1

EPA/540/4-90/054

January 1991

Robert S. Kerr Environmental

Research Laboratory

Ada, OK

Superfund Technology Support Center for Ground Water

Reductive Dehalogenation of Organic

Contaminants in Soils and Ground Water

Judith L. Sims, Joseph M. Suflita, and Hugh H. Russell

United States

Environmental Protection

Agency

Office of Solid Waste

and Emergency

Response

Office of

Research and

Development

EPA Ground Water Issue

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 their toxicity 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,

that there 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
2

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 oxidationreduction 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

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 from these 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 et al. 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
3

Figure 1. Examples of anaerobic dehalogenation mechanisms for aromatic and nonaromatic pesticides (Kuhn and Suflita, 1989a)

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

R

H

R

X

Reduction

X- 2e- + H+

Anaerobic Dehalogenation Mechanisms

Hydrolysis

X

N

N

N

OH

N

N

N

N

OH- XAromatic Compounds

Nonaromatic Compounds

Reduction

(Hydrogenolysis)

X- 2e- + H+

Hydrolysis

(Substitution)

OH- XC X C OH

Vicinal Reduction

(Dihalo-Elimination)

X- 2eC C C

X

C

X

C

X

C

H

Dehydrohalogenation C C

HX

C X C H
4

Class of Halogenated Aromatic Compounds Examples of Specific Compounds

 Carboxylated Benzenes Amiben

Dicamba

2,3,6-trichlorobenzoate

Oxygen-Substituted Benzenes Pentachlorophenol

Chlorinated phenoxyacetates (e.g.,

 2,4-D, 2,4,5-T

Halogenated diphenyl ether herbicides

 (e.g., chloronitrofen)

Nitrogen-Substituted Benzenes 3,4-Dihalogenated aromatic compounds

 (diuron, DCPU, linuron, DCIPC,

 propanil)

Pentachloronitrobenzene

Cyano-Substituted Benzenes 2,4,5,6-tetrachloroisophthalonitrile

 (TPN)

Methylene-substituted Benzenes Benthiocarb

Chlorinated Benzenes Hexachlorobenzene

Polychlorinated biphenyls Araclors (commercial PCB products)

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 persistant 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 et al. 1982, Genthner et al.

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

Table 1. Classes of halogenated aromatic compounds demonstrated to be susceptible to degradation by reductive dehalogenation

processes (Kuhn and Suflita 1989a).

product is methane (Dragun 1988), appear to be optimal for this

type of biotransformation. Genthner et al. (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 inhibition of methanogenesis or inhibition

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
5

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 multihalogenated 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 for the 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 PCB 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 for a 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
6

Oxidant Reduced Product ∆G

(kcal/mol)

Molecular oxygen (O2) Water (H2

O) -676.10

Hexachlorobenzene (C6Cl6) Benzene (C6H6) -410.16

Monochlorobenzene (C6H5Cl) Benzene (C6H6) -369.50

Sulfate (SO4

2-) Reduced Sulfur -131.78

Carbon dioxide (CO2) Methane (CH4) - 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 of s -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 faster than 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. Mesophilic and thermophilic

methanogenic bacteria as well as some thermophilic clostridial

species may catalyze dehalogenation of some aliphatic

compounds. For example, metabolism of

hexachlorocyclohexanes by thermophilic clostridia 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 or vicinal 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 higher than that of the electron

donating reactant. This requirement can limit the number of

available electron donors for dehalogenation of some compounds

(Castro et al. 1985, Vogel et al. 1987, Kuhn and Suflita 1989a).

For example, free ferrous iron (Fe(II)) 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(II) 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(II)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 B1

2. The low

redox potentials of these electron donors allow for their reduction
7

Class of Halogenated Nonaromatic Compound Examples of Specific Compounds

Aliphatic Compounds 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)

Alicyclic Compounds Toxaphene

Mirex

Heptachlor

Hexahalocyclohexanes 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

(Vogel et al. 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(II) porphyrins (e.g., cytochromes),

Co complexes (e.g., vitamin B12), and Ni complexes (e.g., F430) 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(II)porphyrin bound to the cytochrome P-450 complex, but

movement of halogenated compounds across the bacterial

membrane and diffusion to the active iron center can 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
8

Figure 2. Transformation pathways for various chlorinated volatile hydrocarbons in soil systems (Drugun 1988).

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 drycleaning 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

1,1,1-TCA

PCE

1,1,-DCA

1,1,-DCE

TCE

c-1,2-DCE

t-1,2,-DCE

1,2-DCA

Vinyl

Chloride Chloroethane
9

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 et al. 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 et al. 1986).

Management of a bioremediation system to accomplish treatment

of these compounds in a manner to 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

of methanogenic substrate (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 for the

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-

TTCB and 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, deBont et al. 1986, Schraa et al. 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 of the gene would be greatly facilitated by the

isolation and characterization of the requisite dehalogenase.

Summary

Bioremediation of soils and ground waters contaminated with

organic pollutants involves management of the contaminated

system to control and enhance biodegradation of the 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

of the compounds of interest as well as potentially toxic metabolic

degradation products.
10

References

Adrian, N. R., and J.M. 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.

Figure 3. Proposed pathway for HCB dechlorination by an anerobic microbial community (Fathepure et al. 1988).

Cl

Cl

Cl

Cl

Cl

Cl

HCB

Cl

Cl

Cl

Cl

Cl

PCB

Cl

Cl

Cl

Cl

1,2,3,5-TTCB

Cl

Cl

Cl

1,3,5-TCB

Cl

Cl

Cl

Cl

1,2,4,5-TTCB

Cl

Cl

Cl

1,2,4-TCB

Cl

Cl

1,3-DCB

Cl

Cl

1,4-DCB

Cl

Cl

1,2-DCB

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.). Groundwater Pollution

Microbiology. John Wiley & Sons, New York, NY.

deBont, J.A.M., M.J.A.W. Vorage, S. Hartmans, and W.J.J. van

den Tweel. 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.
11

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 prior to 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. III, 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 chlorobenzenedegrading 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, and G.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.

Esaac, E.G., and F. 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 hexachlorobenzene to tri- 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 under a 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.

Hutter, and J. Nuesch (eds.). Microbial Degradation of Xenobiotic

and Recalcitrant Compounds. Academic Press, London, G.B.

Infante, P.F., and T.A. Tsongas. 1982. Mutagenic and oncogenic

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, WI.
12

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. Criddle, 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 orthocleavage pathway. Appl. Environ. Microbiol. 50:447-453.