an anaerobic two-layer permeable reactive biobarrier for the remediation of nitrate-contaminated...

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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

.

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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:




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


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




Fig. 2 e Schematic diagram of the column experimental

setup.


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


(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;




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.


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).


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




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.


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.


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




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)


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.


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;





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).


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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