechlorination of anilines (Kuhn and
Suflita 1989b), though removal of meta halogens from this group
of compounds has also been demonstrated.
Reductive dehalogenation may require the induction of enzymes
responsible for dehalogenation. DeWeerd and Suflita (1990)
have demonstrated reductive dehalogenation of 3-
chlorobenzoate using cell-free extracts of an anaerobic bacterium.
The extracts exhibited the same substrate specificity as whole
cells. Rapid dehalogenation activity was found only in extracts
of cells cultured in the presence of the halogenated molecule,
indicating that the enzymes responsible required induction.
Dehalogenation was inhibited by sulfite, thiosulfate, and sulfide.
Dehalogenation activity was associated with the membrane
fraction and required a low potential electron donor. These
results suggest that a specific enzyme is made by the cells for
dehalogenation of selected halogenated substrates. Research
into the use of enzymes as a potential amendment to enhance
bioremediation should be encouraged.
Further evidence that reductive dehalogenation may depend
upon the induction of enzymes has been presented by Linkfield
et al. (1989). Acclimation periods prior to detectable
dehalogenation of halogenated benzoates in anaerobic lake
sediments ranged from 3 weeks to 6 months. These periods
were reproducible over time and among sampling sites and
characteristic of the specific benzoate compound tested. The
lengthy acclimation period appeared to represent an induction
phase in which little or no aryl dehalogenation was observed.
This was followed by an exponential increase in activity typical
of an enrichment response. Extremely low activities during the
early days of acclimation, coupled with the fact that
dehalogenation yields no carbon to support microbial growth,
suggests that slow continuous growth from time of the first
exposure of the chemical was not responsible forthe acclimation
period. The characteristic acclimation period for each chemical
also argues against nutritional deficiency, inhibitory environmental
conditions, or predation by protozoa or other microbial grazers
as the cause of the acclimation period. The reproducibility of the
findings with time and space and among replicates argues
against genetic changes as the explanation.
The removal of chloride or bromide from an aromatic molecule
proceeds easier when the ring also is substituted with electron
destabilizing groups, such as carboxy, hydroxy, or cyano groups
(Kuhn and Suflita 1989a). Other chemical groups attached to the
aromatic ring by nitrogen or oxygen bonds may have the same
effect on the reductive dehalogenation reaction. However,
recent research has shown that even highly chlorinated, poorly
water soluble aromatic hydrocarbons that do not contain polar
functional groups can also undergo reductive dehalogenation.
Hexachlorobenzene (HCB) has generally been considered
recalcitrant to microbial attack, particularly in the absence of
oxygen (Bouwer and McCarty 1984, Kuhn et al. 1985); however,
HCB was shown to degrade to tri- and dichlorobenzenes by
Fathepure et al. (1988). Brown et al. (1987) performed standard
thermochemical calculations of free-energy changes associated
with the oxidation of organic compounds (in this case, glucose)
coupled with the reduction of chlorobenzene compounds. The
reactions involving HCB and monochlorobenzene offered more
energy to anaerobic bacteria than the reduction of compounds
available naturally in anaerobic environments, such as sulfate
and carbon dioxide (Table 2). Also, more energy could be
obtained from the dehalogenation of hexachlorobenzene to
benzene than the dehalogenation of monochlorobenzene,
indicating that dehalogenation reactions are more likely to occur
with aromatic compounds containing many chloro groups since
they are more highly oxidized and more electronegative than
those containing fewer chloro groups.
Polychlorinated biphenyls (PCBs), commonly thought to be
resistant to biodegradative processes, have also been shown to
be susceptible to degradation by reductive dehalogenation
(Brown et al. 1987, Quensen et al. 1988). Brown et al. (1987)
suggest that dehalogenated products formed were less toxic
than the original PCS congeners and may possible be more
susceptible to oxidative biodegradation by aerobic bacteria.
Hydrolytic Dehalogenation of Aromatic Compounds
Hydrolytic dehalogenation represents a substitution reaction in
which a hydroxyl group replaces a halogen on an organic
molecule (Figure 1). In general, the anaerobic hydrolytic removal
of halogen substituents from homocyclic aromatic compounds is
rare (Kuhn and Suflita 1989a), but has been observed under
aerobic conditions. Also, the enzymes involved have been
shown to be active in reduced media, and some were inhibited
by oxygen (Marks et al. 1984, Thiele et al. 1988). This
transformation has been observed in anaerobic soil fora single
herbicide, flamprop-methyl; however no anaerobic bacteria were
isolated with the ability to catalyze this type of dehalogenation.
A hydrolytic defluorination product of the herbicide was identified
in anaerobic soil incubation studies (Roberts and Standen
1978).
Heterocyclic chloroaromatic compounds, such as chlorinated
triazine herbicides, tend to react more readily with hydroxy,
amino, or sulfhydryl groups than their homocyclic chemical
-------
Oxidant
Reduced Product
AG
(kcal/mol)
Molecular oxygen (O2)
Hexachlorobenzene (C..CL)
v b b'
Monochlorobenzene (C6H5CI)
Sulfate (SO42-)
Carbon dioxide (CO )
Water (H2O)
Benzene (C6H6)
Benzene (C6H6)
Reduced Sulfur
Methane (CH )
-676.10
-410.16
-369.50
-131.78
- 95.63
Table 2. Standard free-energy changes for the oxidation of glucose to CO2 and H2O using various oxidants (Brown et al. 1987).
counterparts. Hydrolytic dehalogenation is, therefore, the
preferred mechanism for removing halogens from hetero-cyclic
aromatic compounds under anaerobic conditions (Adrian and
Suflita 1989).
The hydrolysis of triazine herbicides to form dehalogenated and
less phytotoxic products has been known for many years (Paris
and Lewis 1973). However, there has been controversy over the
involvement of microorganisms in this process. Reactions with
reactive soil surfaces, such as clays and organic matter, appear
to be significant with regard to the rate of hydrolysis (Kuhn and
Suflita 1989a). Dechlorination ofs-triazines has been shown to
be catalyzed by microorganisms. This was demonstrated by
Cook and Huetter (1984, 1986). The organisms studied were
aerobic, but biotransformation of the herbicides did not require
molecular O2 and was functional under anaerobic conditions.
Dehalogenation of Nonaromatic Compounds
Dehalogenation of nonaromatic compounds, particularly
halogenated aliphatic chemicals, is generally better understood
than aryl dehalogenation reactions. The reductive processes of
hydrolysis and dehydrohalogenation have been identified as
anaerobic dehalogenation mechanisms (Figure 1).
In general, biologically mediated anaerobic dehalogenation of
nonaromatic compounds tends to be fasterthan dehalogenation
of aromatic compounds, does not require long adaptation times,
and does not exhibit a high degree of substrate specificity. Some
of these reactions also are not too sensitive to the presence of
oxygen and have been observed in aerobic incubation systems.
The greater variety of reaction mechanisms potentially available
to metabolize nonaromatic halogenated compounds in general
results in rendering these compounds more susceptible to
biodegradation than the haloaromatic compounds (Vogel et al.
1987, Kuhn and Suflita 1989a).
Dehalogenation has been demonstrated with many bacterial
species representing diverse genera. Mesophilicandthermophilic
methanogenic bacteria as well as some thermophilic clostridial
species may catalyze dehalogenation of some aliphatic
compounds. For example, metabolism of
hexachlorocyclohexanes by thermophilicclostridia was reported
by Sethunathan (1973). Dehalogenation reactions are also
sometimes heat resistant, suggesting that some reactions may
not be enzymatically mediated, and therefore not dependent on
intact microorganisms or microbial consortia. The dehalogenation
of nonaromatic compounds can be catalyzed by transition metal
complexes with or without the involvement of enzymes (Kuhn
and Suflita 1989a).
Reductive and Vicinal Dehalogenation of Nonaromatic
Compounds
If a nonaromatic carbon atom in a synthetic molecule is highly
halogenated, dehalogenation is more easily accomplished by
reductive, vicinal reductive or elimination reactions (Vogel et al.
1987). Compounds that have been demonstrated to be degraded
by reduction orvicinal reduction mechanisms are listed in Table
3.
Reductive and vicinal dehalogenation reactions are dependent
on the redox potential of the electron donor and acceptor. To be
thermodynamically feasible, the Eh of the electron accepting
reactant (dehalogenation) must be higherthan that of the electron
donating reactant. This requirement can limit the number of
available electron donors fordehalogenation of some compounds
(Castro et al. 1985, Vogel et al. 1987, Kuhn and Suflita 1989a).
For example, free ferrous iron (Fe(ll)) has a redox potential of +
0.77 V; but most of the halogenated alkanes and alkenes with
lower redox potentials will not react with this transition metal.
However, when Fe(l I) is in a complexed form, such as a porphyrin
or as ferredoxin, the redox potential is dramatically lowered, and
the reaction is possible (Kuhn and Suflita 1989a). As examples,
Fe(ll)deuteroporphin IX and cytochrome P-450 have redox
potentials of 0.00 V and -0.17 V, respectively.
Active transition metal complexes, which include complexes of
iron (Fe), cobalt (Co), nickel (Ni), and perhaps chromium (Cr)
and zinc (Zn), have redox potentials less than zero and can be
as low as -0.8 V for the cobalt complexed vitamin B.,2. The low
redox potentials of these electron donors allowfortheir reduction
-------
Class of Halogenated Nonaromatic Compound
Examples of Specific Compounds
Aliphatic Compounds
Alicyclic Compounds
Hexahalocyclohexanes
Tetrachloromethane (carbon
tetrachloride)
Trichloromethane (chloroform)
Dichloromethane (methylene chloride)
Chloromethane (methyl chloride)
Bromomethane (methyl bromide)
Trichloronitromethane (chloropicrin)
Hexachloroethane
Tetrachloroethene (perchloroethylene)
1,1,1-Trichloroethane
Trichlorethene (trichloroethylene)
1,2-Dichloroethane (ethylene
dichloride, EDC)
1,2-dibromoethane (ethylene
dibromide, EDB)
1,2-dibromo-3-chloropropane
(DBCP)
1,1,1-trichloro-2,2-bis
(p-chlorophenyl)ethane (DDT) (aliphatic
portion)
Toxaphene
Mi rex
Heptachlor
Lindane and its isomers
Table 3. Classes of halogenated nonaromatic compounds demonstrated to be susceptible to degradation by reductive dehalogenation
processes (Kuhn and Suflita 1989a).
to be coupled with dehalogenation of many nonaromatic
compounds having redox potentials which range from 0 to 1.2 V
(Vogeletal. 1987).
Highly halogenated aliphatic compounds have higher reduction
potentials than their lesser halogenated analogues; therefore,
more energy is released by their dehalogenation, indicating a
greater driving force for these reactions. In general, reductive
dehalogenation of tetra- and tri-halogenated carbon atoms is
easier than di- or monohalogenated congeners (Vogel et al.
1987).
In natural environments, Fe(ll) porphyrins (e.g., cytochromes),
Co complexes (e.g., vitamin B.,2), and Ni complexes (e.g., F-
430) are likely to be dominant in the reductive dehalogenation
process. Dead cells can release these stable transition metal
complexes which are then more available for participation in the
dehalogenation process. Such complexes are also active in
living cells, as was demonstrated with Pseudomonas putida by
Castro et al. (1985). Pseudomonas putida contains
Fe(ll)porphyrin bound to the cytochrome P-450 complex, but
movement of halogenated compounds across the bacterial
membrane and diffusion to the active iron centercan limit the rate
of dehalogenation.
Another potential reductant available for dehalogenation of
haloaliphatic compounds in natural environments is the flavin/
flavoprotein complex, which has been shown to mediate many
of the known reductive reactions of xenobiotic compounds in
laboratory studies (Esaac and Matsumura 1980). To date, no
studies have clearly demonstrated the environmental significance
of this reductant. Relative to other dehalogenation reaction
mechanisms, dehalogenation by vicinal reduction appears to be
more tolerant of oxidized conditions and may even be independent
of transition metals or metallo-organic complexes (Kuhn and
Suflita 1989a).
Dehydrohalogenation of Nonaromatic Compounds
Dehydrohalogenation is an elimination reaction in which two
groups are lost from adjacent carbon atoms so that a double
bond is formed, resulting in the release of a halogen and a proton
(HX) and the formation of an alkene (Figure 1). The rate of
dehalogenation is higher when additional chloride ions are
bonded to the carbon atom that loses its chloride ion substituent
(Vogel et al. 1987). Bromine atoms rather than chlorine atoms
are generally more readily eliminated by this reaction. Elimination
reactions can proceed spontaneously (1,1,1-trichloroethane;
1,2-dibromoethane) or can be catalyzed by microbial enzymes
such as the dechlorinase enzyme which is responsible for the
conversion of DDT to DDE—a dechlorination reaction involving
the aliphatic portion of the DDT molecule (Kuhn and Suflita
1989a).
Hydrolytic Dehalogenation
Hydrolysis, a substitution reaction in which one substituent on a
molecule is replaced by another, has been demonstrated with
-------
many aliphatic compounds. Hydrolysis is favored for carbon
atoms with only one or two halogens; however, hydrolytic
dehalogenation has been shown with higher chlorinated
compounds, such as 1,1,1-trichloroethane. This transformation
can be chemically or biologically catalyzed by methanogenic
mixed cultures and by a number of aerobic bacterial isolates.
Bromine loss tends to be favored compared to the corresponding
chlorinated compounds (Kuhn and Suflita 1989a).
Applications And Limitations of Reductive
Dehalogenation of Organic Halogenated Pollutants
The degradation of trichloroethylene (TCE), as shown in Figure
2, may be used to illustrate the potential effectiveness of the
reductive dehalogenation process to remove common pollutants
from the environment, as well as to present some of the cautions
that should be observed (Dragun 1988). TCE is an industrial
solvent used extensively for degreasing metal as well as in dry-
cleaning operations, organic synthesis, refrigerants, and
fumigants. Most septic tank cleaning fluids contain TCE (Craun
1984).
Also illustrated in Figure 2 are possible degradative pathways of
tetrachloroethylene (PCE) and 1,1,1 -trichloroethane (1,1,1 -TCA).
PCE is a solvent widely used in dry cleaning and degreasing
operations; 1,1,1 -TCA is used extensively as an industrial cleaner
and degreaser of metals, spot remover, adhesive, and vapor
pressure depressant (Craun 1984). These compounds have
relatively high water solubility (e.g., 1000 mg/l for TCE) and are
highly mobile in soils and aquifer materials and often are found
in ground waters. Since they are suspected carcinogens (Infante
and Tsongas 1982), they represent a threat to human health.
The degradative pathway for TCE (Dragun, 1988) can be
described as follows:
(1) TCE can undergo reductive dehalogenation, i.e., the
removal of one chloride atom (Cl) and the addition of
one hydrogen (H) atom. Three possible reaction
products can be formed: 1,1-dichloroethylene (1,1-
DCE), cis-1,2-dichloroethylene (c-1,2-DCE), and/or
trans-1,2-dichloroethylene (t-1,2-DCE).
(2) 1,1-DCE can undergo reductive dehalogenation to
form vinyl chloride, or its carbon-carbon double bond
can be reduced to form 1,1-dichloroethane (1,1-DCA).
(3) The two dichloroethylene compounds, c-1,2-DCE and
t-1,2-DCE can undergo reductive dehalogenation to
form vinyl chloride. Their carbon-carbon double bonds
can be reduced to form 1,2-dichloroethane (1,2-DCA).
(4) 1,1-DCA and 1,2-DCA can undergo
dehydrohalogenation to form vinyl chloride. These two
chemicals can also undergo reductive dehalogenation
to form chloroethane.
The degradation pathway of a single compound, TCE, can lead
to the production of six chlorinated volatile hydrocarbons. The
degradation of PCE can lead to the production of seven chlorinated
Figure 2. Transformation pathways for various chlorinated volatile hydrocarbons in soil systems (Drugun 1988).
-------
volatile hydrocarbons, while the degradation of 1,1,1-TCA can
lead to the production of four chlorinated hydrocarbons. Two of
the metabolic products formed, vinyl chloride and 1,1-DCA, have
been classified as a carcinogen and a probable carcinogen,
respectively (Vogel etal. 1987). The dichloroethylene products,
c-1,2-DCE and t-1,2-DCE, and vinyl chloride are also regulated
under the 1986 Safe Drinking Water Amendments (Freedman
and Gossett 1989). Vinyl chloride is the most persistent of the
compounds under anaerobic conditions, but can be rapidly
degraded under aerobic conditions (Hartsmans et al. 1985,
Fogel etal. 1986).
Management of a bioremediationsystemto accomplish treatment
of these compounds in a mannerto protect human health and the
environment should incorporate considerations of detoxification
as well as disappearance of the parent compounds.
Disappearance is not synonymous, however, with mineralization
to inorganic salts, carbon dioxide, and water. Partial degradation
of organic substrates can result in the production of metabolic
products that generate their own environmental and health
consequences. Such contaminants may be of more toxicological
concern than the parent compounds (Suflita et al. 1988).
Fathepure and Boyd (1988) recently suggested that in situ
dechlorination of PCE to TCE could be enhanced by stimulating
methanogenesis. They found that reductive dechlorination of
PCE occurred only during methanogenesis, and no dechlorination
was seen when methane production ceased. There was a clear
dependence of the extent of PCE dechlorination on the amount
ofmethanogenicsubstrate (methanol) consumed. Methanogenic
bacteria are present in a diversity of environmental habitats,
including those where chloroethylenes are commonly found as
contaminants (e.g., soils, ground waters, and aquifers near
landfills).
A bioremediation system for chlorinated ethylenes and ethanes
could consist of maintenance of an anaerobic environment,
followed by aeration to complete the degradation process after
anaerobic degradative processes have reduced the parent
compounds to acceptable levels. Recent research, however, by
Freedman and Gossett (1989) has shown that PCE and TCE can
be degraded to ethylene, a non-chlorinated environmentally
acceptable biotransformation product, under anaerobic
methanogenic conditions if an adequate supply of electron
donors was supplied to a mixed anaerobic enrichment culture.
Methanol was the most effective electron donor, although
hydrogen, formate, acetate, and glucose also served.
Ethylene is sparingly soluble in water and has not been associated
with any long-term toxicological problems (Autian 1980). It is
also a naturally occurring plant hormone. Since complete
conversion of VC to ethylene was not observed in the study, the
authors suggested that further research is required to determine
the concentration of electron donors required to complete the
conversion.
A major operational cost of this method of enhanced anaerobic
bioremediation will be the supply of electron donors. Alternatively,
means to channel more of the donors into the reductive
dechlorination process and less into methane production should
be investigated.
As proposed by Fathepure et al. (1988), a similar potential forthe
use of an anaerobic environment followed by an aerobic
environment, for mineralization and detoxification of halogenated
organic pollutants, is illustrated by the degradation of
hexachlorobenzene (HCB) (Figure 3). HCB is a fungicide used
as a seed coating for cereal crops. Two pathways of dechlorination
were proposed: (1) a major pathway in which 1,3,5-
trichlorobenzene (1,3,5-TCB) is formed via pentachlorobenzene
and 1,2,3,5-tetrachlorobenzene (1,2,3,5-TTCB); and (2) a minor
pathway in which dichlorobenzenes are formed via 1,2,4,5-
TTCBand 1,2,4-TCB.
The authors presented explanations for the existence of two
pathways. One is that there were two populations, each using a
different pathway. The other is that the products reflect the
distribution of reactive ring intermediates in which a chlorine,
between two other chlorines, was lost most readily and
dechlorination ceased when there are no adjacent chlorines as
with 1,3,5-TCB.
Reductive dechlorination appeared to occur in a stepwise fashion
until lower chlorinated compounds accumulated. Most of the
added HCB accumulated as 1,3,5-TCB, which remained
unchanged. Although metabolic products identified in this study
were not further utilized by the anaerobic sludge populations
used to elucidate the metabolic pathways, it is likely that they
would be degraded by aerobic organisms (Reineke and
Knackmuss 1984, deBontetal. 1986, Schraa etal. 1986, Spain
and Nishino 1987) or by facultative anaerobes possessing
dechlorinating activity (Tsuchiya and Yamaha 1984).
The U.S. Environmental Protection Agency is presently
sponsoring research to develop engineered microorganisms
capable of anaerobic reductive dehalogenation of organic
halogenated compounds (Palmer et al. 1989). Desulfomonile
tiedjei (DeWeerd et al, 1990), formerly known as DCB-1, is the
first obligate anaerobe known to accomplish reductive
dehalogenation. Results using this organism indicated that no
plasmid genes responsible for dehalogenating activity could be
detected. Therefore, in order to clone the gene or genes
responsible for the activity, a genomic library of the bacterial
chromosome is being constructed to isolate the dehalogenase
gene. The isolation ofthe gene would be greatly facilitated bythe
isolation and characterization ofthe requisite dehalogenase.
Summary
Bioremediation of soils and ground waters contaminated with
organic pollutants involves management of the contaminated
system to control and enhance biodegradation ofthe pollutants
present (Sims et al. 1989, Thomas and Ward 1989). Reductive
dehalogenation appears to be a potentially powerful process for
achieving bioremediation of a site contaminated with organic
halogenated pollutants, if mechanisms and pathways of
degradation are known and can be managed to achieve removal
ofthe compounds of interest as well as potentially toxic metabolic
degradation products.
-------
Cl
^
Cl
xCI
Cl
Cl
HCB
Cl
T
Cl
PCB
1,2,3,5-TTCB
1,3,5-TCB
Cl
.Cl
1,2-DCB
Cl
Cl
1,2,4,5-TTCB
Cl
Cl
1,2,4-TCB
Cl
Cl
1,4-DCB
1,3-DCB
T
ci
Figure 3. Proposed pathway for HCB dechlorination by an anerobic microbial community (Fathepure et al. 1988).
References
Adrian, N. R., and JM. Suflita. 1989. Reductive dehalogenation
of a nitrogen heterocyclic herbicide in anoxic aquifer slurries.
Appl. Environ. Microbiol. 56:292-294.
Allan, J. 1955. Loss of biological efficiency of cattle-dipping
wash containing benzene hexachloride. Nature (London)
175:1131-1132.
Autian, J. 1980. Plastics, p. 531-556. In: J. Doull, C.D. Klaassen,
and M.O. Amdur (eds.) Casarett and Doull's Toxicology.
Macmillan Publishing Co., Inc. New York, NY.
Bouwer, E.J., and P.L McCarty. 1984. Utilization rates of trace
halogenated organic compounds in acetate-grown biofilms.
Biotechnol. Bioeng. 27:1564-1571.
Brown, J.F., Jr., R.E. Wagner, H. Feng, D.L. Bedard, M.J.
Brennan, J.C. Carnahan, and R.J. May. 1987. Environmental
dechlorination of PCBs. Environ. Toxicol. Chem. 6:579-593.
Castro, C.E., R. S. Wade, and N.O. Belser. 1985.
Biodehalogenation: Reactions of cytochrome P-450 with
polyhalomethanes. Biochemistry 24:204-210.
Cook, A. M., and R. Huetter. 1984. Deethylsimazine: Bacterial
dechlorination, deamination, and complete degradation. J. Agric.
Food Chem. 32:581-585.
Cook, A. M., and R. Huetter. 1986. Ring dechlorination of
deethylsimazine by hydrolases from Rhodococcus corallinus .
FEMS Microbiol Letters 34:335-338.
Craun, G.F. 1984. Health aspects of groundwater pollution, pp.
135-179. In: G.Bitton and C.Gerba(eds.). GroundwaterPollution
Microbiology. John Wiley & Sons, New York, NY.
deBont, J.A.M., M.J.A.W. Vorage, S. Hartmans, and W.J.J. van
denTweel. 1986. Microbial degradation of 1,3-dichlorobenzene.
Appl. Environ. Microbiol. 52:677-680.
DeWeerd, K.A., J.M. Suflita, T. Linkfield, J.M. Tiedje, and P.H.
Pritchard. 1986. The relationship between reductive
dechlorination and other aryl substituent removal reactions
catalyzed by anaerobes. FEMS Microbiol. Ecol. 38:331-340.
DeWeerd, K.A., and J.M. Suflita. 1990. Anaerobic aryl
dehalogenation of halobenzoiates by cell extracts of
"Desulfomonile tiedjei". Appl. Environ. Microbiol. 56: in press
(out in the October issue).
DeWeerd, K.A., L. Mandelco, R.S. Tanner, C.R. Woese, and
J.M. Suflita. 1990. Desulfomonile tiedjei gen. nov. and sp. nov.,
a novel anaeroic, dehalogenating sulfate-reducing bacterium.
Arch. Microbiol. 154:23-30.
Dragun, J. 1988. The Soil Chemistry of Hazardous Materials.
Hazardous Materials Control Research Institute, Silver Spring,
MD.
10
-------
Esaac, E.G., andF. Matsumura. 1980. Metabolism of insecticides
by reductive systems. Pharmac. Ther. 9:1-26.
Fathepure, B.Z., and S.A. Boyd. 1988. Dependence of
tetrachloroethylene dechlorination on methanogenic substrate
consumption by Methanosarcina sp. strain DCM. Appl. Environ.
Microbiol. 54:2976-2980.
Fathepure, B.Z., J.M. Tiedje, and S.A. Boyd. 1988. Reductive
dechlorination of hexachlorobenzenetotri-and dichlorobenzenes
in anaerobic sewage sludge. Appl. Environ. Microbiol. 54:327-
330.
Fogel, M.M., A.R. Taddeo, and S. Fogel. 1986. Biodegradation
of chlorinated ethenes by a methane-utilizing mixed culture.
Appl. Environ. Microbiol. 51:720-724.
Freedman, D.L., and J.M. Gossett. 1989. Biological reductive
dechlorination of tetrachloroethylene and trichloroethylene to
ethylene under methanogenic conditions. Appl. Environ.
Microbiol. 55:2144-2151.
Genthner, B.R.S., W.A. Price, II, and H. P. Pritchard. 1989.
Anaerobic degradation of chloroaromatic compounds in aquatic
sediments undera variety of enrichment conditions. Appl. Environ.
Microbiol. 55:1466-1471.
Gibson, S.A., and J.M. Suflita. 1986. Extrapolation of
biodegradation results to groundwater aquifers: Reductive
dehalogenation of aromatic compounds. Appl. Environ. Microbiol.
52:681-688.
Hartsmans, S., J.A.M. de Bont, J. Tramper, and K.Ch.M.A.
Luyben. 1985. Bacterial degradation of vinyl chloride. Biotechnol.
Letters 7:383-388.
Hutzinger, O., and W. Verkamp. 1981. Xenobiotic chemicals
with pollution potential, pp. 3-46. In: T. Leisinger, A. M. Cook, R.
Mutter, and J. Nuesch (eds.). Microbial Degradation ofXenobiotic
and Recalcitrant Compounds. Academic Press, London, G.B.
Infante, P.F., andT.A. Tsongas. 1982. Mutagenicandoncogenic
effects of chloromethanes, chloroethanes, and halogenated
analogs of vinyl chloride. Environ. Sci. Res. 25:301-327.
King, G.M. 1988. Dehalogenation in marine sediments containing
natural sources of halophenols. Appl. Environ. Microbiol. 54:3079-
3085.
Kuhn, E.P., P.J. Colberg, J.L. Schnoor, O. Wanner, A.J.B.
Zehnder, and R.P. Schwarzenbach. 1985. Microbial
transformation of substituted benzenes during infiltration of river
water to ground water:laboratory column studies. Environ. Sci.
Technol. 19:961-968.
Kuhn, E. P., and J. M. Suflita. 1989a. Dehalogenation of pesticides
by anaerobic microorganisms in soils and groundwater - a
review, pp. 111-180. In: B. L. Sawhney and K. Brown (eds.)
Reactions and Movement of Organic Chemicals in Soils. Soil
Sci. Soc. America Special Publication No. 22. Soil Sci. Soc.
America, Inc. Madison, Wl.
Kuhn, E.P., and J.M. Suflita. 1989b. The sequential reductive
dehalogenation of chloroanilines by microorganisms from a
methanogenic aquifer. Environ. Sci. Technol. 23:848-852.
Linkfield, T.G., J.M. Suflita, and J.M. Tiedje. 1989.
Characterization of the acclimation period priorto the anaerobic
biodegradation of haloaromatic compounds. Appl. Environ.
Microbiol. 55:2773-2778.
Marks, T.S., A.R.W. Smith, and A.V. Quirk. 1984. Degradation of
4-chlorobenzoic acid by an Arthrobacter sp. Appl. Environ.
Microbiol. 48:1020-1025.
Mikesell, M.D., and S.A. Boyd. 1986. Complete reductive
dechlorination and mineralization of pentachlorophenol by
anaerobic microorganisms. Appl. Environ. Microbiol. 52:861-
865.
Palmer, D.T., T. G. Linkfield, J.B. Robinson, B.R.S. Genthner,
and G. E. Pierce. 1989. Determination and enhancement of
anaerobic dehalogenation: Degradation of chlorinated organics
in aqueous systems. EPA/600/S2-88/054, U.S. Environmental
Protection Agency, Cincinnati, OH.
Paris, D.F., and D.L. Lewis. 1973. Chemical and microbial
degradation of ten selected pesticides in aquatic systems.
Residue Rev. 45:95-124.
Quensen, J.F. Ill, J.M. Tiedje, and S.A. Boyd. 1988. Reductive
dechlorination of polychlorinated biphenyls by anaerobic
microorganisms from sediments. Science 242:752-754.
Reineke, W., and H.J. Knackmuss. 1984. Microbial metabolism
of haloaromatics: isolation and properties of a chlorobenzene-
degrading bacterium. Appl. Environ. Microbiol. 47:395-402.
Roberts, T.R., and M.E. Standen. 1978. Degradation of the
herbicide flamprop-methyl in soil under anaerobic conditions.
Pestic. Biochem. Physiol. 9:322-333.
Rochkind, M.L, J.W. Blackburn, andG.S. Sayler. 1986. Microbial
decomposition of chlorinated aromatic compounds. EPA/600/2-
86/090, U.S. Environmental Protection Agency, Cincinnati, OH.
Schraa, G., M.L. Boone, M.M. Jetten, A. R. W. van Neerven, P.J.
Colberg, and A.J.B. Zehnder. 1986. Degradation of 1,4-
dichlorobenzene by Alcaligenes sp. strain A175. Appl. Environ.
Microbiol. 52:1374-1381.
Sethunathan, N. 1973. Microbial degradation of insecticides in
flooded soil and in anaerobic cultures. Residue Rev. 47:143-
165.
Sims, J.L., R.C. Sims, and J.E. Matthews. 1989. Bioremediation
of contaminated soils. EPA/600/9-89/073, Robert S. Kerr
Environmental Research Laboratory, U.S. Environmental
Protection Agency, Ada, OK.
11
-------
Spain, J.C., and S.F. Nishino. 1987. Degradation of 1,4-
dichlorobenzene by Pseudomonas sp. Appl. Environ. Microbiol
53:1010-1019.
Struijs, J. and J.E. Rogers. 1989. Reductive dehalogenation of
dichloroanilines by anaerobic microorganisms in fresh and
dichlorophenol-acclimated pond sediment. Appl. Environ.
Microbiol. 55:2527-2531.
Suflita, J. M., A. Horowitz, D. R. Shelton, and J. M. Tiedje. 1982.
Dehalogenation: A novel pathway for anaerobic biodegradation
of haloaromatic compounds. Science 218:1115-1117.
Suflita, J.M., and G.D. Miller. 1985. Microbial metabolism of
chlorophenolic compounds in groundwater aquifers. Environ.
Toxicol. Chem. 4:751-758.
Suflita, J.M., S.A. Gibson, and R.E. Beeman. 1988. Anaerobic
biotransformation of pollutant chemicals in aquifers. J. Ind.
Microbiol. 3:179-194.
Thomas, J. M., and C. H. Ward. 1989. In situ biorestoration of
organic contaminants in the subsurface. Environ. Sci. Technol.
23:760-766.
Thiele, J., R. Muller, and F. Lingens. 1988. Enzymatic
dehalogenation of chlorinated nitroaromatic compounds. Appl.
Environ. Microbiol. 54:1199-1202.
Tiedje, J.M., and T. O. Stevens. 1987. The ecology of an
anaerobic dechlorinating consortium, pp. 3-14. In: G.S. Omenn
(ed.). Environmental Biotechnology: Reducing Risks from
Environmental Chemicals through Biotechnology. Plenum Press,
New York, NY.
Tsuchiya, T., and T. Yamaha. 1984. Reductive dechlorination of
1,2,4-trichlorobenzene by Staphylococcus epidermis isolated
from intestinal contents of rats. Agric. Biol. Chem. 48:1545-
1550.
Vogel, T. M., C. S. Griddle, and P. L. McCarty. 1987.
Transformations of halogenated aliphatic compounds. Environ.
Sci. Technol. 21:722-736.
Zeyer, J., and P.C. Kearney. 1982. Microbial degradation of
para-chloroaniline as sole carbon and nitrogen source. Pestic.
Biochem. Physiol. 17:215-223.
Zeyer, J., A. Wasserfallen, and K.N. Timmis. 1985. Microbial
mineralization of ring-substituted anilines through an ortho-
cleavage pathway. Appl. Environ. Microbiol. 50:447-453.
12
------- |