Natural and Enhanced Remediation Systems - Chapter 4 pot

Suthersan, Suthan S. “Chapter 4: In Situ Reactive Zones”

Natural and Enhanced Remediation Systems

Edited by Suthan S. Suthersan
Boca Raton: CRC Press LLC, 2001
©2001 CRC Press LLC

CHAPTER

4

In Situ

Reactive Zones

CONTENTS

4.1 Introduction
4.2 Engineered Anaerobic Systems
4.2.1 Enhanced Reductive Dechlorination (ERD) Systems
4.2.1.1 Early Evidence
4.2.1.1.1 Biostimulation vs. Bioaugmentation
4.2.1.2 Mechanisms of Reductive Dechlorination

4.2.1.3 Microbiology of Reductive Dechlorination

4.2.1.3.1 Cometabolic Dechlorination

4.2.1.3.2 Dechlorination by Halorespiring

Microorganisms

4.2.1.4 Electron Donors

4.2.1.4.1 Production of H

2

by Fermentation

4.2.1.4.2 Competition for H

2

4.2.1.5 Mixture of Compounds on Kinetics
4.2.1.6 Temperature Effects

4.2.1.7 Anaerobic Oxidation

4.2.1.8 Electron Acceptors and Nutrients

4.2.1.9 Field Implementation of IRZ for Enhanced Reductive
Dechlorination

4.2.1.10 Lessons Learned

4.2.1.11 Derivation of a Completely Mixed System for
Groundwater Solute Transport of Chlorinated Ethenes

4.2.1.12 IRZ Performance Data

4.2.2

In Situ

Metals Precipitation
4.2.2.1 Principles of Heavy Metals Precipitation
4.2.2.2 Aquifer Parameters and Transport Mechanisms
4.2.2.3 Contaminant Removal Mechanisms

©2001 CRC Press LLC

4.2.3

In Situ

Denitrification
4.2.4 Perchlorate Reduction
4.3 Engineered Aerobic Systems
4.3.1 Direct Aerobic Oxidation
4.3.1.1 Aerobic Cometabolic Oxidation
4.3.1.2 MTBE Degradation
4.4

In Situ

Chemical Oxidation Systems
4.4.1 Advantages
4.4.2 Concerns
4.4.3 Oxidation Chemistry
4.4.3.1 Hydrogen Peroxide

4.4.3.2 Potassium Permanganate
4.4.3.3 Ozone
4.4.4 Application
4.4.4.1 Oxidation of 1,4-Dioxane by Ozone
4.4.4.2 Biodegradation Enhanced by Chemical
Oxidation Pretreatment
4.5 Nano-Scale Fe (0) Colloid Injection within an IRZ
4.5.1 Production of Nano-Scale Iron Particles
4.5.2 Injection of Nano-Scale Particles in Permeable Sediments
4.5.3 Organic Contaminants Treatable by Fe (0)
References

Oxidation-reduction process plays a major role in the mobility, transport, and
fate of inorganic and organic contaminants in natural waters. Hence, the manip-
ulation of REDOX conditions to create an in situ reactive zone (IRZ) to meet
the cleanup objectives was a predictable evolution … .

4.1 INTRODUCTION

The concept of

in situ

reactive zones is based on the creation of a subsurface
zone, where migrating contaminants are intercepted and permanently immobilized
or degraded into harmless end products. Figures 4.1a and b pictorially describe the
concept of

in situ

reactive zones (IRZ). The successful design of these reactive zones
requires the ability to engineer two sets of reactions: 1) between the injected reagents
and the migrating contaminants; and 2) between the injected reagents and the
subsurface environment to manipulate the bio-geo-chemistry to optimize the required
reactions, in order to effect remediation. These interactions will be different at each
contaminated site and, in fact, may vary within a given site. Thus, the major challenge
is to design an engineered system for the systematic control of these reactions under
the naturally variable or heterogeneous conditions found in the field.
The effectiveness of the reactive zone is determined largely by the relationship
between the kinetics of the target reactions and the rate at which the mass flux of
contaminants passes through it with the moving groundwater. Creation of a spatially

©2001 CRC Press LLC

fixed reactive zone in an aquifer requires not only the proper selection of the reagents,
but also the proper mixing of the injected reagents uniformly within the reactive
zone. Furthermore, such reagents must cause few side reactions and be relatively
nontoxic in both its original and treated forms.
When dealing with dissolved inorganic contaminants such as heavy metals, the
process sequence in a pump and treat system required to remove the dissolved heavy
metals present in the groundwater becomes very complex, operation- and mainte-
nance-intensive, and costly. In addition, the disposal of the metallic sludge, in most
cases as a hazardous waste, is also very cost prohibitive. Therefore,

in situ

treatment
methods capable of achieving the same mass removal reactions for dissolved con-
taminants in an

in situ

environment are evolving and gradually gaining prominence
in the remediation industry.
The advantages of an

in situ

reactive zone to address the remediation of ground-
water contamination are as follows:

• An

in situ

technology enables implementation of most ground treatment processes
and eliminates the expensive infrastructure required for a pump and treat system
with no disposal of water or wastes

Figure 4.1a

Pictorial depiction of an

in situ

reactive zone (IRZ) formation.

Figure 4.1b

Cross sectional view of the creation of an IRZ around an individual injection well

at a selected location.
Plan View
Source Area IRZ Grid
Contaminant
Plume
Individual
Reactive Zones
Created by
Individual Injection
Points Providing a
Collective
In Situ

Reactive Zone
(IRZ) Curtain
Cross Sectional View
Contaminant
Zone
Reagent

©2001 CRC Press LLC
• Inexpensive installation because primary capital expenditure for this technology
is the installation of injection wells at appropriate locations
• Inexpensive operation that allows inexpensive reagents to be injected at fairly low
concentrations and, hence, should result in insignificant cost; only sampling
required is for groundwater quality monitoring and performance monitoring
parameters are usually done in the field; remediation of large volumes of contam-
inated water without any pumping or disposal needs
• Can be used to remediate deep sites because cluster injection wells or in-well
mixing systems can be installed to address deeper sites

• Unobtrusive because once the system is installed, site development and operations
can continue with minimal obstructions
•

In situ



degradation of contaminants because organic contaminants and a few
inorganics such as NH

4
+

, NO

3
–

, and CIO

4
–

can be degraded by implementing the
appropriate reactions
• Immobilization of contaminants because once the dissolved heavy metals are
precipitated out, the capacity of the soils and sediments is utilized to adsorb, filter
out, and retain inorganic contaminants

Manipulation of the reduction-oxidation (REDOX) potential of an aquifer is a
viable approach for

in situ

remediation of REDOX-sensitive groundwater contami-
nants. In addition, various microbially induced or chemically induced reactions also
can be achieved in an

in situ

environment. As noted earlier, creation of spatially
fixed reactive zones to achieve these reactions is very cost effective in comparison
to treating the entire plume as a reaction zone.
Since the first IRZ for the precipitation and remediation of hexavalent chromium
(Cr

6+

), was installed in 1993, this technology has advanced by leaps and bounds.

1

Currently the application of this technology can be classified into three categories
based on the creation of specific bio-geo-chemical and REDOX environments: 1)
engineered anaerobic systems, 2) engineered aerobic systems, and 3)

in situ

chemical
oxidation.
The engineered anaerobic systems can be further divided into enhanced reductive
dechlorination (ERD) systems,

in situ

denitrification,

in situ

perchlorate transforma-
tion, and

in situ

heavy metals precipitation. The ERD application has been expanded
to many contaminants since the first trichloroethene (TCE) application site. The IRZ
technology has been successfully applied to remediate the following chlorinated
compounds:

• Chlorinated ethenes: tetrochloroethane (PCE), trichloroethane (TCE), dichloroet-
hene (Cis 1,2 DCE, and 1,1 DCE), vinylchloride
• Chlorinated ethanes: 1,1,2,2 tetrachloroethane (1,1,2,2 PCA), 1,1,1 trichloroet-
hane, (1,1,1 TCA), 1,1,2 trichloroethane (1,1,2 TCA), 1,1 and 1,2 dichloroethane
(DCA), chloroethane (CA)
• Chlorinated phenols: pentachlorophenol (PCP), and tetrachlorophenol
• Chlorinated pesticides
• Perchlorate

In addition, the IRZ technology has been successfully applied to precipitate the
following dissolved metals at contaminated sites: Cr

6+

, Pb

2+

, Cd

2+

, Ni

2+

, Zn

2+

, Hg

2+

.

©2001 CRC Press LLC

4.2 ENGINEERED ANAEROBIC SYSTEMS
4.2.1Enhanced Reductive Dechlorination (ERD) Systems

4.2.1.1 Early Evidence

The first microbially mediated reductive dechlorination of PCE and TCE was
observed in the early 1980s, and this study

2,3

reported the degradation of PCE to
nonchlorinated end products in an acetate-fed, continuous-flow methanogenic glass
bead column. It appeared that the first step in the degradation pathway was dechlori-
nation to TCE. Further anaerobic oxidation of TCE to carbon dioxide and hydrochloric
acid was suggested. In 1984

4

, further evidence of dechlorination of PCE beyond TCE
came in an experiment where sediments from an aquifer recharge basin were incubated
with PCE and methanol as the electron donor. Significant concentrations of TCE,

cis

-
1,2 DCE and VC were observed after three weeks, whereas in sterile controls no
dechlorination had occurred. Another study in the 1980s demonstrated that dechlori-
nation of PCE to VC in a methanogenic column was achievable.

5

Similar studies using

13

C-TCE, showed that TCE was dechlorinated exclusively to

cis

-DCE in soil.

6

In 1989, the first evidence of complete dechlorination of PCE to ethene under
methanogenic conditions with methanol as electron donor was demonstrated.

7

Another study found PCE reduction via ethene to ethane with lactate as electron
donor in a flow-through column filled with a mixture of polluted sediment and
anaerobic granular sludge.

8

Meanwhile, numerous publications showed that micro-
organisms capable of reductively dechlorinating chlorinated ethenes are abundant
in polluted anaerobic environments. (An overview of the biological reductive dechlo-
rination pathway of chlorinated solvents is shown in Figure 4.2.) PCE and TCE are
dechlorinated mainly to

cis

-DCE, although sometimes

trans

-DCE and 1,1-DCE have
also been found as products.

9,10

However, the formation of the 1,1-DCE is believed
to be a result of abiotic dechlorination in the presence of sulfide.

10

Evidence from the earlier studies indicated that the dechlorination of PCE to

cis

-DCE was found to be a relatively fast process, whereas, subsequent rates of
dechlorination of

cis

-DCE to VC and ethene were significantly slower or even
absent.

7,11

Dechlorination of 1,1-DCE and

trans

-DCE was less studied.
In some of the earlier reports and studies the dechlorination of chlorinated
ethenes was often found to be incomplete, both in the laboratory and in field
experiments, resulting in an accumulation of

cis

-DCE and VC. It was not fully
understood at that time why dechlorination beyond these compounds was problem-
atic, other than raising valid questions regarding the required microbial consortia
for complete dechlorination. During that time (late 1980s and early 1990s) micro-
organisms capable of dechlorinating DCE and VC had not been isolated yet, although
several enrichment cultures existed. Little was known about the substrate require-
ments of these bacteria. Later studies reported that PCE dechlorination in a contam-
inated soil down to ethene was only achieved by adding a complex mixture of organic
electron donors. Significant research was focused, during the early to mid 1990s,
on the microbial ecology that could perform complete dechlorination of PCE to

©2001 CRC Press LLC

ethene and the biogeochemical conditions under which this biotransformation could
be achieved.
The choice of a suitable electron donor for the stimulation of

in situ

dechlori-
nation is still a matter of discussion and may be dependent on local conditions; this
will be discussed in detail in a later section. When hydrogen is assumed to be the
major electron donor for dechlorination, its amendment can only be achieved by
using substrates yielding hydrogen after anaerobic degradation.

12

Often, short-chain
organic acids are produced as intermediate products, which may lead to acidification
of the groundwater and soil. Additionally, electron donors that support dechlorination
are generally readily degraded by nondechlorinating microorganisms, leading to
competition for the substrate and excessive bacterial growth in soil pores near the
injection well. As a result, significantly more electron donor mass will be needed
than theoretically necessary to reduce all chlorinated ethenes present to ethene.

4.2.1.1.1 Biostimulation vs. Bioaugmentation

The first level of the treatment hierarchy for chlorinated ethenes is intrinsic
bioremediation, or natural biodegradation, whereby indigenous microflora destroy
the contaminant(s) of concern without any stimulation or enhancements. The second
choice in this hierarchy, biostimulation or enhanced biodegradation, involves stim-
ulating the indigenous microbial populations and thus enhancing microbial activity

Figure 4.2

Biological and abiotic degradation pathways of the common chlorinated com-
pounds encountered at contaminated sites (adapted from McCarty and Semprini,
1994; after Vogel et al., 1987, and Wiedermeier et al., 1999).
PCE CT

TCE CF
cis-1,2 DCE
*
DCM
VC CM
Ethene
Ethane
1,1- DCE
Acetic Acid
Primary ReactionBiotic Reactions
Abiotic Reactions
CO
2
, H
2
O,
CI
-
1,1,1-TCE
1,1- DCA
CA
*

©2001 CRC Press LLC

so that they destroy the target compounds at a rate that meets the cleanup objectives
at the site. At almost every contaminated site a natural population of degradative
microorganisms exists within the contaminated zone; however, specific nutrients,
growth substrates inducers, electron donors, and electron acceptors may be required
to create optimal microbial activity.

12

Thus, through the introduction of required
additional reagents, the native degradative microbial population can be stimulated
to grow, multiply, and destroy the target contaminants. Most environments contain
microorganisms able to grow on and destroy a variety of chlorinated compounds;
at some sites, the persistence of these compounds, is not a consequence of the
absence of organisms but rather of the absence of the full set of conditions necessary
for the indigenous species to function rapidly.

12

In the past there was a significant
debate among remediation experts whether the microorganisms responsible for
cometabolic degradation and dehalorespiration are ubiquitous. Current belief is that
these organisms are nearly ubiquitous.
When intrinsic bioremediation or biostimulation is not feasible at a given site
due to the absence of an appropriate microbial population, bioaugmentation may be
utilized.

Bioaugmentation

involves injection of selected exogenous microorganisms
with the desired metabolic capabilities directly into the contaminated zones along
with any required nutrients to effect the rapid biodegradation of target compounds.
Two distinct bioaugmentation approaches have been developed for remediating
chlorinated ethenes. In the first approach, degradative organisms are added to com-
plement or replace the native microbial population. The added microorganisms can
be selected for their ability to survive for extended periods or to occupy a specific

niche within the contaminated environment. If needed, stimulants or selective cosub-
strates can be added to improve survival or enhance the activity of the added
organism. Thus, the goal of this approach is to achieve prolonged survival and growth
of the added organisms and degradation of the target contaminants.
In the second bioaugmentation approach, large numbers of degradative bacteria
are added to a contaminated environment as biocatalysts which will degrade a
significant amount of the target contaminant before becoming inactive or perishing.

12

Additional microbes can be added as needed to complete the remediation process.
Attempts can be made to increase the production of the degradative enzymes or to
maximize catalytic efficiency or stability, but long-term survival, growth, and estab-
lishment of an active microbial population are not the primary goals of this treatment
approach.
In the past, bioaugmentation has been implemented frequently and successfully
only in bioreactors. The conditions in these bioreactors are controlled and quite
different from those in nature, and prior to start-up, no microorganisms are present
anyway. Hence, the addition of enriched cultures is essential. Furthermore, bioreac-
tors are engineered and controlled systems where conditions can be readily altered
or optimized for a particular process and can be designed to promote the multipli-
cation and activity of the inoculated species — in contrast to contaminated field sites.
The record of success of

in situ

bioaugmentation systems for chlorinated com-
pounds has been rather spotty. On the one hand, the initiation or enhancement of
degradation has been reported (far more commonly in samples of the contaminated
environments in simulated laboratory experiments) following the addition of

©2001 CRC Press LLC

enriched bacterial cultures that can metabolize and grow on chlorinated ethenes. On
the other hand, a number of failures in the field have been reported.
Such reports of failure of bioaugmentation came as no surprise to microbial
ecologists. Without question, a species with a substrate uniquely available to it has
a distinct advantage, yet that advantage may not be sufficient to compensate for
many other traits also necessary for survival, no less multiplication, in a natural
ecosystem. Possessing the requisite enzymes to metabolize a novel compound is a
necessary attribute for the organism, but it is not sufficient for the organism to
succeed. Populations of introduced microorganisms are subject to a variety of abiotic
and biotic stresses, and these must be overcome for these organisms to be able to
express beneficial traits.
The reasons for the frequent failures of bioaugmentation are many:

12

limiting
nutrients and growth factors in the uncontrolled natural environment, suppression
by predators and parasites, inability of the introduced bacteria to penetrate significant
space, metabolism of other nontarget organic compounds present, concentration of
the target chlorinated compound too low to support multiplication, and other inhib-
itory biogeochemical conditions such as pH, temperature, salinity, and toxins.
In summary, the problems usually encountered in scaling up the bioaugmentation
successes achieved in laboratory experiments can be summarized as follows:

12

• Contaminant rates established in controlled laboratory studies may differ substan-

tially from those in pilot-scale, full-scale, or even other laboratory studies.
• Positive biotransformation results from small systems often are not reproduced in
different systems.
• Instantaneous biotransformation rates vary widely and in an apparently stochastic
manner, even in well-operated, steady-state systems.

4.2.1.2 Mechanisms of Reductive Dechlorination

Naturally occurring biological processes can degrade organic contaminants

in
situ

or during transport in the subsurface under aerobic and/or anaerobic conditions.
Microorganisms catalyze degradation reactions to obtain energy for growth, repro-
duction, and cell maintenance. Useable energy is recovered through a series of
REDOX reactions where the microorganisms act as “electron transport mediators”
(Figure 4.3). Biologically mediated electron transfer couples the oxidation of an
electron donor (organic compound) with the reduction of an electron acceptor (inor-
ganic or organic) and results in the production of useable energy for microbial
consortia.

12,13,14

The bulk electron donor acts as a fuel source for the reactions and
the reactions proceed as long as there is a source of bioavailable electrons. Fuel
sources can be the target chlorinated compounds, native organic carbon, co-contam-
inants such as fuel hydrocarbons, or organic compounds such as carbohydrates. In
aerobic environments, the chlorinated compounds act as electron donors and under
anaerobic conditions they act as electron acceptors.

There are two primary mechanisms involved in the biodegradation of chlorinated
organic contaminants (Table 4.1). First, biodegradation may be growth-linked and
provide carbon and energy to support growth when the compound is used as primary

©2001 CRC Press LLC

substrate and directly utilized by the mediating organisms via the processes included
in Category 1. Some chlorinated solvents are used as electron donors and some are
used as electron acceptors when serving as primary growth substrates. When used
as an electron donor (under aerobic and anaerobic conditions) the contaminant is
oxidized. Conversely, when used as an electron acceptor, the contaminant is reduced
via the reductive dechlorination process called halorespiration.

17

In addition to their use as a primary growth substrate, chlorinated solvents can
also be degraded via cometabolic pathways. During cometabolism, microorganisms
gain carbon and energy for growth from metabolism of a primary substrate, and
chlorinated solvents are degraded fortuitously by enzymes present in the metabolic

Figure 4.3

Description of microorganisms acting as electron transport mediators (after
Schwarzenbach et al., 1993).

Table 4.1Summary of the Categories of Degradation Pathways for Chlorinated Organic

Compounds (Adapted from Wiedemeier et al., 1999)

14

Category 1 Category 2

(used as primary substrate)

(used as cometabolic substrate)
Chlorinated
Compound
Halo-
respiration
Direct
Aerobic
Oxidation
Direct
Anaerobic
Oxidation
Aerobic
Cometabolism
(co-oxidation)
Anaerobic
Cometabolism
(reductive
dechlorination)

Tetrachloroethylene
(PCE)
XX
Tr ichloroethylene
(TCE)
XXX

Dichloroethene
(DCE)
XXX X X
Vinyl Chloride (VC) X X X X X
Tr ichchloroethane
(1,1,1 TCA)
XXX
Dichloroethane
(1,2 DCA)
XX X X
Carbontetrachloride
(CT)
XX
Methylenechloride
(MC)
XX X
(bulk)
ox
(bulk)
red
(mediator)
ox
(mediator)
red
+ ne
-
+ ne
-
(Contaminant)
ox

(Contaminant)
red
+ ne
-
- ne
-

©2001 CRC Press LLC

pathways. Cometabolism is a process where the organism receives no direct benefit
from the degradation of the organic compound.

13,16

There are two types of cometa-
bolic reactions: co-oxidation and reductive dechlorination, described as Category 2
in Table 4.1. Cometabolic reactions tend to be incomplete and can possibly lead to
an accumulation of more toxic daughter products. To date, vinyl chloride (VC) and
dichloroethene (cis/trans) are the only chlorinated solvents that can be degraded by
all aerobic and anaerobic pathways.

15

The predominant mechanism for the biodegradation of chlorinated solvents in
anaerobic environments is reductive dechlorination, whether the organic compound
is a primary electron acceptor (halorespiration) or is cometabolized. Before 1994,
reductive dechlorination was thought to be strictly a cometabolic process because
the organisms that cause these reactions are ubiquitous at most contaminated
sites.

14,15

However, research has shown that combetabolic reductive dechlorination
is “sufficiently slow and frequently incomplete.”

15,18

During reductive dechlorination,
the chlorinated solvents act as an electron acceptor and a chlorine atom is replaced
with a hydrogen atom (Figure 4.4).
Cometabolic reduction of the chlorinated solvents is catalyzed by the reductive
dehalogenase and reductase enzymes produced by microorganisms.

14,20

Cometabolic
degradation occurs under iron reducing, manganese reducing, sulfate reducing, and
methanogenic environments.

21

The enzymes of these reducing microorganisms are
induced to reduce abiotic forms of Fe (III) to Fe (II), Mn (IV) to Mn (II), sulphate
to sulfide or hydrogen sulfide, and carbon dioxide to methane. Electrons are trans-
ferred to dissolved contaminants coincidentally during the reducing processes. These
degradation reactions are often incomplete, resulting in an accumulation of toxic
daughter products.
Just as aerobic biodegradation systems utilize oxygen as a terminal electron
acceptor to stimulate microbial activity, oxidative anaerobic systems require other
terminal electron acceptors, such as nitrate or ferric iron (Fe III), to stimulate

biodegradation. Anaerobic oxidation occurs when anaerobic bacteria use the chlo-
rinated contaminant as the electron donor and, in most instances, allow the micro-
organism to derive useful amounts of energy from the reaction. It has been shown
that vinyl chloride can be oxidized to carbon dioxide, water, and chloride ion via
Fe (III) reduction.

22

Significant anaerobic mineralization of DCE, VC, and methylene
chloride also have been reported in the literature.
While in oxidative anaerobic systems the contaminant is used as an electron
donor, in reductive systems highly oxidized contaminants (such as PCE) are used
as electron acceptors. The process begins by supplying excess reduced substrate
(electron donor) to a microbial consortium, i.e., a cooperative community of micro-
bial species (Figures 4.3 and 4.5). The presence of the substrate expedites the
exhaustion of any naturally occurring electron acceptors. As the natural electron
acceptors are depleted, microorganisms capable of discharging electrons to other
available electron acceptors, such as oxidized contaminants, gain a selective advan-
tage. The intricacies of these microbial communities are complex, but recent research
has provided some insight into methods for enhancing populations of contaminant-
degrading microorganisms.

©2001 CRC Press LLC

The reductive dechlorination of PCE to ethene proceeds through a series of
hydrogenolysis reactions (Figure 4.4). Each reaction becomes progressively more
difficult to carry out; subsequently, the DCEs, particularly

cis

-DCE, and vinyl chlo-
ride (VC), tend to accumulate in anaerobic environments under natural conditions
due to the absence of sufficiently reducing conditions.

Figure 4.4

Hydrogenolysis reactions of PCE during reductive dechlorination with H

2

acting
as the electron donor and the chlorinated compounds acting as electron accep-
tors (adapted from Vogel et al., 1987, and Wiedermeier et al., 1999).
+CI
-
H ion
+
2
-
H
Electron
Flow
e
Ethane
h
H
H
H
CC
Ethene

Vinyl Chloride
C
C
H
H
H
CI
H
H
CI
CI
CC
(Limited Biological
Reaction)
1,1-Dichloroethene
CI
H
CI
H
CC
(Limited Biological
Reaction)
trans
-1,2,-Dichloroethene
(Predominant Biological
Reaction)
cis
-1,2,-Dichloroethene
C C
CI

H
H
CI
CI
H
CI
CI
CC
Trichloroethene
Perchloroethene
C
C
CI
CI
CI
CI
H
H
H
H
H
H
CC

©2001 CRC Press LLC

The oxidation-REDOX potential (ORP) affects the thermodynamics of reductive
dechlorination. Microorganisms will facilitate only those oxidation-reduction
reactions that have a net yield of energy. For reductive dechlorination to be thermo-
dynamically favorable the REDOX potential must be sufficiently low, thereby

excluding the presence of oxygen and nitrate as terminal electron acceptors. Fur-
thermore, the presence of nitrate may have an inhibitory effect on PCE dechlorina-
tion.

23

The REDOX potential range for reductive dechlorination is shown in Figure
4.6. It is important to note that the values of Eh ranges shown in Figure 4.6 and the
values of ORP measured in the field by remediation engineers are not the same.
Both parameters have some correlation and do not represent the same conditions.
Figure 4.5 summarizes the mechanisms and the required environmental condition
for the degradation of chlorinated solvents.

4.2.1.3 Microbiology of Reductive Dechlorination

4.2.1.3.1 Cometabolic Dechlorination

A cometabolic process is defined here as a process in which the compound of
interest (e.g., PCE) is converted by a biological enzyme system or cofactor in which
the compound does not serve as a source of carbon or energy.

Pure Microbial Cultures:

Reductive dechlorination is the only biodegradative
conversion known for PCE. This reaction can occur cometabolically or in a metabolic
energy-producing reaction. In both cases, the cofactors of the enzymes involved are
metal-containing porphyrins. Examples

25

of acetogenic and methanogenic bacteria that
dechlorinate PCE cometabolically are listed in Table 4.2. In general, acetogenic bacteria
dechlorinate PCE at higher rates than methanogenic bacteria. Metal-containing cofac-
tors have been found to catalyze the

in vitro

degradation of chlorinated ethenes.

24,25

In general, reductive dechlorination rates decrease with a declining amount of
chlorine atoms in the molecule.

In vivo

experiments with methanogenic and aceto-
genic bacteria indicate that dechlorination rates are low (0.5 to 235 nmol PCE

Figure 4.5

Pictorial description of the conditions which control reductive dechlorination.
Electron Donors
Electron
Acceptor
Processes
Reducing Conditions
Environmental Conditions
ORP
Hydrogen

Generation
Fermentable
Substrates
BTEX
Dissolved
Organic
Compounds
Te mperature pH
Anaerobic
Conditions
Reductive
Dechlorination

©2001 CRC Press LLC

(mg protein)

–1

day

–1

), compared with those of halorespiring bacteria

25,26

(Table 4.3).

In vivo

usually only one halogen atom is removed. An exception is the reductive
dehalogenation of dibromoethene by

Methanobacterium

and

Methanococcus

that
yields acetylene as a product.

27

However, in many studies the possible formation of
nonchlorinated products during dechlorination reactions was not included in the
carbon balance. Complete dechlorination of PCE to ethene by pure cultures of
acetogenic or methanogenic bacteria has not been observed. This is in contrast to

Figure 4.6

Optimal range for reductive dechlorination.

Table 4.2Examples of Acetogenic and
Methanogenic Bacteria that

Dechlorinate PCE
Cometabolic DechlorinationCosubstrate

Methanosarcina

sp.Methanol

Methanosarcina mazei

Methanol

Sporomusa ovata

Methanol

Acetobacterium woodii

Fructose
NO
3
-
Reduction
REDOX Potential (Eh)
in millivolts (mV) at
pH=7 and T=25˚C
1,000mV
500mV
Mn
4+
Reduction
Fe
3+
Reduction

SO
4
2
-
Reduction
Methanogenic
0mV
-500mV
Anaerobic Aerobic
Possible
Range of
Reductive
Dechlorination
Optimum
Range for
Reductive
Dechlorination

©2001 CRC Press LLC

findings with mixed anaerobic cultures in which more extensive dechlorination has
been observed.

7,8,27-29

The latter may be due to the interactions between different
microorganisms. Sometimes, however, it is difficult to distinguish between comet-
abolic and specific dechlorination in these mixed cultures, and often it is not even
clear which microorganisms are responsible for the dechlorination.
The most often observed degradation pathway of PCE is via reductive dechlo-

rination to

cis

-DCE.

10,25,30-32
Several dechlorination rates for chlorinated ethenes have
been reported in the literature, but it is difficult to compare the data because often
the numbers of bacteria involved were not known. Nevertheless, it can be stated that
dechlorination rates in mixed cultures are generally higher than those found for
single acetogenic or methanogenic strains.
There are a few reports on the degradation of PCE by granular methanogenic sludge
from upflow anaerobic sludge blanket reactors. This sludge is enriched with acetogenic
and methanogenic bacteria and contains high concentrations of cofactors. Such bacterial
consortia, therefore, are suitable as a source of cometabolic dechlorinating activity. In
one of these reactors, fed with a mixture of sucrose, lactic acid, propionic acid, and
methanol as primary substrates, granular sludge showed a fast adaptation to high PCE
concentrations. Influent concentrations of 360 to 420 mM PCE were completely dechlo-
rinated to ethene.
26,33
Average removal rates of 7.6 mmol (g VSS)
–1
day
–1
were achieved,
with a maximum removal rate of 28.3 mmol (g VSS)
–1
day
–1

. A bacterial consortium in
a similar reactor operated in batch mode converted PCE to TCE, cis- and trans-DCE
and traces of 1,1-DCE with ethanol as the primary substrate.
4.2.1.3.2 Dechlorination by Halorespiring Microorganisms
Halorespiration is a type of anaerobic respiration in which a chlorinated com-
pound is used as a terminal electron acceptor. In this reductive dechlorination
process, which enables the conservation of energy via electron transport phospho-
rylation, one or more chlorine atoms are removed and replaced by hydrogen. Exam-
ples of halorespiring bacteria species are shown in Table 4.3.
Halorespiration, also referred to as dehalorespiration, occurs when the organic
compound acts as an electron acceptor (primary growth substrate) during reductive
dechlorination. During halorespiration, the chlorinated organic compounds are used
directly by microorganisms (termed halorespirators), such as an electron acceptor
while dissolved hydrogen serves as an electron donor:
15,34
H
2
+ C – Cl Þ C – H + H
+
+ Cl
–
(4.1)
Table 4.3 Examples of Halorespiring Bacteria
HalorespirationElectron Donor
Dehalobacter restrictus H
2
Dehalospirillum multivorans Pyruvate
Desulfitobacterium sp. strain PCE1Lactate
Desulfitobacterium sp. strain TCE1Lactate
Strain MS-1Yeast extract

©2001 CRC Press LLC
where C – Cl represents the chlorine bond to the carbon in the chlorinated ethene
molecule. Halorespiration occurs as a two-step process which results in the inter-
species hydrogen transfer by two distinct strains of bacteria. In the first step, bacteria
ferment organic compounds to produce hydrogen. During primary or secondary
fermentation, the organic compounds are transformed to compounds such as acetate,
water, carbon dioxide, and dissolved hydrogen. Fermentation substrates are either
biodegradable, nonchlorinated contaminants (i.e., BTEX — benzene, toluene, ethyl
benzene, and xylenes — or sugar) or naturally occurring organic carbon. In the
second step, the nonfermenting microbial consortia utilize the hydrogen produced
by fermentation for halorespiration.
35,36
Denitrifiers, iron reducers, sulfate reducers,
methanogens, and halorespirators can all utilize hydrogen as an electron donor.
14
Figure 4.7 shows which reducing environment is favored depending on the hydrogen
concentration. Although compounds produced during fermentation and hydrogen
have been demonstrated to drive halorespiration,
32
hydrogen appears to be the most
important electron donor for this process.
36.37
Halorespiration has been found to be
limited if available nutrients are not present. Direct injection of H
2
is able to serve
as an electron donor for reductive dechlorination of PCE to VC and eventually to
ethene in cultures provided with the proper nutritional supplements.
34,38
Because reductive dechlorination of chlorinated ethenes is a reductive process,

microorganisms may exist that can use chlorinated compounds as a terminal electron
acceptor and possibly conserve the concomitant energy gain into ATP. This hypoth-
esis, developed in the early 1990s, proved to be true.
25,26
The first evidence that
bacteria exist that can couple reductive dechlorination of PCE to growth (halores-
piration) under strict anaerobic conditions was presented in the early 1990s.
25,39
A
highly purified enrichment culture able to grow by the reduction of PCE to cis-DCE
using hydrogen as the electron donor was described. The dechlorinating organism,
later designated Dehalobacter restrictus, uses only hydrogen as the electron donor
and can couple growth to the reduction of PCE or TCE to cis-DCE. A recent study
25,40
described a new isolate, strain TEA, which is closely related to Dehalobacter
restrictus. Another strict anaerobic bacterium, Dehalospirillum multivorans, capable
of coupling dehalogenation of PCE to growth, was identified and described
recently.
25,41
This bacterium is less restricted concerning both electron donors and
acceptors.
Several dechlorinating strains belonging to the genus Desulfitobacterium were
isolated from different sources. The strictly anaerobic D. dehalogenans, able to grow
by the reductive dechlorination of chlorinated phenolic compounds was isolated
recently. Another Desulfitobacterium, strain PCE1, was isolated from polluted soil
and is reported to couple the reduction of both chlorinated phenols and chlorinated
ethenes to growth.
43
This strain dechlorinates PCE only to TCE, whereas other known
halorespiring microorganisms dechlorinate PCE further. The same authors also

described a Desulfitobacterium sp. strain TCE1, which is able to use several electron
donors for the reduction of PCE to cis-DCE.
43,44
Another researcher has described
a Desulfitobacterium sp. (strain PCE-S) that converts PCE to cis-DCE.
45
All isolated
Desulfitobacterium strains are able to use a number of different electron donors and
acceptors for growth. The nature and origin of the dechlorinating enzymes in these
organisms are still unknown.
©2001 CRC Press LLC
A recent study described two facultative aerobic bacteria, strain MS-1 and the
closely related Enterobacter agglomerans biogroup 5, which can reductively deha-
logenate PCE to cis-DCE under anaerobic conditions.
46
It is not clear yet whether
strain MS-1 and E. agglomerans biogroup 5 are actually halorespiring organisms.
Recently, an anaerobic bacterium, Desulfuromonas chloroethenica (strain
TT4B), has been isolated which can not only reductively dechlorinate PCE to cis-
DCE with acetate as an electron donor, but also can reduce Fe (III) and polysulfide.
These are unique features for PCE-dehalogenating organisms.
47
All the above-mentioned organisms are only able to couple growth to the partial
reduction of PCE or TCE. An exception is Dehalococcoides ethenogenes strain 195
that couples growth to rapid dehalogenation of PCE to VC, followed by a substantially
slower reduction to ethene.
37
Growth of this bacterium is restricted to the presence of
hydrogen, which is the only electron donor supporting the dechlorination reactions.
Dechlorination was only sustained by using hydrogen, acetate, vitamin B

12
, anaerobic
digester sludge supernatant, and cell extracts from mixed cultures in the medium.
Figure 4.7Range of hydrogen concentrations for the different anaerobic metabolic pathways
(after Wiedermeier et al., 1999).
15
10
5
0
Denitrifiers Fe(III)
Reducers
Halorespirators Sulfate
Reducers
Hydrogen Concentration (nM)
Methanogens
Possible Reactions
©2001 CRC Press LLC
Chloroethene Reductive Dehalogenases: The biochemistry of PCE dehalores-
piration has been studied with enzymes that were purified from a reductively dechlo-
rinating pure culture or enrichment, which is able to couple dechlorination to energy
conservation (halorespiration). PCE respiration has been studied most extensively
in Dehalobacter restrictus
42,49
and Dehalospirillum multivorans.
26,41,45
In general, a
PCE respiration chain should contain an electron-donating enzyme, electron carriers,
and a reductive dehalogenase as terminal reductase. Studies with D. multivorans
and Desulfitobacterium strain PCE-S indicate that a proton gradient or a membrane
potential may also be essential for chloroethene respiration because several iono-

phores have been found to inhibit dechlorination in whole cell suspensions.
45,50
The
nature of the electron-donating enzyme depends on the electron donor. In D. restric-
tus, which uses hydrogen as electron donor for PCE respiration, hydrogenase activity
has been localized on the membrane, facing the outside.
49
D. multivorans and
Desulfitobacterium strain PCE-S are able to use several electron donors for dechlo-
rination, and different electron-donating activities have been found.
45,50
The electrons
thus generated are transported to the dehalogenase via electron carriers such as
quinones and cytochromes. It was demonstrated that menaquinone is involved as
electron carrier for PCE respiration in D. restrictus,
49
but not in D. multivorans.
45,50
Cytochrome b is present in both organisms, but its involvement in PCE respiration
has not been established.
In contrast to the well studied PCE and TCE dechlorination, little is known about
the mechanism of DCE and VC dechlorination. It was found that the enzymes
catalyzing VC dechlorination in an enrichment culture are membrane bound and, in
contrast to the known PCE reductase, cobalamin independent.
51
It remains unclear
whether this enrichment is able to use VC as terminal electron acceptor. Recently,
an enzyme has been obtained from an enrichment containing D. ethenogenes that
catalyzes the dechlorination of TCE to cis-DCE, VC, and ethane. This cobalamin-
containing TCE-reductive dehalogenase is membrane bound and dechlorinates its

substrates at similar rates, as have been reported for the PCE dehalogenases. More
research is needed to know what determines the difference in substrate specificity
of the cobalt-containing reductive dehalogenases.
4.2.1.4 Electron Donors
The selection of an appropriate electron donor may be the most important design
parameter for developing a healthy population of dechlorinating microorganisms
during implementation of an IRZ for enhanced reductive dechlorination. Recent
studies have indicated a prominent role for molecular hydrogen (H
2
) in the reductive
dechlorination of chloroethenes.
34,39,48
Most known dechlorinators can use H
2
as an
electron donor; some can use only H
2
. Because more complex electron donors are
broken down into metabolites and residual pools of H
2
by other members of the
microbial community, they may also be used to support reductive dechlorination.
From the small but growing pool of knowledge about dechlorinating organisms,
it thus appears that H
2
may serve an important role in reductive dechlorination of
PCE in many environments. The author recently has observed the quick or direct
transformation of PCE or TCE to ethene under very reducing conditions leading to
©2001 CRC Press LLC
Figure 4.8 Conceptual diagram of microbial activity to derive energy for growth and multi-

plication.
Figure 4.9 Distribution of electrons to generation and cell synthesis during the breakdown
of organic electron donors.
PCE
Dechlorinators
Methanogens
Ethene
4 H
+
+ 4CI
-
CO
2
H
2
Acetic Acid
Complex Organics
H electrons
Second Intermediate
Electron Donor
Substrate
H electrons
First Intermediate
Electron Donor
Substrate
Target
Electron Donor
Substrate
Carbon Substrate
Cell Synthesis Reactions

Electron Acceptor Substrate
Energy Generation Reactions
f
s
electrons f
e
electrons
f
s
electrons f
e
electrons
f
s
electrons f
e
electrons
©2001 CRC Press LLC
speculations of the probable effect of high H
2
concentrations or reductive dechlori-
nation. In natural systems, including contaminated aquifers, most H
2
becomes avail-
able to hydrogenotrophic microorganisms through the fermentation of more complex
substrates by other members of the microbial consortium. The dechlorinators must
then compete with other organisms, such as methanogens and sulfate-reducing
bacteria, for the evolved H
2
(Figure 4.8). Figure 4.9 also describes the distribution

of electrons during the microbial breakdown of organic electron donor substrates.
During studies in which ethanol or lactate was used to stimulate dechlorination in
mixed anaerobic enrichment culture, both active dechlorination and methanogenesis
at high H
2
levels was observed; however, when H
2
levels fell, dechlorination con-
tinued, albeit slowly, while methanogenesis ceased entirely. It was speculated that
the addition of electron donors fermented only under low H
2
partial pressures might
give selective advantage to dechlorinators over methanogens.
One school of thought in the past was that the rate and quantity of H
2
made
available to a degrading consortium must be carefully engineered to limit competi-
tion for hydrogen from other microbial groups, such as methanogens and sulfate-
reducers. Competition for H
2
by methanogens is a common cause of dechlorination
failure in laboratory studies. As the methanogen population increases, the portion
of reducing equivalents used for dechlorination quickly drops and methane produc-
tion increases.
17,18,36
Speculation was that the use of slowly degrading nonmethano-
genic substrates would help prevent this. Recent thinking on this issue is evolving
to be different and is discussed later.
Many different compounds may serve as electron donors for the reductive dechlo-
rination of chlorinated solvents (Table 4.4). Several researchers suggest that the

microbial reductive dechlorination of chlorinated ethenes depends on the presence
of molecular hydrogen as the actual electron donor, either directly available or
produced from other substrates by fermentation.
32,34,55,56
Although this statement
applies to many studies, in several cases it does not hold. Acetate, from which usually
no hydrogen is produced during anaerobic metabolism, has been shown to support
reductive dechlorination of chlorinated ethenes both in microcosms and environ-
mental samples
5,7,26,58
and in pure culture.
47
Until recently, most research activities concerning the anaerobic degradation of
chlorinated compounds focused on methanogenic systems. Such systems typically
involve the introduction of a fermentable organic compound, such as acetate, lactate,
hexoses (present in molasses) or even a co-contaminant such as toluene or phenol,
which is fermented to produce hydrogen, among other things. It is now clear that
these systems probably contained at least two distinct strains of bacteria. One strain
fermented the organic carbon to produce hydrogen, and another utilized the hydrogen
as an electron donor for dehalorespiration.
15
Only in the last two or three years have
researchers finally recognized the role of hydrogen as the electron donor in the
reductive dechlorination process.
4.2.1.4.1 Production of H
2
by Fermentation
The production of H
2
by different microorganisms is intimately linked with their

respective energy metabolisms. The production of H
2
is one of the specific
©2001 CRC Press LLC
mechanisms to dispose excess electrons through the activity of hydrogenase present
in H
2
producing microorganisms.
59
All hydrogen producing microorganisms can be
categorized into four groups:
60
• Hetertrophic facultative anaerobes that contain cytochromes and lyse formate to
produce H
2
• Desulfovibrio desulfuricans, which is the only strict anaerobe in this group with
a cytochrome system
• Photosynthetic bacteria with light-dependent evolution of H
2
from reduced NADH
•Strict anaerobic heterotrophs that do not contain a cytochrome system (clostridia,
micrococci, methanobacteria, etc.)
Production of H
2
by obligate anaerobic microorganisms has optimum stoichi-
ometry (1:4, with glucose as substrate) compared with facultative anaerobes (1:2),
although the latter process is comparatively simpler than the former.
60
Under natural conditions, fermentation is the process that generates the hydrogen
used in reductive dechlorination. In the absence of externally available electron

acceptors, many organisms perform internally balanced (different portions of the
same substrate are oxidized and reduced) oxidation-reduction reactions of organic
compounds with the release of energy; this process is called fermentation. Since
only partial oxidation of the carbon atoms of the organic compound occurs, fermen-
tation yields substantially less energy per unit of substrate compared to oxidation
reactions. (Oxidation reactions are those in which external electron acceptors par-
ticipate in the reaction). For instance, the fermentation of glucose to ethanol and
CO
2
has a theoretical energy yield of –57 k cal/mole, enough to produce about
Table 4.4Electron Donors That Have Been Used to Enhance
Reductive Dechlorination and Relative Costs per lb of
PCE
51–53
Electron Donor
Bulk Price
$/lb
$/lb of
PCE
Soluble (Fast Release) Donors
Methanol0.050.64
Milk0.050.18
Ethanol0.20 – 0.25NA
Molasses0.20 – 0.350.16
Sugar (Corn Syrup)0.25 – 0.300.40
Sodium Lactate2.20NA
Slow Release Donors
Whey0.050.04
Edible Oils0.20 – 0.50NA
Flour (Starch)0.300.85

Cellulose0.40 – 0.80NA
Chitin2.25 – 3.00NA
Methyl Cellulose4.00 – 5.00NA
HRC‘ (Regenesis Commercial Material)12.00NA
NA – Not Analyzed.
©2001 CRC Press LLC
6 ATP. However, only 2 ATPs are produced, which implies that the organism operates
at considerably less than maximum efficiency.
59
In any fermentation reaction, there must be a balance between oxidation and
reduction. In a number of these reactions, electron balance is maintained by the
production of molecular hydrogen, H
2
. In H
2
production, protons (H
+
) of the medium,
derived from water, serve as electron acceptor. The energetics of hydrogen produc-
tion are actually somewhat unfavorable, so that most fermentative organisms only
produce a relatively small amount of hydrogen along with other fermentation prod-
ucts. Fermentation reactions that have pyruvate as an intermediate product have the
potential of producing more H
2
. Conversion of pyruvate to acetyl-CoA is an oxidation
process and the excess electrons generated must either be used to make a more
reduced end product, or can be used in the production of H
2
.
Fermentation by bacteria can also be important in controlling the biogeochemical

environment of anaerobic aquifers. Bacterial fermentation can be divided into two
categories:
14,58
Primary fermentation is the fermentation of primary substrates such as sugars, amino
acids, and lipids to yield acetate, formate, CO
2
, and H
2
, but also yields ethanol,
lactate, succinate, propionate, and butyrate. While primary fermentation often yields
H
2
, production of H
2
is not required for these reactions to occur.
Secondary fermentation or coupled fermentation is the fermentation of primary
fermentation products such as ethanol, lactate, succinate, propionate, and butyrate
to yield acetate, formate, H
2
, and CO
2
. Bacteria that carry out these reactions are
called obligate proton reducers because the reactions must produce hydrogen in
order to balance the oxidation of the carbon substrates. These secondary fermenta-
tion reactions are energetically favorable only if hydrogen concentrations are very
low (10
–2
to 10
–4
atm or 8000 to 80 nM dissolved hydrogen, depending on the

fermentation substrate). Thus these fermentation reactions occur only when the
produced hydrogen is utilized by other bacteria, such as methanogens that convert
H
2
and CO
2
into CH
4
and H
2
O. The process by which hydrogen is produced by one
strain of bacteria and utilized by another is called interspecies hydrogen transfer.
It should be noted that the terminal products of anaerobic decomposition, CH
4
, and
CO
2
, respectively, are the most reduced and the most oxidized carbon compounds.
There are a number of compounds besides the ones listed in Table 4.4 that can
be fermented to produce hydrogen (Figure 4.10). While anaerobic degradation of
BTEX compounds has been confirmed for a long time, there is still some controversy
as to whether aromatic compounds (without any oxygen in the molecule) such as
the BTEX compounds can be completely mineralized to CO
2
without alternate
electron acceptors coupled solely by fermentation with methanogenesis.
Based on a number of field observations of the presence of methane, it is well
known that fermentation occurs at almost all sites where BTEX compounds are present
in groundwater.
14,53

Since methane production requires fermentation products as meth-
anogenic substrates, the presence of methane is clear evidence that fermentation is
occurring. Metabolism of BTEX compounds to produce hydrogen probably requires
the involvement of several strains of bacteria. One possible mechanism is a series of
reactions, in which other electron acceptors are used by nonfermenters to break down
the aromatics to simpler compounds that can be used by the fermenters.
©2001 CRC Press LLC
4.2.1.4.2 Competition for H
2
In environments where hydrogen is the most important electron donor for dechlo-
rination of chlorinated solvents, competition for the uptake of hydrogen between
different types of microorganisms, such as methanogenic, homoacetogenic, sulfi-
dogenic, and dechlorinating bacteria, becomes important. In several studies it has been
shown that dechlorinating organisms have a higher affinity for molecular hydrogen
than methanogens.
27,35,55
This indicates that the dechlorinating organisms are able to
survive at lower hydrogen levels, but will possibly be outcompeted by other microor-
ganisms when elevated hydrogen levels are present. These studies suggest that a more
effective dechlorination may be achieved by using an electron donor that generates
low hydrogen concentrations during its fermentation, such as propionate or butyrate.
The speculation is that this would then create more favorable conditions for dechlo-
rinating bacteria than for hydrogen-consuming methanogens.
27,35
Reductive dechlorination of PCE requires the addition of two electrons for each
chlorine removed; for three of the seven recently identified dechlorinating organisms,
H
2
is one of the substrates (and in some cases, the only one) that can serve as the
Figure 4.10 Steps in the process of biodegradation of PCE by reductive dechlorination. As

shown, biodegradable organic matter is required as an electron donor to initiate
the process. Different types of microbes are involved at each stage. The bottom
step shows that PCE must compete for electrons with sulfate, iron, and carbon
dioxide, meaning that a large amount of organic electron donors may be needed
to supply enough electrons. Note: CDCE = cis-dichloroethene. Source: after
McCarty, 1997.
©2001 CRC Press LLC
direct electron donor. Dehalobacter restrictus is another direct dechlorinator that
uses only H
2
as an electron donor, but dechlorinates PCE only to cis-1,2-dichloro-
ethene (cisDCE).
46,47
Dehalospirillum multivorans also dechlorinates PCE to cis-
DCE using H
2
, but has a much more widely varied biochemical repertoire: it is
additionally able to use various organic substrates such as pyruvate, lactate, ethanol,
formate, and glycerol as electron donors.
7,37,44
Other PCE-dechlorinating organisms
have been isolated that do not use H
2
.
34,61
It was later determined that the half-
velocity constant with respect to H
2
for this dechlorinator was one-tenth that of the
methanogenic organisms in the culture. The threshold H

2
level for dechlorination
was also correspondingly lower than values typically reported for methanogens.
Though confirmed thus far with only this one dechlorinator, there are thermodynamic
reasons (i.e., the relatively high free energy available from dechlorination) to suspect
that the threshold for H
2
use by dechlorinators may generally be lower than that for
hydrogenotrophic methanogens.
15,27
This suggests a strategy for selective enhance-
ment of dechlorination — managing H
2
delivery so as to impart a competitive
advantage to dechlorinators.
Numerous microcosm and site studies have shown successful stimulation of
dechlorination with substrates such as methanol, ethanol, lactate, butyrate, and
benzoate.
3,32,36,62,57
However, understanding the fate of the electron donors and that
of the H
2
evolved from their degradation, as well as the extent to which their reducing
equivalents are channeled to desirable dechlorination or competing H
2
sinks, has
important implications for determining how best to effectively stimulate latent
dechlorinating activity for in situ enhanced reductive dechlorination in an IRZ.
Leading to a new school of thought, recent studies have suggested that the type
of substrates and the rate of fermentation may not have an impact on reductive

dechlorination. One study showed the ability of four fermentable substrates to sustain
PCE dechlorination long-term (i.e., approximately four months).
35
The choice of
organic substrates was based upon their rates of fermentation and the H
2
partial
pressures that could be developed and maintained. Despite the difference in the
resulting H
2
partial pressures (ranging approximately 1 ¥ 10
–5
to 3 ¥ 10
–3
atm), no
long-term effect on dechlorination was observed. This result may indicate either that
low H
2
partial pressures were not required to maintain a competitive dechlorinating
community or that several isolated PCE respiring bacteria do not utilize H
2
as an
electron donor.
43,46
H
2
was not the source of PCE-reducing equivalents in all systems
tested. Other laboratory and field studies have also suggested that the steady state
concentration of hydrogen is controlled by the type of bacteria utilizing the hydrogen
and is almost completely independent of the rate of hydrogen production.

As discussed earlier, when hydrogen is produced by fermentative organisms, H
2
is rapidly consumed by other bacteria. This utilization of H
2
by nonfermenters is
known as interspecies hydrogen transfer and is essential for fermentation reactions
to proceed.
59
Note, for example, that a glucose fermenter is unable to utilize glucose
by itself so that both the glucose fermenter and the methanogen benefit from this
symbiotic relationship.
Although H
2
is a waste product of fermentation, it is a highly reduced molecule,
which in turn makes it an excellent, high-energy electron donor. In this symbiotic
relationship, the hydrogen utilizing bacteria gain a high energy electron donor, while,
©2001 CRC Press LLC
for the fermenters, the removal of hydrogen allows continuous fermentation to be
favorable energetically.
In addition to methanogens, a wide variety of bacteria can utilize hydrogen as
an electron donor: denitrifiers, Fe (III) reducers, sulfate reducers and halorespirators.
As discussed earlier, for dechlorination to take place, halorespirators must success-
fully compete against all these hydrogen utilizers.
It was suggested that the competition is mainly controlled by the Monod half-
saturation constant K
s
(H
2
) (the concentration at which a specific bacterial strain can
utilize hydrogen at half the maximum utilization rate).

14,27,56
The measured value of
K
s
(H
2
) for halorespirators was 100 nM and for methanogens 1000 nM.
14
This led
to the suggestion that halorespirators would compete successfully for H
2
only at low
concentrations.
However, a more detailed analysis of halorespiration kinetics and competition
for hydrogen based on the Monod kinetic model was performed recently.
14,27
Using
this model, the ability of hydrogen-utilizing bacteria to compete for hydrogen can
easily be predicted from substrate concentration and two properties of the bacteria,
m
max
(maximum specific growth rate), and K
s
. Table 4.5 lists these parameters for
the various hydrogen-utilizing bacteria.
Table 4.5 illustrates that, from the m
max
term, halorespirators will outcompete
methanogens and sulfate reducers at any hydrogen concentration (since at high
substrate concentration growth rate m ª m

max
and at low substrate concentration m ª
(m
max
.S)/K
s
). However, denitrifiers will probably outcompete halorespirators under
most conditions as their maximum specific growth rate is approximately three times
faster than halorespirators’. Based on these detailed analyses and the synthesis of
wide ranging data from field observations, the following probable sequence takes
place at most sites undergoing halorespiration reactions:
14,27
• Aerobic bacteria consume nonchlorinated organic substrates until the oxygen is
depleted; to implement enhanced reductive dechlorination, oxygen depletion is
forced intentionally in an IRZ.
•Similarly, denitrifying bacteria will consume nonchlorinated organic substrates
until the nitrate is exhausted; nitrate depletion will be forced for enhanced reduc-
tive dechlorination.
• Iron reducing bacteria consume nonchlorinated organic substrates until the avail-
able Fe (III) is expended.
Table 4.5Maximum SpeciÞc Growth Rate
(mm
mm
max
) and Half Saturation
CoefÞcient (K
s
) for Various H
2
-

Utilizing Bacteria (ModiÞed from
Wiedermeier et al., 1999)
Bacterial Strainmm
mm
max
(hr
–1
)K
s
(mg/L)
Halorespirator0.0199500.0002
DenitriÞer0.058080—
Sulfate Reducer0.003936—
Methanogen0.0037920.0019