ENVIRONMENTAL ENGINEERING SCIENCE
Volume 20, Number 5, 2003
© Mary Ann Liebert, Inc.
Microbial Degradation of Perchlorate:
Principles and Applications
Jianlin Xu, Yanguang Song, Booki Min, Lisa Steinberg, and Bruce E. Logan*
Department of Civil and Environmental Engineering
The Pennsylvania State University
University Park, PA 16802
ABSTRACT
Perchlorate (ClO42 ) release into the environment has occurred primarily in association with its manufacture and use in solid rocket propellant. When released into groundwater, perchlorate can spread over large
distances because it is highly soluble in water and adsorbs poorly to soil. Two proven techniques to remove perchlorate from drinking water are anaerobic biological reactors and ion exchange. In this review,
we focus on the application of microbiological systems for degrading perchlorate. Some bacteria can use
perchlorate as an electron acceptor while oxidizing a large range of substrates. Perchlorate-respiring bacteria (PRB) are widely distributed in the environment, and are enriched at perchlorate-contaminated sites.
We review the pathways by which PRB degrade perchlorate, and the different biological treatment processes that have been developed to remove perchlorate from water sources. We also discuss the effects
of alternate electron acceptors in the water, such as oxygen and nitrate, on perchlorate removal. Although
many different biological treatment systems using PRB have so far only been proven at the bench scale,
all pilot scale tests performed to date with a few of these systems have been successful. The success of
these perchlorate bioreactor tests indicates that biological treatment is a suitable method for soil remediation and water treatment of perchlorate-contaminated water.
Key words: perchlorate; application; percolate-respiring bacteria; bioremediation
INTRODUCTION
P
(ClO42 ) is a highly oxidized (17) chlorine oxy-anion manufactured for use as the oxidizer
in solid propellants for rockets, missiles, explosives, and
pyrotechnics (Urbansky, 1998; Gullick et al., 2001; Logan, 2001a). Approximately 90% of all perchlorate salts
are manufactured as ammonium perchlorate for use in
ERCHLORATE
rocket and missile propellants. The periodic replacement
and use of solid propellant has resulted in the discharge
of more than 15.9 million kg of perchlorate salts into the
environment since the 1950s (Motzer, 2001). Perchlorate
salts are highly soluble in water. Sodium perchlorate has
a solubility of about 2 kg/L, allowing large amounts to
be readily transported through surface and ground waters. The U.S. EPA has identified perchlorate users and
* Corresponding
author: Department of Civil and Environmental Engineering, 212 Sackett Bldg, Penn State University, University Park, PA 16802. Phone: 814-863-7908; Fax: 814-863-7304; E-mail: [email protected]
405
XU ET AL.
406
manufacturers in 44 states, and perchlorate releases in at
least 20 states (U.S. EPA, 2002). Such perchlorate releases are estimated to have affected the drinking water
of 15 million people.
Perchlorate can be detected by many methods including ion-selective electrodes, ion chromatography, capillary electrophoresis, HPLC, and spectrophotometry (Urbansky, 1998). In 1997, the California Department of
Health first reported an ion chromatographic method that
was capable of detecting perchlorate concentrations of 4
mg/L. Since then, ion chromatography has been the most
commonly used detection method for perchlorate. Perchlorate is only known to occur in the natural environment in Chilean caliche, a material that is used for some
fertilizers (Ericksen, 1983). The EPA developed a
method for measuring perchlorate concentrations in fertilizers (Collette et al., 2001), and has used this method
to survey large numbers of fertilizers and related materials. It was concluded that most fertilizers did not contain
perchlorate, and therefore, that fertilizers did not contribute to the extensive environmental perchlorate contamination that has been observed (Urbansky et al.,
2001).
Although there is currently no federal drinking water
standard for perchlorate, perchlorate has been included on
the federal Contaminant Candidate List (U.S. EPA, 1998).
High concentrations of perchlorate are known to affect the
function of the thyroid gland in humans by inhibiting the
uptake of iodide (Wolff, 1998), but direct epidemiological evidence that indicates that perchlorate is toxic to exposed humans is lacking (Crump et al., 2000; Li et al.,
2001; U.S. EPA, 2002). Recent studies have indicated that
low concentrations of perchlorate significantly inhibit iodide uptake in humans and animals (Lawrence et al., 2000;
OEHHA, 2002; U.S. EPA, 2002). Perchlorate contamination of the environment poses a threat to indigenous
wildlife as well as human health. Smith et al. (2001) have
found perchlorate contamination of wildlife and vegetation at concentrations that can pose a threat to the normal
growth and development of amphibian populations (Goleman et al., 2002a, 2002b). The Office of Environmental
Health Hazard Assessment in California EPA has proposed a Public Health Goal of 6 mg/L for perchlorate in
drinking water (OEHHA, 2002). In a recent perchlorate
risk assessment draft report, the U.S. EPA (2002) proposed a draft reference dose of 0.03 mg/kg of body weight
per day, which could produce a drinking water equivalent
level of 1 mg/L to protect human health. Based on this information, the California Department of Health Service in
California decreased the action level for perchlorate in
drinking water from 18 to 4 mg/L (DHS, 2002b). In New
Mexico, the action level was set at 1 mg/L (www.cluin.org/studio/perchlorate-060402.
In the last 5 years, several reviews have been published
on various perchlorate issues that include: bacterial
degradation (Herman and Frankenberger, 1998; Logan,
1998); chemistry and analytical chemistry (Urbansky,
1998, 2000a, 2000b; Urbansky and Schock, 1999) toxicological studies and drinking water standards (Wolff,
1998; Urbansky, 2000a; Soldin et al., 2001; OEHHA,
2002; U.S. EPA, 2002); and contamination sources and
occurrence data (Urbansky, 2000a; Cheremisnnoff, 2001;
Gullick et al., 2001; Logan, 2001a). However, there have
been important advances made in the treatment of perchlorate-contaminated water since the microbiology and
existing treatment technologies were reviewed in 1998
(Herman and Frankenberger, 1998; Logan, 1998). One
of the most important developments since these reviews
have been reports of microbial treatment processes capable of removing perchlorate down to levels expected
to be suitable for drinking water (,4 mg/L). These processes are the basis of several recent patents on biological treatment processes for perchlorate treatment (see Attaway et al., 1999; Coppola and McDonald, 2000;
Frankenberger, 2000; Gaudre-Longerinas and Tauzia,
2001; Logan, 2001b; Van Ginkel et al., 1999).
In the current review, we focus on recent developments
that have improved our understanding of the bacteria responsible for perchlorate degradation, the pathways used
for perchlorate degradation, and the treatment systems
developed to use perchlorate respiring bacteria (PRB).
We provide only a brief background on the chemistry,
occurrence, health issues, and drinking water issues for
perchlorate, and refer the reader to other more comprehensive studies and reviews on these subjects. We concentrate on topics related to microbial degradation of perchlorate because biological treatment is a particularly
promising option for remediation of perchlorate-contaminated drinking water. Although these biological perchlorate treatment technologies are quite new, there are
signs that such treatment systems will be accepted for use
in water treatment plants. The California Department of
Health Services recently approved the use of fluidized
bed reactors for the treatment of perchlorate contaminated
groundwater (DHS, 2002a), and biological denitrification
has been used at one site in Oklahoma for the treatment
of nitrate-contaminated groundwater (www.pall.com).
PECHLORATE RESPIRING BACTERIA
Ubiquity and diversity of perchlorate
reducing bacteria
It has been known for several years that the ability to
reduce perchlorate is not limited to a single bacterial
species, although some of this earlier evidence of per-
MICROBIAL DEGRADATION OF PERCHLORATE
chlorate degradation was inferred from information on
biological chlorate degradation (Logan, 1998). Bacteria
capable of chlorate reduction were found to inhabit a variety of environments including rivers, sediments, soils,
and wastewater treatment plants (van Ginkel et al., 1995).
Utilizing media with chlorate as the only electron acceptor, Coates et al. (1999) similarly found that the acetate-oxidizing chlorate reducing bacteria (CRB) represented a significant population whose abundance ranged
from 2.31 3 103 to 2.4 3 106 cells per gram of samples
obtained from a variety of sources, including pristine and
hydrocarbon-contaminated soils, aquatic sediments, paper mill waste sludges, and farm animal waste lagoons.
All 13 isolates obtained in this study were also capable
of growth on acetate using perchlorate, leading to early
speculation, via the abbreviation of (per)chlorate, that all
CRB were also capable of perchlorate reduction (Coates
et al., 1999).
More recent studies have provided evidence that not
all CRB are PRB, although the converse is true, that all
PRB are CRB. Wu et al. (2001) used perchlorate or chlorate in anaerobic growth medium to compare the abundance of PRB and CRB in different environmental samples. They found that when perchlorate was the sole
terminal electron acceptor in the medium, the number of
PRB in a pristine soil was up to 1,000-fold lower than
that found using chlorate in the medium to enumerate
CRB for the same samples (Wu et al., 2001). This provided indirect evidence that not all CRB were PRB, and
that PRB were less abundant than CRB in these envi-
Table 1.
407
ronments. Confirmation that not all CRB were PRB was
provided by three different studies. Only 8 of 10 CRB
isolated from wastewater were found to be capable of
respiring perchlorate (Logan et al., 2001c). In another
study, it was mentioned that a microorganism was isolated that was capable of using chlorate, but not perchlorate (Coates et al., 2000). Wolterink et al. (2002) isolated a CRB strain Pseudomonas chloritidismutansAW-1
from an anaerobic bioreactor treating wastewater containing chlorate, and found that this strain did not reduce
perchlorate. These studies indicate that while most CRB
are able to degrade perchlorate, there is a subset of CRB
that cannot use perchlorate as an electron acceptor for
cell respiration.
Perchlorate reducing isolates
Many PRB have now been isolated (Table 1). These
PRB are all Gram-negative, facultative anaerobes. Analysis of the 16S rRNA sequences of tested strains indicated that all isolates were members of the class Proteobacteria. The majority of these PRB, including GR-1,
perc1ace, KJ, PDX, HZ, JDS5, and 15 other strains
(Achenbach et al., 2001), are located in the b-subclass
of Proteobacteria. Achenbach et al. (2001) proposed two
new genera, Dechloromonas and Dechlorosoma, for
these b-subclass lineages, which represent the predominant PRB in the environment. The phylogenetic tree in
Fig. 1 shows the general positions of several PRB in comparison to other bacteria.
Perchlorate-respiring bacterial isolates.
Isolate
Vibrio dechloraticans Cuznesove B-1168
Wolinella succinogenes HAP-1
Dechloromonas agitata CKB
Dechloromonas sp. SIUL
Dechloromonas sp. MissR
Dechloromonas sp. CL
Dechloromonas sp. NM
Dechlorosoma sp. SDGM
Dechlorosoma sp. PS
Dechloromonas sp. JM
Dechloromonas sp. HZ
Dechlorospirilium anomalous WD
Dechlorosoma sp. KJ
Dechlorosoma sp. PDX
Dechlorosoma sp. GR-1
Dechlorosoma sp. perc lace
Citrobacter sp. IsoCock1
Dechloromonas sp. JDS5
References
Korenkov et al. (1976)
Wallace et al. (1996)
Bruce et al. (1999)
Coates et al. (1999); Coates et al. (2000)
Coates et al. (1999); Coates et al. (2000)
Coates et al. (1999); Coates et al. (2000)
Coates et al. (1999); Coates et al. (2000)
Coates et al. (1999); Coates et al. (2000)
Coates et al. (1999); Coates et al. (2000)
Miller and Logan (2000)
Zhang et al. (2002)
Michaelidou et al. (2000)
Logan et al. (2001c)
Logan et al. (2001c)
Rikken et al. (1996)
Herman and Frankenberger (1999)
Okeke et al. (2002)
Shrout and Parkin (2002)
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
XU ET AL.
408
Figure 1. Neighbor-joining phylogenetic tree based on the 16S
rRNA sequences of PRB (Dechloromonas and Dechlorosoma)
and others. Bootstrap values (100 replicates) are indicated on the
nodes. The prealigned 16S rRNA sequences in TREECON format (Van de Peer and De Wachter, 1994) were downloaded from
the rRNA WWW Server (HTTP://RRNA.UIA.AC.BE/)
Electron donors and acceptors used by PRB
for growth
All PRB are capable of dissimilatory reduction of chlorate to chloride for energy and growth, and many PRB
can also reduce nitrate. For both perchlorate and chlorate, reduction does not occur in the presence of a high
concentration of dissolved oxygen. Most isolates can use
oxygen and many can respire using nitrate as a terminal
electron acceptor. HAP-1 was initially reported to be an
obligate anaerobe (Wallace et al., 1996), but in a later
study it was reported that it was a microaerobic organism (Wallace et al., 1998), although no further details on
oxygen tolerance were provided. It seems reasonable that
PRB should be able to use, or at least tolerate, dissolved
oxygen as it is produced during the decomposition of
chlorate and perchlorate. Although many PRB are capable of complete denitrification, CKB does not grow on
nitrate (Bruce et al., 1999), and Cuznesove B-1168 (Korenkov et al., 1976) and HAP-1 (Wallace et al., 1996)
reduce nitrate only to nitrite and do not produce ammonia or nitrogen gas.
In some cases PRB have been tested for their ability to
use other electron acceptors such as metals and sulfate.
GR-1 can utilize Mn (IV) as an electron acceptor. Most
PRB and CRB cannot reduce sulfate (Riken et al., 1996;
Wallace et al., 1996; Bruce et al., 1999; Coates et al., 1999;
Herman and Frankenberger, 1999; Michaelidou et al.,
2000; Wolterink et al., 2002). The only report of CRB
(Acinetobacter spp.) capable of growth using sulfate was
provided by Stepanyuk et al. (1992), but the ability to
respire perchlorate was not tested and these bacteria could
not use nitrate. PRB have not been found to be capable of
using other electron acceptors, including: Fe (III), selenate,
malate, and fumarate (Coates et al., 1999).
Both heterotrophic and autotrophic PRB have been isolated. Acetate has been most frequently used as a single
substrate for heterotrophic perchlorate reduction (Korenkov et al., 1976; Wallace et al., 1996; Bruce et al.,
1999; Coates et al., 1999; Herman and Frankenberger,
1999; Logan et al., 2001c), but hydrogen or formate was
required as an electron donor for the growth of HAP-1
on acetate (Wallace et al., 1996). Perchlorate reduction
by Citrobacter sp. IsoCock1 (Okeke et al., 2002) was
sustained on acetate, but yeast extract was found to improve growth. A wide variety of organic substrates, including alcohols and carboxylic acids, can be used as
growth substrates by PRB although the use of these substrates is strain-dependent. Dechloromonas sp. JM
(Miller and Logan, 2000), a strain isolated from the bacterial consortium in an autotrophic packed-bed biofilm
reactor, reduced perchlorate using dissolved hydrogen,
but could not grow using hydrogen as the sole electron
donor. Two autotrophic isolates, Dechloromonas sp. HZ
(Zhang et al., 2002) and Dechloromonas sp. JDS5
(Shrout and Parkin, 2002), have recently been isolated
that can grow in a minimal inorganic medium using perchlorate, hydrogen, CO2 , and nutrients.
Nutritional requirements for PRB
There is no detailed information on the best medium
to use for PRB or what trace nutrients or metals are
needed for growth. Several research groups have used a
phosphate buffer system (Attaway and Smith, 1993;
Rikken et al., 1996; Wallace et al., 1996; Herman and
Frankenberger, 1999; Logan et al., 2001c; Wu et al.,
2001), while others (Bruce et al., 1999; Coates et al.,
1999) have used a bicarbonate-buffered freshwater
medium amended with a complex vitamin solution. There
is no evidence that these additional vitamins are necessary for PRB growth, but in one study it was shown that
several PRB strains could not grow without the trace
metal solution (Michaelidou et al., 2000). Iron, molybdenum, and selenium appear to be important for PRB
growth and perchlorate degradation. Perchlorate reductase purified from GR-1 was found to contain 11 mol of
iron, 1 mol of molybdenum, and 1 mol of selenium per
mol of heterodimer (Kengen et al., 1999). Bender et al.
MICROBIAL DEGRADATION OF PERCHLORATE
(2002) also reported that molybdenum was important for
PRB growth. Wallace et al. (1996) used yeast extract and
peptone in their medium for HAP-1. Zhang et al. (2002)
found that yeast extract improved the growth of the autotrophic PRB strain HZ, but that yeast extract was not
needed for growth.
It is likely that the different trace nutrients used for
laboratory media are not needed for bioremediation of
low perchlorate concentrations in natural systems. As explained below, several field studies have achieved perchlorate degradation only through the addition of an oxidizable substrate (acetate, ethanol, etc.), nitrogen, and
phosphorus (Green and Pitre, 2000; Hatzinger et al.,
2000; Logan et al., 2001a; Evan et al., 2002; Min et al.,
2003).
THE PERCHLORATE
DEGRADATION PATHWAY
Much of what is known about bacterial degradation of
perchlorate has resulted from earlier studies on chlorate
reduction. Quastel et al. (1925) found that chlorite was
produced from chlorate by one strain of Escherichia coli
without further reduction of chlorite (ClO22 ). The failure of chlorate respiration in this strain was likely due to
the toxic effects of chlorite and an absence of the enzyme
chlorite dismutase. Later, it was found that some bacteria could reduce chlorate to chloride (Aslander, 1928;
Bryan and Rohlich, 1954). The reduction of perchlorate
to chloride by several species of heterotrophic bacteria
was first demonstrated with the use of 36Cl-labeled perchlorate (Hackenthal et al., 1964). Hackenthal (1965)
concluded that chlorate was the product of perchlorate
reduction by cell free extracts obtained from nitrate-
Figure 2.
409
adapted cells of Bacillus cererus. They also found that
chlorate could be reduced by the same cell free preparation, and that it competitively inhibited perchlorate
reduction. The first proposed perchlorate reduction pathway was ClO42 ClO3 2 ClO2 2 Cl2 O2 Cl2 (Hackenthal et al., 1964; Hackenthal, 1965). The reduction of
perchlorate or chlorate to chloride by bacteria was subsequently confirmed by other researchers (Korenkov et
al., 1976; Malmqvist et al., 1991; Rikken et al., 1996;
Coates et al., 1999; Wu et al., 2001).
Research on the perchlorate reduction pathway did not
make further progress until a new enzyme, chlorite dismutase, was purified from the PRB strain GR-1 and found
to produce oxygen from chlorite (Rikken et al., 1996; van
Ginkel et al., 1996). Rikken et al. (1996) proposed a
three-step mechanism of perchlorate reduction (Fig. 2) in
which chlorate, chlorite, and dissolved oxygen were sequentially produced. This pathway, which is now widely
accepted for bacterial respiration using perchlorate and
chlorate, is: ClO4 2 ClO3 2 ClO2 2 O2 1 Cl2 . Chlorite dismutase has been isolated from two other bacteria,
Ideonella dechloratans, and strain CKB. Stenklo et al.
(2001) purified chlorite dismutase from I. dechloratans
that had catalytic properties that were similar, but not
identical, to those found for the chlorite dismutase enzyme obtained from GR-1. Stenklo et al. (2001) determined the 22-residue N-terminal amino acid sequence for
this enzyme and found no homologue in the protein sequence of the purified enzyme from I. dechloratans.
Coates et al. (1999) also purified a chlorite dismutase
from the strain CKB, which had characteristics similar to
the enzyme purified from strain GR-1. The finding by
Coates et al. (1999) that all 13 CRB isolates obtained in
their laboratory could disproportionate chlorite into chloride and oxygen makes it likely that chlorite dismutase
Perchlorate reduction pathway (adapted from Rikken et al., 1996)
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
XU ET AL.
410
is the central enzyme for the dissimilatory reduction of
perchlorate to chloride in PRB.
The impact of nitrate on perchlorate reduction is important because nitrate is a common contaminant of perchlorate-contaminated waters. Many PRB are capable of
partial or complete denitrification, and the presence of
nitrate usually decreases the rate of perchlorate reduction.
The similarities between the perchlorate reduction and
denitrification pathways led to an early suggestion that
the perchlorate and nitrate pathways were catalyzed by
the same nitrate reductase (Hackenthal et al., 1964;
Stouthamer, 1967). In addition, it was found that perchlorate reductase obtained from GR-1 reduced nitrate as
well as chlorate and perchlorate (Kengen et al., 1999).
The concept of a shared enzyme for nitrate and perchlorate reduction is unlikely given more recent findings.
It has been found that some PRB, for example
Dechloromonas agitata CKB (Bruce et al., 1999), can
respire using perchlorate but not nitrate. In a PRB capable of denitrification (strain perclace), it was found that
nitrate and perchlorate reductases were located in different cell fractions (membrane and periplasmic fractions,
respectively), indicating that these two enzymes were
separate (Giblin and Frankenberger, 2001). Further evidence of completely separate denitrifying and perchlorate pathways was provided by Xu et al. (2002). These
researchers found that perchlorate degradation and denitrification in Dechlorosoma sp. KJ and PDX were separately induced. When cells were grown on only perchlorate, there was minimal nitrate reduction. Also, there was
no perchlorate or chlorate reduction when cells were initially grown on only nitrate. However, the presence of
both perchlorate and nitrate in the medium stimulated the
degradation of both electron acceptors. Their kinetic results were supported by whole cell protein profiles using
SDS-PAGE that showed the appearance of bands in the
gels at locations expected for perchlorate reductase, chlorite dismutase, and nitrate reductase under conditions that
agreed with the kinetic studies (Xu et al., 2002).
It is still not clear if only a single enzyme is used by
PRB for chlorate and perchlorate reduction, or if there
are separate enzymes used for perchlorate and chlorate
reduction. Research by Kengen et al. (1999) suggests
only one enzyme is needed for chlorate and perchlorate
respiration. They found that a single enzyme could catalyze both chlorate and perchlorate reduction, and that
this oxygen-sensitive enzyme was located in the
periplasm. However, the maximum reaction rates (measured with methyl viologen) for perchlorate (3.8 U/mg)
were actually less than those with either nitrate (6.2
U/mg) or chlorate (11.3 U/mg) (Kengen et al., 1999). Because it appears likely that separate enzymes are used for
nitrate and perchlorate (as explained above), it may also
be that there are separate chlorate and perchlorate enzymes. Additional evidence for separate enzymes is provided indirectly by the fact that not all CRB are capable
of respiration with perchlorate, although this question
will require further research to resolve.
MICROBIAL TREATMENT PROCESSES—
BENCH SCALE PROCESSES
The first bench-scale treatment system reported to degrade perchlorate was a suspended growth system developed to treat high concentrations of perchlorate (Attaway and Smith, 1993); this system has been adequately
described in previous reviews (Herman and Frankenberger, 1998; Logan, 1998). More relevant for drinking
water treatment and bioremediation studies are reports on
fixed and fluidized bed bioreactors primarily designed to
treat low concentrations of perchlorate, and laboratory
batch tests to evaluate the feasibility of in situ bioremediation. These different bench scale studies are summarized in Table 2.
Heterotrophic fixed-bed reactor studies
The first reported fixed-bed bioreactor was an up-flow
anaerobic reactor inoculated with a mixed culture containing Wollinella succinogenes HAP-1 (Wallace et al.,
1998). This system was designed to treat high perchlorate concentrations (about 10%) found in rocket washout
wastewaters. The reactor was made of acrylic tubing
(1.17 m in length; 7.6 cm inside diameter) filled with diatomaceous earth pellets (mean pore diameter of 20 mm).
The substrate for the mixed culture (BYF-100) consisted
of 54% naturally occurring protein, peptides, free amino
nitrogen, vitamins, and trace elements. Perchlorate was
reduced to ,300 mg/L over the entire 43-day period of
operation at hydraulic retention times (HRT) and influent perchlorate concentrations of 1.17 h and 1500 mg/L,
and 0.46 h and 500 mg/L, respectively. For both influent
concentrations, the effluent concentration was below 100
mg/L (the detection limit in this system) for 95% of the
samples taken during the reactor operation.
Perchlorate was found to be degraded to lower concentrations in a sand media bioreactor inoculated with
perc1ace, which was isolated from wastewater biosolids
(Herman and Frankenberger, 1999). The reactor was a
glass column (2.8 cm diameter; 14 cm long) packed with
sterilized, oven-dried sand (40 to 70 mesh size). Following a start up period, it was found that perchlorate could
be degraded from a concentration of 130 mg/L to below
the detection limit (4 mg/L) at an HRT of 3 h. It was also
reported that a celite-packed bioreactor inoculated with
the same isolate completely removed perchlorate from
MICROBIAL DEGRADATION OF PERCHLORATE
Table 2.
411
Bench-scale bioreactors for treating perchlorate in water.
Reactor type (media)
Suspended growth
Inoculum
Wolinella
HAP-1 in
Wolinella
HAP-1 in
succinogenes
mixed culture
succinogenes
mixed culture
Perchlorate
(mg/L)
Substrates
Protein nutrientsa
7750
BYF-100b
500, 1500
Perclace
Acetate
0.13, 0.738
Fixed bed (sand
GAC)
Fixed bed (GAC)
Mixed culture
Acetate
20
Tapwater
Fixed bed (sand)
Fixed bed
(cylindrical
pall rings)
Fixed bed
(Celite pellets)
Fixed bed
Dechlorosomas sp. KJ
Primary sludge and
effluent from wastewater treatment plant
Perclace
A mixture of acetate,
lactate, and pyruvate
Acetate
Acetate
Mixed culture
Hydrogen
0.740
Fixed bed (Celite
R-635)
Biosolids from the
municipal water
treatment plant
Mixed culture
Hydrogen
0.740
Hydrogen
0.073
Biological solids from
an anaerobic digester
Ethanol, methanol,
or a mixture of the
two alcohols
Acetic acid and ethanol
Hydrogen
Fixed bed
(diatomacous
earth pellets)
Fixed bed (sand
or Celite)
Fixed bed (glass
beads)
Fluidized bed (sand
or GAC)
Fludized bed (GAC)
Hollow-fiber
membrane
bioreactor
Mixed culture
Ralstonia eutropha
Acetate
0.051–0.55
20
0.1, 1
0.8
25
11–23
0.006–0.100
References
Attaway and Smith
(1993)
Wallace et al. (1998)
Herman and
Frankenberger (1999);
Giblin et al. (2000a)
Kim and Logan (2000)
Brown et al. (2000)
Kim and Logan (2001)
Burns et al. (2001)
Losi et al. (2002)
Miller and Logan
(2000)
Giblin et al. (2000b)
Logan and LaPoint,
(2002)
Greene and Pitre,
(2000); Hatzinger
et al. (2000)
Togna et al. (2001)
Nerenberg et al.
(2002)
a Aged
brewer’s yeast, cottonseed protein, or whey power; bNaturally occurring protein (54%), peptides, free amino nitrogen,
vitamins, and trace elements.
738 mg/L at a flow rate of 1 ml/min (Giblin et al., 2000a).
When the flow was increased to 2 mL/min, 92 to 95%
of the perchlorate was removed.
Perchlorate concentrations of 20 mg/L were found to
be completely removed in a sand-packed bioreactor (2.5
cm diameter; 28 cm long) inoculated with a perchloratedegrading enrichment (Kim and Logan, 2000). The enrichment was developed with acetate and perchlorate using primary digestor solids from a wastewater treatment
plant. Perchlorate was removed to nondetectable levels
(,4 mg/L) over 35 days, after a 10-day startup period, at
HRTs ranging 18 to 51 min. The same sand column system was also used to examine perchlorate degradation
rates achieved when the reactor was inoculated with a
pure culture of Dechlorosoma sp. KJ vs. that obtained
using a perchlorate-degrading enrichment (Kim and Logan, 2001). The reactor inoculated with the pure culture
removed perchlorate from 20 mg/L to less than 4 mg/L
at a much shorter HRT of ,1 min than the mixed culture reactor (12 min). Acetate was used as the substrate
for the culture, and the molar ratio (6.6, n 5 156) of acetate-to-perchlorate for the pure culture was more than
twice the ratio (2.9, n 5 6) of the mixed culture.
Perchlorate degradation using granular-activated carbon (GAC) in a bioreactor was compared to reactor performance with sand media in fixed-bed reactors (Kim and
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
XU ET AL.
412
Logan, 2000). The GAC media (12 3 40 mesh-sieved)
had a bulk density of 0.44 g/cm3 (NORIT Americas Inc.,
Atlanta, GA), while the sand had an average size of 0.425
mm (density of 2.67 g/cm3 ). Both reactors were fed the
same perchlorate-degrading enrichment. The GACpacked reactor completely removed perchlorate for the
first 4 days of operation, but after backwashing the media on day 6, the reactor performed erratically with complete perchlorate removal on some days and detectable
perchlorate concentrations on others. The poor operation
of the GAC reactor was attributed to desorption of perchlorate from the GAC in the lower parts of the reactor
where biological activity was insufficient to degrade the
perchlorate. Based on these results, the researchers recommended that GAC would not be suitable for fixed-bed
bioreactors that would need backwashing to control
biofilm formation and prevent clogging.
A slightly different GAC system was examined by
Brown et al. (2000) for perchlorate degradation. They examined the use of biologically active carbon (BAC) filters for perchlorate removal at low influent perchlorate
concentrations (Brown et al., 2000). The BAC filters
were constructed of glass pipes 2.54 cm in diameter and
Teflon™ endcaps. The filters were rendered biologically
active by feeding 600 bed volumes (BV) of dechlorinated
tapwater containing 50 mg/L perchlorate and 20 mg/L
bromate. Deionized water containing 50–55 mg/L perchlorate was then fed to the column for 5,800 empty BV,
with perchlorate removal occurring mainly by sorption
to the carbon. Perchlorate removal was then stimulated
by adding a mixture of acetate, lactate, and pyruvate (2
mg/L as carbon) to the water that also contained 2.5 mg/L
of dissolved oxygen and nitrate. Perchlorate removal was
found to be a function of the concentration of nitrate in
the water. Perchlorate was removed from 52 mg/L to below the detection limit, for an influent nitrate concentration of 1.4 mg/L (0.07 mg/L in the effluent). However,
perchlorate was not appreciably removed (from 51 to 50
mg/L) at a higher influent nitrate concentration of 4.5
mg/L (2.0 mg/L in the effluent). The effects of nitrate on
perchlorate reduction are further discussed below.
Burns et al. (2001) reported that two fixed-bed bioreactors placed in series could remove perchlorate from an
influent concentration in the range of 0.1 to 1 mg/L, to
below the detection limit even in the presence of high
concentrations of nitrate (50 mg/L). Both reactors (70
mm diameter; 460 mm in a height) initially inoculated
with primary sludge were fed a synthetic solution containing acetate (250 mg/L) and sodium sulfite (to remove
dissolved oxygen in the reactor influent). HRTs of 1 h
(0.1 mg/L) and 10 h (1 mg/L) were needed for complete
removal of perchlorate.
Another bench-scale packed-bed biological reactor
was examined in a 30-day study for its efficiency at removing low concentrations of perchlorate (,1 mg/L)
from groundwater (Losi et al., 2002). The Plexiglas column reactor (total volume of 3,062 mL) was filled with
Celite pellets and inoculated with the perchlorate-reducing bacterium perc1ace (Herman and Frankenberger,
1999). Groundwater from a perchlorate-contaminated site
in California was pumped (upflow mode) into the reactor along with nutrients (N and P) and acetate (approximately 500 mg/L). Influent perchlorate concentrations of
about 800 mg/L were degraded to nondetectable levels
(,4 mg/L) at residence times as low as 0.3 h. Acetate
concentrations were less than 50 mg/L in the effluent. Nitrate (20 mg/L NO3 2 -N) was also completely removed
while sulfate reduction was not observed.
Heterotrophic fluidized bed bench scale reactors
Fluidized bed reactors (FBRs) have also been examined for perchlorate removal in bench scale studies
(Greene and Pitre, 2000; Hatzinger et al., 2000). The
FBRs contained either sand or GAC with particle sizes
of 0.2 to 0.6 mm and 0.9 to 1.4 mm, respectively, as the
media. The reactors were fed ethanol, methanol, or a mixture of the two alcohols as substrates for perchlorate
biodegradation. Ethanol or the ethanol–methanol mixture
produced better performance than just methanol. The
FBR packed with GAC showed better performance than
the sand reactor, and achieved consistent perchlorate removal from 25 mg/L to less than the detection limit of 4
mg/L (Greene and Pitre, 2000; Hatzinger et al., 2000).
A laboratory-scale FBR was used to treat perchloratecontaminated groundwater from the Longhorn Army Ammunition Plant (Togna et al., 2001). The glass column reactor containing GAC as the fluidization medium was fed
both acetic acid and ethanol as the electron donors. Effluent perchlorate concentrations were less than the detection limit except during a low substrate study conducted to determine the point of treatment failure.
Autotrophic H2 reactors
Nitrate has been degraded in hydrogen-fed bioreactors
for many years (Gayle et al., 1989; Kapoor and Viraraghavan 1997), but only very recently there have been
reports of using autotrophic bioreactors for the degradation of perchlorate. Miller and Logan (2000) developed
an unsaturated-flow, packed-bed bioreactor for the
degradation of perchlorate in drinking water. The bioreactor was inoculated with a hydrogen-oxidizing, autotrophic, and perchlorate degrading enrichment culture
developed in a chemostat. The packed-bed reactor was
fed with a gas mixture of hydrogen (5%) and carbon dioxide and a feed stream containing only perchlorate and in-
MICROBIAL DEGRADATION OF PERCHLORATE
organic nutrients. The reactor was initially operated for
a 10-day start up period using a high concentration of
perchlorate (50 mg/L) to build up the biofilm, and was
then switched to a lower level of perchlorate (740 6 110
mg/L) for the next 145 days of operation. During this operation period, perchlorate was reduced to 460 6 80 mg/L
at a constant hydraulic loading rate of 0.45 cm/min
(1.1–1.3 min detention time).
A slightly modified version of the reactor used by
Miller and Logan (2000) was tested to examine the removal of perchlorate from a perchlorate contaminated
groundwater (Logan and LaPoint, 2002). The reactor was
also fed a gas mixture of hydrogen (5%) and carbon dioxide as the electron donor and carbon source, respectively.
An average of 25 6 5% of perchlorate was removed at a
HRT of 1.5 min from a groundwater containing 73 6 2
mg/L of perchlorate and 21 6 5 mg/L of nitrate. It was
estimated that a reactor with a detention time of 8 to 42
min (95% C.I.) a length of 0.65 to 3.4 m would be necessary to completely remove perchlorate in this type of
unsaturated-flow system.
A 120-mL bioreactor packed with Celite R-635 (Celite
Corporation, Lompoc, CA) was used to degrade perchlorate with hydrogen (Giblin et al., 2000b). The autotrophic consortium of bacteria was fed a pressurized
water stream containing dissolved hydrogen and bicarbonate. Perchlorate (740 mg/L) was removed to below
the detection limit at a 2-h detention time. Perchlorate
removal was incomplete at shorter detention times, with
perchlorate concentrations of 100 and 200 mg/L and detention times of 1 and 0.67 h, respectively. It was speculated that the perchlorate breakthroughs at the shorter
detention times resulted from nonuniform distribution of
biomass, pH variations of the groundwater, and limited
hydrogen transport to the bacteria.
The degradation of perchlorate in a groundwater has
also been demonstrated using a hollow-fiber membrane
bioreactor (Nerenberg et al., 2002). The membrane module in the reactor contained 98 hollow fibers (93 cm long;
280 mm outside diameter) in a PVC pipe shell. Hydrogen gas fed through the fiber pores was used to support
a biofilm growing on the membrane and in the liquid in
the reactor. The groundwater contained both perchlorate
(6–100 mg/L) and nitrate (2.5–3 mg/L). It was found that
perchlorate could be removed to below the detection limit
when the nitrate concentration from the effluent was less
than 30 mg/L in the presence of at least 300 mg/L of dissolved hydrogen in the bulk solution.
Effects of nitrate on perchlorate removal
Both perchlorate and nitrate are common contaminants
in surface and ground waters in the United States. As
most of the PRB are also denitrifiers, some papers have
413
discussed the simultaneous removal of perchlorate and
nitrate from contaminated waters. Nitrate has been found
to have different effects on perchlorate removal rates. Nitrate has inhibited the perchlorate reduction rate in some
studies (Herman and Frankenberger, 1999; Brown et al.,
2002) but not in others (Burns et al., 2001; Logan and
LaPoint, 2002). A PRB isolate, perc1ace, was able to
completely remove perchlorate (130 mg/L) and to simultaneously remove more than 95% of the nitrate (20 mg/L)
in a sand-packed bioreactor (Herman and Frankenberger,
1999). Coppola and McDonald (2000) reported that a
mixed culture of W. succinogenes could completely degrade perchlorate (1.2–1.5 g/L), chlorate (3–3.5 g/L), and
nitrate (0.2 g/L) in a 7-L continuously stirred tank reactor at a HRT of 16 h. Nerenberg et al. (2002) showed
that perchlorate and nitrate reduction to nondetectable
levels could be concurrently achieved using a hydrogenfed, autotrophic membrane bioreactor. In pilot-scale tests,
Min et al. (2003) found that perchlorate and nitrate were
both completely removed in a fixed-bed bioreactor. Thus,
it appears likely that perchlorate and nitrate can be simultaneously removed in bioreactors. However, the concentration of nitrate will increase the total amount of electron donor needed by the system for complete perchlorate
removal.
Soil bioremediation feasibility tests
Important questions for soil bioremediation are
whether the added substrate will be used for perchlorate
degradation or lost to other biochemical routes (such as
methanogenesis), and whether there are sufficient microorganisms in the soil to achieve perchlorate degradation
using added substrate. The first indication that perchlorate bioremediation could be easily accomplished were
studies that demonstrated perchlorate reduction could be
achieved solely by adding substrate and nutrients to
wastewater and soil samples. For example, it was shown
that by adding acetate to headspace-free BOD bottles
50% of 20 mg/L perchlorate was degraded within 4 days,
and 74% was removed within 9 days (Kim and Logan,
2000). The cultures obtained after 9 days of acclimation
were transferred (5% by volume) to freshly prepared media containing acetate (200 mg/L) and either perchlorate
(50 mg/L) or chlorate (50 mg/L), or same concentrations
of both anions (100 mg/L). The perchlorate and chlorate
degradation rates were similar (75 and 78%, respectively)
within 5 days using the acclimated cultures (Kim and Logan, 2000).
Other researchers have shown that CRB are ubiquitous
in soil samples, providing indirect evidence that bioremediation to remove perchlorate would not require
bioaugmentation. For example, van Ginkel et al. (1995)
found CRB in rivers, sediments, soils, and wastewater
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
XU ET AL.
414
treatment plants (van Ginkel et al., 1995) and Coates et
al. (1999) found CRB at concentrations of 2.31 3 103 to
2.4 3 106 cells per gram in pristine and hydrocarboncontaminated soils, aquatic sediments, paper mill waste
sludges, and farm animal waste lagoons.
Direct evidence for the feasibility of perchlorate bioremediation with different substrates was provided by Wu
et al. (2001). They examined perchlorate degradation using samples from soils, natural waters, and wastewaters.
Complete degradation of 210 mg/L perchlorate was
achieved within 4 to 7 days and 8 to 29 days for raw
wastewater and creek water, respectively, using four different substrates (acetate, lactate, citric acid, or molasses).
PRB were nonmeasurable using a five-tube MPN method
in some soil samples, but were present at high concentrations at perchlorate-contaminated sites. For example,
perchlorate reduction was not achieved with 10 g of “pristine” soil, but it was observed with 100 g of soil. Microorganisms in 10 g of a perchlorate-contaminated soil
(from a site in Texas) completely removed perchlorate
from 100 mg/L within only 7 days using polylactate or
lactate.
Hunter (2001) demonstrated that injecting sand
columns with a slowly degradable substrate (vegetable
oil) could remove perchlorate without the need for
bioaugmentation with perchlorate-acclimated bacteria.
Columns were seeded with bacteria from soil, and perchlorate (20 mg/L) was added to water and pumped into
the columns at a rate of ,25 mL/day. After 14 days of
operation, 0.47 mg of soybean oil was injected onto the
treatment columns. Although perchlorate levels remained
high in the control columns (no soybean oil), perchlorate
in the effluent of the treatment columns decreased by
,99% over a 17-week operation in the columns containing the vegetable oil.
MICROBIAL TREATMENT PROCESSES—
PILOT AND FULL-SCALE PROCESSES
Fixed-bed reactors
A pilot-scale bioreactor was operated as an ex situ
treatment process to treat perchlorate-contaminated
groundwater at the Naval Weapons Industrial Reserve
Plant in McGregor, TX (Perlmutter et al., 2000). The reactor was constructed of steel (5-ft diameter, 18 ft in
height) and fed water containing 7 to 20 mg/L of perchlorate. It was found that adding only acetate (acetate:
ClO4 2 ratio of 5:1) and nutrients (nitrogen and phosphorus) at a flow rate of 20 gpm, the perchlorate concentration could be completely removed (,20 mg/L, the
detection limit at that site).
Two pilot-scale bioreactors were operated for about 7
months to treat perchlorate-contaminated groundwater
(average 76 mg/L) at the Texas Street Well Facility in
Redlands, CA (Logan et al., 2001a; Evans et al., 2002;
Min et al., 2003). The 7-ft tall reactors contained two rectangular beds each 2 ft 3 1 ft. One reactor contained sand
media (1 mm diameter) and the other one plastic media
(3.76 cm diameter). The reactors were inoculated with a
pure culture of Dechlorosoma sp. KJ. Acetic acid (47 6
9 mg/L for the sand media reactor; 53 6 10 mg/L for the
plastic media reactor) and an ammonium phosphate solution (C:N ratio of 5:1) were added to the groundwater.
The groundwater contained nitrate (4 mg/L), oxygen (7
mg/L), and sulfate (33 mg/L) as potential terminal electron acceptors in addition to perchlorate. Perchlorate was
completely and consistently removed to nondetectable
levels (4 mg/L) at flow rates of 1 gpm in the plastic media reactor, and 2 gpm in the sand reactor. Backwashing
(weekly) was critical to maintain consistent perchlorate
removal rates in the sand reactor. In the absence of regular backwashing, there was short circuiting in the sand
reactor, resulting in insufficient groundwater detention
time for consistent perchlorate removal.
Fluidized bed bioreactor
A pilot-scale FBR was tested in 1996 at the Aeroject
facility in Sacramento, CA, using water containing 7,000
to 8,000 mg/L of perchlorate and 1.5 mg/L of nitrate-nitrogen (Catts and McCullough, 1998). There are few details available on this system, but it was reported that effluent concentrations of ,400 mg/L were achieved for
perchlorate, and ,50 mg/L for nitrate-nitrogen. Ethanol
was added in proportion to perchlorate at a molar ratio
of 4:1.
In 1998, a second FB bioreactor was tested at the Baldwin Park site in the San Gabriel Basin in California. The
reactor (20 in high, 15 ft long) contained activated carbon (10 3 30 mesh), and was similar to one used at the
Aerojet site, except this reactor was inoculated with
sludge from a food processing industry (vs. a wastewater
treatment plant). The reactor was fed groundwater (30
gpm, 114 L/min) amended with ethanol (40 to 70 mg/L)
and nutrients. Effluent ethanol concentrations ranged
from ,10 mg/L to nondetectable concentrations. Perchlorate (50 to 100 mg/L) removal was generally greater
than 90% when dissolved oxygen was low (near 1 mg/L).
At other times, removal varied from 23 to 45%. Nitrate
(5 to 6 mg/L NO 32 -N) removal was generally greater
than 99%. The denatured ethanol used for the study contained low concentrations of methanol, methyl isobutyl
ketone and isopropyl alcohol (Catts and McCullough,
1998). These chemicals can be toxic to humans, and
therefore the addition of these chemicals to drinking water sources should be avoided. Nitrosodimethylamine
MICROBIAL DEGRADATION OF PERCHLORATE
present in the water at 70 to 80 mg/L was not affected by
the bioreactor treatment.
A full-scale FBR was operated to treat perchlorate in
groundwater at a site in Rancho Cordova, CA (www.envirogen.com/perchlorate.htm; Green and Pitre, 2000;
Hatzinger et al., 2000). The system consisted of four
FBRs each 14-ft in diameter treating 4,000 gpm of
groundwater. The reactors were filled with GAC, inoculated with a perchlorate-degrading enrichment, and fed
ethanol as the electron donor. Following a startup period
of 3 weeks, the effluent concentration of perchlorate was
removed from 6 to 8 mg/L to less than the detection limit
(4 mg/L) for 8 weeks of operation during periods of excess ethanol addition. Periods of incomplete perchlorate
removal occurred at lower ethanol doses (Green and Pitre,
2000; Hatzinger et al., 2000).
FBRs have recently been approved for treatment of
perchlorate contaminated drinking water by the California Department of Health Service (DHS, 2002a). However, the presence of methanol and other impurities in
ethanol may prohibit the use of ethanol as a substrate in
these bioreactors if they are to be used for drinking water treatment.
In situ bioremediation
Pilot-scale tests for perchlorate remediation have
shown that perchlorate reduction can be achieved by
adding various electron donors to create anoxic conditions in the subsurface. An excess of an electron donor
can be used for the degradation of perchlorate, nitrate,
and other pollutants. In situ bioremediation field tests of
perchlorate were conducted at the Naval Weapons Industrial Reserve Plant in McGregor, TX (Perlmutter et
al., 2000). Trenches were dug at the site to collect groundwater and direct the flow of this water through mounds
of material containing various organic sources such as
compost (from an edible mushroom production facility)
and cottonseed meal. Perchlorate in the water was removed from 16–27 mg/L to less than 100 mg/L (detection limit in this study) after only 2 weeks of operation.
Nitrate was also reduced from 15 mg/L to nondetectable
concentrations.
Cox et al. (2001) reported that perchlorate in situ
biodegradation was successfully achieved in a deep
aquifer at the Aerojet Superfund site in California. Acetate was added into the aquifer to support perchlorate
reduction. Perchlorate concentrations in the groundwater
decreased from 12 mg/L to less than the detection limit
(4 mg/L) within 15 feet of the electron donor delivery
well. In addition to perchlorate, trichloroethylene, another common contaminant of groundwater, was completely dechlorinated to ethane.
415
MICROBIAL TREATMENT PROCESSES—
SYSTEMS USED IN CONJUNCTION WITH
OTHER PROCESSES
Processes such as anion exchange (Venkatesh et al.,
2000; Gu et al., 2001), membrane filtration (Urbansky
and Schock, 1999), and activated carbon (Brown et al.,
2002) have been studied for the removal of perchlorate,
especially for treating waters contaminated with low
(,100 mg/L) concentrations of perchlorate. For example,
anion exchange resins have been shown to be highly effective for perchlorate removal (Urbansky, 2000a). A
field experiment has shown that one bed volume of a
unique bifunctional resin can treat greater than 100,000
BV of groundwater with an initial perchlorate concentration of 50 mg/L (Gu et al., 2000). Because these physical/chemical techniques do not degrade perchlorate, further treatment using chemical or biological processes is
necessary. Perchlorate in the brine wastes can be reduced
to chloride by chemical methods using catalytic air-sensitive metal cations such as titanium or ruthenium (AbuOmar et al., 2000; Espenson, 2000), or by electrochemical reduction (Urbansky, 1998). One company (Calgon
Carbon) has patented a treatment process (Venkatesh et
al., 2000) for perchlorate removal that uses ion exchange
to remove perchlorate and a bioreactor or catalytic reactor for complete perchlorate destruction. This treatment
process has been approved for treatment of perchlorate
contaminated water by the California Department of
Health Service (http://www.calgoncarbon.com/industry/productdata.php?id511).
Biological treatment of concentrated waste streams
from ion exchange processes can be difficult to treat using biological systems due the high salt content of the
brine. In the case of an anion exchange process, the regeneration of the resin typically generates a 7–12% NaCl
brine solution enriched in perchlorate. Gingras and
Batista (2002) were unable to adapt a PRB culture to degrade perchlorate in an ion exchange brine. As little as
1% NaCl reduced perchlorate reduction rates by their perchlorate-degrading culture by half (Gingras and Batista,
2002). High NaCl systems may not be required for certain types of specialized ion exchange resins. In one system recently developed by Gu et al. (2001), the resin was
regenerated using a tetrachloroferrate displacement technique. Using this method, only 5 BV of regenerant solution were needed to achieve nearly 100% recovery of the
resin. Large-scale field application of this technique is
reported to be under way (Gu et al., 2001).
Although biological degradation of perchlorate in ion
exchange brines can be difficult, some bacteria can grow
and degrade perchlorate even in highly saline solutions.
A culture containing primarily W. succinogenes HAP-1
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
XU ET AL.
416
was reported to degrade perchlorate in brines derived
from the regeneration of ion exchange resins (7% NaCl,
180 mg/L perchlorate) (Coppola and McDonald, 2000).
However, the 7% brine solution had to be diluted to 3%
with water to allow biodegradation of the perchlorate.
This dilution is undesirable, as it increases the total volume of salt solution requiring disposal. Logan et al.
(2001b) demonstrated that the biological reduction of
perchlorate was possible at salinities of up to 11% NaCl.
The perchlorate-degrading enrichment was developed using samples from The Great Salt Lake. Although perchlorate was degraded in these highly saline solutions,
the degradation rates were slow. Okeke et al. (2002) isolated a salt tolerant PRB strain, Citrobacter sp. IsoCock1,
that was able to achieve a 32% reduction of perchlorate
in a 7.5% solids solution in 1 week at an initial perchlorate concentration of 0.5 g/L. A 46% reduction of perchlorate was achieved by a mixture of this bacterium and
the strain perc1ace (Okeke et al., 2002).
Reverse osmosis (RO) can also be used to remove perchlorate from drinking water. The waste stream containing perchlorate that was produced from an RO process
contained much lower amount of salts (,1%) than the
NaCl brine generated using an ion exchange process. A
packed bed bioreactor, inoculated with the pure culture
perc1ace, was tested for its ability to remove perchlorate
from a simulated RO rejectate (Giblin et al., 2002). It
was found that this system removed 98% of perchlorate
from a twice-concentrated rejectate (total dissolved solids
of 0.4%) with an influent perchlorate concentration of 8
mg/L and a residence time of 2 h. Nitrate was removed
simultaneously with perchlorate from an initial concentration as high as 900 mg/L to below 4 mg/L. Despite the
efficiency of perchlorate removal, the system suffered
from clogging due to the high total dissolved solids of
the twice-concentrated rejectate.
It is possible to simultaneously remove and degrade
perchlorate using BAC. BAC has been used in drinking
water treatment to remove trace amounts of halo-acetic
acids in drinking water (Xie and Zhou, 2002). Kim and
Logan (2000) reported that GAC could be used to treat
water containing 10 mg/L of perchlorate. They found that
perchlorate was completely removed by the reactor at the
beginning of operation, but that after backwashing reactor efficiency was substantially reduced due to perchlorate desorption from the carbon in regions of low biological activity near the end of the column. At much lower
influent perchlorate concentrations of 50 mg/L, Brown et
al. (2002) found, over a 103-day study, that it was possible to remove perchlorate to below the detection limit
in a BAC operated at a 25-min empty bed contact time.
Although the BAC was not bioaugmented with PRB, they
found that Dechloromonas and Dechlorosoma were pres-
ent in the carbon bed during the period when perchlorate
removal was successful (Lin et al., 2002).
PHYTOREMEDIATION
Perchlorate contamination of the environment may affect agricultural plants as well as naturally occurring flora
(U.S. EPA, 2002). Chlorate has been used as a defoliant
(van Wijk and Hutchinson, 1995), and therefore, it is not
surprising that perchlorate can also be taken up by plants.
The accumulation of perchlorate in plants is of concern
for several reasons. Perchlorate can be toxic to some
plants, if the perchlorate accumulates and is not degraded,
and the death of the plant may release perchlorate back
into the environment that could be toxic to other plants
or wildlife (Urbansky et al., 2000). Perchlorate accumulation in food plants could present another route of human exposure to perchlorate. Perchlorate contaminated
water, such as Lake Mead or the Colorado River, is
presently used for irrigating food crops (Susarla et al.,
1999; Urbansky et al., 2000). Hutchinson et al. (2000)
are currently studying the accumulation of perchlorate in
lettuce irrigated with perchlorate-tainted water.
Because perchlorate can accumulate in plants, phytoremediation has been suggested as a potential mechanism for degrading perchlorate in soil systems. Phytoremediation may occur by phytoextraction (accumulation
in the branches and leaves), phytodegradation, or rhizotransformation (degradation in the root sphere primarily
due to microbial activity). Although many plants have
shown the ability to accumulate perchlorate, some plants
can drive perchlorate degradation completely to chloride
(Nzengung et al., 1999; Susarla et al., 1999; Hutchinson
et al., 2000; Nzengung and Wang, 2000). Nzengung and
Wang (2000) found that willow trees could degrade 100
mg/L of perchlorate in 53 days, and that minced spinach
and tarragon leaves could degrade 7 mg/L of perchlorate
in 30 days. There were no lag times for perchlorate degradation in either experiment. Perchlorate degradation by
plants was found to occur in two stages (Nzengung and
Wang, 2000). The first stage consisted of an initial uptake of perchlorate proportional to the water uptake by
the plant, and a slow transformation of perchlorate to
chloride in the plant tissues. The second stage was characterized by a rapid removal of perchlorate by degradation in the root zone with little perchlorate taken up by
the plant. This second stage was assumed to arise from
the stimulation of growth of PRB in the rhizosphere
(Nzengung et al., 1999; Nzengung and Wang, 2000).
Several phytoremediation studies have been conducted
to study the rate of perchlorate uptake and degradation,
and the effects of other chemicals on degradation rates.
MICROBIAL DEGRADATION OF PERCHLORATE
Nzengung et al. (1999) examined removal and degradation of perchlorate in water amended with acetate and
differing amounts of nitrate using sand or hydroponic systems cultivated with willow (Salix nigra). Both system
types were capable of removing perchlorate from the water. Degradation was found to occur due to transformation in the leaves and stems, and degradation by the microbial population in the plant rhizosphere. Multiple
dosing of sand systems with 100 mg/L perchlorate, and
less than 100 mg/L of nitrate, achieved a maximum
degradation rate of perchlorate of 2.35 mg/L-h (Nzengung et al., 1999). Krauter (2001) constructed a wetland
system to degrade both nitrate and perchlorate in groundwater. The system consisted of four tanks containing
gravel and planted with a variety of indigenous wetland
plants. Influent water contained nitrate (68 mg/L), perchlorate (4.5 mg/L), and trichloroethene (54 mg/L), but it
was further amended with sodium perchlorate (0.1 mg/L).
It was discovered that although the young wetland system was capable of supplying sufficient carbon to support denitrification, perchlorate reduction was carbonlimited. When acetate (260 mg/L) was added to the
influent, both nitrate and perchlorate were completely degraded within 52 h. The maturity of the plants in the
bioreactor and the time of year were found to greatly influence the nitrate and perchlorate degradation rates in
the bioreactor (Krauter, 2001).
CONCLUSIONS
Many heterotrophic biological treatment systems have
been tested to degrade perchlorate including suspended,
fixed-bed, and fluidized-bed reactors (Attaway and
Smith, 1993; Wallace et al., 1998; Attaway et al., 1999;
Green and Pitre, 2000; Coppola and McDonald, 2000;
Hatzinger et al., 2000; Perlmutter et al., 2000; Logan et
al., 2001a; Evans et al., 2002). Organic electron donors
that have been used include simple compounds such as
acetate and ethanol, as well as more complex organic substrates such as those found in compost piles. Perchlorate
degradation has also been achieved in bioreactors using
only inorganic amendments. These reactors are sustained
by hydrogen gas delivered by pressurization, gas transfer across liquid films, or synthetic membranes (Giblin
et al., 2000b; Miller and Logan, 2000; Nerenberg et al.,
2002). These hydrogen-based technologies are promising
technologies for water treatment because less biomass is
produced by autotrophic processes than heterotrophic
processes. Large-scale tests are needed to evaluate process efficiency and the economics of these different hydrogen-based systems.
Although at least one biological treatment process has
417
been approved for use in the state of California for drinking water treatment (DHS, 2002a), little has been done
to study the removal of PRB from the treated water.
Membrane bioreactors can be used to keep the bacteria
separated from the contaminated water (Batista and Liu,
2001), but these systems are at a less advanced stage of
development than other biological perchlorate treatment
systems. It has been suggested that reactors based on enzymes to reduce perchlorate could avoid the potential
health problems associated with biological treatment
(Betts, 1999). Perchlorate reductase, which can reduce
both perchlorate and chlorate (Kengen et al., 1999), and
chlorite dismutase (van Ginkel et al., 1996; Coates et al.,
1999; Stenklo et al., 2001) have both been purified. However, no such enzyme-based systems have been reported
in the literature for treatment of perchlorate contaminated
water.
The presence of alternate electron acceptors in perchlorate contaminated water will be an issue for all types
of biological reactors. Oxygen is an important intermediate in the perchlorate degradation pathway (Rikken et
al., 1996). It is well known that for PRB oxygen is a preferential electron acceptor to perchlorate, and that high
concentrations of dissolved oxygen inhibit perchlorate reduction (Kengen et al., 1999; Song and Logan, 2002). It
is not clear what concentration of dissolved oxygen will
completely inhibit perchlorate reduction, how long bacteria can withstand exposure to high concentrations of
oxygen before losing the ability to reduce perchlorate, or
how long it would take oxygen-exposed bacteria to regain the ability to reduce perchlorate. However, the presence of oxygen, nitrate, or sulfate in bioreactor feed
streams does not appear to be a problem for the steady
operation of such systems. In a pilot-scale test for ex situ
groundwater treatment, it was found that oxygen, nitrate,
and perchlorate were all completely reduced but that sulfate was not measurably degraded (Logan et al., 2001a;
Evans et al., 2002). Thus, it is likely that the main impact of oxygen and nitrate on a treatment system will be
to increase the requirement of substrate (such as acetate
or hydrogen) that is oxidized by the bacteria.
One of the most important issues for designing a perchlorate treatment system will be the regulatory requirement for perchlorate removal. It now appears likely that
the removal of perchlorate to very low levels will be necessary. The U.S. EPA is expected in the near future to finalize a draft of its final assessment of its toxicological
effects of perchlorate, which could lead to recommendations for perchlorate removal to ,1 mg/L for drinking
water (Renner, 2002). The action level of perchlorate for
drinking water in several states has been set at levels of
1–4 mg/L based on the release of the draft report
(www.clu-in.org/studio/perchlorate-060402). If enacted,
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003
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418
these regulatory levels could require perchlorate treatment systems for users of large water sources such as
Lake Mead and the Colorado River that serve major cities
in California, Utah, and Arizona (Motzer, 2001). Both biological and chemical treatment systems can be used to
treat perchlorate contaminated water to low levels. The
most appropriate system will likely be site and case-specific, with economic, social, and political factors playing
a role in the selection of each treatment system.
BETTS, K.S. (1999). A wide variety of bugs can break down
perchlorate. Environ. Sci. Technol. 33, 515A.
ACKNOWLEDGMENTS
BRUCE, R.A., ACHENBACH, L.A., and COATES, J.D.
(1999). Reduction of (per)chlorate by a novel organism isolated from paper mill waste. Environ. Microbiol. 1, 319–
329.
The authors thank Husen Zhang for constructing the
phylogenetic tree, and support from the National Science
Foundation (Grants BES9714575 and BES0001900), and
support via a grant from the United States Environmental Protection Agency, the East Valley Water District, and
the American Water Works Association Research Foundation (AwwaRF Grant No. 2557).
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