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An anaerobic two-layer permeable reactivebiobarrier for the remediation of nitrate-contaminated groundwater
She-Jiang Liu*, Zhi-Yuan Zhao, Jie Li, Juan Wang, Yun Qi
School of Environmental Science & Engineering, Tianjin University, Tianjin 300072, China
a r t i c l e i n f o
Article history:
Received 4 March 2013
Received in revised form
4 June 2013
Accepted 16 June 2013
Available online xxx
Keywords:
Nitrate
Remediation
Groundwater
Permeable reactive barrier
Denitrification
* Corresponding author. Tel.: þ86 22 2789001E-mail address: [email protected] (S
Please cite this article in press as: Liu, S.-nitrate-contaminated groundwater, Wat
0043-1354/$ e see front matter ª 2013 Elsevhttp://dx.doi.org/10.1016/j.watres.2013.06.028
a b s t r a c t
In this paper, an anaerobic two-layer permeable reactive biobarrier system consisting of an
oxygen-capturing layer followed by a biodegradation layer was designed firstly for evalu-
ating the remediation effectiveness of nitrate-contaminated groundwater. The first layer
filling with granular oxygen-capturing materials is used to capture dissolved oxygen (DO)
in groundwater in order to create an anaerobic condition for the microbial denitrification.
Furthermore, it can also provide nutrition, such as carbon and phosphorus, for the normal
metabolism of immobilized denitrifying bacteria filled in the second layer. The second
layer using granular activated carbon as microbial carrier is able to biodegrade nitrate
entering the barrier system. Batch experiments were conducted to identify the effect of DO
on microbial denitrification, oxygen-capturing performance of zero valent iron (ZVI)
powder and the characteristics of the prepared oxygen-capturing materials used to stim-
ulate growth of denitrifying bacteria. A laboratory-scale experiment using two continuous
upflow stainless-steel columns was then performed to evaluate the feasibility of this
designed system. The first column was filled with granular oxygen-capturing materials
prepared by ZVI powder, sodium citrate as well as other inorganic salts, etc. The second
column was filled with activated carbon immobilizing denitrifying microbial consortium.
Simulated nitrate-contaminated groundwater (40 mg NO3eN/L, pH 7.0) with 6 mg/L of DO
content was pumped into this system at a flow rate of 235 mL/d. Samples from the second
column were analyzed for nitrate and its major degradation byproduct. Results showed
that nitrate could be removed more than 94%, and its metabolic intermediate, nitrite, could
also be biodegraded further in this passive system. Further study is necessary in order to
evaluate performance of its field application.
ª 2013 Elsevier Ltd. All rights reserved.
1. Introduction Michalsen et al., 2006; Morrison et al., 2006). A PRB consist-
In the last decade, there has been an explosion of activities
directed at the development and implementation of a most
promising remediation technology-permeable reactive barrier
(PRB) (Bartzas et al., 2006; Basu and Johnson, 2012; Ebert et al.,
2006; Flury et al., 2009; Ludwig et al., 2009; Mak and Lo, 2011;
7; fax: þ86 22 27891291..-J. Liu).
J., et al., An anaerobic twer Research (2013), http:
ier Ltd. All rights reserved
ing of permanent or replaceable reactive media is placed in
the subsurface across the flow path of contaminated
groundwater, which must move through it as it flows, typi-
cally under its natural gradient, thereby creating a passive
treatment system. PRB is not a barrier to the groundwater, but
a barrier to the contaminants (Nooten et al., 2008; Phillips
o-layer permeable reactive biobarrier for the remediation of//dx.doi.org/10.1016/j.watres.2013.06.028
.
wa t e r r e s e a r c h x x x ( 2 0 1 3 ) 1e92
et al., 2010; Thiruvenkatachari et al., 2008; Van Nooten et al.,
2007). The main advantage of permeable reactive barrier is
the lower cost. Once installed, PRB does not need above
ground facilities or energy inputs, and it can take advantage of
the in situ groundwater flow to bring the contaminants in
contact with the reactive materials (Liu et al., 2006).
Agricultural runoff has been identified as the principal
source of groundwater contamination by nitrate. Additional
sources of nitrates contamination include landfill leachate,
leaking septic tanks, treated wastewater discharged to rivers,
and municipal storm water runoff (Savard et al., 2010; Suthar
et al., 2009; Tarkalson et al., 2006; Wick et al., 2012). In addi-
tion, climate changes such as changes in temperature, pre-
cipitation amounts and distribution, and the underlying
increases in atmospheric CO2 concentrations will impact on
both soil processes and agricultural productivity. Studies of soil
processes suggest climate change is likely to lead to increased
nitrate leaching from the soil. Climate change will also affect
the hydrological cycle with changes to recharge, groundwater
levels and resources and flowprocesses. The predicted impacts
are variable but many predictions suggest an overall decrease
in recharge and a fall in water levels and almost all predict an
enhanced seasonal variation in water levels. This will impact
on concentrations of nitrate in abstracted water and other
possibly more-sensitive receptors such as groundwater
dependent wetlands on an annual timescale (Stuart et al.,
2011). In recent decades, a lot of projects succeeded on
reducing nitrate pollution, nevertheless in most places nitrate
concentration in groundwater is still on the rise in varying
degrees (Chen et al., 2010; Fenech et al., 2012; Majumder et al.,
2008; Rivett et al., 2008). The EuropeanUnion andWorld Health
Organization (WHO) have both set the standard for nitrate in
potable water at 11.3 mg N/L (50 mg-NO3/L) (WHO, 2004).
Excessive ingested nitrites and nitrates from polluted drinking
waters can induce methemoglobinemia in humans, and also
have a potential role in developing cancers (Camargo and
Alonso, 2006; Fewtrell, 2004; Suthar et al., 2009). Many tech-
nologies are available for treating nitrate from groundwater,
such as reverse osmosis; ion exchange; chemical denitrifica-
tion; electrodialysis and distillation (McAdam and Judd, 2007;
Ricardo et al., 2012; Schnobrich et al., 2007). Although these
techniques are effective in moving nitrate from water, most of
them are limited in factual application for the remediation.
Themain products of such chemical reduction are ammonium
ions that are potential toxic to aquatic organisms at high
concentrations (Hwang et al., 2011; Li et al., 2010; Shin and Cha,
2008; Suzuki et al., 2012). Research is carried out toward nitrate
removal from water resources, whereas the most promising
approach being studied is biological denitrification. Microcosm
studies demonstrated that nitrate may be biodegrable with
special bacterial strains on natural isolates under aerobic and
anaerobic condition (Aslan and Cakici, 2007; Liu et al., 2009;
Wang et al., 2009; Zhou et al., 2011). The pathway for nitrate
reduction is
NO3/NO2/NO/N2O/N2 (1)
The reaction of complete microbial denitrification is
commonly shown in general equation (reaction 2) where
microbially available carbon is simplified as carbohydrate
(Rivett et al., 2008).
Please cite this article in press as: Liu, S.-J., et al., An anaerobic twnitrate-contaminated groundwater, Water Research (2013), http:
4NO�3 þ 5CH2O/2N2 þ CO2 þ 4HCO3� þ 3H2O (2)
Biological denitrification is considered to be the most
economical strategy among other conventional techniques
like physicochemical. The denitrifying bacteria using nitrate
as sole source of nitrogen under anaerobic conditions grow
slowly with low yields of biomass and are sometimes unsta-
ble. As a result, an effective bioremediation process for nitrate
from groundwater has not been fully developed so far.
The aim of the recent studies was to select a suitable nat-
ural organic substrate as a potential carbon source for use in a
denitrification PRB (Gibert et al., 2008). However, few re-
searchers focus their attention on the negative influence of
microbial denitrification caused by dissolved oxygen (DO) level
in the groundwater so far. In generally, DO content is not low
enough in the groundwater (Schnobrich et al., 2007). As a
result, an anaerobic condition is difficult to achieve for deni-
trifying bacteria used in the bioremediation of nitrate-
contaminated groundwater. This paper attempted to treat
groundwater contaminated by nitrate using a two-stage
removal system. The first is to capture DO in order to create
artificially an anaerobic environment in groundwater, and the
second is to degrade nitrate using the denitrifying bacteria.
Studies demonstrated that zero valent iron (ZVI) has a chem-
ical reaction with O2 dissolved in the water (Su and Puls, 2007).
2Fe0 þ 2H2OþO2/2Fe2þ þ 4OH� (3)
2Fe0 þ 4Hþ þO2/2Fe2þ þ 2H2O (4)
Therefore, ZVI was selected as the potential oxygen-
capturing reagent in this paper. In addition, the following
reasons also make it good candidate for this study: (1) It is
nontoxic to aquatic organisms; (2) Nitrate can be degraded
chemically by ZVI; (3) It is available and cheap.
Besides indigenous microbe, injection of special bacterial
strains as well as nutrient salts is usual measure for
enhancing remediation efficiency of groundwater (Ito et al.,
2012). For preventing special microorganism injected into
the aquifer from losing with groundwater flow, it may be
necessary to immobilize bacteria in the barrier (Ha et al., 2009).
In this paper, activated carbon was selected as microbial
carrier because of the following reasons: (1) the surface of
activated carbon is porous and coarse, which helpsmicrobe to
adsorb and immobilize; (2) activated carbon doesn’t bring any
new contaminant into groundwater when it is placed in the
barrier; (3) activated carbon is relatively inexpensive.
Based on the above discussions, we designed an anaerobic
two-layer permeable reactive biobarrier system containing
oxygen-capturing and biodegradation layers to evaluate the
remediation effectiveness of nitrate-contaminated ground-
water. The first layer that filled with granular oxygen-
capturing materials can capture oxygen dissolved in ground-
water, and provide carbon source as well as other nutrition for
the metabolism of denitrifying bacteria. The second layer that
filled with immobilized denitrifying bacteria can enhance the
removal efficiency of nitrate in the groundwater. The sche-
matic diagram of this designed system is shown in Fig. 1.
The principle of this work was to design a passive treat-
ment system to bioremediate groundwater contaminated by
nitrate. Experiments were conducted as follows:
o-layer permeable reactive biobarrier for the remediation of//dx.doi.org/10.1016/j.watres.2013.06.028
Fig. 1 e Schematic diagram of the designed biological
barrier system.
wat e r r e s e a r c h x x x ( 2 0 1 3 ) 1e9 3
(1) Effects of DO on biodegradation efficiency of nitrate;
(2) Oxygen-capturing performance of ZVI;
(3) Preparation and characteristics of oxygen-capturing
materials;
(4) A column experiment for evaluating the feasibility and
potential by using this designed barrier system.
2. Material and methods
2.1. Denitrifying Bacteria’s collection, enrichment andacclimation
The original experimental microbes were collected from a
cornfield soil located 25e40 cm deep at Xiqing District, Tianjin
city, China. Microbes were enriched and acclimated in a bio-
chemical culture bottle for the next experiments. The com-
ponents of liquid mineral salts mediumwere as follows (units
are in mg/L of water): sodium citrate, 5000; KNO3, 2000;
KH2PO4, 1000; K2HPO4, 1000; CaCl2, 180; MgSO4, 100. KNO3 was
used as the sole nitrogen source in themedium. It is sufficient
to mention that the DO was removed from the medium by
sparging nitrogen in this present work. A previous study has
reported that the optimum pH for microbial denitrification
should be keep in the range of 7.0e9.0 (Tang et al., 2011).
Thereby, the pH of cultures was checked every 3 day, and it
was regulate to about 7.5 using KH2PO4 and K2HPO4 when the
pH value changed. Microbes were enriched and acclimated at
room temperature for 2 months. Then, the suspension of
acclimated denitrifying bacteria was used for the subsequent
experiments.
2.2. Batch experiments
2.2.1. Effects of DO on the denitrification efficiency of nitrateBecause of the effect of DO on the nitrate reductase activity
and metabolism pathways, the control of DO is crucial to the
nitrate biodegradation. In the present experiments, the effect
of DO on denitrification was investigated by varying the initial
Please cite this article in press as: Liu, S.-J., et al., An anaerobic twnitrate-contaminated groundwater, Water Research (2013), http:
DO value in the medium from 0.02 mg/L to 4 mg/L. The ex-
periments were performed in four 600 mL enclosed reactors
containing 400 mL of the mineral salts medium described
above except KNO3. In order to control the DO content, ni-
trogen was sparged into the medium before the experiment.
40 mg NO3eN/L was added into the each reactor firstly, and
then, 100 mL of the suspension of acclimated denitrifying
bacteria was inoculated, respectively. The four reactors, once
inoculated, were enclosed and placed in a reciprocal shaker at
constant temperature (20 �C) and rotate speed (130 r/min). The
control case was also maintained with the same concentra-
tion of nitrate and abiotic denitrifying bacteria. The enclosed
reactors were incubated for 60 h with shaking. The concen-
trations of NO3eN in the reactor were analyzed at the begin-
ning and end of the experiments.
2.2.2. Oxygen-capturing performance of ZVITechnical ZVI powder was purchased from Tianjin Jiangtian
Chemical Corporation Ltd, whose main impurity was about
0.5% of sulfuric acid insoluble material by weight. ZVI can
consume oxygen upon contact with water according to the
equation (3) or (4). For evaluating oxygen-capturing perfor-
mance of ZVI, the experiments were performed by adding
1000mL of sterile deionized water into four enclosed reactors,
and then 208, 503, 804, 1000 mg of ZVI powder were also
added, respectively. Here, it is sufficient to mention that 6mg/
L of DO content in sterile deionized water was controlled by
sparging nitrogen at the beginning of experiments. In addi-
tion, the same adding weights of ZVI powder (1000 mg/L)
under the condition of different initial DO values was also
studied. All above experiments were conducted in a reciprocal
shaker at constant temperature (20 �C) and rotate speed (130 r/
min). A portable DO meter (HQ30D, HACH) was used to online
monitor the DO change.
2.2.3. Preparation and characteristics of oxygen-capturingmaterialsThe oxygen-capturing materials were prepared by blending
ZVI powder, sodium citrate, KH2PO4, K2HPO4, CaCl2, MgSO4,
cement, quartz sand at a ratio of 0.10:0.20:0.04:0.04:0.006:
0.004:0.30:0.31 by weight. ZVI powder was used as oxygen-
capturing reagent for the anaerobic denitrifying bacteria, so-
dium citrate was used as the carbon source, KH2PO4 and
K2HPO4 were used to provide nutrients for in situ nutrient
supplement, CaCl2 andMgSO4 were used to provide necessary
elements for the microbial growth, cement was used as the
binder, and quartz sand (40e80 mesh) was used to increase
the permeability of the materials, which may make nutrients
(carbon, phosphorus and so on) easy to release from the ma-
terials’ interior, and ZVI located in the materials’ interior
sufficient to react with O2 dissolved in groundwater. Above
powdery components were blended, and a certain amount of
water was added under the condition of 15e18 mL/100 g of
liquidesolid ratio. A pharma-ceutical extruder-rounder was
used to turn the powdery components into granular materials
of about 5 mm in diameter. The prepared granular materials
were kept in a laboratory vacuum freeze dryer for 24 h. And
then, the oxygen-capturing materials were obtained.
Two Erlenmeyer flasks (500 mL) used for the growth of
anaerobic denitrifying bacteria were prepared. 400 mL of
o-layer permeable reactive biobarrier for the remediation of//dx.doi.org/10.1016/j.watres.2013.06.028
Fig. 2 e Schematic diagram of the column experimental
setup.
wa t e r r e s e a r c h x x x ( 2 0 1 3 ) 1e94
sterile deionized water with DO value of 6 mg/L was added
into a flask containing 8 g of prepared oxygen-capturing ma-
terials and 40mg NO3eN/L. The same volume of mediumwith
DO value of 6 mg/L was added into the other flask containing
40 mg NO3eN/L, in which the medium had the same mineral
salts components described above except KNO3. The flasks
were then enclosed when 100 mL of denitrifying bacteria was
inoculated, respectively. The Erlenmeyer flasks, once inocu-
lated, were incubated at 20 �C in a reciprocal shaker (130 r/
min). Optical density measurements were made to express
the growth of denitrifying bacteria with a spectrophotometer
at 600 nm.
2.3. Microbial immobilization
The laboratory column of 50 cm length and 4 cm internal
diameter made of stainless-steel was homogeneously packed
with activated carbon. Activated carbon was purchased from
Tianjin Jiangtian Chemical Corporation Ltd, whose main im-
purity was about 2% of ethanol soluble material by weight.
The other characteristics of activated carbon are as following:
particle size of (4e5) mm � F (1.5e2.5) mm; specific surface
area of about 1400 m2/g. The suspension of acclimated deni-
trifying bacteria was injected into this column to submerse
the activated carbon. Column feed solution consisting of
40 mg NO3eN/L and above mineral salts medium was pump
into the column by using a peristaltic pump at a flow rate of
0.5 L/d. The aim of this process is to maintain the microbial
metabolism and permit the development of microbial film on
the surface of activated carbon. In order to ensure this system
under the anaerobic condition, nitrogen was introduced into
the feed solution before entering this column. TheDO levels in
the effluent, the concentrations of nitrate in the influent and
effluent were measured respectively every 2e3 days. The
whole process of microbial immobilization lasted 20 days.
2.4. Column experiment
A laboratory-scale barrier system was designed by using two
continuous upflow stainless steel columns. The first oxygen-
capturing column (50 cm length and 4 cm internal diameter)
was filled with granular oxygen-capturing materials prepared
above. The column that had immobilized denitrifying bacteria
described above was used as the biodegradation column,
which was equipped with 4 sampling ports positioned every
10 cm. These portswere numbered 1e4 from the bottom to top
of this column. In this study, NO.1 and NO.3 were monitored
for the change of nitrate concentration versus time. Simulta-
neously, the influent and effluent of the biodegradation col-
umn were also analyzed. In addition, as an appropriate
parameter to assess the success of remediation effort, nitrite
was also monitored during the period of column experiment.
Simulated nitrate-contaminated groundwater (40 mg NO3eN/
L, pH 7.0) with 6 mg/L of DO content was continuous pump
into the first columnwith an upflowmode by peristaltic pump
at a flow rate of 235mL/d. After the solution passed through it,
designated hour 0, the biodegradation column was connected
to the first column. It is emphasized here that the feed solu-
tion using for microbial immobilization in the biodegradation
column had been replaced by sterile deionized water
Please cite this article in press as: Liu, S.-J., et al., An anaerobic twnitrate-contaminated groundwater, Water Research (2013), http:
(DO ¼ 0 mg/L) before column experiment. Fig. 2 presents the
schematic diagram showing the laboratory anaerobic two-
layer permeable reactive biobarrier system.
2.5. Analytical methods
The column system was operated for about 730 h at room
temperature (w20 �C). UV/Vis spectrophotometer was used to
analyze the nitrate and nitrite concentrations in the sampling
ports, influent and effluent of biodegradation column. The pH
and DO content in the influent and effluent of biodegradation
columnwere determined respectivelywith a PHS-3C pHmeter
and a portable DO meter (HQ30D, HACH).
3. Results and discussion
3.1. Effects of DO on biodegradation efficiency of nitrate
Most denitrifying bacteria were anaerobic and greatly influ-
enced by DO content in the solution. As a result, the DO
content can cause the significant difference in the biodegra-
dation efficiency of nitrate. Fig. 3 illustrates the effect of DO
content on the biodegradation efficiency of nitrate. The
degradation experiment was operated for about 60 h. As the
DO content decreased from 4 mg/L to 0.02 mg/L, the biodeg-
radation efficiency of nitrate increased from 70.1% to 85.3%.
That showed negative correlation between the degradation
efficiency of nitrate and DO value. According to the results of
this batch experiment, we can conclude that biodegradation
efficiency of nitrate can be enhanced with DO content
decrease. This is the just reason that an oxygen-capturing
layer was designed in this barrier system before ground-
water contaminated by nitrate enters into the biodegradation
layer.
3.2. Oxygen-capturing performance of ZVI
Many works have been done by utilizing ZVI for pollutants
removal in groundwater (Flury et al., 2009; Hwang et al., 2011;
o-layer permeable reactive biobarrier for the remediation of//dx.doi.org/10.1016/j.watres.2013.06.028
60%
65%
70%
75%
80%
85%
90%
0 1 2 3 4 5
DO (mg/L)
Deg
rada
tion
effi
cien
cy
Fig. 3 e Effects of DO on the biodegradation efficiency of
nitrate.
1.0
2.0
3.0
4.0
5.0
6.0
0 200 400 600 800
Time (min)
DO
(m
g/L
)
(a)
1.0
3.0
5.0
7.0
9.0
0 50 100 150 200
Time (min)
DO
(m
g/L
)
(b)
1000mg/L804mg/L
503mg/L208mg/L
8mg/L
7mg/L
5mg/L
Fig. 4 e Oxygen-capturing performance of ZVI (a) DO
change with different adding weights of ZVI (b) DO change
with differently initial DO value.
wat e r r e s e a r c h x x x ( 2 0 1 3 ) 1e9 5
Ludwig et al., 2009; Morrison et al., 2006; Phillips et al., 2010).
For example, chemical reactions in the process of remediation
of nitrate-contaminated groundwater using ZVI are concluded
as following.
5Feþ 2NO�3 þ 6H2O/5Fe2þ þN2 þ 12OH� (5)
4FeþNO�3 þ 7H2O/4Fe2þ þNHþ
4 þ 10OH� (6)
FeþNO�3 þH2O/Fe2þ þNO�
2 þ 2OH� (7)
In recent years, some researchers have developed oxygen-
releasing compounds, such as calcium peroxide and magne-
sium peroxide, to increase passively DO in groundwater for
improving aerobic biodegradation (Kunukcu, 2007; Liu et al.,
2006; Yeh et al., 2010). However, there are seldom studies
about how to decrease DO in groundwater in order to enhance
anaerobic biodegradation. According to chemical reaction
equation (3) or (4), O2 dissolved in the water can be consumed
effectively by Fe. Thereby, ZVI was used as a good oxygen-
capturing reagent in this work.
A previous study indicated that microbial denitrification
can occur when DO is below 1 mg/L, or even below 2 mg/L
(Rivett et al., 2008). Oxygen-capturing performance of ZVI
powder is illustrated in Fig. 4, in which Fig. 4a presents DO
change with different adding weights of ZVI, and Fig. 4b
presents the effect of initial DO content on such performance
under the condition of same adding weights of ZVI. As seen
from Fig. 4a, DO value in water decreased quickly once ZVI
was added, and the different adding weights of ZVI had
obvious influence on the oxygen-capturing rates. This effect,
however, was no long obvious when the adding weight of ZVI
was more than 800 mg/L. From Fig. 4b, it was observed that
the decrease rates of DO became slow with the initial DO
content decrease. Field studies provided evidence of a diverse
microbial population within and in the vicinity of the iron
barrier. Microbial populations are important not only for
nutrient cycling, but also for contaminant remediation.
Adding ZVI powder had no deleterious effect on total bacte-
rial abundance in the microcosms (Gu et al., 2002; Van Nooten
et al., 2010).
Please cite this article in press as: Liu, S.-J., et al., An anaerobic twnitrate-contaminated groundwater, Water Research (2013), http:
3.3. Effects of oxygen-capturing materials on microbialgrowth
Fig. 5 presents growth curves of denitrifying bacteria with
oxygen-capturing materials and without oxygen-capturing
materials. As depicted in Fig. 5, growth of denitrifying bacte-
ria in the sterile deionized water containing oxygen-capturing
materials was better than that of denitrifying bacteria in the
mineral salts medium. Because the denitrifying bacteria
require a period of time to adapt to new environment and
synthesize denitrifying enzymes, there was no significant
difference during the first 1 h under the both conditions. Cell
counts, however, at the end of growth were obvious differ-
ence. The results suggested that the prepared oxygen-
capturing materials can satisfy all requirements for deni-
trifying bacteria. That is, thematerials can create an anaerobic
environment for denitrifying bacteria and provide enough
nutrients for microbial metabolism. In addition, it proved
further that effects of DO on denitrifying bacteria cannot be
neglected.
3.4. Microbial immobilization
Microbial immobilization was determined by measuring DO
value in the effluent, and the concentration of nitrate in the
o-layer permeable reactive biobarrier for the remediation of//dx.doi.org/10.1016/j.watres.2013.06.028
0.5
0.7
0.9
1.1
1.3
0 2 4 6 8 10 12
Time (h)
OD
600n
m
Fig. 5 e Growth curves of denitrifying bacteria with
oxygen-capturing materials (-) and without oxygen-
capturing materials (:).
6.0
6.5
7.0
7.5
8.0
8.5
9.0
0 100 200 300 400 500 600 700 800
Time (h)
pH
Fig. 6 e Variation of pH values in the influent and effluent
of biodegradation column.
wa t e r r e s e a r c h x x x ( 2 0 1 3 ) 1e96
influent and effluent throughout 20 days. During the begin-
ning phase (0e4 days), the effluentwas slight yellow because a
few of denitrifying bacteria were washed out. But with the
time passing, the effluent became gradually clear. It indicated
that denitrifying bacteria had been adsorbed on the microbial
carrier. During thewhole process ofmicrobial immobilization,
DO levels in the effluent were in the range of 0.05e0.10 mg/L,
and the concentration of nitrate in the effluent was below the
influent’s (data not shown). As described above, it suggested
that there was formation of microbial film on the carriers, and
the attached denitrifying bacteria could maintain the anaer-
obic metabolism. For the field application, immobilization of
denitrifying bacteria offers several advantages over freely
suspended cells, such as highly dense microbial mass in
special remediation area and avoidance of microbe washout
when the groundwater flows through it.
0 100 200 300 400 500 600 700 800
Time (h)
DO
(m
g/L
)
Fig. 7 e Variation of DO levels in the influent and effluent of
biodegradation column.
3.5. Column experiment
In the column experiment, the samples of biodegradation
column influent, effluent and specific sampling ports were
monitored and analyzed for pH, DO, nitrate and nitrite.
Fig. 6 presents the variation of pH values in the influent and
effluent of biodegradation column. The pH values in the
influent of biodegradation column varied slightly from 7.92 to
7.65 when the simulated nitrate-contaminated groundwater
flowed through the oxygen-capturing column, which were
still in the suitable pH range of 7.0e9.0, and then closed to
about 7.5 slowly after 80 h. The influent of biodegradation
column was alkalescence that should be attributed to the re-
action of ZVI with oxygen (equation (3)) in the oxygen-
capturing column. Due to the gradual replacement of deion-
ized water by solution from the oxygen-capturing column, the
observed pH values in the effluent of biodegradation column
increased rapidly from 7.02 to 7.47 before about 80 h. By
analyzing experimental data about pH, we concluded that the
biodegradation column was always in a suitably pH condition
for microbial denitrification during the whole experimental
period.
Please cite this article in press as: Liu, S.-J., et al., An anaerobic twnitrate-contaminated groundwater, Water Research (2013), http:
The variation of DO levels in the influent and effluent of
biodegradation column is shown in Fig. 7. According to the
equation (3), the reaction of O2 with ZVI occurred in the
oxygen-capturing column because DO of simulated nitrate-
contaminated groundwater entering this system was 6 mg/L.
As a result, DO concentration in the influent of biodegradation
column dropped quickly to 1.4 mg/L in the beginning of col-
umn experiment (0e20 h), and then, as the reaction reached
equilibrium gradually, DO levels decreased slowly to 1.05 mg/
L. Because the feed solution using for microbial immobiliza-
tion in the biodegradation column was replaced by sterile
deionized water (DO ¼ 0 mg/L) before connecting it to the
oxygen-capturing column, the observed DO levels in the
effluent of biodegradation column went up slowly to 0.1 mg/L
in the beginning of column experiment, and then reached a
quasi-steady state and remained constant (about 0.1 mg/L).
According to the variation of DO levels in the influent and
effluent of biodegradation column, it can be concluded that
the biodegradation column was always in anaerobic
o-layer permeable reactive biobarrier for the remediation of//dx.doi.org/10.1016/j.watres.2013.06.028
0.005
0.007
0.009
0.011
0.013
0.015
0 100 200 300 400 500 600 700 800
Time (h)
mg
NO
-N
/L 2
mg
NO
-N
/L 2
(a)
0.04
0.08
0.12
0.16
0.20(b)
wat e r r e s e a r c h x x x ( 2 0 1 3 ) 1e9 7
environment, and microbial denitrification may occur in this
designed system.
The concentrations of nitrate and its degradation byprod-
uct (nitrite) versus time in the influent, effluent and specified
sampling ports of biodegradation column are shown in Figs. 8
and 9, respectively. Before the column experiment started,
concentrations of nitrate in each specified sampling port were
below the detection limit.
As seen from Fig. 8a, a slight decrease in nitrate concentra-
tion from 40 mg NO3eN/L (initial concentration of simulated
groundwater) to 39.08 mg NO3eN/L was observed after the
simulated nitrate-contaminated groundwater flowed through
the oxygen-capturing column. It implied that nitrate was
degradedchemicallybyZVIcomponent intheoxygen-capturing
materials, and the removal efficiency caused by the chemical
reaction was about 2.43%. As indicated in Fig. 8b, the variation
trendofnitrate concentration in theeffluentwassimilar to each
specific sampling port. That is, the nitrate concentration
increased gradually during the beginning phase of the experi-
ment, and then reached a quasi-steady state and remained
constant. For the effluent, nitrate concentration increased
gradually to 2.32mgNO3eN/L during the beginning phase of the
experiment (0e216 h), and it was no longer variable after this
period. The results suggested that the microbial denitrification
reached equilibrium in the biodegradation column.
35.0
37.0
39.0
41.0
43.0
45.0
0 100 200 300 400 500 600 700 800
Time (h)
mg
NO
-N
/L3
mg
NO
-N
/L3
(a)
Time (h)
(b)
Fig. 8 e Concentrations of nitrate in the influent, effluent
and specified sampling ports (a) the influent (b) NO.1; NO.3;
the effluent.
0.000 100 200 300 400 500 600 700 800
Time (h)
Fig. 9 e Concentrations of nitrite in the influent, effluent
and specified sampling ports (a) the influent (b) NO.1; NO.3;
the effluent.
Please cite this article in press as: Liu, S.-J., et al., An anaerobic twnitrate-contaminated groundwater, Water Research (2013), http:
Fig. 9a presents the nitrite concentration in the influent of
biodegradation column. Nitrite (0.0097 mg NO2eN/L) was
detected at the beginning of the experiment because of the
chemical reaction of nitrate with ZVI, and then it decreased
gradually. Finally, the nitrite concentration reached a quasi-
steady state and remained constant (0.0081 mg NO2eN/L).
Here, the detection of nitrite was also a proof that the nitrate
was degraded chemically by ZVI in the oxygen-capturing
column. Fig. 9b presents the concentrations of nitrite in the
effluent and specified sampling ports of biodegradation col-
umn. As seen from Fig. 9b, significant accumulation of nitrite
was found before 200 h, and then the nitrite started to
decrease. As the nitrate byproduct, the transient accumula-
tion of nitrite in the early operational periodmay be attributed
to the relative difference of nitrate and nitrite degraded rates.
For the effluent, nitrite concentration decrease finally to about
0.08 mg NO2eN/L after 400 h. All the experimental data indi-
cated that the metabolic intermediate of nitrate, nitrite, could
also be degraded further in the biodegradation column.
The removal efficiencies of nitrate in the specified sam-
pling ports (NO. 1, NO. 3) and the effluent of biodegradation
column were about 74.3%, 82.5% and 94.1%, respectively.
Compared with other studies on bioremediation of nitrate-
contaminated groundwater, higher efficiency of nitrate
biodegradation was obtained in this study (Huang et al., 2012;
o-layer permeable reactive biobarrier for the remediation of//dx.doi.org/10.1016/j.watres.2013.06.028
wa t e r r e s e a r c h x x x ( 2 0 1 3 ) 1e98
Zhou et al., 2011). It demonstrated that the anaerobic two-
layer permeable reactive biobarrier system indeed enhanced
the bioremediation effectiveness of nitrate-contaminated
groundwater. It is necessary to mention that some studies
have discussed the longevity of field scale PRB. Considering
the usually slow groundwater movement, PRB has to function
properly for decades. But, within time, the accumulation of
mineral precipitates and hydrogen gas can reduce barrier
reactivity and permeability (Flury et al., 2009; Phillips et al.,
2010; Robertson et al., 2008). In this study, the laboratory-
scale barrier system was operated for about 730 h at room
temperature. The removal efficiency of nitrate has remained
steady for a long time. It suggests that the anaerobic two-layer
permeable reactive biobarrier system for field remediation of
nitrate-contaminated groundwater is practical and achiev-
able, and further study is necessary to evaluate performance
of its field application.
4. Conclusions
In the batch experiments, as the DO content decreased from
4 mg/L to 0.02 mg/L, the denitrification efficiency of nitrate
increased from 70.1% to 85.3%. Thereby, the effect of DO on
denitrifying bacteria cannot be neglected. Based on the
experiment of oxygen-capturing performance of ZVI, it can be
concluded that ZVI can consume O2 dissolved in the water
efficiently. As a result, ZVI can be used as an available oxygen-
capturing reagent when groundwater needs to be treated
using an anaerobic biotechnology. According to the growth
curves of denitrifying bacteria, the prepared oxygen-capturing
materials can satisfy all requirements for denitrifying bacte-
ria. That is, the granular materials can create an anaerobic
environment for denitrifying bacteria and provide enough
nutrients for microbial metabolism.
An anaerobic two-layer permeable reactive biobarrier
system was designed to bioremediate nitrate-contaminated
groundwater. Based on the results of the column experi-
ment, occurrence of anaerobic degradation in the designed
system can be verified by the reduction of nitrate. The deni-
trification efficiency of the column experiment was estimated
to be more than 94%. As the nitrate byproduct, the transient
accumulation of nitrite in the early operational period may be
attributed to the relative difference of nitrate and nitrite
degraded rates. After the accumulation phase, nitrite started
to be biodegraded. Finally, the concentrations of NO3eN and
NO2eN in the simulated groundwater treated by this passive
system were below the standards set by the USEPA (10 mg
NO3eN/L and 1.0 mg NO2eN/L). Results from this study will be
useful in designing an anaerobic two-layer permeable reactive
biobarrier system for field remediation of nitrate-
contaminated groundwater. Further study is necessary to
evaluate performance of its field application.
Acknowledgments
This work is supported by the Natural Science Foundation of
Tianjin, Tianjin City, China (No. 10JCYBJC05500).
Please cite this article in press as: Liu, S.-J., et al., An anaerobic twnitrate-contaminated groundwater, Water Research (2013), http:
Appendix A. Supplementary data
Supplementary data related to this article can be found at
http://dx.doi.org/10.1016/j.watres.2013.06.028.
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