Artículo en PDF
How to cite
Complete issue
More information about this article
Journal's homepage in redalyc.org
Sistema de Información Científica
Red de Revistas Científicas de América Latina y el Caribe, España y Portugal
REVISTA MEXICANA DE INGENIERIA QUIMICA
Vol. 4
(2005)
273-287
Publicado por la Academia Mexicana de Investigación y Docencia en Ingeniería Química, A. C.
273
ANAEROBIC BIOREMEDIATION–AN UNTAPPED POTENTIAL
BIORREMEDIACION ANAEROBICA-UN POTENCIAL SIN EXPLOTAR
N. Balagurusamy
Universidad Autónoma de Coahuila, Facultad de Ciencias Químicas, Departamento de Biotecnología, Coahuila,
Saltillo. C .P. 25280.
Received May 16 2005; Accepted October 17 2005
Abstract
Recently there has been considerable interest in employing bioremediation technologies for treatment of wastes and
for reclamation and restoration of contaminated ecosystems. In this technology, microorganisms or their constituents
such as enzymes are used to degrade or transform the wastes. Though, contaminated ecosystems lack oxygen and
favor the growth and activity of anaerobic bacteria, most of the bioremediation technology employs aerobic
microorganisms. However, treatment of industrial wastewaters is the only area in which anaerobic technology is
widely employed presently. Recent advances in molecular ecology helps us in understanding the diversity of
anaerobic bacteria, their processes and their important role in global cycle of carbon, nitrogen, sulfur. In addition,
different anaerobic bacterial groups possess the ability to use different types of electron acceptors such as nitrate,
sulfate, and carbonate for degradation of organic contaminants or for biotransformation heavy metals. Still, the
potential of anaerobic bacteria for bioremediation of various contaminants is not capitalized. This paper deals with
application of anaerobic bioremediation for biorestoration of land and water ecosystems contaminated with
hydrocarbons, chlorinated compounds and heavy metals.
Keywords
: bioremediation, anaerobic bacteria, hydrocarbons, chlorinated compounds, heavy metals.
Resumen
Recientemente ha habido un interés considerable por el empleo de tecnologías del biorremediación para el
tratamiento de basuras y para la restauración de ecosistemas contaminados. En esta tecnología, los microorganismos
o algunos de sus componentes como las enzimas son usados para degradar o transformar las basuras. Dado que a los
ecosistemas contaminados les falta oxígeno, se favorece el crecimiento y actividad de bacterias anaerobias, aunque la
mayoría de la tecnología del biorremediación emplea los microorganismos aerobios. Sin embargo, el tratamiento de
aguas residuales industriales representa la única área en que la tecnología anaerobia es extensamente aplicada.
Logros recientes en ecología molecular nos ayudan a entender la diversidad de bacterias anaerobias, sus procesos y
su papel importante en el ciclo global de carbono, el nitrógeno y el azufre. Además, los diferentes grupos de
bacterias anaerobias poseen la habilidad de usar tipos diferentes de aceptores de electrones como el nitrato, el sulfato
y carbonato para la degradación de contaminantes orgánicos o para la biotransformación de los metales pesados.
Todavía, el potencial de bacterias anaerobias para la biorremediación de varios contaminantes no se capitaliza. Este
artículo revisa la aplicación de la biorremediación anaerobia para la biorrestauración de ecosistemas de tierra y agua
contaminados con hidrocarburos, compuestos clorados y metales pesados.
Palabras clave
: biorremediación, bacterias anaerobias, hidrocarburos, compuestos tratados con cloro, metales
pesados.
1. Introduction
Bioremediation technology uses the
metabolic potential of the microorganisms to
clean up the contaminated land and water.
The microorganisms may be indigenous to a
contaminated area or they may be isolated
from
elsewhere
and
brought
to
the
contaminated site. Contaminant compounds
are transformed by living organisms through
reactions that take place as a part of their
metabolic processes. As bioremediation can
be effective only where environmental
conditions permit microbial growth and
activity. Its application often involves the
manipulation of environmental parameters to
*Corresponding author:
E-mail:
bnagamani@mail.uadec.mx
Phone:
Tel. (84) 44155752 Ext.22, Fax: (84) 44159534
AMIDIQ
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
274
allow microbial growth and degradation to
proceed at a faster rate (Vidali, 2001). A
large number of microorganisms have been
isolated in recent years that are able to
degrade compounds that are previously
considered to be non-degradable. This shows
that
under
selective
pressure
of
environmental
pollution,
microorganisms
develop catabolic capacity either to degrade
or convert them to innocuous products
(Timmis and Pieper, 1999). However most of
the processes do not function optimally. The
design of improved biocatalysis involves
different aspects of optimization, such as;
creating new metabolic routes, expanding the
substrate ranges of existing pathways,
improving the process-relevant properties of
microorganisms, etc. Until recently, practical
applications of
in situ
bioremediation have
focused mostly on aerobic microorganisms,
which gain energy by oxidizing organic
compounds to carbon dioxide with oxygen
serving as the electron acceptor. However,
this approach has had limited success, not
least
because
oxygen,
an
absolute
requirement for aerobes, is scarce in almost
all contaminated environments. The scarcity
of oxygen in many contaminated sub surface
environments has raised interest in the
bioremediation potential of anaerobes.
The most ancient of all life processes is
anaerobic microbial metabolism. Billions of
years ago, our entire planet was anaerobic.
Living organisms that began to evolve in the
warm oceans had no choice but to follow
anaerobic
pathways.
Free
atmospheric
oxygen, as we know it today, simply did not
exist. All molecular oxygen was bound in
water, carbon dioxide, carbonates, and
sulfates. The world arose anaerobically, and
much of it remains so in terms of the
numbers of living things rather than their
size. Anaerobic bacteria are present in soil
and are a part of the normal flora of humans
and all other animals, as well as the insects
examined so far. This microbial life in the
absence of oxygen is beginning to show
significant potential for solving one of the
important
present
day
problems
of
environmental pollution and degradation
(Coates and Anderson, 2000). Lu
et al
.
(1999) indicated that anaerobic processes
naturally found in subsurface account for
bulk of organic contaminant degradation in
aquifers.
Anaerobes
oxidize
organic
compounds to carbon dioxide but use
electron acceptors such as nitrate, sulfate, or
Fe
3+
oxides instead of oxygen. The diverse
metabolic capabilities of anaerobes represent
a potentially potent force in the fight against
contamination. In sediments in which
anaerobic
processes
are
established,
stimulating the anaerobic community has the
advantage of enhancing an active and
acclimated microbial population. If the
natural rates of contaminant degrada
tion are
too slow, increased levels of alternate
electron acceptors may accelerate the rate of
anaerobic contaminant degradation.
Electron
acceptors such as sulfate and nitrate do not
have these limitations because they are highly
soluble and are not consumed by non-
biological processes (Lovley, 2001). In this
review,
potential
and
applications
of
anaerobes in remediation of different
contaminants is discussed.
2. Anaerobic bioremediation of
hydrocarbon contamination
Hydrocarbons are one of the most
important groups of chemicals to mankind
because of their natural abundance, industrial
importance and their extensive use as a
primary
energy
source.
Petroleum
hydrocarbon contamination is becoming a
great concern due to the toxicity and
recalcitrance of many of the fuel components.
The majority of bioremediation strategies for
removal of petroleum hydrocarbon are
aerobic respiration. Prior to the 1980s, it was
accepted
that
microbial
hydrocarbon
degradation occurs mainly under aerobic
conditions due to favorable energetics and
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
275
that anaerobic hydrocarbon degradation was
negligible (Atlas, 1981). However there are
several considerations in the use of aerobic
bioremediation technologies. Normally the
hydrocarbon contaminated soils lack oxygen.
Further urea and ammonia-based fertilizers
are added sometimes for biostimulation,
which also can potentially exert an oxygen
demand due to biological ammonia oxidation.
In addition, mass transfer of oxygen may not
be sufficient to replenish oxygen consumed
by microbial metabolism. Under such
conditions
anaerobic
hydrocarbon
degradation may be of relevance. Recently
anaerobic hydrocarbon metabolism and the
importance
of
anaerobic
hydrocarbon
metabolism
in
contaminated
anoxic
environments is being widely reported
(Coates
et al
., 1997; Caldwell
et al
., 1998;
Wilkes
et al
., 2002, Chakraborty and Coates,
2004). Different mechanisms for microbial
utilization of aromatic compounds are
presented in Fig. 1. Several denitrifying,
manganese-iron and sulfate-reducing bacteria
that have the ability to degrade simple
aromatic or aliphatic hydrocarbons under
anoxic conditions have been isolated (Leahy
and Olsen, 1997; Rabus and Widdell, 1995;
Rueter
et al
., 1994; Zhou
et al
., 1995).
Gibson and Harwood (2002) reviewed the
microbial diversity of anaerobic bacteria
utilizing aromatic hydrocarbons.
Chemotrophic,
aerobic
Phototrophic,
O
2
H
2
O
a
n
o
x
y
g
e
n
i
c
C
O
2
Cell mass
Light
Cell mass
----------------------------------
Hydrocarbons
-------------------
CO
2
C
n
H
m
NO
3
-
N
2
CO
2
Cell mass
Fe (III)
Fe (II)
CO
2
Cell mass
Chemotrophic,
SO
4
2-
H
2
S
anaerobic
CO
2
Cell mass
CH
4
CO
2
Cell mass
Fig. 1. Microbial utilization of hydrocarbons with different terminal e
-
acceptors. Reprinted with
permission from Widdel and Rabus (2001), Copyright 2001. Elsevier Science Ltd.
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
276
Several studies have also demonstrated
that monoaromatic hydrocarbons such as
benzene, toluene, and xylene (BTEX) and
hexadecane can be biodegraded in the
absence of oxygen (Coates
et al
., 1996a;
Coates
et al
., 1996b; Edwards and Grbic-
Galic, 1992; Kazumi et al., 1995; Rabus and
Widdell, 1995). Anaerobic degradation of
aliphatic hydrocarbon has also been reported
and has been linked to denitrification
(Bregnard
et al.,
1997; Chayabutra and Ju,
2000; Wilkes
et al
., 2002), sulfate reduction
(Coates
et al
., 1997; Kropp
et al.,
2000;
Rueter
et al.,
1994; So and Young, 1999) and
methanogenesis (Anderson and Lovley,
2000; Zengler
et al
., 1999). Most of the
reports related to the anaerobic mineralization
of
aliphatic
hydrocarbons
are
with
enrichments or pure cultures in laboratory
scale (Chayabutra and Ju 2000; So and
Young 1999; Zengler
et al.,
1999). Metabolic
pathways
of
degradation
of
aromatic
compounds by different groups of anaerobic
bacteria can be referred in these review
articles (Chakraborty and Coates, 2004; Diaz,
2004; Gibson and Harwood, 2002; Heider
and Fuchs, 1997; Widdel and Rabus, 2001).
Significance
of
these
results
in
the
bioremediation of contaminated soils and
sediments is not yet completely investigated
and only in few cases the anaerobic
degradation of alkanes in environmental
samples have been reported (Anderson and
Lovley, 2000).
Polycyclic
aromatic
hydrocarbons
(PAHs) are of most concern due to their
toxicity,
low
volatility,
resistance
to
microbial degradation, and high affinity for
sediments. Although several authors have
reported that PAHs are not degraded under
strict anaerobic conditions (Heitkamp
et al
.,
1987;
Mihelcic
and
Luthy,
1988),
degradation of PAHs in the absence of
oxygen with nitrate as the apparent electron
acceptor has also been demonstrated (Leduc
et al
., 1992). Though degradation of
polycyclic aromatic hydrocarbons has been
confirmed, the rate of anaerobic hydrocarbon
degradation was generally lower (Coates
et
al
., 1997; Caldwell
et al
., 1998) than
equivalent
aerobic
degradation
rates.
Nonetheless, removal of the contaminant
hydrocarbons was just as extensive under
anoxic conditions (Coates
et al
., 1997;
Caldwell
et al
., 1998) and substantial
degradation of high molecular weight
n
-
alkanes and the isoprenoid hydrocarbons
pristane and phytane was also observed
(Caldwell
et al
., 1998).
Though there are several pure culture
examples of nitrate-reducing, Fe (III)-
reducing, and sulfate-reducing bacteria that
are capable of completely oxidizing some of
these hydrocarbon contaminants to carbon
dioxide (Coates
et al
., 1996a; Rabus and
Widdell, 1995; Rueter
et al
., 1994), there are
only few reports on the efficiency of these
cultures in the degradation of complex
hydrocarbon
mixtures
under
anaerobic
conditions. In marine environments, the most
important
terminal
electron-accepting
processes are iron, manganese and sulfate
reduction (Canfield
et al
., 1993), and the
limited number of studies indicated that the
process
of
anaerobic
hydrocarbon
degradation in marine environments is
associated primarily with sulfate reduction
(Coates
et al
., 1997; Caldwell
et al
., 1998). In
contrast nitrate was observed to stimulate
hydrocarbon degradation in terrestrial and
freshwater environments. Coates
et al.
(1997)
demonstrated that a wide variety of
hydrocarbon contaminants can be degraded
under sulfate-reducing conditions, and they
suggested that it is possible to use sulfate
reduction rather than aerobic respiration as a
treatment
strategy
for
hydrocarbon-
contaminated dredged sediments. In addition,
the authors observed that inoculation of high
activity sediment samples to low activity
sediments stimulated anaerobic hydrocarbon
degradation significantly.
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
277
3. Anaerobic bioremediation of
chlorinated compounds
Chlorinated
aliphatic
hydrocarbons
(CAH) are manmade organic compounds.
CAH are used in a wide variety of
applications, such as solvents and degreasers
and in the manufacturing of raw materials.
Some of the most frequently occurring CAHs
in soil and groundwater are tetrachloroethene
(PCE),
trichloroethene
(TCE),
carbon
tetrachloride
(CT),
chloroform
(CF),
methylene chloride (MC), trichloroethane
(TCA), dichloroethane (DCA), chloroethane
(CA),
dichloromethane
(DCM),
chloromethane (CM),
etc
. Of the various
CAHs,
TCE
is
the
most
prevalent
contaminant.
Under anaerobic conditions, direct
metabolism by reductive dechlorination,
cometabolism and fermentation are the three
mechanisms employed by anaerobic bacteria
for effective biodegradation of CAHs.
Reductive dechlorination generally involves
the sequential replacement of a chlorine atom
on a CAH with a hydrogen atom and has
been observed to occur both directly and
cometabolically. Reductive dechlorination
theoretically is expected to occur under most
anaerobic conditions, but has been observed
to be most effective under sulfate-reducing
and methanogenic conditions (USEPA,
1998). As in the case of aerobic oxidation,
the direct mechanisms may biodegrade CAHs
faster
than
cometabolic
mechanisms
(McCarty
and
Semprini,
1994).
In
cometabolism, often the amount of primary
substrate required is a factor of 100 to 1,000
times the amount of the CAH. In direct
metabolism (respiration with only the
chlorinated solvent as the electron acceptor),
the stoichiometry is much more favorable,
and a much smaller amount of supplemental
chemical is required (Bouwer, 1994). In the
case
of
fermentation,
the
chlorinated
hydrocarbon serves as an electron donor,
electron acceptor and carbon source. Mägli
et
al
. (1996) isolated a strict acetogenic
anaerobic bacterium capable of growth on
dichloromethane and reported that the
fermentation of dichloromethane to formate
and acetate. Fermentative dechlorination is an
energetically favorable process (Table 2).
The diversity of dechlorinating anaerobic
bacteria (El Fantroussi
et al
., 1998) and their
energy metabolism (Holliger
et al
., 1999)
have been reviewed earlier.
4. Direct anaerobic reductive
dechlorination
In
direct
anaerobic
reductive
dechlorination bacteria gain energy and grow
as one or more chlorine atoms on a
chlorinated hydrocarbon are replaced with
hydrogen (Gerritse
et al
., 1999). In that
reaction, the chlorinated compound serves as
the electron acceptor, and hydrogen serves as
the direct electron donor (Fennel
et al
.,
1997). Hydrogen used in the reaction
typically is supplied indirectly through the
fermentation of organic substrates. The
reaction when coupled with growth is also
referred
to
as
halorespiration
or
dehalorespiration
(Gossett
and
Zinder,
1997). The Gibbs free energy change for all
major half-reactions of the reduction of
chlorinated methanes, ethanes and ethenes
and electron acceptors such as sulfate, nitrate
and carbon dioxide are given in Table 1.
From these data, it can be seen that the
reductive dechlorination of nearly all CAHs
is an energetically favorable (exergonic)
reaction under standard conditions. This
implies
that
biological
reductive
dechlorination processes may occur in nature.
The anaerobic reductive dechlorination
of the more chlorinated CAHs (PCE and
TCE)
occurs
more
readily
than
the
dechlorination of CAHs that already are
somewhat reduced,
viz
., dichloroethene
(DCE) and vinyl chloride (VC).
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
278
Table 1. Half reactions of reductive dechlorination of some chlorinated hydrocarbons in comparison with
other commonly used terminal electron acceptors.
Electron
acceptor
Half reaction of reductive transformations
Δ
Gº’/ electron
(kJ)
*
O
2
O
2
+ 4 H
+
+ 4e
-
2 H
2
O
-78.7
MnO
2
MnO
2
+ HCO
3
-
+ 3H
+
+ 2e
-
MnCO
3
+ 4H
2
O
-58.9
NO
3
-
NO
3
-
+ 2H
+
+ 2e
-
NO
2
-
+ H
2
O
-41.7
Fe(OH)
3
Fe(OH)
3
+ 3H
+
+ e
-
Fe
2-
+ 3H
2
O
-11.4
SO
4
2-
SO
4
2-
+ 9H
+
+ 8e
-
HS
-
+ 4H
2
O
+20.9
HCO
3
-
HCO
3
-
+ 9H
+
+ 8e
-
CH
4
+ 3H
2
O
+23.0
PCE
C
2
Cl
4
+ H
+
+ 2e
-
C
2
HCl
3
+ Cl
-
-55.3
TCE
C
2
HCl
3
+ H
+
+ 2e
-
cis
-C
2
H
2
Cl
2
+ Cl
-
-53.0
TCE
C
2
HCl
3
+ H
+
+ 2e
-
trans
-C
2
H
2
Cl
2
+ Cl
-
-50.9
TCE
C
2
HCl
3
+ H
+
+ 2e
-
1,1-C
2
H
2
Cl
2
+ Cl
-
-50.8
cis
-DCE
cis
-C
2
H
2
Cl
2
+ H
+
+ 2e
-
C
2
H
3
Cl + Cl
-
-38.3
trans
-DCE
trans
-C
2
H
2
Cl
2
+ H
+
+ 2e
-
C
2
H
3
Cl + Cl
-
-40.4
1,1-DCE
1,1-C
2
H
2
Cl
2
+ H
+
+ 2e
-
C
2
H
3
Cl + Cl
-
-40.5
VC
C
2
H
3
Cl + H
+
+ 2e
-
C
2
H
4
+ Cl
-
-43.4
DCM
CH
2
Cl
2
+ H
+
+ 2e
-
CH
3
Cl + Cl
-
-47.5
CM
CH
3
Cl + H
+
+ 2e
-
CH
4
+ Cl
-
-45.2
TCA
C
2
H
3
Cl
3
+ H
+
+ 2e
-
C
2
H
4
Cl
2
+ Cl
-
-54.1
1,2-DCA
CH
2
Cl-CH
2
Cl + 2e
-
C
2
H
4
+ 2Cl
-
-71.3
DCA
C
2
H
4
Cl
2
+ H
+
+ 2e
-
C
2
H
5
Cl + Cl
-
-38.3
1,2-DCA
CH
2
Cl-CH
2
Cl + H
+
+ 2e
-
C
2
H
5
Cl + Cl
-
-36.2
CA
C
2
H
5
Cl + H
+
+ 2e
-
C
2
H
5
+ Cl
-
-44.5
CT
CCl
4
+ H
+
+ 2e
-
CHCl
3
+ Cl
-
-65.0
CF
CHCl
3
+ H
+
+ 2e
-
CH
2
Cl
2
+ Cl
-
-54.0
* Calculated on the basis of data from Thauer
et al
. (1977), Dolfing and Janssen (1994) and Vogel
et al
. (1987) under the following standard
conditions; H
+
= 10
-7
M; Cl
-
= 10
-3
M; T=25ºC.
Table 2. Fermentative dechlorination of few chlorinated hydrocarbons and Gibb´s free energy of the
reactions.
Compound
Reaction
Δ
Gº’ (kJ)
TCM
CH
3
CCl
3
+ 2H
2
O
CH
3
COOH + 4HCl
-379.5
DCM
3CH
2
Cl
2
+ CO
2
+ 4H
2
O
CH
3
COOH + 2HCOOH + 6HCl
-685.2
CM
4CH
3
Cl + 2CO
2
+ 2H
2
O
3CH
3
COOH + 4HCl
-455.5
CT
CCl
4
+ 2H
2
O
CO
2
+ 4HCl
-619.7
CF
4CHCl
3
+ 6H
2
O
CH
3
COOH + 2CO
2
+ 12HCl
-416.6
Accumulation of DCE and VC are
observed in anaerobic environments. It also
has been observed that, while VC can be
effectively dechlorinated, the presence of
PCE in groundwater may inhibit the
anaerobic reductive dechlorination of VC
(Tandoi
et al
., 1994). VC is more commonly
remediated using aerobic mechanisms than
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
279
anaerobic
mechanisms.
In
anaerobic
environments in which VC accumulates,
enhanced aerobic bioremediation can be
implemented to degrade the VC. Recent
studies
have
demonstrated
significant
anaerobic oxidation of VC to carbon dioxide
under Fe (III)-reducing conditions and of
DCE to VC and VC to carbon dioxide under
humic acid-reducing conditions (Bradley and
Chapelle, 1998). These studies suggest the
possibility of alternative biotransformation
mechanisms under anaerobic conditions.
Hydrogen has been observed to be an
important electron donor in anaerobic
reductive dechlorination (Fennell
et al
.,
1997). The presence of hydrogen establishes
a competition between the bacteria that
mediate
the
anaerobic
reductive
dechlorination
(such
as
Dehalococcus
ethenogenes
and
Dehalospirillium
multivorans
) and methanogenic bacteria that
also use hydrogen as an electron donor.
However, it has been observed that the
dechlorinating bacteria can survive at a
partial pressure of hydrogen ten times lower
than that at which the methanogenic bacteria
can survive (Smatlak
et al
., 1996), thus
providing an opportunity to support the
dechlorinating
bacteria
by
providing
hydrogen at a slow rate (Fennell
et al
., 1997).
In
addition,
in
some
subsurface
environments, competition from nitrate or
sulfate-reducing bacteria may limit both
methanogenic activity and the extent of
anaerobic reductive dechlorination. Studies
have shown that anaerobic reduction of
CAHs can occur by reductive dechlorination
in a variety of environmental conditions
(Beeman
et al
., 1994; Semprini
et al
., 1995).
However, the efficiency of the anaerobic
dechlorination processes at high redox
potential values is
limited; efficiency
improves as the redox potential decreases.
Pilot studies have been conducted at a variety
of sites to examine the feasibility of
stimulating
in situ
anaerobic reductive
dechlorination by providing to the subsurface
simple organic substrates, such as lactate,
butyrate, methanol, ethanol, and benzoate
(Harkness
et al
., 1999).
5. Cometabolic anaerobic reductive
dechlorination
In cometabolic anaerobic reductive
dechlorination, a chlorinated hydrocarbon is
fortuitously degraded by an enzyme or
cofactor
produced
during
microbial
metabolism of another compound. In such a
case, biodegradation of the chlorinated
compound does not appear to yield any
energy
or
growth
benefit
for
the
microorganism
mediating
the
reaction
(Gossett and Zinder, 1997). In addition,
enzyme systems such as iron-sulfur clusters,
cobalamins, factor F430 or hematin can take
part in side reactions that yield cometabolic
transformation of CAHs. All these enzyme
systems contain redox-active metal centers
and are referred to as transition-metal
cofactors. These transition-metal cofactors
act as electron carriers in anaerobic bacteria.
As electron transfer by these carriers is not
very specific, a wide range of CAHs can be
transformed (Gantzer and Wackett, 1991).
Cometabolic
anaerobic
reductive
dechlorination has been observed for PCE,
TCE, DCE, VC, DCA, and CT under
anaerobic conditions (Fathepure
et al
., 1987;
Yager
et al
., 1997). In pilot- and full-scale
applications, it is generally difficult to
distinguish between direct and cometabolic
anaerobic reductive dechlorination reactions.
Both biodegradation mechanisms are referred
to more generally as anaerobic reductive
dechlorination.
In
laboratory-scale
applications,
direct
and
cometabolic
anaerobic reductive dechlorination reactions
can be distinguished. Several investigators
have suggested that the most efficient
bioremediation of CAHs will occur in
aquifers that are characterized by an up-
gradient anaerobic zone and a down-gradient
aerobic zone (Bouwer, 1994). In the up-
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
280
gradient
zone,
anaerobic
reductive
dechlorination of PCE might degrade to TCE,
and eventually to VC. VC could then be
degraded in the down-gradient aerobic zone
of the CAH plume.
6. Anaerobic bioremediation of metal
contamination
In case of toxic metals and metalloids,
they are often soluble, and thus mobile in
aerobic systems. However, under anoxic
conditions, microorganisms reduce them to
insoluble forms and immobilize them as
precipitates. Different mechanisms used by
anaerobic
bacteria
in
the
metals
biotransformation are given in Fig. 2.
Elements that may be immobilized in this
way include chromium, uranium, technetium,
cobalt,
and
selenium.
Until
recently,
investigations of terrestrial bioremediation
primarily focused on the treatment of soils
and waters contaminated with organic
pollutants (Blackburn, 1998; Head, 1998).
Anaerobic Respiration
Precipitation
e
-
acceptor (M)
H
2
S + M
MS
Biosorption
M
oxd
M
red
Enzymatic reduction
Fig. 2. Schematic representation of the mechanisms of metal transformation by different anaerobic
bacteria (M-Metal; MS-Metal sulfide; M
oxd
– Metal oxidized form; M
red
– Metal reduced form).
There has been little investigation into
the
use
of
microorganisms
for
the
remediation of metal contamination. In
recent years there is considerable interest on
the exploitation of microorganisms to
ameliorate toxic metal contamination in
terrestrial environments. Though, use of
anaerobic microorganisms for remediation of
toxic
metal
contamination
is
mostly
laboratory-based research (Schmieman
et al
.,
1997; Tucker
et al
., 1998), field level studies
utilizing microorganisms for toxic metal
remediation are becoming more prevalent.
Lovley and Coates (1997) observed that
remediation is due to the result of changes in
the
redox
state
of
metal
ions.
Microorganisms can remove toxic metals
and metalloids by converting them to forms
that are precipitated or volatilized from
solution. In other instances, microbial
alteration of the redox state of either the
contaminants or the Fe3+ and Mn4+ oxides,
which bind most heavy metals, can make
metals and metalloids more soluble. This can
aid in the leaching of these contaminants
from soils. The adsorption of metals and
metalloids onto microbial biomass can also
prevent
further
migration
of
these
contaminants. Furthermore, microbe–metal
interactions can play an important role in the
remediation
of
organic
contamination
because microorganisms that use Fe3+ or
sulfate as the terminal electron acceptor can
remove organic contaminants from the
environment (Lovley and Coates, 1997).
Chromium is recognized as a serious
pollutant among heavy metals in the
environment. Indeed, it is one of the most
widely used metals in industry. Hexavalent
chromium is considered to be much more
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
281
toxic than trivalent chromium. Current
treatment
techniques
for
chromium-
containing wastes involve aqueous reduction
of Cr (VI) to Cr (III) by means of a chemical
reducing agent to precipitate the less soluble
Cr (III) (Eary and Rai, 1988).
In case of
bioremediation of chromium, a wide variety
of microorganisms, such as
Escherichia coli
,
Enterobacter cloacae
(Shen and Wang,
1993),
Deinococcus
radiodurans
(Frederickson
et al
., 2000),
Pseudomonas
fluorescens
(DeLeo and Ehrlich, 1994), iron
and manganese reducing
Pyrobaculum
islandicum
(Kashefi and Lovley, 2000),
denitrifying bacterial consortia (Schmieman
et al
., 1997) and anaerobic sulfate reducing
bacteria (SRB), are able to reduce Cr (VI)
and thereby able to detoxify the polluted
environments (Gadd, 2000; Lovley, 1994,
1995; Lovley and Coates, 1997; White
et al.
,
1997). Marsh and McInerney (2001)
developed a consortium dependent on
hydrogen for growth and Cr (VI) reduction.
This technique appears to be efficient,
environmentally friendly and cheaper than
currently used processes such as treatment
with lime (Barkay and Schaefer, 2001; Sen
and Johnson, 1999). Due to their resistance
to high heavy metal concentrations, SRB
appear to be the best candidates for this
process (Lovley, 2001). Cr (VI) is a known
oxidizing agent (Mahan, 1967) and causes
stress to the bacteria. The length of the lag
phase increases with Cr (VI) concentration.
But when substrate is not limiting, growth
occurred in presence of chromium, leading
to the removal of Cr (VI) from chromium-
polluted effluents (Michel
et al
., 2001).
Turick
et al
. (1996) reported that 92% of
anaerobic Cr (VI)-reducing bacteria from
soils were capable of greater than 30% Cr
(VI)-reduction.
The use of sulfate as terminal electron
acceptor, associated with the presence of
very low redox potential cytochromes
(cytochromes c3), produces sulfide, which
can easily reduce a large number of heavy
metals and precipitate them as metallic
sulfides (Kim
et al
., 2001; White and Gadd,
1996).
White
et al
. (1997) leached metals
from the contaminated soil and stripped the
toxic metals by a mixed culture of sulfate-
reducing bacteria, which precipitated the
metals as solid metal sulfides. Using this
integrated procedure, 69% of the toxic
metals
present
in
an
industrially
contaminated soil were removed in 75 days.
The procedure produced a liquid effluent of
sufficiently low metal concentration that it
could be safely discharged into the
environment. This integrated procedure will
probably be more commonly applied to
ex
situ
applications (e.g. slurry reactors) but its
successful use in soil is a prime example of
advances in this field.
Chromate reduction is controlled not
only by chemical reduction but also by an
enzymatic process. T
he enzymatic reduction
of Cr (VI) by cytochrome c
3
and
hydrogenase has recently been demonstrated
in
Desulfomicrobium
norvegicum
and
Desulfovibrio vulgaris
strain Hildenborough
(Lovley and Phillips, 1994; Tucker
et al
.,
1998).
It has also been demonstrated that the
metal reductase activity of
Desulfuromas
acetoxidans
and
Desulfovibrio vulgaris
is
associated with the polyhemic
c
-type
cytochrome (Assfalg
et al
., 2002).
Further heavy metals are used as
terminal electron acceptors and the ability to
use this reaction for energy conservation and
growth is strain-dependent. For instance,
Dv
H can reduce Cr (VI) using several
enzymes involved in the electron chain
transfer, but reduction of this metal does not
support growth (Chardin
et al.
, 2002).
Ganesh
et al
., (1999) reported that
there was effective precipitation of uranium
from U (VI)-containing waste streams by
D.
desulfuricans
. Technetium (
99
Tc), a fission
byproduct of
235
U, is a contaminant in waste
streams of the nuclear-fuel cycle. The highly
soluble pertechnetate anion (TcO
4
-
) has been
reduced and precipitated from solution in a
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
282
flow-through bioreactor using
Desulfovibrio
desulfuricans
(Lloyd
et al
., 1999). Tc(VII)
precipitation rates by this organism were
higher than Fe(III)-reducing
Shewanella
and
Geobacter
species (Lloyd and Macaskie,
1996). This implies that sulfate-reducing
organisms are the optimal candidates for Tc
(VII) removal from waste streams. Thus,
direct and indirect anaerobic microbial
reduction offers potential mechanisms for
immobilizing these contaminants because
many metals, such as uranium (U) and
chromium, are less soluble in the reduced
valence state (Lovley and Coates, 1997). In
addition, bioremediation technologies using
anaerobic bacteria are feasible alternatives to
physical cleansing of soils and chemical or
physical concentration of metals in polluted
waters. Since the toxic metal seemed to
undergo a degree of precipitation around the
bacterial cells (Valls
et al
., 2000), metal
immobilization in soil using bacteria (and
the
corresponding
decrease
in
bioavailability) could have a longer term
effect, not unlike addition of chemical
amendments, such as zeolite, beringite and
hydroxyapatite (Oste
et al
., 2001; Seaman
et
al
., 2001). Microorganisms could therefore
be employed to immobilize metals in
moderately polluted fields, thus allowing
their use in agriculture (Valls
et al
., 2000). It
has also been shown that inoculation of
metal-resistant bacteria into soils protected
the indigenous bacterial community from the
effects of heavy metals (Stephen
et al
.,
1999).
Conclusions
Anaerobic strategies for
in situ
bioremediation
are
promising,
but
substantial research remains to be done
before they can be widely adopted.
Studies
of
in situ
bioremediation of hydrocarbons
coupled
to
Fe
3+
reduction,
reductive
dechlorination,
and
stimulated
metal
reduction
suggest
that
the
anaerobic
microorganisms involved in these processes
in the contaminated aquifers are closely
related to those that carry out these
bioremediation reactions in pure culture in
the lab. But s
ome reports suggest that the
anaerobic processes are effective at some
sites, but not so in other cases. It is not clear
whether this is due to the variability or
heterogeneity in the distribution of the
anaerobes or in environmental factors
controlling
their
activity.
A
clear
understanding on the diversity, distribution
of anaerobic bacteria in diverse natural
environments
and
their
metabolic
mechanisms will shed more light on the
above observed discrepancies. These studies
will provide better knowledge on the natural
attenua
tion and promote rational design of
strategies for accelerating bioremediation
using anaerobes.
In addition, understanding
the physiology and genetics of anaerobic
populations will be very useful in assessing
their potential and to give a thrust on
bioremediation using anaerobic bacteria.
References
Anderson, R.T. and Lovley, D.R. (2000).
Hexadecane decay by methanogenesis.
Nature
404
, 722–723.
Assfalg, M., Bertini, I., Bruschi, M., Michel, C.
and Turano, P. (2002). The metal
reductase activity of some multiheme
cytochrome
c
:
chromium
(III)
by
cytochrome
c
7
.
Proceedings of the
National Academy of Science USA 99,
9750–9754.
Atlas, M. (1981). Microbial degradation of
petroleum hydrocarbons: an environmental
perspective.
Microbiological Reviews
45,
180-209.
Barkay, T. and Schaefer, J. (2001). Metal and
radionuclide
bioremediation:
issues,
considerations and potentials.
Current
Opinion in
Microbiology
4,
318–323.
Beeman, R.E., Howell, J.E., Shoemaker, S.H.,
Salazar, E.A. and Buttram, R. (1994). A
field evaluation of in situ microbial
reductive
dehalogenation
by
the
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
283
biotransformation of chlorinated ethenes.
In:
Bioremediation of Chlorinated and
Polycyclic
Aromatic
Hydrocarbon
Compounds.
(R.E.Hinchee, A. Leeson, L.
Semprini and S.K. Ong, eds.). Pp. 14-27.
Lewis Publishers. Boca Raton. Blackburn,
J.W. (1998). Bioremediation scale-up
effectiveness: a review.
Bioremediation
Journal 1,
265-281.
Blackburn, J.W. (1998). Bioremediation scale-up
effectiveness: a review.
Bioremediation
Journal
1, 265-282.
Bouwer,
E.J.
(1994).
Bioremediation
of
chlorinated
solvents
using
alternate
electron acceptors. In:
Handbook of
Bioremediation
. (R.D. Norris, ed.). Lewis
Publishers, Inc. Boca Raton, Florida.
Bradley, P.M. and Chapelle, F.H. (1998). Effect
of contaminant concentration on aerobic
microbial mineralization of DCE and VC
in stream-bed sediments.
Environmental
Science and Technology
32,
553-557.
Bregnard, T.P-A., Häner, A., Höhener, P. and
Zeyer, J. (1997). Anaerobic degradation of
pristane in nitrate-reducing microcosms
and enrichment cultures.
Applied and
Environmental Microbiology
63
, 2077–
2081.
Caldwell, M.E., Garrett, R.M., Prince, R.C. and
Suflita,
J.M.
(1998).
Environmental
Environmental Science and Technology
63,
2077–2081.
Caldwell, M.E., Garrett, R.M., Prince, R.C. and
Suflita,
J.M.
(1998).
Anaerobic
biodegradation of long-chain
n
-alkanes
under
sulfate-reducing
conditions.
Environmental Science and Technology
32
, 2191-2195.
Canfield, D., Thamdrup, B. and Hansen, J.W.
(1993). The anaerobic degradation of
organic
matter
in
Danish
coastal
sediments: Fe reduction, Mn reduction and
sulfate
reduction.
Geochimica
et
Cosmochimica
57,
3867-3883.
Chakraborty, R. and Coates, J.D. (2004).
Anaerobic degradation of monoaromatic
hydrocarbons.
Applied
Microbiology
Biotechnology
64,
437-446.
Chardin, B., Dolla, A., Chaspoul, F., Fardeau,
M.L., Gallice, P. and Bruschi, M. (2002).
Bioremediation
of
chromate:
thermodynamic analysis of the effects of
Cr(VI)
on
sulfate-reducing
bacteria.
Applied Microbiology Biotechnology
60,
352–360.
Chayabutra, C. and Ju, L. (2000). Degradation of
n
-hexadecane and its metabolites by
Pseudomonas aeruginosa
under micro
aerobic
and
anaerobic
denitrifying
conditions.
Applied and Environmental
Microbiology
66
, 493–498.
Coates, J. D., Anderson, R. T. and Lovley, D. R.
(1996a).
Anaerobic
hydrocarbon
degradation in petroleum-contaminated
harbor sediments under sulfate reducing
and artificially imposed iron-reducing
conditions.
Environmental Science and
Technology
30,
2784–2789.
Coates, J. D., Anderson, R. T. and Lovley, D. R.
(1996b).
Anaerobic
oxidation
of
polycyclic aromatic hydrocarbons under
sulfate-reducing conditions.
Applied and
Environmental Microbiology
62,
1099–
1101.
Coates, J.D. and Anderson, R.T. (2000).
Emerging
techniques
for
anaerobic
bioremediation
of
contaminated
environments.
Trends in Biotechnology
18,
408-412.
Coates, J.D., Woodward, J., Allen, J., Philp, P.
and Lovley, D.R. (1997). Anaerobic
degradation
of
polycyclic
aromatic
hydrocarbons and alkanes in petroleum-
contaminated marine harbour sediments.
Applied and Environmental Microbiology
63,
3589-3593.
DeLeo, P.C. and Ehrlich, H.L. (1994). Reduction
of hexavalent chromium by
Pseudomonas
fluorescens
LB300
in
batch
and
continuous cultures.
Applied Microbiology
and Biotechnology
40
, 756–759.
Diaz, E. (2004). Bacterial degradation of
aromatic pollutants: a paradigm of
metabolic
diversity.
International
Microbiology
7,
173-180.
Dolfing, J. and Janssen, D.B. (1994). Estimates
of Gibbs free energies of formation of
chlorinated
aliphatic
compounds.
Biodegradation
5
, 21-28.
Eary, L. E. and Rai, D. (1988). Chromate
removal from aqueous wastes by reduction
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
284
with ferrous iron.
Environmental Science
and Technology
22,
972–977.
Edwards, E.A. and Grbic-Galic, D. (1992).
Complete mineralization of benzene by
aquifer microorganisms under strictly
anaerobic
conditions.
Applied
and
Environmental Microbiology
58,
2663–
2666.
El Fantroussi, S., Naveau, H. and Agathos, S.N.
(1998). Anaerobic dechlorinating bacteria.
Biotechnology Progress
14,
167-188.
Fathepure, B.Z., Nengu, J.P. and Boyd, S.A.
(1987).
Anaerobic
bacteria
that
dechlorinate perchloroethene.
Applied and
Environmental Microbiology
53,
2671-
2674.
Fennell, D.E., Gossett, J.M. and Zinder, S.H.
(1997). Comparison of butyric acid,
ethanol, lactic acid, and propionic acid as
hydrogen
donors
for
the
reductive
dechlorination
of
trichlorothene.
Environmental Science and Technology
31,
918-925.
Frederickson, J.K., Kostandarithes, H.M., Li,
S.W., Plymale, A.E. and Daly, M.J.
(2000). Reduction of Fe(III), Cr(VI),
U(VI), and Tc(VII) by
Deinococcus
radiodurans
R1.
Applied
and
Environmental Microbiology
66,
2006–
2011.
Gadd, G.M. (2000). Bioremedial potential of
microbial
mechanisms
of
metal
mobilization and immobilization.
Current
Opinion in Biotechnology
11,
271–279.
Ganesh, R., Robinson, K.G., Chu, L., Kucsmas,
D. and Reed, G.D. (1999). Reductive
precipitation of uranium by
Desulfovibrio
desulfuricans
: evaluation of cocontaminant
effects and selective removal.
Water
Research
33,
3447-3458.
Gantzer, C.J. and Wackett, L.P. (1991).
Reductive dechlorination catalyzed by
bacterial
transition-metal
coenzymes.
Environmental Science and Technology
25,
715-722.
Gerritse, J., Drzyzga, O., Kloetstra, G., Keijmel,
M., Wiersum, L.P., Hutson, R., Collins,
M.D. and Gottschal, J.C. (1999). Influence
of different electron donors and acceptors
on dehalospiration of tetrachloroethene by
Desulfitobacterium
Frappiere
TCE1.
Applied and Environmental Microbiology
65,
5212-5221.
Gibson, J. and Harwood, C.S. (2002). Microbial
diversity in aromatic compound utilization
by anaerobic microbes.
Annual Review of
Microbiology 56,
345-369.
Gossett, J.M. and Zinder, S.H. (1997).
Microbiological aspects relevant to natural
attenuation
of
chlorinated
ethenes.
Proceedings of the US EPA Symposium in
Natural
Attenuation
of
Chlorinated
Organics in Ground Water
. Pp. 12-15.
Dallas, Texas. EPA/540/R-97/504/1997.
Harkness, M.R., Bracco, A.A., Brennan Jr., M.J.,
DeWeerd, K.A. and Spivack, J.L. (1999).
Use of bioaugmentation to stimulate
complete reductive dechlorination of
trichloroethene in Dover soil columns.
Environmental Science and. Technology
33,
1100-1109.
Head, I.M. (1998). Bioremediation: towards a
credible technology.
Microbiology
144,
599-608.
Heider, J. and Fuchs, G. (1997). Microbial
anaerobic aromatic metabolism.
Anaerobe
3,
1-22.
Heitkamp, M. A., Freeman, J. P. and Cerniglia,
C. E. (1987). Naphthalene biodegradation
in environmental microcosms: estimates of
degradation rates and characterization of
metabolites.
Applied and Environmental
Microbiology
53,
129–136.
Holliger, C., Wohlfarth, G. and Diekert, G.
(1999). Reductive dechlorination in the
energy metabolism of anaerobic bacteria.
FEMS Microbiology Reviews
22
, 383-398.
Kashefi, K. and Lovley, D.R. (2000). Reduction
of Fe(III), Mn(IV), and toxic metals at
100°C
by
Pyrobaculum
islandicum
.
Applied and Environmental Microbiology
66,
1050–1056.
Kazumi, J., Häggblom, M.M.
and Young, L.Y.
(1995). Degradation of mono-chlorinated
and non-chlorinated aromatic compounds
under iron-reducing conditions.
Applied
and Environmental Microbiology
61,
4069–4073.
Kim, C., Zhou, Q., Deng, B., Thornton, E.C. and
Xu, H. (2001). Chromium(VI) reduction
by hydrogen sulfide in aqueous media:
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
285
stoichiometry and kinetics.
Environmental
Science and Technology
35,
2219–2225.
Kropp, K.G., Davidova, I.A. and Suflita, J.M.
(2000). Anaerobic oxidation of n-dodecane
by an addition reaction in a sulfate-
reducing bacterial enrichment culture.
Applied and Environmental Microbiology
66,
5393–5398.
Leahy, J.G. and Olsen, R.H. (1997). Kinetics of
toluene degradation by toluene oxidizing
bacteria
as
a
function
of
oxygen
concentration, and the effect of nitrate.
FEMS Microbiology Ecology 23,
23-30.
Leduc, R., Samson, R., Al-Bashir, B., Al-
Hawari, J. and Cseh, T. (1992). Biotic and
abiotic disappearance of four PAH
compounds from flooded soil under
various redox conditions.
Water Science
and Technology
26,
51–60.
Lloyd, J.R. and Macaskie, L.E. (1996). A novel
phosphoimager
based
technique
for
monitoring
microbial
reduction
of
technetium.
Applied and Environmental
Microbiology
62
, 578-582.
Lloyd, J.R., Ridley, J., Khizniak, T., Lyalikova,
N.N. and Macaskie, E. (1999). Reduction
of
technetium
by
Desulfovibrio
desulfuricans
: biocatalyst characterization
and use in a flowthrough bioreactor.
Applied and Environmental Microbiology
65,
2691-2696.
Lovley, D.R. (1994). Microbial reduction of
iron,
manganese
and
other
metals.
Advances in Agronomy
54,
175–229.
Lovley, D.R. (1995). Bioremediation of organic
and metal contaminants with dissimilatory
metal reduction.
Journal of Industrial
Microbiology
14,
85–93.
Lovley, D.R. (2001). Anaerobes to the rescue.
Science
293, 1444–1446.
Lovley,
D.R.
and
Coates,
J.D.
(1997).
Bioremediation of metal contamination.
Curr. Opin. Biotechnol.
8,
285–289.
Lovley, D.R. and Phillips, E.J.P. (1994).
Reduction of chromate by
Desulfovibrio
vulgaris
and its c
3
cytochrome.
Current
Opinion in Biotechnology
60,
726–728.
Lu, G., Clement, T.P., Zhen, C. and Wiedemeier,
T.H. (1999). Natural attenuation of BTEX
compounds model development and field
application.
Groundwater
37,
707-717.
Mägli, A., Wendt, M. and Leisinger, T. (1996).
Isolation
and
characterization
of
Dehalobacterium formicoaceticum
gen.
nov. sp. nov., a strictly anaerobic
bacterium utilizing dichloromethane as
source of carbon and energy.
Archives of
Microbiology
166,
101-108.
Mahan, B.H. (1967). The chromium family. In:
University chemistry, 2nd edn. Pp 557–
559. Addison Wesley, London.
Marsh, T. and McInerney, M.J. (2001).
Relationship of hydrogen bioavailability
and
chromate
reduction
in
aquifer
sediments.
Applied and Environmental
Microbiology
67,
1517-1521.
McCarty, P.L. and Semprini, L. (1994). Ground-
water treatment for chlorinated solvents.
In:
Handbook of Bioremediation
. (R.D.
Norris, ed.). Pp. 87-116. Lewis Publishers.
Boca Raton, Florida.
Michel, C., Brugna, M., Aubert, C., Bernadac,
A. and Bruschi, M. (2001). Enzymatic
reduction
of
chromate:
comparative
studies using sulfate-reducing bacteria.
Applied Microbiology and Biotechnology
55, 9
5–100.
Mihelcic, J. R. and Luthy, R. G. (1988).
Microbial degradation of acenaphthene
and naphthalene under denitrification
conditions in soil-water systems.
Applied
and Environmental Microbiology
54,
1188–1198.
Oste, L.A., Dolfing, J., Ma, W.C. and Lexmond,
T.M. (2001). Effect of beringite on
cadmium and zinc uptake by plants and
earthworms: more than a liming effect?
Environmental Toxicology and Chemistry
20,
1339-1345.
Rabus, R. and Widdel, F. (1995). Anaerobic
degradation of ethylbenzene and other
aromatic
hydrocarbons
by
new
denitrifying
bacteria.
Archives
of
Microbiology
163,
96–103.
Rueter, P., Rabus, R., Wilkes, H., Aeckersberg,
F., Rainey, F.A., Jannasch, H.W. and
Widdel, F. (1994). Anaerobic oxidation of
hydrocarbons in crude oil by new types of
sulphate reducing bacteria.
Nature
372,
455-4458.
Schmieman, E.A., Petersen, J.N.,
Yonge, D.R.,
Johnstone, D.L., Bereded-Samuel, Y.,
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
286
Apel, W.A. and Turick, C.E. (1997).
Bacterial Reduction of Chromium.
Applied
Biochemistry and Biotechnology
63,
855–
864.
Seaman, J.C, Meehan, T. and Bertsch, P.M.
(2001). Immobilization of cesium-137 and
uranium in contaminated sediments using
soil
amendments.
Journal
of
Environmental Quality
30,
1206-1213.
Semprini, L., Kitanidis, P.K., Kampbell, D. and
Wilson,
J.T.
(1995).
Anaerobic
transformation of chlorinated aliphatic
hydrocarbons in a sand aquifer based on
spatial chemical distributions.
Water
Resources Research
31,
1051-1062.
Sen, A.M. and Johnson, B. (1999). Acidophilic
sulphate-reducing bacteria: candidates for
remediation of acid mine drainage. In:
Biohydrometalurgy and the environment
toward the mining of the 21st century. Part
A: Bioleaching, microbiology
. (R. Amils
and A. Ballester, eds.). Pp. 709-718.
Elsevier, Amsterdam.
Shen,
H.
and
Wang,
Y.T.
(1993).
Characterization of enzymatic reduction of
hexavalent chromium by
Escherichia coli
ATCC 33456.
Applied and Environmental
Microbiology 59,
3771–3777.
Smatlak, C.R., Gossett, J.M. and Zinder, S.H.
(1996). Comparative kinetics of hydrogen
utilization for reductive dechlorination of
tetrachloroethene and methanogenesis in
an
anaerobic
enrichment
culture.
Environmental Science and Technology
30,
2850-2858.
So, C.M. and Young, L.Y. (1999). Isolation and
characterization of a sulfate-reducing
bacterium that anaerobically degrades
alkanes.
Applied
and
Environmental
Microbiology
65,
2969–2976.
Stephen, J.R., Chang, Y.J., Macnaughton, S.J.,
Kowalchuk, G.A., Leung. K.T., Flemming,
C.A. and White, D.C. (1999). Effect of
toxic metals on indigenous soil beta-
subgroup
proteobacterium
ammonia
oxidizer
community
structure
and
protection against toxicity by inoculated
metal-resistant bacteria.
Applied and
Environmental Microbiology
65,
95-101.
Tandoi, V., DiStefano, T.D., Bowser, P.A.,
Gossett, J.M. and Zinder, S.H. (1994).
Reductive dehalogenation of chlorinated
ethenes and halogenated ethanes by a
high-rate anaerobic enrichment culture.
Environmental Science and Technology
28,
973-979.
Thauer, R.K., Jungermann, K. and Decker, K.
(1977).
Energy
conservation
in
chemotrophic
anaerobic
bacteria.
Bacteriological Review
41,
100-180.
Timmis, K.N. and Pieper, D.H. (1999). Bacteria
designed for bioremediation.
Trends in
Biotechnology
17,
201-204.
Tucker, M.D., Barton, L.L. and Thomson, B.M.
(1998). Reduction of Cr, Mo,
Se and
U
by
Desulfovibrio
desulfuricans
immobilized in polyacrylamide gels.
Journal of Industrial Microbiology and
Biotechnology
20,
13–19.
Turick, C.E., Apel, W.A. and Carmiol, N.S.
(1996). Isolation of hexavalent chromium
reducing
anaerobes
from
hexavalent
chromium
contaminated
and
non-
contaminated
environments.
Applied
Microbiology and Biotechnology 44,
683–
688.
U.S. Environmental Protection Agency. (1998).
Technical Protocol for Evaluating Natural
Attenuation of Chlorinated Solvents in
Ground Water
. Office of Research and
Development. Washington, D.C.
Valls, M., Atrian, S., de Lorenzo, V. and
Fernandez, L.A. (2000). Engineering a
mouse metallothionein on the cell surface
of
Ralstonia
eutropha
CH34
for
immobilization of heavy metals in soil.
Nature Biotechnology
18
, 661- 665.
Vidali, M. (2001). Bioremediation. An overview.
Pure and Applied Chemistry
73,
1163-
1172.
Vogel, T.M., Criddle, C.S. and McCarty, P.L.
(1987). Transformations of halogenated
aliphatic
compounds.
Environmental
Science and Technology
22,
722-736.
White, C. and Gadd, G.M. (1996). A comparison
of carbon/energy and complex nitrogen
sources for bacterial sulphate-reduction:
potential applications to bioprecipitation of
toxic metals as sulphides.
Journal of
Industrial Microbiology
17,
116–123.
White, C., Sayer, J.A. and Gadd, G.M. (1997).
Microbial
solubilization
and
N. Balagurusamy
/
Revista Mexicana de Ingeniería Química
Vol. 4
(2005)
273-287
287
immobilization of toxic metals: key
biogeochemical processes for treatment of
contamination.
FEMS
Microbiology
Reviews
20,
503–516.
Widdel, F. and Rabus, R., (2001). Anaerobic
biodegradation of saturated and aromatic
hydrocarbons.
Current
Opinion
in
Biotechnology
12,
259-276.
Wilkes, H., Rabus, R., Fischer, T., Armstroff, A.,
Behrends, A. and Widdel. F. (2002).
Anaerobic degradation of n-hexane in a
denitrifying
bacterium:
Further
degradation of the initial intermediate (1-
methylpentyl) succinate via C-skeleton
rearrangement.
Archives of
Microbiology
177,
235–243
Yager, R.M., Bilotta, S.E., Mann, C.L. and
Madsen, E.L. (1997). Metabolic and
in situ
attenuation of chlorinated ethenes by
naturally occurring microorganisms in a
fractured dolomite aquifer near Niagara
Falls, New York.
Environtal Science and
Technology
31,
3138-3147.
Zengler, K., Richnow, H.H., Rossello-Mora, R.,
Michaelis, W. and Widdel, F. (1999).
Methane
formation
from
long-chain
alkanes by anaerobic microorganisms.
Nature
401,
266–269.
Zhou, J., Fries, M.R., Chee-Sanford, J.C. and
Tiedje, J.M. (1995). Phylogenetic analyses
of a new group of denitrifiers capable of
anaerobic
growth
on
toluene
and
description of
Azoarcus tolulyticus
sp. nov.
International Journal of
Systematic
Bacteriology
45,
500-506.
logo_pie_uaemex.mx