Remediation of Nitrate and Chromium Contaminated Groundwater by Zero-valent Iron PRB

Frontier of Environmental Science June 2015, Volume 4, Issue 2, PP.39-45

Remediation of Nitrate and Chromium Contaminated Groundwater by Zero-valent Iron PRB Fansheng Meng 1,2#, Yeyao Wang 3, Liping Bai 2 1. Research Center for Water Pollution Control Technology, Chinese Research Academy of Environmental Sciences, Beijing, China 2. State Key Laboratory of Environmental Criteria and Risk Assessment, Chinese Research Academy of Environmental Sciences, Beijing, China 3. China National Environmental Monitoring Center, Beijing, China #

Email: mengfs@craes.org.cn

Abstract Through continuous flow experimentation, the reactivity characteristics of zero-valent iron (Fe0)-PRB with ground water contaminated by nitrate, chromium and the combination of nitrate and chromium were investigated. The results showed that nitrate could be effectively deoxidized by zero-valent iron. NO2--N was the transitional deoxidization product, while NH4+-N was the main final product in the effluent. Chromium could be deoxidized by zero-valent iron more effectively for the chromium contaminated ground water which was treated by PRB. The redox products such as Fe 3+ and Cr(III) precipitated on the packing media during the process. For the treatment of ground water contaminated by both nitrate and chromium, the results showed that the Cr(VI) removal efficiency by the zero-valent iron was not affected by the co-existence of NO3--N, while the NO3--N removal efficiency decreased with the existence of Cr(VI).

Keywords: Zero-valent Iron; Permeable Reactive Barrier(PRB); Ground Water; Nitrate; Chromium

1 INTRODUCTION Groundwater is the most important water resource on Earth, and it is a major source of water for domestic, industrial and agricultural uses in many countries (Alemaw et al. 2004; Brito et al. 2006). The contribution from groundwater is vital, perhaps as many as two billion people depend directly upon aquifers for drinking water, and 40% of the world’s food is produced by irrigated agriculture that relies largely on groundwater (Thiruvenkatachari et al. 2008).

In China, groundwater provides nearly 70% drinking water and 40% irrigation water. With the social and economic development, huge amount of discharged industrial wastewater and sewage, randomly stacked municipal solid waste and disordered exploration of the groundwater come along. As a result, groundwater was contaminated. Nitrate and chromium contamination of groundwater has become an environmental and health problem in developed and developing countries. Nitrate pollution is caused by the intensive use of nitrogen fertilizers, crop irrigation with domestic wastewater and use of manure, therefore, it is concern of diffuse pollution. The non-point sources of nitrate contamination make really difficult to apply the ex-situ approach to groundwater remediation (Rocca et al. 2007).

Contamination of groundwater by Cr at numerous localities primarily resulted from uncontrolled or accidental release of Cr-bearing solutions, used in various industrial applications, into the subsurface environment (Mukhopadhyay et al. 2007; Xu et al. 2007).

Developed countries have put a lot of effort into the study of contaminated groundwater remediation, and in-situ remediation technology has been developing rapidly. Permeable Reactive Barrier (PRB) is an in-situ remediation technology that has been developed rapidly in the past 20 years. PRBs are an emerging alternative to pump-and-treat systems for treating groundwater contamination. PBR is usually installed in underground aquifers, perpendicular to - 39 http://www.ivypub.org/fes

the direction of groundwater flow. The concentration of pollutants declines by means of adsorption, precipitation, oxidation reduction and biological degradation, while the contaminated groundwater flows through the reaction medium (Burmeier 1998; Wilkin et al. 2003; Mak et al. 2011). PRB can effectively remediate the groundwater contaminated by various contaminants, including organic compounds, heavy metals, inorganic ions and radioactive element (Guerin et al. 2002; Ludwig et al. 2002; Blowes et al. 2000; Barton et al. 2004; Thiruvenkatachari et al.

2008). Zero-valent iron (Fe0) is widely used as the reaction medium for PRB (Gavaskar 1999; Cundy et al. 2008). At present, the researchers have paid much attention to the remediation of single pollutant contamination for Fe0-PRB system. This paper employed Fe0-PRB continuous flow simulation experiments to remedy nitrate and chromium combined contaminated groundwater; discussed the reactivity characteristics and products of the reduction of nitrate and chromium by Fe0-PRB; studied the interaction effects of coexisted pollutants, so as to provide technical support to the application of Fe0-PRB in China.

2 MATERIAL AND METHODS 2.1 Designs of experiment The experiment equipment consists of solution reservoir, peristaltic pump, PRB simulator and sampling well. The PRB simulator is made of engineering plastics. As shown in Fig.1, the PRB simulator is a cuboid with L×W×H=1000mm×500mm×400mm.The illustrations of the three components of the PRB simulator are shown in Tab.1.

4

Ⅰ 1

Ⅱ

Ⅲ

2

Ⅱ

Ⅰ 2

3

Effluent water

1. Solution reservoir 2.Peristaltic pump 3.PRB system 4.Sampling well FIG. 1 SCHEMATIC OF EXPERIMENT APPARATUS TABLE 1 COMPOSITION OF PART Ⅰ, Ⅱ, Ⅲ IN PRB APPARATUS Section

Composition

Particle size

Thickness

Ⅰ

Silica sand

0.45～1.0mm

10cm

Ⅱ

Sandy soil

<0.25mm

35cm

Ⅲ

Reaction medium

10cm

Illustration For injecting and sampling, silica sand plays a role of filtering, buffering and protecting Simulating aquifer, the sandy soil was sampled from orchard soil in CRAES PRB reactor, according to pre-experiment (Meng et al. 2007), reaction medium was prepared by mixing an equal weight of Fe0 and active carbon, the particle size of Fe0 ranges from 0.15-0.42mm, the particle size of active carbon is 0.15mm.

2.2 Experiment materials Fe0 was collected from industrial scrap iron, and the content of Fe was above 97% by element analyzing. The scrap iron in the size range of 0.15-0.42mm was screened out, and then it was wiped off the oxide film covering the surface with 1mol/L hydrochloric acid, rinsed with ultrapure water to neutral, dried with nitrogen, and stored in a vacuum environment as a stand-by. The measured surface area was 3.43 m2/g. The activated carbon was industrial activated carbon and the measured surface area was 856.17 m2/g.

All chemical reagents were analytically pure grade. The experiment water was prepared with tap-water, potassium dichromate and sodium nitrate. The concentration of hexa-valent chromium (Cr(Ⅵ)) and NO3--N was 10mg/L and 20mg/L respectively. The value of pH ranged from 6.8-7.4. The flow velocity (50-80cm/d) of groundwater was - 40 http://www.ivypub.org/fes

controlled by peristaltic pump and the hydraulic gradient of PRB device was 3‰. The artificial polluted water prepared for simulation was pumped into simulation system. The liquid level of sampling well holed at a certain height. The treated water was pumped out by peristaltic pump. Treated water samples were collected from sampling reservoir at appropriate time intervals and analyzed on the concentrations of Cr(Ⅵ), total chromium(TCr), NO3--N, NO2--N, NH4+-N and Fe as well as the pH. The influent water was stopped when the concentration of Cr(Ⅵ) and NO3--N in effluent water reached half of that of influent water.

2.3 Analytical method The concentration of Hexavalent chromium and total chromium was measured with the methods of dipehenylcarbohydrazide spectrophotometry and potassium permanganate oxidation-dipehenylcarbohydrazide spectrophotometry respectively. The concentration of NO3--N, NO2--N and NH4+-N was measured by the way of ultraviolet spectrophotometry, N (1 naphthyl) ethylenediamine spectrophotometric method and Nessler's Reagent spectrophotometry respectively. The concentration of Fe was measured by the mean of o-Phenanthroline spectrophotometric method.

3 RESULTS AND DISCUSSION 3.1 Remediation of groundwater contaminated by NO3−-N with Fe0-PRB At the initial stage, the concentration of NO3--N in effluent was about 10mg/L in effluent. During the process of reaction, the concentration of NO3--N decreased to about 1mg/L (the lowest was 0.32mg/L).The removal efficiency was 95% and water quality of effluent was relatively stable, which indicate that Fe0 can remove nitrate nitrogen visibly. When volume of effluent reached up to 425 L, the concentration of NO3--N kept increasing until reaching the concentration of influent. It stated that Fe0 consumed gradually and lost activity. The main removal mechanism of nitrate by Fe0 is redox. In reaction conditions, NO3- can be deoxidized into NO2- 、NH4+ and N2 by Fe0. The conversion equations that may exist can be expressed as follows (Haugen et al. 2002; Hao et al. 2005; Tai et al. 2009; Rodrí guez-Maroto et al. 2009).

Fe  2H   Fe2  H 2   3



Fe  NO  2H  Fe  3

2



(1)  2

 NO  H 2 O

5Fe  2 NO  12H  5Fe

2

 N 2  6H 2 O

4Fe  NO3  10H   4Fe 2  NH 4  3H 2 O  2



2

(2) (3) (4)

 4

(5) 3Fe  NO  8H  3Fe  NH  2H 2O + It can be seen from the above equations that NO3 is deoxidized into NO2 first, and then deoxidized into NH4 and other nitrogen-containing gases. The reduction product NO2- exists as transition reduction product of NO3-, which can be deoxidized by Fe0 and its corrosion products. Therefore, the concentration of NO2--N is low in sampling solution and NH4+ is the main product (Westerhoff et al. 2003; Yong et al. 2005). Reduction product, NO2- and NH4+, both are contaminants. NO2- does harm to humans body and has highly ecological risk. NH4+ can be absorbed easily by soils with negative charge. Comparing with nitrate, migration capacity of NH4+ is relatively lower; meanwhile, NH4+ can be removed easily. NH4+ retained in the soils can be absorbed by plants as nitrogenous fertilizer.

The concentration of NO2--N in effluent varied in the interval of nd-3.33mg/L. At the initial stage, Fe0 was sufficient and the concentration of NO2--N was relatively low; NO2- could be further deoxidized. Fe0 was consumed gradually with the reaction, and the hydraulic retention time didn’t change much. Because there was competitive adsorption leaded by NO3- and NO2-, the concentration of NO2--N increased. The concentration of NH4+-N was at a low value, which varied in the interval of nd-1.00mg/L. It may be due to soil adsorption in the process of NH4+ downward migration. During the experimental time of 150d, denitrifying bacteria was domesticated and cultivated, which took NH4+ as electron donor and NO3- as electron acceptor. The concentration of NH4+-N was low, because the denitrifying bacteria consumed NH4+ and NO3- and generated N2( 5NH 4  3NO3  4 N2  9H 2O  2H  )(Kuenen et al.

1994; Mulder et al. 1995).

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3.2 Remediation of groundwater contaminated by chromium with Fe0-PRB Cr(Ⅵ) can be removed by Fe0 efficiently, and its concentration maintained at a relatively low level, which was lower than 0.40mg/L. When the volume of effluent reached to 300 L, the Cr(Ⅵ) concentration in effluent increased and exceeded 1mg/L gradually, finally, reached the concentration of Cr(Ⅵ) in influent. This indicates that Fe0 became invalid. The difference between the concentration of Cr(Ⅵ) and Cr(Ⅲ) was small, which means the Cr(Ⅲ) concentration in effluent was low. The main removal mechanism of Cr(Ⅵ) by Fe0 was oxidation reduction and coprecipitation. Cr(Ⅵ) was deoxidize to Cr(Ⅲ) , which is less toxic, following the process experienced as follows (Powell et al. 1995; Blowes et al. 1999).

Fe 0  CrO4

2

 8H   Fe 3  Cr 3  4H 2 O

2Fe 0  Cr2 O7

2

 14H   2Fe 3  2Cr 3  7 H 2 O

(1  x) Fe

3

 xCr

3

 3H 2 O  Crx Fe1 x (OH ) 3  3H

(6) (7) 

(8)

According to above equations, oxidation product Fe3+ of Fe0 and the reduction product Cr(Ⅲ) of Cr(Ⅵ) can precipitate separately or co-precipitate. There was no Fe3+ and Cr( Ⅲ ) detected in downstream, because the precipitations were attached to the reaction medium. Corrosion-cell reaction （also called micro-electrolysis） happened to activate carbon and iron in the reaction medium (In the action, the low-potential Fe0 acted as the anode, the high-potential of C acted as the cathode, and simulation sewage acted as the electrolyte) (Wang et al. 2006; Fan et al. 2009). Oxidation-reduction reaction occurred immediately between Cr(Ⅵ) and reaction product Fe2+, which was faster than the direct reduction reaction between Fe0 and Cr(Ⅵ). Therefore, the reaction capacity of ferrous powder was enhanced. The surface of activate carbon was adhered by a large number of acidic groups or basic groups, such as carbonyl and phenolic hydroxyl, which made the activate carbon not only acquired adsorptive capacity, but also acted as catalyst for oxidation-reduction reaction.

10 days after stopping simulative contaminated groundwater injection, deionized water was used to wash the simulation device. When the volume of deionized water was 0.6 L, the concentration of TCr in effluent was 2.20 mg/L; when the volume of deionized water was 1.2 L, no TCr was detected; when the experiment finished, the volume of deionized water was 20 L and there was still no TCr detected. It indicated that Cr(Ⅲ), what Cr(Ⅵ) was deoxidized to be by Fe0, precipitated as Cr(OH)3, so there was no Cr(Ⅵ) washed away as the water flowing to the downstream. This ensures that the Cr(Ⅵ) can be removed safely by reduction precipitation methods rather than runs off to the downstream after species changing (Pratt et al. 1997). 928.6 milligram Cr(Ⅵ), obtained theoretically by calculation ( Fe0  Cr 6  Fe3  Cr 3 ), can be reduced by 1 grams zero-valent iron. On the basis of the computational formula and meaning of iron-to-chromium ratio, it is showed that when the concentration of Cr(Ⅵ) in effluent was 1.00mg/L, only 0.62 milligram Cr(Ⅵ) were reduced by 1 grams zero-valent iron, this means scrap iron was used inefficiently in accordance with the theoretically calculated iron-to-chromium ratio value. Further studies are needed to improve the iron utilization efficiency.

3.3 Remediation of groundwater contaminated by nitrate and chromium with Fe0-PRB 20 Cr(Ⅵ) TCr

3

Concentration (mg/L)

Concentration (mg/L)

4

2 1 0

0

100

200

300

400

15 10 5 0

0

100

Volume of effluent (L)

200 Volume of effluent (L)

300

400

FIG. 2 HEXAVALENT CHROMIUM, TOTAL CHROMIUM

FIG. 3 NO3--N MASS CONCENTRATION CURVE OF

MASS CONCENTRATION CURVE OF EFFLUENT IN SIMULATED EXPERIMENTATION OF NITRATE AND CHROMIUM CONTAMINATION

EFFLUENT IN SIMULATED EXPERIMENTATION OF NITRATE AND CHROMIUM CONTAMINATION

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Fig.2 and Fig.3 show the concentration variation of Cr(Ⅵ), TCr and NO3--N when groundwater contaminated by nitrate and chromium compound. It can be seen from Fig.3 that there is not much difference on removal efficiency

by Fe0-PRB in the case of chromium removal. The high removal efficiency of Cr(Ⅵ) by PRB emerged at the beginning of the reaction. The concentration of chromium was relatively low in the initial stage, and the concentration of TCr was equal to that of Cr(Ⅵ) on the whole, which indicates that Cr(Ⅵ) became stable in the form of precipitation after it was deoxidized to be Cr(Ⅲ).

It can be seen from Fig.3 that when the volume of effluent is same, the concentration of NO3--N in effluent is relatively high due to the existence of Cr( Ⅵ ). Meanwhile, when the volume of effluent reached 300 L, the concentration of NO3--N increased always until reached to the level in the influent. However, if only dealing with nitrate, the concentration of NO3--N began to increase when the volume of effluent reached up to 425 L. It indicates that the simultaneous removal of nitrate nitrogen and Cr(Ⅵ) by Fe0 from contaminated groundwater made no difference on Cr(Ⅵ) removal, but made the removal efficiency and removal rate of NO3--N decrease comparing with the efficiency of only removing nitrate nitrogen from contaminated groundwater. This maybe because that Cr(Ⅵ) did not catalyze the reaction of NO3- removal by Fe0, but competed activity space on zero-valent iron surface with NO3-. So the removal efficiency of nitrate nitrogen decreased. On the contrary, NO3- has no effect on removing Cr(Ⅵ) by Fe0. According to the analyzing of iron-to-chromium ratio, it shows that the coexistence of nitrate nitrogen and chromium obtained the effective use of Fe0.

During the experiment process, the pH value in effluent was generally increasing. When NO3--N was processed, the value rose from 7.1 to about 9.0, and the effluent was alkaline. When Cr(Ⅵ) was experimented, the influent pH was about 7.0. Due to the consumption of H+ after a period of time, the effluent pH value was increasing gradually.

Meanwhile, the precipitation of Fe3+ and Cr(Ⅲ) would consume some OH-, the pH value in Cr(Ⅵ) oxidationreduction system increased smaller than that in NO3- oxidation-reduction system. Therefore, the pH value of effluent stayed around 8.0. Also, the pH value of the effluent maintained between 7.5 to 8.5 under the co-existence of NO3--N and Cr(Ⅵ).

The concentration of Fe maintained a relative low level in the effluent of the experiment. When NO3--N was processed, the peak concentration of Fe in the effluent was 0.58mg/L. During the process, oxidation products Fe2+ and Fe3+ can be precipitated in the forms of Fe(OH)2, FeCO3, Fe(OH)3 and Fe2(CO3)3. And the precipitations are porous surface substance, which can enhance their adsorption capacity to Fe2+ and Fe3+, so as to reduce the secondary pollution of Fe. In addition, when Cr(Ⅵ) was experimented, the concentration of total iron was lower than 0.30 mg/L and was lower than that in NO3- oxidation-reduction system. It means that Fe3+ was the main reduction product of Fe0. The vast majority of Fe3+ precipitated in the form of Fe(OH)3 or co-precipitated with Cr. And the precipitation retained in the reaction medium, therefore the Fe3+ was almost never found in downstream.

The concentration of total iron in effluent was lower than 0.30 mg/L. Therefore, the corrosion of scrap iron does not bring significant pollutants, and scrap iron is the appropriate reaction medium material (Cantrell et al. 1995).

4 CONCLUSIONS (1) For the remediation of groundwater contaminated by nitrate with Fe0-PRB, the main reduction products were NO2--N, N2 and NH4+, among which, NO2- was a transition state and NH4+ was the main product. For the remediation of groundwater contaminated by Cr(Ⅵ) with Fe0-PRB, the main removal mechanisms of Cr(Ⅵ) were oxidoreduction and coprecipitation. The Cr(Ⅵ) was deoxidized to Cr(Ⅲ) by Fe0. There were precipitations of chromium or iron or iron-chromium co-precipitation.

(2) For simultaneous removal of nitrate nitrogen and hexavalent chromium from contaminated groundwater, the coexistence of nitrate nitrogen did not affect the removal of hexavalent chromium. Due to competition of the chromium nitrate nitrogen for the active site on the surface of the zero-valent iron, the removal efficiency and removal rate in this experiment were lower than that in the case that only nitrate nitrogen was dealt with.

ACKNOWLEDGMENT - 43 http://www.ivypub.org/fes

This work was supported by Chinese Research Academy of Environmental Sciences special fund for basic scientific research business of central public research institutes (2013-YKY-006).

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AUTHORS Fansheng Meng (1979-),

male, Ph.D.

Chief field of research is water chemistry and environmental risk control.

Yeyao Wang (1965-), male, Ph.D. Email: yeyaowang@163.com Liping Bai (1979-), male, Ph.D. Email: bcrlp@163.com

Email: mengfs@craes.org.cn

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