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4 Early Work on Reaction Pathways and Organisms Involved

4 Early Work on Reaction Pathways and Organisms Involved

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P.L. McCarty

halogens, ease and speed of transformation tended to be best with brominated, followed with chlorinated, and then iodinated compounds. Fluorinated compounds

have been the most difficult to degrade, indeed fluorinated compounds are among

the most persistent synthetic organic chemicals produced, either under aerobic or

anaerobic conditions.

4.5 Oxidative Transformation of Chlorinated Solvents

While oxidative transformation of halogenated compounds is not the subject of this book, it is worth mentioning such reactions briefly as they are a significant alternative to reductive dehalogenation. Wilson and Wilson (1985), EPA

scientists, reported on the aerobic cometabolism of TCE by methanotrophic organisms. Because of Stanford’s growing experience with pilot groundwater studies,

EPA sought Stanford’s interest in pursuing a field pilot study of this process for

destroying TCE. This was the beginning of extensive laboratory and field research

on aerobic cometabolism, leading to two successful full-scale evaluations of aerobic TCE cometabolism. Methanotrophs use methane monooxygenase to initiate

the conversion of methane to methanol, which they then oxidize to obtain energy

for growth. Methane monooxygenase had previously been reported to fortuitously

oxidize halogenated methanes (Haber et al. 1983), and this report stimulated the

Wilsons to find if longer chain halogens might be transformed by this reaction.

They found that TCE could be oxidized by methane monooxygenase to the end

product CO2. This was an exciting finding as it avoided the formation of VC.

Through extensive laboratory and field studies, it was found that cometabolism of TCE resulted in the formation of TCE epoxide, which was lethal to the

methanotrophs. However, PCE was not cometabolized by methanotrophs or by

other oxygenase-producing organisms. We found that at low TCE concentrations,

the growth of methanotrophs on methane exceeded their death rate from TCE

epoxide formation, but at TCE concentrations above about 1 mg/L, the death rate

exceeded the growth rate, and thus the process would fail (Alvarez-Cohen and

McCarty 1991; Anderson and McCarty 1996). The oxidation products formed

from CF (Alvarez-Cohen and McCarty 1991) and 1,1-DCE (Dolan and McCarty

1995) through cometabolism were even more lethal. Also, the rate of TCE degradation by oxygenase was low, and the removal process was slow. However, methanotrophs were found to be quite efficient in the cometabolism of cis-DCE and

VC, intermediates often found from reductive dehalogenation at PCE and TCEcontaminated sites.

In later studies, toluene was found to be a much more effective substrate than

methane for TCE cometabolism (Hopkins and McCarty 1995). Applications of

this substrate to a field site at Edwards Air Force Base in Southern California provided successful demonstrations of the process, where TCE concentrations ranged

below 1 mg/L in an aquifer that had essentially no other organic contaminants

and was aerobic (McCarty et al. 1998; Gandhi et al. 2002). However, there was

4  Discovery of Organohalide-Respiring Processes …


concern by some over the use of toluene as a substrate, even though the remaining

concentrations were orders of magnitude lower than health or ascetic concentrations. Also, to be effective, toluene had to be continuously added to the aquifer,

and this involves what has been considered to be high operation and maintenance

costs. Unfortunately, little interest in applying this effective process has been demonstrated by practitioners, even though it is quite a practical approach that also

renders the aquifer waters aerobic.

4.6 Organohalide Respiration

All early studies on reductive dehalogenation provided no suggestion that organisms were obtaining energy for growth from these reactions. As with aerobic

studies in which oxygenases were implicated in the cometabolism of TCE and

other chlorinated solvents, cometabolism was generally assumed to be associated with reductive dehalogenation as well. However, this belief changed in 1986

when Dolfing and Tiedje reported on a three-organism anaerobic symbiotic conversion of 3-chlorobenzoate (Dolfing and Tiedje 1986, 1987). A new organism,

strain DCB-1, was found to reductively remove the chlorine atom from the benzene ring while replacing it with hydrogen from H2 produced by another organism

that anaerobically oxidized benzoate into hydrogen, acetate, and carbon dioxide.

The third organism converted the extra H2 plus carbon dioxide into methane.

The important finding was that strain DCB-1 obtained energy for growth from

the reductive dehalogenation of 3-chlorobenzoate. This was the first observation

of energy yield from reductive dehalogenation. The DCB-1 isolate now has the

generic name Desulfomonile tiedjei.

Several years later, the growth of Dehalobacter restrictus on energy obtained

from reductive dehalogenation of PCE and TCE leading to the formation of cisDCE was reported by Holliger et al. (1993, 1998). This observation changed the

outlook for reductive dehalogenation of chlorinated solvents, suggesting that the

anaerobic process might be carried out much more efficiently than aerobic cometabolism as large excesses of primary substrate should not be required to drive the

reaction. A broader search for organisms that could carry out reductive dehalogenation of all chlorinated aliphatic compounds was then begun.

4.7 Reductive Dehalogenation of Vinyl Chloride and the

Associated Genus, Dehalococcoides

In 1989, a major breakthrough came with the report by Freedman and Gossett

(1989) of an anaerobic mixed culture that reductively dehalogenated PCE through

VC into ethene. That resulted in a strong re-interest among practitioners in

the anaerobic process. Studies were begun by many to exploit this new finding.


P.L. McCarty

DiStefano et al. (1991) demonstrated, with an enrichment culture grown on methanol, the complete transformation of PCE to ethene without the involvement of

methanogens, thus confirming that methanogens were not involved in reductive

dehalogenation. They further demonstrated that the reductive process occurred

only when H2 was available as the donor substrate (DiStefano et al. 1992).

A field study at the time allowed us to demonstrate that complete anaerobic

conversion of TCE to ethene could also occur naturally. This study came about

through a venture into a possible application of aerobic cometabolism to treat a

TCE-contaminated aquifer in St. Joseph, Michigan, that had been brought to our

attention by John Wilson, an employee of the EPA Laboratory in Ada, Oklahoma.

From a few sample wells, the downgradient presence of TCE and its dechlorination products cis-DCE and VC had been found, compounds that would be readily susceptible to aerobic cometabolism using methane as a substrate. A treatment

strategy was developed (McCarty et al. 1991), and a detailed site characterization was carried out by EPA in preparation for a field application of the process

(Semprini et al. 1995). However, the results of the detailed characterization were

unexpected, and the actual concentrations of TCE and other contaminants were an

order of magnitude above those observed from the earlier limited investigation. It

was apparent, first, that the TCE concentrations were well above the toxic concentrations for aerobic cometabolism to be applied, and second, that complete

transformation of TCE to ethene was abundant and evident. The organic substrate

that was the source of the H2 necessary for reductive dehalogenation appeared to

come from a factory dump that had been leaching organics into the groundwater

for many years, forming an organic plume that intersected with the downgradient

spill of TCE, thereby slowly producing through organic fermentation the H2 that

was needed for in situ remediation to occur. While aerobic cometabolism was now

known not to be a viable answer at this site, the site became one of the first where

“natural” attenuation through reductive dehalogenation was found to be an acceptable remediation alternative. Complete anaerobic transformation of high concentrations of TCE to ethene in the field was demonstrated not only to be possible, but

was actually occurring on its own. Many such sites have now been found.

While many organisms had been identified that were capable of converting

PCE and TCE to cis-DCE, isolating an organism capable of converting VC to

ethene was elusive. The breakthrough came from the Cornell group with the publication in 1997 of the isolation of an organism (strain 195) that could reductively

dehalogenate PCE all the way through to ethene (Maymó-Gatell et al. 1997).

Hydrogen was the only substrate strain 195 was capable of using, and was oxidized through reductive dehalogenation, using the chlorinated compounds as electron acceptors. The organism’s 16S ribosomal RNA gene sequence did not cluster

it with any of the known phylogenetic lines, thus the authors suggested naming

the organism “Dehalococcoides ethenogenes” strain 195. While complete conversion of PCE to ethene was possible by this organism, growth occurred only while

dechlorinating PCE, TCE, and cis-DCE, but VC dehalogenation appeared to be

only through cometabolism, which was relatively slow.

4  Discovery of Organohalide-Respiring Processes …


Another Dehalococcoides organism (strain CBDB1) was soon after isolated by

Adrian et al. (2000) that could partially reductively dehalogenate a range of triand tetra-chlorobenzenes, but not PCE, TCE, DCEs, nor VC. Hendrickson et al.

(2002) analyzed samples from 24 chloroethene-dechlorinating field sites throughout North America and Europe to determine the presence of the Dehalococcoides

group using PCR analysis, and found Dehalococcoides to be present only at

the 21 sites where complete dechlorination to ethene was occurring, but not at

three others were dechlorination stopped at cis-DCE. While to date, organisms

from many different genera can dehalogenate PCE and TCE to cis-DCE, only

Dehalococcoides appears capable at this time of carrying the process all the way

through to ethene.

Based upon small differences in 16S rRNA gene sequences, Hendrickson

et al. (2002) noted three different clades of Dehalococcoides, which he designated as the Cornell group (strain 195), the Pinellas group (strains FL2 and

CBDB1), and the Victoria group (strain vic). We at Stanford had been working

with samples from the Victoria, TX, site for several years (Rosner et al. 1997;

Yang and McCarty 1998; Haston and McCarty 1999) and eventually found from

VC enrichment cultures that growth of the Victoria strain was occurring on VC

reductive dehalogenation (Cupples et al. 2003), which was different from the

finding that strain 195 dehalogenated VC cometabolically only. About the same

time, He et al. (2003) isolated a new Dehalococcoides strain called BAV1, reporting that it also obtained growth from VC reduction. Somewhat differently, the

Victoria strain obtained growth also from the reduction of TCE and cis-DCE,

while BAV1 obtained it from all DCE isomers, but not TCE. Thus, the different

Dehalococcoides strains all use hydrogen as their primary electron donor, but use

different chlorinated compounds as their electron acceptor in energy metabolism.

Also found was that the VC reductase in stain Victoria termed VcrA (Müller et al.

2004) differed from that in BAV1, which was termed BvcA (Krajmalnik-Brown

et al. 2004).

Subsequently, additional Dehalococcoides strains have been isolated, strain

FL2 that obtains energy for growth using TCE and both cis-DCE and trans-DCE

as electron acceptors (He et al. 2005), and strain GT that obtains energy while

using TCE, cis-DCE, and VC as electron acceptors for growth (Sung et al. 2006).

The genomes of these six Dehalococcoides strains have all been sequenced and

the differences and similarities between them have been characterized (Löffler

et al. 2013). Interesting is that the various Dehalococcoides strains have multiple

and different dehalogenation genes, but the function of many of them is not yet

known. They all use hydrogen as the electron donor and organohalides as electron acceptors in energy metabolism. They are widespread throughout the world.

Just what they were using as electron acceptors in energy metabolism before

humans evolved to supply them with synthetic halogenated compounds is not

known. While many other species are capable of reductive dehalogenation, it is

the Dehalococcoides strains that have often been the ones capable and necessary

for completing the process, allowing anaerobic dehalogenation to become the


P.L. McCarty

mainstream biological processes for in situ remediation of many of the organohalide-contaminated sites.

4.8 Summary Statement

The many bacterial species that are now known to reductively transform organohalides, either through energy metabolism or cometabolism, have been greatly

beneficial for ridding the environment of numerous toxic organic compounds that

humans have synthesized. Many such compounds tend to resist aerobic biodegradation, and thus the finding of anaerobic organisms in particular that can use

organohalides as electron acceptors has been most fortunate. Such anaerobic degradation requires an electron donor, that for some species may be an organic compound such as acetate, but most often, it is H2 that is preferred or necessary. When

conditions are right and the necessary electron donor is present, complete degradation may occur. But most often, sufficient suitable electron donor is not naturally

present, and thus must be added. Adding H2 itself presents difficulties, particularly

since elevated levels of H2 presents a highly attractive energy source for many

other organisms, such as homoacetogens, sulfate and Fe(III) reducers, and methanogens, all of which tend to be ubiquitous in the environment. Dehalogenating

organisms may dominate if H2 is kept at very low nanomolar concentrations,

which is difficult to achieve when H2 is added directly to groundwater, the most

commonly contaminated environment. Nevertheless, such low H2 concentrations

can be achieved by adding a slowly biodegradable organic such as propionate,

which is oxidized anaerobically to H2 and acetate. Carrying out an effective in

situ bioremediation system for organohalide compounds is often an engineering

challenge, but this challenge is now being effectively met by careful application

of the extensive knowledge about dehalogenating organisms that has been gained

through the research conducted by numerous investigators over the past several

decades, many of whom are authors or coauthors of the chapters in this book. I am

most honored to be among them.


Adrian L, Szewzyk U, Wecke J, Görisch H (2000) Bacterial dehalorespiration with chlorinated

benzenes. Nature 408:580–583

Alvarez-Cohen L, McCarty PL (1991) Product toxicity and cometabolic competitive inhibition

modeling of chloroform and trichloroethylene by methanotrophic resting cells. Appl Environ

Microbiol 57(4):1031–1037

Anderson J, McCarty PL (1996) Effect of three chlorinated ethenes on growth rates for a methanotrophic mixed culture. Environ Sci Technol 30(12):3517–3524

Bouwer EJ, McCarty PL (1983a) Transformations of 1- and 2-carbon halogenated aliphatic

organic compounds under methanogenic conditions. Appl Environ Microbiol 45:1286–1294

Bouwer EJ, McCarty PL (1983b) Transformations of halogenated organic compounds under denitrification conditions. Appl Environ Microbiol 45:1295–1299

4  Discovery of Organohalide-Respiring Processes …


Bouwer EJ, Rittmann BE, McCarty PL (1981) Anaerobic degradation of halogenated 1- and

2-carbon organic compounds. Environ Sci Technol 15(5):596–599

Cupples AM, Spormann AM, McCarty PL (2003) Growth of a dehalococcoides-like microorganism on vinyl chloride and cis-dichloroethene as electron acceptors as determined by competitive PCR. Appl Environ Microbiol 69(2):953–959

DiStefano TD, Gossett JM, Zinder SH (1991) Reductive dechlorination of high concentrations of

tetrachloroethene to ethene by an anaerobic enrichment culture in the absence of methanogenesis. Appl Environ Microbiol 57(8):2287–2292

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Dolan ME, McCarty PL (1995) Methanotrophic chloroethene transformation capacities and

1,1-dichloroethene transformation product toxicity. Environ Sci Technol 29(11):2741–2747

Dolfing J, Tiedje JM (1986) Hydrogen cycling in a three-tiered food web growing on the methanogenic conversion of 3-chlorobenzoate. FEMS Microbiol Ecol 38:293–298

Dolfing J, Tiedje JM (1987) Growth yield increase linked to reductive dechlorination in a defined

3-chlorobenzoate degrading methanogenic coculture. Arch Microbiol 149:102–105

Freedman DL, Gossett JM (1989) Biological reductive dechlorination of tetrachloroethylene

and trichloroethylene to ethylene under methanogenic conditions. Appl Environ Microbiol


Gandhi RK, Hopkins GD, Goltz MN, Gorelick SM, McCarty PL (2002) Full-scale demonstration of in situ cometabolic biodegradation of trichloroethylene in groundwater - 2.

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38(4) 11.1–11.19

Haber CL, Allen LN, Zhao S, Hanson RS (1983) Methylotrophic bacteria: biochemical diversity

and genetics. Science 1147–1153

Haston ZC, McCarty PL (1999) Chlorinated ethene half-velocity coefficients (K-s) for reductive

dehalogenation. Environ Sci Technol 33(2):223–226

He JZ, Ritalahti KM, Aiello MR, Löffler FE (2003) Complete detoxification of vinyl chloride by

an anaerobic enrichment culture and identification of the reductively dechlorinating population as a Dehalococcoides species. Appl Environ Microbiol 69(2):996–1003

He J, Sung Y, Krajmalnik-Brown R, Ritalahti KM, Löffler FE (2005) Isolation and characterization of Dehalococcoides sp strain FL2, a trichloroethene (TCE)- and 1,2-dichloroethenerespiring anaerobe. Environ Microbiol 7(9):1442–1450

Hendrickson ER, Payne JA, Young RM, Starr MG, Perry MP, Fahnestock S, Ellis DE, Ebersole

RC (2002) Molecular analysis of Dehalococcoides 16S ribosomal DNA from chloroethene-contaminated sites throughout north America and Europe. Appl Environ Microbiol


Hill DW, McCarty PL (1967) Anaerobic degradation of selected chlorinated hydrocarbon pesticides. J Water Pollut Control Fed 39:1259–1277

Holliger C, Schraa G, Stams AJM, Zehnder AJB (1993) A highly purified enrichment culture

couples the reductive dechlorination of tetrachloroethene to growth. Appl Environ Microbiol


Holliger C, Hahn D, Harmsen H, Ludwig W, Schumacher W, Tindall B, Vazquez F, Weiss N,

Zehnder AJB (1998) Dehalobacter restrictus gen. nov. and sp. nov., a strictly anaerobic bacterium that reductively dechlorinates tetra- and trichloroethene in an anaerobic respiration.

Arch Microbiol 169(4):313–321

Holliger C, Wohlfarth G, Diekert G (1999) Reductive dechlorination in the energy metabolism of

anaerobic bacteria. FEMS Microbiol Rev 22:383–398

Hopkins GD, McCarty PL (1995) Field-evaluation of in-situ aerobic cometabolism of trichloroethylene and 3 dichloroethylene isomers using phenol and toluene as the primary substrates.

Environ Sci Technol 29(6):1628–1637

Horvath RS (1972) Microbial co-metabolism and the degradation of organic compounds in

nature. Bacteriological Rev 36(2):146–155


P.L. McCarty

Krajmalnik-Brown R, Holscher T, Thomson IN, Saunders FM, Ritalahti KM, Loffler FE (2004)

Genetic identification of a putative vinyl chloride reductase in Dehalococcoides sp strain

BAV1. Appl Environ Microbiol 70(10):6347–6351

Löffler FE, Yan JL, Ritalahti KM, Adrian L, Edwards EA, Konstantinidis KT, Müller JA,

Fullerton H, Zinder SH, Spormann AM (2013) Dehalococcoides mccartyi gen. nov., sp.

nov., obligately organohalide-respiring anaerobic bacteria relevant to halogen cycling

and bioremediation, belong to a novel baterial class, Dehalococcoidia classis nov., order

Dehalococcoidales ord nov and family Dehalococcoidaceae fam nov., within the phylum

Chloroflexi. Int J Syst Evol Microbiol 63:625–635

Maymò-Gatell X, Chien YT, Gossett JM, Zinder SH (1997) Isolation of a bacterium that reductively dechlorinates tetrachloroethene to ethene. Science 276:1568–1571

McCarty PL, Semprini L, Dolan ME, Harmon TC, Tiedeman C, Gorelick SM (1991) In Situ

methanotrophic bioremediation for contaminated groundwater at St. Joseph, Michigan. In:

Hinchee RE, Olfenbuttel RF (eds) On-site bioreclamation, processes for xenobiotic and

hydrocarbon treatment. Butterworth-Heinemann, Boston, pp 16–40

McCarty PL, Goltz MN, Hopkins GD, Dolan ME, Allan JP, Kawakami BT, Carrothers TJ (1998)

Full-scale evaluation of in-situ cometabolic degradation of trichloroethylene in groundwater

through toluene injection. Environ Sci Technol 32(1):88–100

Mohn WW, Tiedje JM (1992) Microbial reductive dehalogenation. Microbiol Rev 56(3):482–507

Müller JA, Rosner BM, von Abendroth G, Meshulam-Simon G, McCarty PL, Spormann

AM (2004) Molecular identification of the catabolic vinyl chloride reductase from

Dehalococcoides sp. strain VS and its environmental distribution. Appl Environ Microbiol


NAS (1978) Nofluorinated halomethanes in the evironment, Washington D.C, pp 36–51

Parsons F, Wood PR, DeMarco J (1984) Transformations of tetrachloroethylene and trichloroethylene in microcosms and groundwater. J Am Water Works Assoc 76:56–59

Roberts PV, Reinhard M, McCarty PL (1980) Organic contaminant behavior during groundwater

recharge. J Water Pollut Control Fed 52:161–172

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groundwater recharge in the Palo Alto Baylands. Water Res 16:1025–1035

Rosner BM, McCarty PL, Spormann AM (1997) In vitro studies on reductive vinyl chloride

dehalogenation by an anaerobic mixed culture. Appl Environ Microbiol 63(11):4139–4144

Semprini L, Kitanidis PK, Kampbell DH, Wilson JT (1995) Anaerobic transformation of chlorinated aliphatic-hydrocarbons in a sand aquifer based on spatial chemical-distributions. Water

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Sung Y, Ritalahti KM, Apkarian RP, Loffler FE (2006) Quantitative PCR confirms purity of

strain GT, a novel trichloroethene-to-ethene-respiring Dehalococcoides isolate. Appl Environ

Microbiol 72(3):1980–1987

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dechlorination of PCBs. Biodegradation 4(4):231–240

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dichloroethylene, vinyl chloride, and carbon dioxide under methanogenic conditions. Appl

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Yang YR, McCarty PL (1998) Competition for hydrogen within a chlorinated solvent dehalogenating anaerobic mixed culture. Environ Sci Technol 32:3591–3597

Chapter 5

Overview of Known Organohalide-Respiring

Bacteria—Phylogenetic Diversity

and Environmental Distribution

Siavash Atashgahi, Yue Lu and Hauke Smidt

Abstract  To date, organohalide respiration (OHR) has been restricted to the bacterial domain of life. Known organohalide-respiring bacteria (OHRB) are spread

among several phyla comprising both Gram-positive and Gram-negative bacteria.

As a unique trait, OHRB benefit from reductive dehalogenase enzymes enabling

them to use different organohalides as terminal electron acceptors and occupy a

wide range of terrestrial and aquatic environments. This chapter comprises three

sections: First, we give an overview of phylogeny of known OHRB and briefly

discuss physiological and genetic characteristics of each group. Second, the environmental distribution of OHRB is presented. Owing to the application of molecular diagnostic approaches, OHRB are being increasingly detected not only from

organohalide-contaminated groundwaters and sediments but also from pristine

environments, including deep oceanic sediments and soils that are ample sources

of naturally occurring organohalides. Finally, we highlight important factors that

impact the ecology of OHRB and their interaction with other microbial guilds.

Siavash Atashgahi and Yue Lu equally contributed.

S. Atashgahi · Y. Lu · H. Smidt (*) 

Laboratory of Microbiology, Wageningen University,

Stippeneng 4, 6708 WE Wageningen, The Netherlands

e-mail: hauke.smidt@wur.nl

S. Atashgahi

e-mail: siavash.atashgahi@wur.nl

Y. Lu

e-mail: yue.lu@wur.nl

© Springer-Verlag Berlin Heidelberg 2016

L. Adrian and F.E. Löffler (eds.), Organohalide-Respiring Bacteria,

DOI 10.1007/978-3-662-49875-0_5



S. Atashgahi et al.

5.1 Introduction

The vast functional diversity of microorganisms and their metabolic capabilities

have made them successful mediators of electron liberation from the oxidation

of inorganic and organic matter coupled to the reduction of a wide array of inorganic and organic electron acceptors including organohalides (Leys et al. 2013).

Although most well-known representatives of organohalides are considered to be

man-made products of industrial origin, thousands of naturally occurring organohalides have been reported from geogenic and biogenic sources (Gribble 2010).

Numerous reports exist on natural production of organohalide compounds from

biogenic sources such as a broad range of seaweeds, sponges, terrestrial plants,

fungi as well as through geogenic processes such as volcanic activity, forest fires

and other geothermal processes, some of which predate the industrialization era

(see Chap. 2). Such ancient natural production of organohalides might have contributed to the development of biochemical strategies capable of unlocking the

chemically stable carbon–halogen bond in organohalides. This is particularly

important for non-oxygenolytic dehalogenation processes that probably have

developed in the originally oxygen-free atmosphere on Earth. Taking advantage of

organohalides as thermodynamically favourable electron acceptors under anoxic

conditions, reductive dehalogenation is used as a terminal electron-accepting process by organohalide-respiring bacteria (OHRB). These microbes have greatly

contributed to global cycling of halogens by breathing (rather) toxic organohalides

and preventing their accumulation in the environment. Exploiting reductive dehalogenases (RDase in case of functionally characterized enzymes and RdhA for yet

uncharacterized reductive dehalogenases predicted from genomes and molecular

surveys, see also Chap. 16) dedicated to organohalide respiration (OHR), OHRB

occupy a wide range of niches/environments. Hence, understanding factors governing their evolution, distribution and ecology will help to unravel their role in

the fate of organohalides. It should be noted that organohalide degradation processes have been described for a wide range of redox conditions (from highly oxidizing to strictly reducing), and mediated by a wide variety of (micro)organisms

in co-metabolic and/or energy-yielding modes; however, in this chapter we specifically discuss the phylogeny and environmental distribution of OHRB that are

known to gain energy and grow using organohalides as electron acceptors.

5.2 Phylogeny of OHRB

Since the description of Desulfomonile tiedjei as the first isolated OHRB

(DeWeerd et al. 1990), numerous bacterial strains capable of OHR have been

obtained in axenic culture (Fig. 5.1), providing indispensable insights into

their phylogenetic, physiological and biochemical traits. Members of the genus

Dehalococcoides comprise the biggest groups of isolates to date (19 isolates)

5  Overview of Known Organohalide-Respiring Bacteria …


Fig. 5.1  Number of OHRB available in axenic culture. The data is based on the number of

isolates as of July 2015

Fig. 5.2  Phylogenetic tree of known OHRB based on bacterial 16S rRNA gene sequences.

Alignment and phylogenetic analysis were performed with MEGA and the tree was constructed

using the neighbour-joining (NJ) method. The reference bar at the bottom indicates the branch

length that represents 2 % sequence divergence. Colour Key: Chloroflexi (red), Firmicutes

(green), Deltaproteobacteria (blue), Betaproteobacteria (violet), Epsilonproteobacteria (brown)


S. Atashgahi et al.

followed by strains of Desulfitobacterium (17 isolates) (Fig. 5.1). The known

OHRB are spread among several distinct phyla comprising both Gram-positive

and Gram-negative bacteria (Fig. 5.2). The known isolates can be and large

be divided into facultative and obligate groups based on whether OHR is their

only energy-gaining metabolism (Maphosa et al. 2010). The members of the

facultative OHRB are characterized by a more versatile metabolism, in general

have the ability to grow on a wide range of electron acceptors, and include proteobacterial OHRB such as Geobacter, Desulfuromonas, Anaeromyxobacter,

Desulfomonile, Desulfovibrio, Desulfoluna, Sulfurospirillum, Comamonas,

Shewanella as well as Desulfitobacterium from the phylum Firmicutes. The fact

that some facultative OHRB such as Comamonas, Geobacter and Shewanella

belong to phylogenetic groups that mostly comprise non‐OHRB points towards

horizontal acquisition of reductive dehalogenase genes. The obligate OHRB on

the other hand are restricted to OHR for energy conservation and growth and

include Dehalobacter (phylum Firmicutes) and the OHRB belonging to the

Dehalococcoidia class (phylum Chloroflexi) including strains of Dehalococcoides

mccartyi, Dehalogenimonas spp. and the single isolate ‘Dehalobium chlorocoercia’ DF-1 (Maphosa et al. 2010; Löffler et al. 2013; May et al. 2008). However,

recent studies showed fermentative growth of Dehalobacter spp. on chloromethane (Justicia-Leon et al. 2012; Lee et al. 2012), suggesting that at least some of

the isolates previously considered obligate OHRB might harbour additional

modes of metabolism beyond the canonical OHR. It is also interesting to note that

recent single-cell genomic studies of marine Dehalococcoidia did not reveal any

evidence for catabolic reductive dehalogenation, indicating that microorganisms

closely related to known obligate OHRB do not rely on OHR for energy conservation, but rather utilize organic matter degradation pathways (Kaster et al. 2014;

Wasmund et al. 2014).

With few exceptions, OHRB from different phyla, i.e. members of the

Firmicutes, Chloroflexi and different classes of Proteobacteria, benefit from similar enzymes/pathways for OHR indicating that these genes could be acquired

from common ancestors via transposon mediated dissemination (see Chap. 16 for

more details). However, to date little to no correlation could be found between the

type of organohalides used as electron acceptor and the phylogenetic affiliation of

OHRB, and haloaliphatic and haloaromatic compounds can be dehalogenated by

isolates of taxonomically different genera (Hug et al. 2013b). Furthermore, isolates with similar phylogeny and physiology have been obtained from very different environments. For example, Desulfuromonas michiganensis strain BB1

was isolated from pristine river sediment while the closely related strain BRS1

was obtained from chloroethene-contaminated aquifer material (Sung et al.

2003). Similarly, the closely related uncultured Chloroflexi Lahn and Tidal Flat

Clusters were both capable of dechlorination of perchloroethene (PCE) to transDCE while being enriched from sediments from river and marine environments,

respectively (Kittelmann and Friedrich 2008a, b). Another example is the case

of D. mccartyi strains KS and RC. Both bacteria grow with 1,2-dichloropropane

(1,2-D) as an electron acceptor in enrichment cultures while being derived from

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4 Early Work on Reaction Pathways and Organisms Involved

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