ecotoxicology and environmental safety · a ranjan plant physiology and biochemistry laboratory,...

Ecotoxicology and Environmental Safety 112 (2015) 247–270

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Ecotoxicology and Environmental Safety

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Review

Arsenic contamination, consequences and remediation techniques:A review

Rachana Singh a, Samiksha Singh b, Parul Parihar a, Vijay Pratap Singh c,n,Sheo Mohan Prasad a,n

a Ranjan Plant Physiology and Biochemistry Laboratory, Department of Botany, University of Allahabad, Allahabad 211002, Indiab Department of Environmental Science, University of Lucknow, Lucknow 226025, Indiac Govt. Ramanuj Pratap Singhdev Post Graduate College, Baikunthpur, Korea 497335, Chhattisgarh, India

a r t i c l e i n f o

Article history:Received 12 August 2014Received in revised form6 October 2014Accepted 6 October 2014Available online 26 November 2014

Keywords:Arsenic contaminationArsenic sourcesHealth hazardsRemediation techniques

x.doi.org/10.1016/j.ecoenv.2014.10.00913/& 2014 Elsevier Inc. All rights reserved.

esponding authors.ail addresses: [email protected] (V.P.

a b s t r a c t

The exposure to low or high concentrations of arsenic (As), either due to the direct consumption of Ascontaminated drinking water, or indirectly through daily intake of As contaminated food may be fatal tothe human health. Arsenic contamination in drinking water threatens more than 150 millions peoples allover the world. Around 110 millions of those peoples live in 10 countries in South and South-East Asia:Bangladesh, Cambodia, China, India, Laos, Myanmar, Nepal, Pakistan, Taiwan and Vietnam. Therefore,treatment of As contaminated water and soil could be the only effective option to minimize the healthhazard. Therefore, keeping in view the above facts, an attempt has been made in this paper to review Ascontamination, its effect on human health and various conventional and advance technologies which arebeing used for the removal of As from soil and water.

& 2014 Elsevier Inc. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2482. Sources of arsenic in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 248

2.1. Groundwater/drinking water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2482.2. Freshwaters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2492.3. Marine waters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2502.4. Arsenic concentration in soil. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2502.5. Arsenic concentration in food stuffs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 250

3. Health hazards. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2514. Remediation of arsenic contamination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 252

4.1. Arsenic revomal by oxidation techniques. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 253
4.1.1. Oxidation and filtration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2534.1.2. Photochemical oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2534.1.3. Photocatalytic oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2544.1.4. Biological oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2544.1.5. In-situ oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 254
4.2. Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2554.3. Coagulation–flocculation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2554.4. Electrocoagulation (EC) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2564.5. Electro-chemical arsenic remediation (ECAR) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2564.6. Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 256

4.6.1. Activated alumina . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2564.6.2. Iron based sorbents (IBS). . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2574.6.3. Zero valent iron . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 258

Singh), [email protected] (S.M. Prasad).

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4.6.4. Indigenous filters and cartridges. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2584.6.5. Miscellaneous adsorbents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 259

4.7. Ion exchange . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2604.8. Electrokinetics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2604.9. Membrane technology. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 261

4.9.1. As removal using microfiltration. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.9.2. As removal using ultrafiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.9.3. As removal using Nanofiltrations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.9.4. As removal using reverse osmosis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 262

4.10. As removal by advanced hybrid and integrated technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 262
4.10.1. As removal using membrane distillation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2624.10.2. As removal using forward osmosis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 262
4.11. Disposal of As laden sludges and wastes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2635. Conclusion and future perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 263Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 263References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 263

1. Introduction

Recently, the environmental fate and behavior of arsenic (As) isreceiving increased attention due to the arsenic (As) pollution inSouth-East Asia. Although As contamination in the environmenthas been reported worldwide (Sohel et al., 2009; Li et al., 2011),however, As pollution in groundwater has been a serious healththreat to the human beings in South-East, South-West and North-East USA, inner Mongolia (China), South-West Taiwan coastal re-gions, Sonora (Mexico), Pamplonian Plain (Argentina), West Ben-gal (India), Northern Chile, and Bangladesh (Argos et al., 2010). TheWorld Health Organization (WHO) deemed the As in Bangladeshigroundwater to be “the largest mass poisoning of a population inhistory” (Argos et al., 2010).

Arsenic is ubiquitous in the environment and highly toxic to allforms of the life. It is a crystalline “metalloid”, a natural elementwith features intermediate between metals and non-metals, oc-curs naturally as an element, ranks as the 20th most occurringtrace element in the earth's crust, 14th in seawater, and 12th in thehuman body (Mandal and Suzuki, 2002). Arsenic exists mainly infour oxidation states – arsenate (AsV), arsenite (AsIII), arsenic (As0),and arsine (As� III) and its solubility depends on the pH and ionicenvironment. Among them, the AsV being the most stable form(Sharma and Sohn, 2009; Zhao et al., 2010; Gupta et al., 2011). AsV

is thermodynamically stable state in aerobic water, while AsIII ispredominant in reduced redox environment. Arsenic can be pre-sent in the environment in various chemical forms such asmonomethylarsonic acid [MMA; CH3AsO(OH)2], dimethylarsinicacid [DMA; (CH3)2AsOOH], trimethylarsine oxide [TMAO;(CH3)3AsO], arsenobetaine [AsB; (CH3)3AsþCH2COOH], arseno-choline [AsC], arsenosugars [AsS], arsenolipids etc. (Tangahu et al.,2011). In general, inorganic arsenicals are more toxic than organicones (Meharg and Hartley-Whitaker, 2002). AsIII is usually moretoxic than AsV (Abedin et al., 2002a, 2002b; Schat et al., 2002), anddimethylarsinous acid (DMAAIII) and monomethylarsonous acid(MMAAIII) are more toxic than their parent compounds (Petricket al., 2000; Mass et al., 2001). Methylated As compounds, such asMMA, DMA and TMAO are found sometimes as a minor compo-nent in the soil (Huang and Matzner, 2006), but can reach highconcentrations (Abedin et al., 2002a, 2002b). Both MMA and DMA(also known as cacodylic acid) have been widely used as pesticidesand herbicides, the DMA also as a cotton defoliant. Arsenobetaine,the dominant As species in marine animals, was found to bepresent in an acidic fen soil with unclear origin (Huang andMatzner, 2006). Arsenolipid, a lipid-soluble As compound, main-ly found in the marine organism, and its concentration may reachupto 16 mg As/kg fish oil (Sele et al., 2012). Recent findings

suggested the following order in terms of acute As toxicity: MMA(III)4As(III)4As(V)4DMA(V)4MMA(V), where the MMA(III)metabolite is the most toxic compound and some researchersconsidered it to be the central As mode of action (EFSA, 2009; Kileet al., 2011; Wen et al., 2011). In this review, we have summarizedAs contamination and its remediation techniques in water and soil.

2. Sources of arsenic in the environment

The primary source of As in the environment (hydrosphere,pedosphere, biosphere and atmosphere) is the release of As fromAs-enriched minerals. The sources of As includes both natural i.e.through dissolution of As compounds adsorbed onto pyrite oresinto the water by geochemical factors and anthropogenic i.e.through use of insecticides, herbicides and phosphate fertilizers,semi-conductor industries, mining and smelting, industrial pro-cesses, coal combustion, timber preservatives etc. (Mondal et al.,2006; Bundschuh et al., 2011). A survey of occurrence of As ingroundwater/drinking water, fresh waters, marine waters, soil andfood stuffs is given below.

2.1. Groundwater/drinking water

According to the WHO guidelines, the recommended limit ofarsenic in drinking water is 0.01 mg L�1. However, the levels of Asin unpolluted surface water and groundwater vary typically from1–10 μg L�1. Groundwater concentrations of As is reported to bevery large range from less than 0.5–5000 mg L�1 covering naturalAs contamination found in more than 70 countries (Ravenscroftet al., 2009). The As contamination in groundwater in differentparts of the world is summarized in Table 1. Large areas of Ban-gladesh, West Bengal and other states of India and Vietnam rely onAs contaminated groundwater for irrigation of staple crops such asrice (Nickson et al., 1998; Berg et al., 2001; Abedin et al., 2002a,2002b). On applying the WHO provisional guideline for drinkingwater of 10 μg L�1 of As, a worldwide population of more than 100millions people are at risk, and out of these more than 45 millionspeople mainly in developing countries from Asia are at risk ofbeing exposed to more than 50 μg L�1 of As, which is the max-imum concentration limit in drinking water in most of the coun-tries in Asia (Ravenscroft et al., 2009). Contamination of drinkingwater is the main source of As for human being but for the po-pulation not exposed to elevated As in drinking water, consump-tion of food grown in As-contaminated soil or irrigated with As-contaminated water represents the main sources of As intake forhumans, which causes a life-threatening problem for millions of

Table 1Concentrations of arsenic in groundwater of the arsenic-affected countries.

Location (Countries) Concentration in μg L�1 Arsenic sources References

Argentina, Pampa, Cordoba 100–3810 2–15 m, 61°45′–63°W; 32°20′–35°00′S Nicolli et al., 1989Argentina, Cordoba 4100 Astolfi et al., 1981Bangladesh o10–41000 Well waters Dhar et al., 1997Calcutta, India o50–23,080 Arsenic-rich sediments Mandal et al., 1996Chile 470–770 United Nations, 2001Fukuoka, Japan 0.001–0.293 Natural origin Kondo et al., 1999Hanoi, Vietnam 1–3050 Arsenic-rich sediments Berg et al., 2001Hungary 1–174 Deep groundwater Sancha and Castro, 2001Inner Mongolia, China 1–2400 Drinking water; bores Guo et al., 2001Lagunera region, Mexico 8–624 Well waters Razo et al., 1990Mekong River floodplain, Cambodia 1–1340 Groundwater Buschmann et al., 2007Nakhon Si Thammarat Province, Thailand 1.25–5114 Shallow (alluvial) groundwater, mining Williams et al., 1996Nepal 8–2660 Drinking water Shrestha et al., 2003Northeastern Ohio o1–100 Natural origin Matisoff et al., 1982Peru 500 Drinking water Sancha and Castro, 2001Romania 1–176 Drinking water bores Gurzau and Gurzau, 2001Ronpibool, Thailand 1–5000 Water contaminated by tin mining waste Choprapwon and Porapakkham, 2001Shanxi, PR China 0.03–1.41 Well water Yinlong, 2001South-west Finland 17–980 Well waters; natural origin Kurttio et al., 1998West Bengal, India 3–3700 Arsenic-rich sediments Mandal et al., 1996Western USA 1–48,000 Drinking water Welch et al., 1988Xinjiang, PR China 0.05–850 Well water Yinlong, 2001

R. Singh et al. / Ecotoxicology and Environmental Safety 112 (2015) 247–270 249

people in large areas of South-East Asia. For example, recentlyinorganic As exposure drawn attention through food, since somefood items especially rice and vegetables were reported to containhigh inorganic As concentrations in areas with elevated As in soiland irrigation water (EFSA, 2009; Mondal et al., 2010; Zhao et al.,2010; Fu et al., 2011; Rahman and Hasegawa, 2011a, 2011b; WHO,2011; and Bhattacharya et al., 2012).

2.2. Freshwaters

In freshwater systems (rivers and lakes), the variation in Asconcentration is in the range of 0.15–0.45 μg L�1 (Bissen andFrimmel, 2003a, 2003b) depending on the source, availability,

Table 2Arsenic concentrations in some major aquatic systems (rivers, lakes, est

Aquatic systems and location Arsenic concentrations

LakesBiwa Lake, Japan 2.2 (0.6–1.7)Moira Lake, Ontario, Canada 20.4 (22.0–47.0)Mono Lake, California, USA 10,000–20,000Marine and EstuariesBunnefjord, Norway 0.5–1.9Coastal Malaysia 1.0 (0.7–1.8)Coastal Nakaminato, Japan 3.1Deep Pacific and Atlantic 1.0–1.8Krka Estuary, Yugoslavia 0.1–1.8Rhone Estuary, France 2.2 (1.1–3.8)Saanich Inlet, B.C., Canada 1.2–2.5Schelde Estuary, Belgium 1.8–4.9Southern coast, Australia 1.3 (1.1–1.6) (inorganic)Southeast coast, Spain 1.5 (0.5–3.7)Tamar Estuary, UK 2.7–8.8Uranouchi Inlet, Japan 22.0–32.0Vestfjord, Norway 0.7–1.0RiversAshanti, Ghana 284(o2–7900)Cordoba, Argentina 7–114Dordogne, France 0.7Madison and Missouri rivers, USA 44 (19–67), 10–370Mole River, NSW, Australia 110–600 (up to 13900)Owens River, CA, USA 85–153Po River, Italy 1.3Ron Phibun, Thailand 218 (4.8–583)Waikato, New Zealand 32 (28–36)

and geochemistry of the catchments (Smedley and Kinniburgh,2002). Basic As concentrations in the water of various con-taminated rivers range between 0.1 to 2.1 mg L�1 with an averageof 0.8 mg L�1 (Table 2), which might be due to the source of con-tamination, surface recharge, base flow, and the bedrock lithology.The geothermal inputs, evaporation, and groundwater con-tamination are the main cause of high concentrations of As inrivers. For example, in Lao River of Northern Chile the extremelyhigh concentrations of As (upto 21,000 mg L�1) is due to the above-mentioned processes (Cáceres et al., 1992). Mining activity can alsocontribute in the occurrence of high As concentrations in riverwaters. For example, water of Mole River, New South Wales,Australia contains high levels of As (from 110–600 mg L�1 upto

uaries and marine) of the world.

(average/range (mg L�1) References

Hasegawa et al., 2010Azcue and Nriagu, 1995Maest et al., 1992

Abdullah et al., 1995Yusof et al., 1994Ishikawa et al., 1987Cullen and Reimer, 1989Seyler and Martin, 1991Seyler and Martin, 1990Peterson and Carpenter, 1983Andreae and Andreae, 1989Maher, 1985aNavarro et al., 1993Howard et al., 1988Hasegawa, 1996Abdullah et al., 1995

Smedley et al., 1996Lerda and Prosperi, 1996Seyler and Martin, 1990Robinson et al., 1995Ashley and Lottermoser, 1999Wilkie and Hering, 1998Pettine et al., 1997Williams et al., 1996McLaren and Kim, 1995


13,900 mg L�1) due to the mining and processing of arsenopyriteores (Ashley and Lottermoser, 1999). However, according to theWHO guidelines for irrigation purpose the permissible limit ofarsenic in water is 0.10 mg L�1.

As concentrations in lake waters are close to or lower than thatreported for river waters (Table 2). Azcue et al. (1994) and Azcueand Nriagu (1995) studied that the As concentrations in lakesaround British Columbia and Canada ranged between 0.2 and2.08 mg L�1, has been transported from the abandoned CaribooGold Quartz mine tailings of that area, and gets accumulated inhigh concentration (upto 1104 mg g�1) in the bottom sediments ofthe lakes. Geothermal sources and mining activities have also in-creased the concentrations of As in lake waters (Smedley andKinniburgh, 2002). In mine affected lake waters, the As con-centrations are relatively low because of its adsorption onto Fe-oxides under neutral pH (Smedley and Kinniburgh, 2002), and alsodue to its accumulation in bottom sediments (Azcue and Nriagu,1995).

Several studies have reported thermal stratification of As con-centrations in lake waters (Azcue and Nriagu, 1995; Hasegawa,1996; Hasegawa et al., 2010). Azcue and Nriagu (1995) studied thatthe dissolved As concentration was highest during summer in theMoira Lake, Ontario, Canada with an average concentration of47 mg L�1 in surface water, compared to 22 mg L�1 in winter.Hasegawa et al. (2009) have also reported similar trends in theoccurrence of As concentrations in lake waters. Smedley andKinniburgh (2002) and Hasegawa et al. (2010) reported thatthermal stratification in lake water also causes the release of Asinto the water column from bottom sediments due to depletion ofO2 levels in the hypolimnion (due to increased biological activities)and its subsequent redistribution throughout the lake.

Recently, the major As affected regions are found in large deltasand along major rivers emerging from the Himalayas with theBengal delta being the worst affected area where 488% of the 45millions inhabitants are at high risk of exposure to As concentra-tions 450 mg L�1 (Acharyya and Shah, 2007; Ravenscroft et al.,2009; Uddin et al., 2011). The other affected river deltas and riverbasins in South and South-East Asia are the Red River Delta andthe Mekong Delta (410 millions exposed) (Berg et al., 2007;Buschmann et al., 2008) and river basins of Chindwin-Irrawady,Salween; Brahmaputra, Ganges, Indus, Chenab (Chakraborti et al.,2003; Nickson et al., 2005, 2007; Stanger, 2005; Thakur et al.,2011).

2.3. Marine waters

In seawater, the As concentration is usually less than 2 μg L�1

(Ng, 2005), and its concentrations in Atlantic and deep Pacificwaters are between 1.0–1.8 mg L�1 (Cullen and Reimer, 1989),3.1 μg L�1 in marine waters of the Pacific coast near Nakaminato(Ibaraki, Japan) (Ishikawa et al., 1987), and 1.1–1.6 mg L�1 in coastal

Table 3Concentrations of arsenic in soil of the arsenic-affected countries.

Country Region Soil As concentration in mg/

Bangladesh Noakhali 3.6–26 mg/kg (Meghna RiveBrazil Minas Gerais (Southeastern Brazil) 200–860 mg/kgChile Esquiña Up to 489 mg/kg (Río CaritaIndia Uttar Pradesh 16–417 mg/kg (Central IndiaMexico Lagunera 2215–2675 mg/g (Highly poPoland Lower Silesia, (Southwestern Poland) Up to 18,100 mg/kg (HighlySpain Duero Cenozoic Basin 23 mg/kg (Mean)Turkey Simav plain (Kutahya) Up to 660 mg/kg (Highly poUnited Kingdom Cornwall 2–17 mg/kg (Bioaccessible)USA Tulare lake average 280 mg/kg (Hawaii)

waters of southern Australia (Maher, 1985) (Table 2). Arsenicconcentrations in estuarine waters are more uniform than those ofopen marine waters. Smedley and Kinniburgh (2002) reportedthat As concentrations in the estuarine waters may be affected byindustrial and mining effluents and geothermal water. The physi-cal mixing of the fresh and seawater masses and salinity may in-fluence the concentration of dissolved As in estuaries and con-tinental shelves. For example, a linear increase in total As con-centrations, ranging from 0.13 mg L�1 in freshwaters to 1.8 mg L�1

in offshore waters, with increase in the salinity has been reportedin Krka Estuary, Yugoslavia (Seyler and Martin, 1991).

2.4. Arsenic concentration in soil

According to the U.S. Environmental Protection Agency, thepermissible limit of arsenic in soil is 24 mg kg�1. In the case of soil,there are also numerous pathways for propagating the con-tamination of As. The major sources of its contamination in soil areidentified to include many man-made activities e.g. the use ofinsecticides, herbicides and phosphate fertilizers, semi-conductorindustries, mining and smelting, industrial processes, coal com-bustion, timber preservatives etc. (Mondal et al., 2006; Bundschuhet al., 2011). Average arsenic concentration in European topsoil isestimated at 7.0 mg kg�1 (Stafilov et al., 2010) but the backgroundconcentration can significantly differ depending on soil conditions.Arsenic contamination in soil in different parts of the world issummarized in Table 3. In lower Silesia, Southwestern Poland upto18,100 mg kg-1 of As was reported in soil of Au-enriched me-tallogenic zones (Karczewska et al., 2007).

2.5. Arsenic concentration in food stuffs

The interest in rice as a potential source of exposure to arsenicis very recent. Rice is a staple food for more than half the world'spopulation beacuse it is a good source of carbohydrates, thiamin,vitamin B6, and some essential elements like magnesium, zinc andcopper. The world's total production of rice in 2009 was estimatedto be 682 million metric tons (FAO, 2010). Interestingly, many ofthe rice-producing countries suffer from arsenic contamination intheir groundwater or soil (Rahman and Hasegawa, 2011a, 2011b).However, rice may accumulate hazardous levels of toxic elementssuch as arsenic. Due to its large daily consumption, it accumulatesin human body and poses serious threat (Meharg et al., 2008;Shraim, 2014). Besides rice, other cereals such as wheat, corn, oat,buckwheat are also source of arsenic exposure. Vegetables andmeat products have also been reported a good source of arsenicexposure to human beings (EFSA, 2014). The amount of arsenicingested daily by humans via food is greatly influenced by theamount of food in the diet. The amount of arsenic in various foodstuffs and its safe limit is summarized in Table 4.

kg References

r) Nriagu et al., 2007Bundschuh et al., 2012

ya region) Bundschuh et al., 2012; Nriagu et al., 2007) 5.40–15.43 ppm (Uttar Pradesh) Das et al., 2013; Srivastava and Sharma, 2013lluted area) Nriagu et al., 2007polluted area) Karczewska et al., 2007

Gómez et al., 2006lluted area) Gunduz et al., 2010

Palumbo-Roe et al., 2005Nriagu et al., 2007

Table 4Arsenic concentrations in various food stuffs and its recommended limit.

Food stuff(s) Arsenic concentra-tion (mg As kg�1)

Recommended limit ofarsenic in food (FAO)

Reference

Rice 153.1 1 mg kg�1 EFSA, 2014Wheat 22.0 do EFSA, 2014Oat 27.3 do EFSA, 2014Corn 49.3 do EFSA, 2014Vegetables 90–3900 do Das et al.,

2004Pulses 1300 do Santra et al.,

2013Chicken meat 286 do Islam et al.,

2013Fish 3000 do Lin and Liao,

2008


In poultry, use of arsenic compounds is very common. For in-stance, arsenic is used in the forms of roxarsone as an additive inthe feed of conventionally-raised broilers. It is used to controlprotozoan parasites known as coccidians and to enhance weightgain (Miller et al., 2000) Feeding arsenic to laying hens is pro-hibited. Organic regulations prohibit feeding arsenic to birdsraised for organic certification. It is estimated that roxarsone isadded to poultry feed at the rate of 22.7–45.4 g per ton, or 0.0025–0.005 percent (Miller et al., 2000; Bellows, 2005). Most of theroxarsone passes through the birds and is excreted unchanged(Kpomblekou et al., 2002). Each broiler excretes about 150 mg ofroxasone during the 42-day growth period in which it is ad-ministered (Sims and Wolf, 1994). Moreover, Rosal et al. (2005)have also investigated fate of roxarsone and its possible transfor-mation products (arsenite, arsenate, monomethylarsonate, di-methylarsinate, 3-amino-4-hydroxyphenylarsonic acid, and 4-hy-droxyphenylarsonic acid) in chicken manure and found that thesecompounds are responsible for arsenic contamination of chicken.

Large amount of pig manure is produced all over the worldwhich can be used as organic fertilizers on agricultural lands.

Fig. 1. Schematic diagram is showing transfer of arsenic from soil and water to human be

Besides, the organic arsenic compounds have been used as feedadditives for swine disease control and weight improvement (Liand Chen, 2005). Once the excessive additives are released in theenvironment, arsenic may compromise food safety and environ-mental quality as it contains arsenic in range of 0.42–119.0 mg kg�1. Therefore, there is a growing public concern aboutthe arsenic residues accumulation in pig manure. Silbergeld andNachman (2008) also observed arsenicals in pig manure that maylikely increase the burden of global human arsenic exposure andrisk.

3. Health hazards

Numerous studies have been conducted to assess the toxicity ofAs and its effects on human health in various As-contaminatedregions (Kongkea et al., 2010; Maity et al., 2012). Arsenic enters inhuman beings through two pathways; first, direct consumption ofAs contaminated drinking water and second, for populations notexposed to elevated As in drinking water, foods represent the mainsources of As intake for humans (Fig. 1). Arsenic accumulation invegetables followed by ingestion may result in a significant con-tribution on to the daily human intake of inorganic As(Fontcuberta et al., 2011). Arsenic contamination in drinking waterthreatens health risk for more than 150 millions people all overthe world (Ravenscroft et al., 2009). Around 110 millions of thosepeople live in 10 countries in South and South-east Asia: Bangla-desh, Cambodia, China, India, Laos, Myanmar, Nepal, Pakistan,Taiwan and Vietnam (Brammer, 2008). In total 88,750 km² in WestBengal has been identified as As contaminated zone among which38,861 km² area has been identified as highly affected zones, thisinclude Nadia, North and South 24 Parganas, Murshidabad andKolkata districts (Chakraborti et al., 2009). Groundwater is used forthe irrigation to cultivate a variety of crops and vegetables, andthus irrigation with As-enriched groundwater is the main pathwayfor As to enter the human food chain (Das et al., 2004; Chatterjeeet al., 2010; Samal et al., 2011) which is the potential human health

ings through food chains. Intake of arsenic by human beings causes several diseases.


risk have recently received attention, especially for rice (Zhaoet al., 2010). The accumulation of As is also important in wetlandspecies, where there is a potential for reducing conditions to re-lease As into the soil solution (Ha et al., 2009; Rahman andHasegawa, 2011a, 2011b), or for co-deposition of As with Fe hy-droxides adsorbed to the plant's surface (Zhao et al., 2002). InWest Bengal (India) and Bangladesh, most of the crop fields arecontaminated with As (Roychowdhury et al., 2002; Alam et al.,2003).

The various effects of As poisoning on human health are mel-anosis, leuco-melanosis, keratosis, hyperkeratosis, dorsum, non-petting edema, gangrene and skin cancer (Fig. 1). Melanosis andkeratosis are the most common presentations among the affectedpeoples (Karim, 1999). Patients of leuco-melanosis and hyperker-atosis have been found in many cases. Few cases of skin cancerhave also been identified among the patients seriously affected byarsenite and arsenate (Hsueh et al., 1995). Chronic exposure toinorganic As causes several disorders upon different biologicalsystems like digestive system, respiratory system, cardiovascularsystem, hematopoietic system, endocrine system, renal system,neurological system, and reproductive system which ultimatelylead to cancer (Mandal et al., 1996; Maharjan et al., 2005).

Excessive and long-term (such as 5–10 years) human intake oftoxic inorganic As having concentration above 0.05 mg L�1 leadsto arsenicosis, which is a common term used for As related healtheffects including skin problems, skin cancers, internal cancers(bladder, kidney, lung), diseases of the blood vessels of the legsand feet, and possibly diabetes, high blood pressure and re-productive disorders (WHO, 2011). The global extent and severityof appearing arsenicosis is probably not yet fully revealed. Over-whelming evidence of non-occupational chronic As exposure isthrough ingestion of drinking water including food and beveragesprepared from drinking water. Inhaled amounts of As may be highand important in poorly ventilated huts where As rich coal or cowdung is used as a fuel (Lin et al., 2010) with high As concentrationshas been reported over the last two decades and is now

Fig. 2. Schematic diagram is showing various techniques used for removal of arse

recognized to be one of the world's greatest environmental ha-zards (Sohel et al., 2009; Chakraborti et al., 2010; Bundschuh et al.,2012; Naujokas et al., 2013). The major arsenicosis regions arepresently found in large deltas and/or along major rivers emergingfrom the Himalayas (Fendorf et al., 2010) such as in the Bengaldelta (Chakraborti et al., 2010), other parts of India (Saha, 2009;Shukla et al., 2010), Nepal (Thakur et al., 2011), Pakistan (Maliket al., 2009), Myanmar, Vietnam, Cambodia (Berg et al., 2007;Polya et al., 2008) and China (He and Charlet, 2013). A recent studyhas shown that nutritional deficiency is an important promoter ofarsenicosis especially for women (Deb et al., 2013).

4. Remediation of arsenic contamination

Arsenic contaminated drinking water is a major threat tomankind. Although in small quantities it is necessary, however, itis known to be highly toxic if ingested in large dose. Its elevatedconcentrations are found in groundwater in some areas of India,Bangladesh, Chile, China, Argentina, Mexico, Hungary, Taiwan,Vietnam, Japan, New Zealand, Germany and the United States dueto naturally occurring arsenic in the aquifer sediment (Bang et al.,2005a). Thus, in order to reduce the health risk arising due to thedirect consumption of As contaminated drinking water or due tothe consumption of food/vegetables, grown in soil irrigated withAs contaminated water, there is need to develop strategies thatcould alleviate toxicity and availability of As from soil to edibleportions of food/vegetables. Removal of As highly depends on thechemistry and composition of the As contaminated water. In mostof the major reported incidences As occurs as AsIII and oxidation ofAsIII to AsV is deemed necessary to achieve the satisfactory resultsof As removal. Here dealing with the presently available technol-ogies, especially oxidation, phytoremediation, coagulation–floc-culation, adsorption, ion exchange, electrokinetics and membranetechnologies will be discussed (Fig. 2).

nic from soil and water. These techniques are discussed in detail in the text.


4.1. Arsenic revomal by oxidation techniques

These techniques use various processes and are discussedbelow.

4.1.1. Oxidation and filtrationThe main purpose of oxidation is to convert the soluble AsIII to

AsV, which is then followed by precipitation of AsV. This is es-sential for anaerobic groundwater because AsIII is the predominantform of As at neutral pH (Masscheleyn et al., 1991). Generally,oxidation and filtration refer to the processes which are designedto remove the naturally occurring iron and manganese fromwater.The process involves the oxidation of the soluble forms of iron andmanganese to their insoluble forms and then removal by filtration.Firstly, soluble AsIII is oxidized. Arsenic is mainly present as AsV

and, as such, is likely to be in the solid phase. Therefore, in suchsoils, As in groundwater used for irrigation is quickly adsorbed byiron hydroxides and becomes largely unavailable to plants. Inanaerobic soil conditions such as in flooded paddy fields, As ismainly present as AsIII and is easily dissolved in the soil-pore water(the soil solution) (Xu et al., 2008). It is thus more readily availableto plant roots. Arsenite has a low affinity to mineral surfaces, whilearsenate adsorbs easily to solid surfaces. Thus, for the removal ofAs from water, oxidation/precipitation technology is very effective(Ghurye et al. 2004; Leupin and Hug, 2005). The oxidation of AsIII

into AsV is carried out by traditional chemical oxidants (Ox) suchas chlorine (Cl2), chlorine dioxide (ClO2), ozone (O3), hydrogenperoxide (H2O2), chloroamine (NH2Cl), permanganate (MnO4

�),and ferrate (FeO4

2�) and have been published in many studies(Emett and Khoe, 2001; Johnston et al., 2001; Bissen and Frimmel,2003b; Lee et al., 2003; Ghurye et al., 2004; Vasudevan et al.,2006; Dodd et al., 2006; Sharma et al., 2007; Mondal et al., 2013).

Following equations expresses the stoichiometries of the oxi-dation reactions:

Cl2: As(OH)3þHOCl-AsO43�þClþ4Hþ (1)

ClO2: As(OH)3þ2ClO2þH2O-AsO43�þ2ClO2þ5Hþ (2)

O3: As(OH)3þO3-AsO43�þO2þ3Hþ (3)

H2O2: As(OH)3þH2O2-AsO43�þ3HþþH2O (4)

NH2Cl: As(OH)3þNH2ClþH2O-AsO43�þNH4

þþCl�þ3Hþ (5)

MnO4�: 3As(OH)3þ2MnO4

--3AsO43-þ2MnO2þ7HþþH2O (6)

FeO42�: 3As(OH)3þ2FeO4

2�þH2O-3AsO43�þ2Fe(OH)3þ

5Hþ (7)

When used for oxidation of AsIII to AsV, the reaction is very fastfor Cl2, O3 and MnO4

� compared to the H2O2 and NH2Cl (Lee et al.,2003; Ghurye et al., 2004; Dodd et al., 2006). By using air and pureoxygen, about 54–57% of AsIII can be oxidized to AsV in con-taminated groundwater (Kim and Nriagu, 2000) whereas, com-plete oxidation of AsIII can be achieved with O3. The NH2Cl andH2O2 are sluggish in reacting with AsIII while Cl2 and O3 react veryrapidly. Free Cl2 or hypochlorite is effective for the oxidation ofAsIII, but chlorination creates and leaves disinfectant by-products(DBPs) in treated water. Reduction in the levels of trihalo me-thanes (THMs) and halo acetic acids (HAAs) was seen with O3, butit can form the potent carcinogenic bromate ion by reacting withbromide present in the water (Gunten, 2003; Richardson, 2006). Itis suggested that treatment with NH2Cl produces N-ni-trosodimethylamine (NDMA) which is a suspected human carci-nogen (Mitch and Sedlak, 2002). For the treatment of high qualitywater such as surface water, the use of ClO2 is restricted anddosing of ClO2 must be kept low. For example, in the United States,dosages ranging from 1 to 1.4 mg L�1 are used mainly for the

preoxidation of surface water (Gates, 1998). Removal of As usingH2O2 and NH2Cl oxidants would take hours as they oxidizes AsIII

very slowly whereas Cl2, O3, and FeO42� would react with AsIII in

millisecond time scale. The scavenger substances present in waterwill affect the fast kinetics of AsIII oxidation with Cl2, O3, andFeO4

2� . However, specific selection of oxidants can reduce theeffect of scavengers on effectiveness of oxidant. For example, inorder to remove As from water that contains excess ammonia, it isbetter to use ozonation rather than chlorination because O3 reactsslowly with ammonia. FeO4

2� can be related to the use of otherchemical oxidants for removing As (Sharma, 2007a). FeO4

2� doesnot react with bromide ion and thus carcinogenic bromate ionwould not be produced in the treatment of bromide-containingwater (Sharma, 2007a). Moreover, FeIII, a by-product of FeVI is non-toxic, and acts as a powerful coagulant (Sharma, 2002, 2004;Sharma et al., 2005a, 2005b; Yngard et al., 2008) which is suitablefor the removal of AsV in water (Lee et al., 2003; Sharma et al.,2007). Thus, FeVI acts as multifunctional chemical oxidant, disin-fectant, and coagulant in a single mixing (Sharma, 2007b). Re-cently, it has been demonstrated that the oxidation of AsIII byMnO2 coated PEEC-WC nanostructured capsules and demon-strated that when water contains a low concentration of As, theyhave a higher efficiency than conventional oxidation methods(Criscuoli et al., 2012).

4.1.2. Photochemical oxidationThe most widely tested chemical oxidant in presence of natu-

rally occurring iron in the field is UV-light assisted oxidation of AsIII

(Ryu et al., 2013). The oxidation rate of AsIII in the water can beincreased by UV irradiation in the presence of oxygen. UV/solarlight helps to generate hydroxyl radicals through the photolysis ofFeIII species: (FeOH2þ) and in the presence of both hydroxyl radicalsand O2, the oxidation rate becomes faster (Yoon and Lee, 2005;Sharma et al., 2007). Several studies have investigated the photo-chemical oxidation of AsIII using UV light irradiation. In perchlorate/perchloric solution at pH 0.5–2.5, addition of FeIII to As-con-taminated water followed by exposure to UV/solar light enhancedthe removal of As (Emett and Khoe, 2001). In this study, FeIII-hy-droxide and chloride species absorb photons to give highly oxidiz-ing hydroxyl and dichloro radicals which converts AsIII to AsV

(Emett and Khoe, 2001). Although, this systemwas also found to beuseful under natural water conditions (Hug et al., 2001). An oxi-dation of AsIII solution containing 0.06–5 mg L�1 FeII and FeIII using90Wm�2 UV-A light removed more than 90% of the 500 μg L�1

total As in 2–3 h. Addition of citrate to this solution strongly ac-celerated the oxidation of iAsIII (inorganic arsenite) (Hug et al.,2001). Instead of UV-light, solar-light can also remove As fromnatural water upon addition of iron and citrate (Lara et al., 2006).Addition of a few drops of lime or lemon juice (citrate) inwater mayalso be helpful for the enhancement of photochemical oxidation ofAsIII to the less harmful AsV (Hug et al., 2001; Kocar and Inskeep,2003; Lara et al., 2006). The cyclic reaction of lemon juice (citrate)with strongly oxidizing radicals gives rise to further radicals due towhich the removal rate was higher but excessive concentration ofcitrate has a negative impact due to the formation of acid com-plexants (Bissen and Frimmel, 2003b). Recently, oxidation of AsIII

was accomplished by using vacuum-UV lamp irradiation at 185 and254 nm wavelengths (Yoon et al., 2008). In this study, the effects ofFeIII, H2O2, and humic acids (HA) were examined. Both FeIII andH2O2 increased oxidation efficiency but humic acid did not showany influence on the oxidation. In order to achieve effective oxi-dation of AsIII, intense UV light source was used with the potassiumperoxydisulphate (KPS) system (Neppolian et al., 2008).


4.1.3. Photocatalytic oxidationThe efficient oxidation of AsIII to AsV can be achieved by pho-

tocatalytic oxidation (PCO) (Bissen et al., 2001). The PCO of AsIII toAsV followed by adsorption of As on TiO2 was investigated (Duttaet al., 2004, 2005; Miller et al., 2011). The PCO of AsIII in suspen-sions with low TiO2 loadings followed by subsequent adsorption ofAsV onto TiO2 surfaces in slightly acidic media reduced As con-centrations below the WHO drinking water limit of 10 μg L�1 inwater (initial [As]¼66.7 μM) (Dutta et al., 2005). Miller andZimmer-man (2010) synthesized a TiO2-impregnated chitosanbead (TICB) and it was used for oxidation as well as removal of Asfrom aqueous solution. The adsorption of AsV onto TiO2 is influ-enced by pH, initial concentration of As, type of TiO2, presence ofanions (e.g. CO3

2� and PO43�) and NOM (natural organic matter)

(Dutta et al., 2004; Bang et al. 2005b; Ferguson and Hering, 2006;Pena et al., 2005, 2006; Liu et al., 2008; Miller et al., 2011). Theyobserved a higher amount of adsorption of As [6400 mg AsIII g�1

TICB and 4925 mg AsV g�1] followed by UV radiation compared tothe solution that was not exposed to UV light [2198 mg AsIII g�1

TICB and 2050 mg AsV g�1]. A nanocrystalline Al2O3 and TiO2 im-pregnated chitosan for As removal was prepared by Yamani et al.(2012) and their study proposed a mechanismwhere AsIII is photo-oxidized to AsV by TiO2, then adsorbed by Al2O3. When a very lowamount of TiO2 is present, the TiO2/UV system has an inefficientremoval of As due to incomplete oxidation of AsIII to AsV (Guanet al., 2012). The effect of competitive anions and organic matterthat are commonly found in the groundwater are an importantdisadvantage of using UV/TiO2. Bicarbonate and humic acid affectthe PCO of AsIII, while the adsorption of As on to the TiO2 basedadsorbent was affected by silicate, fluoride, phosphate and humicacid (Guan et al., 2012). The PCO increases in the presence of NOMbut at higher concentration of NOM (2–15 mg L�1), the adsorptionof AsV on the surface of TiO2 decreases (Dodd et al., 2006; Liu et al.,2008). Sharma and Sohn (2009) recently reviewed the possiblereasons for the decreased adsorption of AsV on TiO2, that are thecompetitions between NOM and AsV for available binding siteson TiO2 surface and/or the adsorption of NOM, which modifiesthe surface charge of TiO2. Recently, it has been demonstrated thatthe PCO of pentavalent MMAV (monomethylarsonic acid) andDMAV (dimethylarsinic acid) using Degussa P25 and nanocrystal-line TiO2 (Xu et al., 2007, 2008). In the use of Degussa P25 TiO2,both MMAV and DMAV were readily mineralized to iAsV (Xu et al.,2007). The MMAV was formed as an intermediate of the PCO ofDMAV, which was subsequently oxidized to iAsV.

4.1.4. Biological oxidationIt is relatively a new method of oxidation of iron and manga-

nese as a treatment method for As removal. Biological treatmentmethods exploit natural biological processes that allow certainplants and micro-organisms to help in the remediation of metalsin soil and groundwater. This process is based upon the fact that Ascontaminated groundwater is usually reducing and containingiron and manganese concentrations. In the treatment system, thefollowing sequence of reactions have been found to occur:(i) oxidation of MnII to MnIV and FeII to FeIII, (ii) oxidation of AsIII toAsV, (iii) precipitation of manganese oxides, (iv) abiotic oxidationof AsIII by manganese oxides, and (v) AsV sorption by manganeseoxides, where steps (i) and (ii) are biotic and steps (iii) to (v) areabiotic. Katsoyiannis and Zouboulis (2004) reported that the mi-croorganisms Gallionella ferruginea and Leptothrix ochracea werefound to support biotic oxidation of iron. They performed someexperiments in the laboratory where iron oxides and above givenmicroorganisms were deposited in the filter medium, offering afavorable environment for the adsorption of As because As in theform of AsIII cannot be efficiently sorbed onto iron oxides. Prob-ably, these microorganisms oxidized AsIII to AsV, which got

adsorbed in FeIII resulting in overall As removal of up to 95% evenat high initial As concentrations of 200 mg L�1. The kinetics ofbacterial AsIII oxidation and subsequent removal of AsV by sorptiononto biogenic manganese oxides during ground water treatmentwas studied by Katsoyiannis et al. (2004). Their findings suggestedthat the rate of oxidation of AsIII was comparatively higher thanthe rates reported for abiotic AsIII oxidation by manganese oxides,supporting that bacteria play an important role in both the oxi-dation of AsIII and the generation of reactive manganese oxidesurfaces for the removal of AsIII and AsV from solution. Thus,bacteria play an important role in both the AsIII oxidation and thegeneration of reactive manganese oxide surfaces for the removal ofdissolved AsIII and AsV. Later, Katsoyiannis and Zouboulis (2006)reviewed the use of iron and manganese oxidizing bacteria for thecombined removal of iron, manganese and As from contaminatedground water. According to report, iron oxidizing bacteria removeAs more efficiently than those of manganese oxidizing bacteria.The rates of oxidation of iron, manganese and As are faster thanthose reported for physicochemical oxidation, indicating the cat-alytic role of bacteria in As removal. Leupin and Hug (2005) passedaerated artificial ground water with high As and iron concentra-tion through a mixture of 1.5 g iron fillings and 3–4 g quartz sandin a vertical glass column. By using dissolved oxygen, FeII wasoxidized to hydrous ferric oxides (HFO) while AsIII was partiallyoxidized and AsV adsorbed on the HFO. This principle was suc-cessfully applied in the field by Sen Gupta et al. (2009), wherewithout using any chemical they reversed the bacterial As reduc-tion process, by recharging calculated volume of aerated water(DO44 mg L�1) in the aquifer to create an oxidized zone. Thisboosted the growth of iron oxidizing bacteria and suppressed thegrowth of As reducing anaerobic bacteria and promoted thegrowth of chemoautotrophic As oxidizing bacteria (CAOs) over aperiod of six to eight weeks. Saalfield and Bostick (2009) de-monstrated a process in the laboratory, where the mobility of Aswas affected by biologically mediated redox processes by bindingit to iron oxide through dissimilatory sulphate reduction andsecondary iron reduction processes, in reducing aquifers. Incuba-tion experiments were conducted using AsIII/V-bearing ferrihydritein carbonate-buffered artificial groundwater enriched with sul-phate (0.08–10 mM) and lactate (10 mM) and inoculated withDesulfovibrio vulgaris (ATCC 7757), which reduces only sulphatebut not Fe or As. The end product formed through sulphidizationof ferrihydrite was magnetite, elemental sulfur and trace Fe sul-phides. It was suggested that only AsIII species got released underreducing conditions and bacterial reduction of AsV was necessaryfor As sequestration in sulphides.

4.1.5. In-situ oxidationThis method is mostly popular for removal of Fe from

groundwater (Appelo et al., 1999). In-situ oxidation can beachieved by pumping the oxygenated water into the groundwateraquifer to reduce the As content in the pumped groundwater. Itspotential for removal of As is investigated in very few studies al-though the results show that As concentrations can be reduced inthe groundwater zone before water extraction (Sen Gupta et al.,2009; van Halem et al., 2010). In-situ oxidation of As and iron inthe aquifer has been tried under Danida Arsenic Mitigation PilotProject (DPHE, 2001). The aerated tube well water is stored in atank and released back into the aquifers through the tube well byopening a valve in a pipe connecting the water tank to the tubewell pipe under the pump head. The dissolved oxygen content inwater oxidizes AsIII to less mobile AsV and also the ferrous iron toferric iron in the aquifer causing the reduction in As content oftube well water.


4.2. Phytoremediation

Phytoremediation is the plant based environmental-friendlytechnology, for the remediation of As contaminated sites, usingplants and microbes to clean up contaminated air, soil and water(Lasat, 2002; Cherian and Oliveira, 2005; Dickinson et al., 2009;Behera, 2014). The Pteris vittata (Chinese brake fern) was found tobe resistant to As, having the capability of hyperaccumulatinglarge amounts of As in its fronds (Ma et al., 2001) by area con-taminants are picked up by the roots of plants and transported totheir overground parts, and then removed together with the crops(phytostabilization, phytoextraction and phytovolatilization). Thebrake fern can accumulate between 1442–7526 mg kg�1 As infronds from contaminated soils, and up to 27,000 mg kg�1 As in itsfronds in hydroponics culture. The As hyperaccumulation capacityhas also been demonstrated in other plants (Meharg, 2003; Duet al., 2005; Keller et al., 2007; Tripathi et al., 2007; Gonzaga Mariaet al., 2008; Zhang et al., 2008). It is hypothesized that hyper-accumulation is associated with the interaction of As with high-affinity chelating molecules present in the cytoplasm of the plant.For example, an arsenate activated glutaredoxin from the fern P.vittata L. regulates intracellular arsenite (Sundaram et al., 2008).The molecular studies have shown that many gene products areinvolved in the process of As hyperaccumulation (Dhankar et al.,2002), hence single gene and multigenic engineering approachesmay be applicable to enhance the efficiency of phytoremediation(Padmavathiamma and Li, 2007; Tripathi et al., 2007). Besidesphytoremediation, phytostabilization methods using plants canalso be applied for long-term remediation of As. This methodlimits uptake and excludes mobilization of As. The major benefit ofphytostabilization is that the vegetative biomass above ground isnot contaminated with As, thus reduces the risk of As transferthrough food chains (Madejon et al., 2002).

Furthermore, the bioremediation techniques, including a vari-ety of sulfate reducing bacteria and other species such asPaenibacillus, Pseudomonas, Haemophilus, Micrococcus, and Bacillusmay be involved to remediate As from contaminated environ-ments (Yamamura et al., 2003, Kirk et al., 2004, Ike et al., 2008).The basic principle of bioremediation is change in the redox re-actions, increasing/decreasing the solubility using different com-plexation reactions, pH changing and adsorption/uptake of asubstance from the environment (Smith et al., 1994). Still, thecurrent bioremediation techniques fail mainly because of thelimitations of phytoremediation in arid area, re-release of im-mobilized or adsorbed heavy metals by some bacteria in the en-vironment, microbial sensitivity to redox potential change andchanges into the valence state of particular toxic metal.

During the last two decades, phytofiltration, a very en-vironmentally friendly and low-cost alternative technique, is apromising emerging alternative to conventional cleanup. After thediscovery of As hyperaccumulating and tolerating plants, it ispossible to phytoremediate the As contaminated substrates. Phy-tofiltration involves several steps, (i) the selection of the mostpromising plants capable of removing the contaminant fromwaterand retaining it in their roots, and (ii) plants are then transplantedinto a constructed wetland, where As from the polluted water willbe removed. These plants mainly absorb and concentrate the As intheir roots, but can also translocate low quantities to their shoots(Dushenkov et al., 1995; Salt et al., 1998). In recent years, onlyplants which are able to hyperaccumulate As were discovered, likeP. vittata (Ma et al., 2001) and other ferns (Francesconi et al., 2002;Zhao et al., 2002; Srivastava et al., 2005). Aquatic macrophyteshave been particularly considered for As removal from con-taminated surface water bodies. Several studies were performedfor the removal of As from contaminated surface water bodiesusing different species of aquatic macrophytes: water hyacinth

(Eichhornia crassipes), lesser duckweed (Lemna minor) (Alvaradoet al., 2008), dried algae (Lessonia nigrescens) (Hansen et al., 2004)and dried macro-algae (Spyrogira spp.) (Bundschuh et al., 2007).

Recently, use of native biomasses (powdered) was reported toremove As from surface water. For example, biomass from thestem of a thorny Acacia nilotica was used for the removal of Asfrom As contaminated water bodies (Baig et al., 2010). Earlier,other biomasses derived from fish scales, coconut fiber, dried rootsof water hyacinth plant, seed powder of Moringa oleifera, Mo-mordica charantia, powdered eggshell, human hair, rice husk, ricepolish; without chemical treatment have been reported(Wasiuddin et al., 2002; Al-Rmalli et al., 2005; Kumari et al., 2006;Nurul-Amin et al., 2006; Oke et al., 2008; Rahaman et al., 2008;Pandey et al., 2009; Ranjan et al., 2009). However, there is still astrong challenge in developing economical and commonly avail-able biosorbents for the As removal.

4.3. Coagulation–flocculation

In As removal processes, coagulation and flocculation areamong the most common method ever employed. The addition ofa coagulant followed by the formation of a floc is a potential wayfor the removal of As from groundwater. Coagulants change thesurface charge properties of solids to allow the agglomeration orenmeshment of particles into a flocculated precipitate. The finalproducts are larger particles or floc, which settle under the influ-ence of gravity or filtered more readily. The destabilization ofcolloids by neutralizing the forces that keep them apart, is thepurpose of coagulation. Positively charged cationic coagulantsprovide positive electric charges to reduce the negative charge(zeta potential) of the colloids and as a result, larger particles areformed due to the aggregation of particles (Choong et al., 2007).Flocculation is the action of polymers to form the bridges betweenthe larger mass particles or flocs and bind the particles into thelarge agglomerates or clumps (Choong et al., 2007). In this tech-nique, commonly used chemicals are aluminum salts such asaluminum sulfate [Al2(SO4)3 �18H2O], and ferric salts such as ferricchloride [FeCl3] or ferric sulfate [Fe2(SO4)3 �7H2O] because of theirlow cost and relative ease of handling. In As removal by this pro-cess, chemicals transform As (dissolved) into solid (insoluble)which is precipitated later. Dissolved As may also be adsorbed onthe solid hydroxide surface site and be coprecipitated with otherprecipitating species (Mondal et al., 2006). The solids can be re-moved through sedimentation and/or filtration. Removal of Asfrom water by coagulation using ferric or aluminum salts havebeen reported in several studies (Zouboulis and Katsoyiannis,2002; Yuan et al., 2003; Wickramasinghe et al., 2004; Songet al., 2006; Andrianisa et al., 2008; Lakshmanan et al., 2010;Lacasa et al., 2011). Coagulants such as alum [Al2(SO4)3.18H2O],ferric chloride [FeCl3] and ferric sulfate [Fe2(SO4)3 �7H2O] arefound to be effective in removing As from water (Edwards, 1994).Ferric salts have been found to be more effective than alum inremoving As on a weight basis and effective over a wider pH range(Cheng et al., 1994; Hering et al., 1997). At pH 7.6 or lower, iron andaluminum coagulants are of equal effectiveness in removing AsV.However, iron coagulants are advantageous if pH is above 7.6, ifsoluble coagulant metal residuals are problematic, or if AsIII ispresent in the raw water. Generally, with increasing coagulantdosages higher As removal efficiencies can be achieved. The ef-fectiveness of iron coagulants in removing AsIII diminishes at pH6.0. Wickramasinghe et al. (2004) used ferric chloride and ferricsulfate as a coagulant and their study suggested that the rate of Asremoval was dependent on the raw water quality and pH adjust-ment before coagulation. On the other hand, studies investigatingthe effects of cationic (Wickramasinghe et al., 2004; Zouboulis andKatsoyiannis, 2002; Han et al., 2002) and anionic (Zouboulis and


Katsoyiannis, 2002) polymers for increasing As removal are lim-ited in numbers. Pallier et al. (2010) used kaolinite and FeCl3 as acoagulant/flocculent and they observed over 90% and 77% re-movals of AsV and AsIII, respectively, using 9.2 mg L�1 of Fe3þ . In arecent study, Hu et al. (2012) used three aluminum based coagu-lants (aluminum chloride and two types of polyaluminium chlor-ide) and all of them were able to reduce the concentration of Asbelow the MCL (Maximum Contaminant Level) with an initial AsV

concentration of 280 mg L�1. Further their study suggests that thealuminum species regulates the removal of As and therefore, theefficiency of As removal can be improved by adjusting the pH. Thepresence of sulfates significantly decreases AsIII removal, but onlyslightly affects AsV removal. At pH higher than 7.0, removal of AsV

increases in the presence of calcium. The major limitation of thecoagulation/flocculation process is the production of a largeamount of sludge with a considerable concentration of As.

Recently, instead of conventional Al and Fe salts, titanium tet-rachloride (TiCl4) was used (Shon et al., 2007) to remove theparticulate and dissolved organic matter from wastewater insewage treatment plants. TiCl4 successfully achieved high organicmatter removal to the same extent as Al and Fe salts and the re-sulting flocs are with better settleability. The most significant ad-vantage of using TiCl4 as a coagulant is that sludge recovery pro-duces a valuable by-product, namely titanium dioxide (TiO2) (Shonet al., 2007; Lee et al., 2009), which is the most widely used metaloxide, whose applications include cosmetics, electronic paper,paints, photocatalysts, and solar cells (Hoffmann et al., 1995; Obeeand Brown., 1995). Furthermore, the residual Ti salt concentrationin the treated water satisfied the requirement of the World HealthOrganization's (WHO) guidelines (0.5–15 mg L�1) for drinkingwater (Ravenscroft et al., 2009). Therefore, TiCl4 is expected to be apromising alternative coagulant for conventional Al and Fe salts.

4.4. Electrocoagulation (EC)

It is an alternative process to CF (coagulation/flocculation). In-stead of adding a chemical reagent as ferric chloride, metallic ca-tions are directly generated in the effluent to be treated by ap-plying a current between iron electrodes to dissolve soluble an-odes. In EC, electrolytic oxidation of a sacrificial iron anode pro-duces FeIII oxyhydroxides/precipitates in As contaminated water.With FeIII precipitates As forms binuclear, inner-sphere complexes(van Genuchten et al., 2012), which aggregate to form a floc. Me-tallic cations and hydroxides formed neutralize negatively chargedcolloids allowing them to coagulate (Matteson et al., 1995). DuringEC with iron electrodes, the amount of iron cations experimentallydissolved from the anode corresponds to the value predicted bythe second Faraday's law (Vik et al., 1984; Pretorius et al., 1991)which is used to calculate the treatment dose to apply.

4.5. Electro-chemical arsenic remediation (ECAR)

The ECAR is a form of electrocoagulation (EC) that has beendeveloped to support a community scale micro-utility businessmodel (Amrose et al., 2013). In ECAR, the high capacity adsorbentmedia are generated in-situ, removing the need for a central ad-sorbent manufacturing plant or importing media from abroad. InECAR, the As-laden flocs are separated from clean water throughgravitational settling aided by a small amount of alum as a coa-gulant. The effectiveness of ECAR has been demonstrated byAmrose et al. (2013), using synthetic groundwater in lab studies,real contaminated groundwater from Bangladesh and Cambodia,and in short-duration field trials of two 100 L batch reactors inWest Bengal.

The ECAR was found to lower initial As concentrations as highas 3000 μg L�1 to below the WHO-MCL of 10 μg L�1, and easily

reached below 5 μg L�1. Strong oxidants produces during theFenton-type reactions were found to oxidize AsIII to AsV (Li et al.,2012). This is a key reaction for high effectiveness in the Bengalregion because AsIII does not adsorb as strongly as AsV to FeIII

oxyhydroxide surfaces in natural water at neutral pH (Dixit andHering, 2003), and both AsIII and AsV are present in the ground-water (Kinniburgh and Smedley, 2001). Although the initial as-sessments of reliability, robustness, consumables cost, and sludgeproduction from 100 L reactor field trials were promising (Amroseet al., 2013), however, is very limited in scope due to the smallsystem size and short duration.

4.6. Adsorption

Adsorption is a process that uses solids for removing sub-stances from either gaseous or liquid solutions. Adsorption processhas been used most widely because of its high removal efficiency,easy operation and handling, low cost and sludge-free. Recently,several studies have focused in the development of novel materialsbased on alumina (Han et al., 2013), activated carbon (Zhang et al.,2007; Oliveira et al., 2008), iron oxides (Giménez et al., 2010; Sunet al., 2013), zeolites (Swarnkar and Tomar, 2012), clays (Anjumet al., 2011) etc. to adsorb As from water. Adsorption has attractedmuch attention due to the following advantages: (i) it usually doesnot need a large volume and additional chemicals, (ii) it is easier toset up as a POE/POU (point of entry/point of use) As removalprocess (Jang et al., 2008), and (iii) it does not produce harmfulbyproducts (Genc et al., 2004, Zhang et al., 2005) and can be morecost effective (Zhang et al., 2007). Generally, the removal of As byadsorption techniques depends on pH and the speciation of AsV

thus, at pH lower than 7 showing better AsV removals compared tothe AsIII (Zhu et al., 2013). The capacity and adsorption rate furtherdepends on the presence of other ions like phosphate, silicate,HCO3

- and Ca2þ competing for the adsorption sites (Giles et al.,2011, Zhu et al., 2013). There are some common adsorption tech-niques used for the efficient As removal from water are discussedbelow.

4.6.1. Activated aluminaActivated alumina (AA) is a physical/chemical process by which

ions in the feed water are sorbed to the oxidized AA surface. It isthe most widely tested aluminium oxide (Lin and Wu, 2001; Singhand Pant, 2004; Giles et al., 2011). AA was the first adsorptivemedium to be successfully applied for the removal of As fromwater supplies (EPA, 2000a, 2000b). It is prepared by the thermaldehydration of aluminium hydroxide Al(OH)3 at high temperature.It is a porous, granular material having typical diameter of 0.3–0.6 mm and a high surface area for good sorption properties. TheAA is used in packed beds to remove contaminants such as As,fluoride, NOM, selenium and silica. Under pressure, feed water iscontinuously passed through the beds to remove the con-taminants. The contaminant ions are exchanged with the surfacehydroxides on the alumina. When adsorption sites on the AAsurface become filled, the bed must be regenerated.

The As removal capacity of activated alumina is pH sensitivethus requires pre-and post-pH adjustment using caustic soda andsulfuric acid. AsV is strongly adsorbed on AA at pH 5–6 whereasAsIII is best adsorbed at pH 7–8 (Singh and Pant, 2004). The ad-sorption capacity of AA ranges from 0.003 to 0.112 g of As g�1 ofAA. The factors which have significant effects on the As removalachieved with AA are pH, As oxidation state, competing ions,empty bed contact time (EBCT), and regeneration. Other factorsinclude spent regenerant disposal, alumina disposal, and second-ary water quality.

The AA media can either be regenerated on-site or disposed ofand replaced with fresh media. Regeneration is achieved through a


sequence of rinsing with regenerant, flushing with water, andneutralizing with acid. The regenerant is a strong base, typicallysodium hydroxide (NaOH) and the neutralizer is a strong acid,typically sulfuric acid (H2SO4). The regeneration of saturated alu-mina is carried out by exposing the medium to 4% NaOH, either inbatch or by flow through the column resulting in high As con-taminated caustic wastewater. Onsite regeneration of AA mediatypically produces 37–47 bed volumes of caustic soda waste (EPA,2000a, 2000b). The caustic soda residue is then washed out andthe medium is neutralized with 2% solution of H2SO4 rinse. It hasbeen reported that at pH range 2.8–11.5, alum-impregnated AA ismuch better adsorbent for AsV than untreated AA and when em-ployed in batch mode AsV concentration could be brought downfrom 10 mg L�1 (10,000 ppb) to 40 ppb (Tripathy and Raichur,2008).

The adsorption capacity and rate of conventional activatedalumina (CAA) for As removal is relatively low and slow, which hasbeen attributed to its ill-defined pore structure along with smallsurface area as it is reported that the maximum AsV adsorptioncapacity of granular AA with a surface area of 118 m2 g�1 (Ma-cherey-Nagel, Germany) is 15.9 mg g�1 (Lin and Wu, 2001). On theother hand, mesoporous materials have attracted much attentionin the field of chemical engineering, optics, electromagnetics,biomedicine, industrial catalysis and adsorption, environmentalprotection, and fabrication of novel nano-object materials due tothe high surface area, well-defined and adjustable pore diameterof 2–50 nm (Ying et al., 1999; De Vos et al., 2002). Therefore, itcould be concluded that mesoporous alumina (MA) should be anideal adsorbent for removing As. Recently, efforts have been madeto synthesize appropriate mesoporous alumina for removing AsV,which included high-temperature crystallization (Patra et al.,2012) or the use of expensive aluminum alkoxide (e.g. aluminumtri-sec-butoxide) and organic solvent (2-butanol) (Yu et al., 2008).

4.6.2. Iron based sorbents (IBS)Adsorption on IBS is an emerging treatment technique for As

removal. Removal has been attributed to ion exchange, specificadsorption to surface hydroxyl groups or coprecipitation. Cur-rently, IBS products available in the market are granular ferrichydroxide (GFH), iron coated sand, modified iron and iron oxidebased adsorbents. Selvin et al. (2000) have described the sorptionprocess as chemisorption, which is typically considered to be ir-reversible. It has been reported that AsV adsorption on hydrousiron (III) oxide strongly depended on the system's and pH (Ranjanet al., 2003), while AsIII adsorption was pH insensitive. The AsIII

required less contact time to attain equilibrium and sulphate;phosphate and hydrogen carbonate did not compete strongly withthe AsIII adsorption. At acidic to neutral pH, adsorption of AsV isgenerally more effective than the adsorption of AsIII (Arienzo et al.,2005; Dixit and Hering, 2003; Leupin et al., 2005; Singh et al.,2007; Sharma et al., 2007; Su and Puls, 2008; Abdallah andGagnon, 2009; Burton et al., 2009). On the basis of chemistry ofthe remediation process, iron based technologies can be dividedinto two overlapping groups; first is when iron acts as sorbent, co-precipitant or contaminant immobilizing agent and the second iswhen iron behaves as a reductant (convert contaminants intolower oxidizing state or used as an electron donor). Two importantiron based materials are hydrous ferric oxide (HFO) and goethite(a-FeOOH) which are used as sorbent but goethite is less reactivethan HFO due to the lack of sufficient surface area (Smedley andKinniburgh, 2002). The mechanism of AsV adsorption on GFH wasstudied in detail by Guan et al. (2008), and they proposed that at7.4 pH, bidentate binuclear complexes with GFH are formed asevidenced by an average Fe–AsV bond distance of 3.32 Å by EXAFSanalyses. The impacts of temperature on adsorption kinetics andequilibrium capacities for AsIII and AsV on GFH have been reported

by Banerjee et al. (2008) and they showed that overall adsorptionreaction rate constant values for both AsV and AsIII increased withincrease in the temperature. The thermodynamic parameters ex-amination revealed that the adsorption of AsV and AsIII by GFH wasa spontaneous endothermic process.

Polymorphs of iron (III) hydroxy-oxide mineral [FeO(OH)] suchas goethite (a-FeOOH), akaganèite (b-FeOOH) and lepidocrocite (c-FeOOH) have been described as good adsorbents for AsV. Puregoethite was synthesized by Mohapatra et al. (2007) with differentdopant cations (Cu, Ni and Co) for adsorption of AsV. With330 m2 g�1 surface area, akaganèite exhibited adsorption capacityas high as 120 mg g�1 (Deliyanni et al., 2003). The adsorptioncapacity of akaganèite (b-FeOOH) can be enhanced by isomorphicsubstitution of Fe3þ by Zr4þ on its structure (Sun et al., 2013). AtpH 7.0, the maximum adsorption capacity for AsV by akaganèitewas 60 mg g�1. Adsorption capacities for lepidocrocite have beenreported 25.17 mg g-1 with specific area of 103.9 m2 g�1 (Repoet al., 2012). Another less known polymorph constitutes d-FeOOH,having similar crystallographic structure as CdI2. Crystalline d-FeOOH can be easily prepared in the laboratory by a simplemethod, with small particle size, high specific area and narrowpore size distribution. The uniqueness of this FeOOH polymorph isthat it is ferrimagnetic, thus after use in catalysis or adsorptionprocess it can be easily recovered by using a simple magnet (Pintoet al., 2012). These characteristics make d-FeOOH a suitable can-didate for use as an adsorbent for heavy metals in aqueous med-ium. Among the natural minerals, Fe containing natural magnetite(Fe3O4), siderite and hematite have given much attention (Dixitand Hering, 2003; Giménez et al., 2007; Jönsson and Sherman,2008). Guo et al. (2007a, 2007b) found that natural siderite andhematite removed As through electrostatic attraction and surfacecomplexation with the Fe hydroxides in the minerals. However,the rate of reaction was slow and common anions such as bi-carbonate and phosphate decreased the adsorption capacity of As.

In recent years, Fe3O4 (magnetite), a magnetic nanoparticles ofiron oxide nature and c-Fe2O3 (i.e., maghemite) have been foundapplicable in many practical branches of human activity. Based onthese, a promising technique was devised by mixing magnetiteand maghemite nanoparticles which can adsorb As from aqueoussolution and fit to use for the groundwater treatment. Under acidicpH conditions, 96–99% As uptake was recorded. The maximumadsorption occurred at 2.0 pH with values of 3.69 and 3.71 mg g�1

of adsorbent for AsIII and AsV, respectively at the initial con-centration of 1.5 mg L�1 solution of both species. However, inpresence of phosphate in the solution, the efficiency of the ad-sorption process suffered. Less than 60% As uptake was achievedfrom the natural groundwater containing more than 5 mg L�1

phosphate and 1.13 mg L�1 of As. This is a practical problem to befaced in the field application of the technology (Chowdhury andYanful, 2010).

Highly stable amorphous mesoporous iron oxides, prepared bythermal decomposition of ferric nitrate-oxalic acid complex, showedto be promising for the adsorption of AsV (Muruganandham et al.,2010). To increase the adsorption capacity of iron oxides, iron oxidescomposites and carbonaceous materials have been used as a strategy.For example, composites of Fe3O4-reduced graphite oxide-MnO2 re-moved 12.22 mg AsV per gram of catalyst (Luo et al., 2012) while in thepresence of Fe3O4/graphene composite, a high adsorption capacity(180.3 mg g�1) for AsV was achieved (Mishra and Ramaprabhu, 2012).Composites of iron oxides and TiO2 or Al2O3 presented adsorptioncapacity for AsV of 7.8 and 54.55mg g�1, respectively (D'Arcy et al.,2011; Basu et al., 2012)). On the other hand, natural iron oxidesshowed very low adsorption capacity (0.02–0.4 mg g�1) because of itslow specific surface area (Zhang et al., 2004; Guo et al. 2007a, 2007b;Giménez et al., 2010).


Besides micrometer-sized particle assemblies, iron oxide na-noparticles are frequently employed for the removal of As due totheir high surface area. Recently, Sabbatini et al. (2010) have de-veloped tubular ceramic adsorbers based on iron oxide nano-particles that have been deposited on porous alumina tubes. Ironoxide nanoparticles with combined ozonization-reactive ceramicmembranes have been proved a good candidate for AsIII removal(Park and Choi, 2011). In a study, Pan et al. (2010) have reportedthat on using iron-oxide-coated diatomite, a granular adsorbent,for adsorption of AsV from aqueous solution, effectively working atpH very close to the neutral value; here an industrial grade dia-tomite served as the iron oxide support. Recently, Chen et al.(2011) have suggested ferric-impregnated volcanic ash as a re-presentative of low-cost adsorbent for removal of AsV from anaqueous medium.

4.6.3. Zero valent ironOver the last decade, a great deal of research has been focused

on the removal of contaminants by zero valent iron (ZVI) becauseZVI is non-toxic, abundant, cheap, easy to produce, and littlemaintenance is required by its reduction process. The method isbased on the principle is when the oxygenated water comes incontact with ZVI, it results in corrosion of ZVI and formation ofvarious byproducts like FeII and FeIII hydroxides, that can oxidizeand remove As (Farrell et al., 2001; Manning et al., 2002). Manygroups have extensively investigated the removal of As by usingZVI in the laboratory (Leupin and Hug, 2005; Katsoyiannis et al.,2008; Klas and Kirk, 2013) as well as in the field (Hussam andMunir, 2007; Chiew et al., 2009; Neumann et al., 2013). Based onthis technology, approximately 350,000 filters are in use in India,Bangladesh, Pakistan, Nepal and Egypt with better results of Asremovals in field (Hussam and Munir, 2007, 2013; Neumann et al.,2013). It is desirable to analyze the effect of key variables, such aspH, dissolved oxygen (DO), hardness, and humic acid (HA) on Asremoval. Generally it is accepted that the mechanism of As re-moval by ZVI involves adsorption, reduction, surface precipitation,and co-precipitation with various iron corrosion products such asferrous/ferric hydroxides (Mak et al., 2009). The removal rates ofAsIII were investigated in the presence of various ions and underdifferent pH values and results indicated that most of theseparameters affected negatively the removal of AsIII (Biterna et al.,2010). Sun et al. (2011) reported that in anaerobic conditions whenAsV reacted with commercial ZVI or acid-treated ZVI both ad-sorption and reduction of AsV were fast; commercial ZVI reducedAsV to AsIII while acid-treated ZVI reduced AsV to As0.

The DO plays an important role in As removal by ZVI. The re-action of ZVI with DO leads to the formation of reactive inter-mediates (e.g. HO2/–O2

� , H2O2, –OH) and Fe(II). Fe(II) is oxidizedto Fe(III) and forms hydrous ferric oxides with large sorption ca-pacities (Joo et al., 2004), such that there is a potential for thetransformation and removal of a range of inorganic and organiccontaminants. It was found that low pH and high DO would favorthe As removal (Tanboonchuy et al., 2011). In aerobic condition,Bang et al. (2005b) observed a rapid rate of As removal fromaqueous solution via precipitation/co-precipitation on iron hy-droxide precipitation product and at pH 7.0 the uptake of AsV wasfaster than AsIII. However, in anaerobic conditions, the rate of re-action was considerably slower as compared to the aerobic en-vironment and at pH 4.0 and 7.0 the removal rate of AsIII was fasterthan AsV. Furthermore, their study also demonstrated that in ab-sence of O2, reduction of AsV to AsIII and AsIII to metallic As takesplace upon reaction with ZVI. Thus, different reaction mechanismsof As with ZVI predominate in aerobic and anaerobic environ-ments. In the absence of O2, ZVI is able to act as a reductant and toreduce AsV and AsIII.

Mak et al. (2009) had reported that the removal rate of As byZVI increased with increasing concentrations of Ca2þ or HCO3

�

because CaCO3 can form and acts as a nucleation seed for thegrowth of iron hydroxides. In the co-existence of Ca2þ , HCO3

� andhumic acid, the presence of HA diminished the positive role ofCa2þ due to the formation of Fe-humate complexes in solutionand delaying of the formation of CaCO3. The effects of HA on AsV

removal by ZVI from groundwater was investigated by Rao et al.(2009) and they observed that the removal rate of As was inhibitedin the presence of HA probably because of the formation of solubleFe-humate in the groundwater which hindered the production ofiron precipitates. When Fe-humate (complexation of HA withdissolved Fe) was saturated, further corrosion of ZVI acceleratedthe removal of As from groundwater via adsorption and co-pre-cipitation. The AsIII is more mobile and more toxic than AsV,however, the removal of AsIII is more difficult than the removal ofAsV. To remove high AsIII concentration of 10 mg L–1, Wan et al.(2010) conducted the combined processes of biological oxidationof AsIII, the removal of AsIII and AsV by ZVI obtained a very high Asremoval capacity (more than 70 mg As g�1 Fe). The kinetics andmechanism of AsIII oxidation and removal by ZVI at pH 3–11 inaerated water was investigated by Katsoyiannis et al. (2008) andfound that AsIII was oxidized by the Fenton reaction and removedby the sorption on newly formed hydrous ferric oxides and hy-droxyl radicals were the main oxidant for AsIII at low pH.

Recently nZVI, a nanoparticle, is effective and extensively usedfor the removal of As from contaminated water (Ludwig et al.,2009; Morgada et al., 2009; Tanboonchuy et al., 2011, 2012). Forexample, a novel process, air and/or CO2 bubbling nZVI was testedfor the removal of high concentration of AsV (3000 mg L�1) anddemonstrated outstanding performances. Tanboonchuy et al.(2012) investigated the influence of background species on theremoval of AsIII and AsV in the groundwater by nZVI process. Theyfound that Ca2þ plays a promoting role while PO4

3� and HA playan inhibiting role on removal of As. As for Cl� and HCO3

� , theformer enhances AsIII removal, whereas the later inhibits AsIII re-moval; AsV removal was affected slightly in the presence of Cl�

and HCO3� . Although nZVI is effective for the removal of As from

contaminated water, but transformations and translocation of Asat and within the nanoparticles are not clearly understood.

4.6.4. Indigenous filters and cartridgesPresently, several indigenous materials are available that are

used as filters for As adsorbent. Red soil rich in oxidized iron, ironore, clay minerals, iron scrap or fillings and processed cellulosematerials are known to have the capacity for As adsorption. Thetypical clay minerals such as kaolinite, montmorill ionite and illiteare used for the As adsorption (Lin and Puls, 2000; Goldberg,2002; Mohapatra et al., 2007). Manning and Goldberg (1996, 1997)and Mohapatra et al. (2007) reported that As is adsorbed on thesurface of clay through inner sphere surface complexation. Toenhance the adsorption of As on clay, Li et al. (2007) and Su et al.(2011) prepared surfactant-modified clay to remove As fromaqueous solution. Su et al. (2011) in a study used octadecyl tri-methylammonium chloride, octadecyl benzyl dimethylammoniumchloride and dioctadecyl dimethyl ammonium chloride as cationicsurfactants for the modification of bentonite. The charge on theclay surface get reversed i.e. from negative to positive by the re-placement of exchangeable Naþ and Ca2þ cations on clays withthe cationic surfactants and resulting organoclay becomes a po-tential adsorbent to adsorb the oxy-anions of As from aqueoussolution (Li and Bowman, 2001).

Zeolites are tectosilicate minerals with 3D aluminosilicatestructure containing water molecules, alkali and alkaline earthmetals in their structural framework. Due to high ion exchange,adsorbing, catalytic and molecular sieving capacities these have


potentials to be used as treatment mineral in the permeable re-active barriers (PRBs) (ITRC, 2005; Roehl et al., 2005). Removal ofAs from different type of water having varying mineralizationdegree can also occur through natural zeolitic rocks such as cha-bazite–phillipsite, clinoptilolite, and volcanic glass. The As removalefficiency for chabazite–phillipsite was 60–80% and for clin-optilolite-bearing rocks was 40–60%. Removal of As is influnced bythree key factors first, mineralogy of the zeolites occurring in thevolcanic rock, second, zeolite content of the zeolitic rock and lastly,the degree of water mineralization (Ruggieri et al., 2008).

By using above mentioned materials, some of the filters aremanufactured including Sono 3-Kolshi Filter, Granet Home-madeFilter, Chari Filter, Adarsha Filter, Shafi Filter, and Bijoypur Clay/Processed Cellulose filters. The Garnet home-made filter containsrelatively inert materials like brick chips and sand as filteringmedia without added any chemical to the system. Air oxidationand adsorption on iron-rich brick chips and flocs of naturallypresent iron in ground water could be the possible reason for Asremoval from ground water. The study demonstrates that three-pitcher filters are an effective option as a short-term measure forAs removal. However, the three-pitcher filters that are not effec-tive option for long duration. The Chari filter also uses brick chipsand inert aggregates in different Charis as filter media. The Shafiand Adarshs filters use clay material as filter media in the form ofcandle. Although the Shafi filter was reported to have better Asremoval capacity but clogging of filter media is still a problem.Khair (2000) found that Bijoypur clay and treated cellulose werealso able to adsorb As from water. These units/filters remove Aslike any other dissolved ions present in the water but are notsuitable for water having high impurities and iron content inwater. In early 90s, the development of Bio-Sand Filter (BSF) forhomebased drinking water treatment has received much attentionbecause of its high pollutants removal, technical simplicity, costeffectiveness and least maintenances (Bajpai and Chaudhuri, 1999;Gene-Fuhrman et al., 2005; Ngai et al., 2006; Guo et al., 2007a,2007b; Mahmood et al., 2011). Later Ngai et al. (2006), Ahammedand Davra (2011) and Noubactep et al. (2012) made certain mod-ifications to BSF design and filter media for improving its perfor-mance. In a recent study, Noubactep et al. (2012) employed threecompartment model of BSF having extended reactive layer of ZVIand reported the significant pollutants removal for safe drinkingwater production. Kanchan arsenic filter (KAF), a saturated sandfilter was designed by Ngai et al. (2006), and found to successfullyremove As and pathogens from drinking water. Sand is consideredas a non-reactive material and only removed suspended particlesduring filtrations (Noubactep, 2010).

Metal oxides/hydroxides coated adsorbents such as iron oxidecoated sand, manganese-coated sand, Fe3þ impregnated activatedcarbon, siderite-coated quartz and hematite coated quartz havebeen extensively used to improve the filter efficiency (Jessen et al.,2005; Leupin et al., 2005; Guo et al., 2007a, 2007b; Chang et al.,2008; Mondal et al., 2008; Chiew et al., 2009; Noubactep, 2010;Noubactep and Care, 2010; Maji et al., 2011; Noubactep et al.,2012). Recently, Rahman et al. (2011) employed sub-surface wet-land and soil filter systems for As removal. Many emergent ad-sorbents being used in filter-based treatments are not sufficient toremove total As (AsIIIþAsV) from water (Xu et al., 2007; Guanet al., 2012). Therefore, in order to provide As free drinking wateron sustainable basis, pre-oxidation process is essential to convertAsIII into better adsorbable AsV (Chang et al., 2008). In many de-veloping countries, cinders generated from the combustion of coalhoney comb briquette (HBC) in decentralized cylindrical stoveshave mostly been used for the civil applications. Recently, Yueet al. (2011) and Sheng et al. (2014) experimentally proved HBC tobe useful for the pollutants removal. Sheng et al. (2014) confirmedthat iron-amended HBC efficiently removed AsV (961.5 mg g�1) in

aqueous solutions. Furthermore, in China, over 60% of ruralhouseholds are directly depends on coal honeycomb briquette forcooking and heating (Sinton et al., 2004), that ultimately producemassive cinders.

4.6.5. Miscellaneous adsorbentsIn last years, a wide variety of adsorbent systems have been

developed for the removal of As. Activated carbon (Huang and Fu,1984), fly ash (Diamadopoulos et al., 1993), aluminum-loaded corallimestone (Ohki et al., 1996), modified fly ash (Goswami and Das,2000), iron oxide minerals (Suvasis and Janet, 2003), activatedneutralized red mud (Hulya et al., 2004), chitosan (Chen andChung, 2006), chitosan derivatives (Laurent et al., 2002), iron hy-droxide-coated alumina (Hlavay and Polyak, 2005), modified fun-gal biomass (Pokhrel and Viraraghavan, 2006), iron containingmesoporous carbon (Zhimang and Baolin, 2007), iron oxide-im-pregnated activated carbon (Ronald et al., 2007), nanoparticles ofhydrous iron oxide (Sylvester et al., 2007) etc. were used as ad-sorbents for the removal of As from aqueous environments. Chit-osan is transformed polysaccharide obtained from the de-acet-ylation of chitin, which makes the shells of crustaceans such ascrabs and shrimps. It is biodegradable, biocompatible, and non-toxic, making it environment friendly. Guibal (2004) reportedchitosan as an efficient heavy metal scavenger due to the presenceof hydroxyl and amino group with high activity as adsorption site.Although, chitin/chitosan have been used for the removal of AsV

from water but capacity was found to be very low(0.13 mmol g�1). Recently, a novel composite chitosan bioadsor-bent (CCB) have been developed (Boddu et al., 2008) for the re-moval of AsIII and AsV from aqueous solutions by coating naturalbiopolymer, chitosan, on ceramic alumina, using a dip-coatingprocess. Their study reported a very high adsorption capacity atpH 4.0 (56.50 and 96.46 mg g�1 for AsIII and AsV, respectively).Chen et al. (2008) used the chitosan impregnated with molybdateto remove AsIII and AsV from contaminated water.

A high As sorption capacity (2860 mg g�1) was observed onactivated carbon by Rajakovic (1992). Rajakovic (1992) found thatcarbon pretreated with Agþ or Cu2þ ions improved the AsIII ad-sorption but reduced the AsV adsorption. In order to improve theAs adsorption carbon was impregnated with different metal ionssuch as iron oxide. For example, Gu et al. (2005) prepared the ironcontaining granular activated carbon (GAC) adsorbents for As ad-sorption from drinking water. Factors such as solution pH, carbontype and carbon pretreatment and elution of the As from loadedcarbon that affect the mechanism of the adsorption of As specieson activated carbons were studied (Lorenzen et al., 1995). Theyobserved that AsV is more effectively removed from solution byusing activated carbon with high ash content and pre-treatment ofthe carbon with Cu(II) solutions also improved its As removal ca-pacity. Gu and Deng (2006) prepared iron containing mesoporouscarbon (IMC) from a silica template (MCM-48) and used for theremoval of As from drinking water. The maximum adsorption ca-pacities were 5.96 mg As g�1 for AsIII and 5.15 mg As g�1 for AsV.

Focusing the morphology and size of adsorbents it is aimed togenerate high surface areas and high density of adsorptive sites.For example, three-dimensional flower-like, urchin-like and hier-archical superstructure adsorbents have been found to well suitthe adsorption application (Zhong et al., 2006; Wang et al., 2012;Zhang et al., 2012). Pore structure also influences the adsorptionbehavior of an adsorbent (Drisko et al., 2009; Kimling et al., 2012).Hierarchically porous materials, possessing macro-and/or meso-porous networks, facilitate rapid As species diffusion as well aspromote access to the active sites, resulting in a high As uptakeand kinetics (Zhang et al., 2010a, 2010b; Wu et al., 2012). Chemicalcomposition variation can change the properties of adsorbentmaterials; consequently affect the adsorption performance in


applications (Kimling et al., 2012). In addition, the speciation of Asand the surface charge of the adsorbents are related strongly to thewater pH value. Thus, pH effect plays a great role in the As ad-sorption process (Sofia Tresintsi et al., 2012).

In the past few years, ceria (Ce) and zirconia (Zr), highly re-active rare-earth metal oxides, are found incorporated into theadsorbents that significantly can improve the adsorption capacityof As because they have a unique selectivity for polyoxy anions(Biswas et al., 2008). For example, granular Fe–Ce oxide (Zhanget al., 2010a, 2010b), Fe–Zr binary oxides (Ren et al., 2011), Ce–Tioxide (Deng et al., 2010), Zr (IV)-loaded ligand exchange fiber(Rabiul Awual et al., 2012), Zr (IV)-loaded orange waste gel (Biswaset al., 2008), etc. showed enhanced As adsorption performancebecause of their increased surface areas, surface hydroxyl group,and pore accessibility. In aqueous solutions, Ce and Zr can formtetranuclear or octanuclear species, which have abundant hydro-xyl groups and water molecules to be involved in ligand sub-stitution with As species (Zhong et al., 2007; Biswas et al., 2008; Liet al., 2010). Thus, composites combined CeO2 and ZrO2 might bean ideal candidate for the removal of As.

4.7. Ion exchange

In the past, ion exchange has been used for removal of con-taminants from water (Oehmen et al., 2006). Ion exchange is aphysical/chemical process by which an ion on the solid resin phaseis exchanged for an ion in the feed water. The solid resin is typi-cally an elastic three-dimensional hydrocarbon network contain-ing a large number of ionizable groups electrostatically bound tothe resin. These groups are exchanged for ions of similar charge insolution that have a stronger exchange affinity (i.e. selectivity) forthe resin. Typically, strong-base anion exchange resins are com-monly used for the removal of As where the oxy-anionic species ofAsV (such as H2AsO4

� , HAsO4�2 and AsO4

�3) are effectively ex-changed with the anionic charged functional group of the resin,thus produces effluents with low concentration of AsV (Choonget al., 2007). Over a larger range of pH, strong-base anion resinstend to be more effective than weak-base resins. The order ofexchange for most strong-base resins is given below:

HCrO4�4CrO4

2�4ClO4�4SeO42�4SO4

2�4NO3�4Br�4

(HPO42� , HAsO4

2� , SeO32� , CO3

2�)4CN� 4NO2�4Cl�4

(H2PO4� , H2AsO4

� , HCO3�)4OH�4CH3COO�4F

This technology in drinking water treatment is commonly usedfor the softening and nitrate removal. Before passing the As con-taminated water, the resin bed are usually flushed with HCl so asto implant labile Cl� on the surface of the resin, which later easilyexchanged with As. The AsV can be easily removed through the useof strong-base anion exchange resin either in the form of chlorideor hydroxide. Sarkar et al. (2007), Wan et al. (2010) and Donia et al.(2011) have reported As removal by using strong base anion ex-change resins. The efficiency of ion exchange process is improvedby pre-oxidation of AsIII to AsV but before the ion exchange, theexcess of oxidant often needs to be removed in order to avoid thedamage of sensitive resins. Therefore, the efficiency of the ionexchange process for AsV removal strongly depends on the solu-tion pH and the concentration of competing ions most notablysulfates and nitrates, resin type, alkalinity, and influent. Otherfactors include the affinity of the resin for the contaminant, spentregenerant and resin disposal requirements, secondary waterquality effects, and design operating parameters. The performanceof an ion exchange system can be adversely affected by high levelsof total dissolved solids (TDS). Nitrate, sulfate and phosphate,common competitive anions, play a significant role for the removalof As via ion exchange. When an ion is preferred over AsV, higherAs level in the product water than exist in the feed water can beproduced. Höll (2010) suggested if a resin prefers sulfate over AsV,

for example, sulfate ions may displace previously sorbed AsV ions,the resulting levels of As in the effluent is greater than the As levelin the influent. In general, ion exchange for As removal is onlyapplicable for low-TDS, low-sulfate source waters. Removal of Ascan also be affected by the presence of iron, FeIII, in feed water. Inthe presence of FeIII, As may form complexes with iron. Thesecomplexes are not removed by ion exchange resins and thereforeAs is not removed.

4.8. Electrokinetics

Electrokinetic (EK) remediation is a technique that already hadproven its value, especially in contaminated fine-grain soils.Virkutyte et al. (2002) reported three phenomena occurs duringelectrokinesis are electro-osmosis, electromigration and electro-phoresis. This method uses a low-level direct current as the“cleaning agent”, several transport mechanisms (electroosmosis,electromigration and electrophoresis) and electrochemical reac-tions (electrolysis and electrodeposition) are induced (Acar andAlshawabkeh, 1993). When a direct electrical field is applied acrossa wet mass of contaminated soil, the migration of non-ionic porefluids by electro-osmosis and the ionic migration of dissolved ionstake place towards the electrodes. This technique has certain ad-vantages over the conventional methods: (1) it is efficient in lowpermeability soil, which is difficult to treat by using other meth-ods; (2) is possible to set-up in situ for sites that are impossible toexcavate, such as residential areas and railway soil and (3) it canremove organic and inorganic pollutants from soil simultaneously.

The electro-remediation is considered to be the most effectivein treating near saturated, clay soils polluted with metals, wherebyremoval is 490% (Virkutyte et al., 2002). Sims (1990) andCauwenberghe (1997) reported that in sub-surface the electro-migration rate is dependent upon the density of water current, soilpore, ionic mobility, grain size, concentration of contaminant andtotal ionic concentration. In turn, it is governed by advectionwhich is generated by electroosmotic flow and externally appliedhydraulic gradients, diffusion of the acid front to the cathode andthe migration of anions and cations towards the respective elec-trodes (Zelina and Rusling, 1999).

During the EK process, the dominant electron transfer reactioni.e. electrolysis of water occurs at electrodes is as follow:

H2O-2Hþþ½ O2(g)þ2e� (8)

2H2Oþ2e� -2OH-þH2(g) (9)

The hydrogen ions produced in the above given process de-creases the pH near the anode causing desorption of metalliccontaminants from the soil solid phases. The dissolved metallicions are then removed from the soil solution by ionic migrationand precipitation at the cathode (Acar and Alshawabkeh, 1993). Onthe other hand, when hydroxide ion concentration increases thepH near the cathode also increases. Electro-kinetic remediationtechniques demonstrated 85–95% As removal efficiency, from low-permeability soils such as clay, peat, kaolinite, high-purity finequartz, Na and sand montmorillonite mixtures, as well as fromargillaceous sand (Yeung et al., 1997). However, kaolinite showedmore than 90% removal efficiencies of heavy metals (Pamukcu andWittle, 1992).

During EK remediation process, various chemicals such aschelating agents, surfactants, etc. have been investigated to facil-itate the mobility of pollutants in soil. To remediate the soil con-taminated with gasoline Bhattacharya (1996) applied surfactanteffectively using EK system. To restore the diesel-contaminatedsandy soil, Kim and Lee (1999) adopted an anionic surfactant,sodium dodecyl sulfate (SDS), in the EK process. In their experi-ments, the effects of electrophoretic transportation of SDS fed into


a catholyte chamber, showed the dominant mechanism for theremoval of diesel and compared them to those of electroosmoticflow. Yang et al. (2005) tested surfactants that had anionic or non-ionic characteristics and observed the effects of their ionic char-acteristics and concentrations during the EK process. On the otherhand, to increase removal of heavy metals from the soil, Yeunget al. (1996) and Wong et al. (1997) tested a chelating agent, EDTA,in EK remediation system.

4.9. Membrane technology

Membranes are typically synthetic materials with billions ofpores or microscopic holes that act as a selective barrier; thestructure of the membrane allows some constituents to passthrough, while others are excluded or rejected. The movement ofmolecules across the membrane needs a driving force, such aspressure difference between the two sides of the membrane. Thistechnology can reduce As concentrations to less than 50 mg L�1

and in some cases to below 10 mg L�1. It produces large residualvolumes and is more expensive than other As treatment technol-ogies. Researchers have explored various types of pressure drivenmembrane such as microfiltration (MF), ultra-filtration (UF), nano-filtration (NF) and reverse osmosis (RO) for the removal of As fromcontaminated water. The separation by these processes dependson the pore size of the membrane; for MF and UF membranes,mechanical sieving is responsible for separation while for NF andRO membranes, separation is achieved via capillary flow or solu-tion diffusion (Shih, 2005; Choong et al., 2007).

4.9.1. As removal using microfiltrationThe MF is a low pressure driven membrane process for separ-

ating colloidal and suspended particles in the range 0.1–10 μm.The MF alone cannot remove the dissolved AsV and AsIII speciesfrom As contaminated water. The As removal by MF membranecan only be achieved by increasing particle size of As bearingspecies therefore prior to MF, coagulation and flocculation pro-cesses could be effective to increase the particle size of As bearingspecies and were found to remove As species from As con-taminated water (Meng et al., 2000; Han et al., 2002; Chwirkaet al., 2004; Ghurye et al., 2004). For example, Han et al. (2002)used FeCl3 and Fe2(SO4)3 as flocculants and studied the removalrate, which was dependent on the adsorption of As on to the FeIII-complex and their subsequent removal from the solution. Theirresults showed that by the combination of flocculation and MFtechnique, the As removal efficiency are higher than MF alone anddepends on the effectiveness of As adsorption onto the FeIII-complex present and also on the rejection of the As containingflocs formed by the MF membrane. In turn, adsorption of As on theFeIII complex was found to be affected by the pH of the solution aswell as the presence of other ions in the solution.

Shih (2005) reported that in the pH range 4.0–10.0, negativelycharged AsV anions got effectively adsorbed by forming surfacecomplex while AsIII removal was poor because in the pH range4.0–10.0, it remained as neutral species and could not get ad-sorbed. Therefore, complete removal of As from water could havebeen achieved by completely oxidizing AsIII to AsV prior to coa-gulation–microfiltration process. The size of As containing parti-cles increases via coagulation–flocculation process and thus makesit possible to remove As species using low-pressure membranetechnology like MF.

Recently, Ghosh et al. (2011) studied the electrocoagulation (EC)followed by MF by using a ceramic membrane was found to be ef-fective in the removal of As from feed solution having concentrationof 200 μg L�1 in presence of fluoride and iron contaminant to a Ascontent of 8.7 μg L�1. The EC experiment, consisted of a bath with

four aluminium sheets of 0.15 m�0.05 m�0.002 m, were continuedupto 45 minutes with a current density of 625 Am�2.

4.9.2. As removal using ultrafiltrationUltrafiltration (UF) is a size exclusion-based low pressure-dri-

ven membrane separation process having pore sizes in the rangefrom 10 to 1000 Å and is capable of retaining species in the mo-lecular weight ranging from 300 to 5,00,000 Da. The rejection ofAs by charged membrane explored the influence of co-occurringdivalent ions and natural organic matter (NOM). In presence ofdivalent cations such as Ca2þ , Mg2þ , AsV rejection reduced almostupto zero. This reduction in AsV rejection probably due to theformation of ion pairs between counter ions and the fixed chargegroup in the membrane matrix locally neutralizes the membranecharges. Brandhuber and Amy (2001) investigated the effect ofcharge on the UF membranes and reported that mechanism of Asremoval was mainly due to the electrostatic interaction betweenthe As ions and the negatively charged membrane surface, con-sistent with the Donnan theory. In their study they found amoderate rejection of 53% and 65% for AsIII and AsV, respectively.Thus, alone UF is not able to remove the As species directly due tothe pore size, which easily passes the dissolved As through themembrane.

The presence of NOM improved AsV rejection in presence ofdivalent cations. This might be due to the complexion of divalentions whose presence in the solution reduces AsV rejection. Anotherpossibility might be due to the adsorption of NOM onto themembrane surface to form a negatively charged layer. Solutionwith higher concentration of NOM led to higher charge density inthe adjacent membrane layer causing more rejection of the ne-gatively charged As species.

Iqbal et al. (2007) studied the effect of co-occurring inorganicsolutes (such as HCO3

-, HPO42-, H4SiO4 and SO4

2-) on the removalof AsV in feed water and permeate flux were investigated by usinga cationic surfactant cetlylpyridinium chloride (CPC) and a flatsheet hydrophilic polyethersulfone (PES) ultrafiltration membrane(Millipore, Bedford, MA). The PES membrane without surfactantmicelles was found to be ineffective for the removal of As whilethe addition of surfactant significantly increased the efficiency ofAs removal. Removal of As with surfactant was found to be 78–100% while As removal in the presence of inorganic solutes wasonly 25%.

The removal characteristics of As with ground water was in-vestigated by using micellar-enhanced ultrafiltration (MEUF)(Iqbal et al., 2007). The different cationic surfactants used by themwere benzalkonium chloride (BC), hexadecylpyridinium chloride(CPC), hexadecyltrimethyl ammonium bromide (CTAB) and octa-decylamine acetate (ODA). Among these four cationic surfactantsused during MEUF, the highest As removal efficiency was obtainedin case of CPC i.e. 96%, with CTAB was 94% and the removal effi-ciency with ODA was over 80% while the As removal efficiencywith BC was the lowest i.e. 57% due to the higher critical micelleconcentration (CMC) of BC compared to those of other surfactants.

4.9.3. As removal using NanofiltrationsNanofiltrations (NFs), high pressure techniques and are able to

remove the dissolved As from the contaminated water to an ap-preciable level provided that the feed water contains a very lowamount of suspended solids (Figoli et al., 2010). The nominal poresize of the membrane is about 1 nm and molecular weight is ty-pically less than 1000 Da.

Generally, NF membranes are applied for the separation ofmultivalent ions from the monovalent ions. These membraneshave slightly charged surfaces and charge interaction plays adominant role in separation of molecules. The NF membranes areusually asymmetric and negatively charged at neutral and alkaline


media but lose their charge in acidic pH. Waypa et al. (1997) re-ported that both dissolved AsV and AsIII can be effectively removed(up to 99%) from water due to the size exclusion. However, Uraseet al. (1998) demonstrated that charged membranes generallyhave a higher rejection for charged solutes than for unchargedsolutes and concluded that the pore size of the membrane doesnot have a significant effect on As rejection, but rather chargeexclusion is predominant over the size exclusion mechanism.

Seidel et al. (2001) and Velizarov et al. (2004) demonstratedthat separation of As is caused due to the electrostatic repulsionbetween anionic As species and the charge of membrane (Donnanexclusion) as well as the small sized pores of the membrane. Al-though the charges of the NF membranes are not fixed but maychange due to adsorption of anions onto the membrane, andtherefore are dependent on the bulk anion concentration(Velizarov et al., 2004 ). Moreover, at the isoelectric point of themembrane, the charge of the membrane is zero, which changes topositive values for a lower pH. Thus, pH selection is importantwith respect to the target ions are to be separated. The perfor-mances of two commercially available NF membranes (NF90 andNF200) were studied by Uddin et al. (2007) for removal of bothAsIII and AsV from contaminated water. They demonstrated thatthe rejection of AsV was better than AsIII and even when the feedsolutions contain 50 mg L�1, AsIII could not be reduced to MCL andthey concluded that oxidation of AsIII to AsV is an essential pre-treatment step. Figoli et al. (2010) demonstrated that the AsV re-moval efficiency was increased with increasing of solution pHalong with decrease of operating temperature and the feed Asconcentration, while the transmembrane pressure had little/noeffect. Sato et al. (2002) and Uddin et al. (2007) reported thatwithout oxidative pretreatment, NF is not sufficient for the re-moval of total As from aqueous solution and therefore is not ableto provide water quality below the MCL.

In a recent study, As removal by employing self-made PMIA(poly m-phenylene isophthalamide) nanofiltration membrane wasinvestigated with aqueous solution of As salt (Na2HAsO4.7H2O)and more than 90% AsV rejection was reported (Zhao et al., 2012).Furthermore, AsV rejection increased from 83% at pH 3 to 99% atpH 9. In presence of NaCl, AsV rejection increased in the feed rangestudied and was reported to be 99% when feed concentration was100 μg L-1.

4.9.4. As removal using reverse osmosisReverse osmosis (RO) is the oldest and identified as the best

available technology for the small water treatment systems toremove As from water. The RO membrane contains extremelysmall pores (o0.001 μm) (Schneiter and Middlebrooks, 1983) anda very high (often close to 100%) rejection of low-molecular masscompounds and ions can be achieved (Velizarov et al., 2004).Moreover, the process can easily be automated and controlled. Inthe 1980s, with the invention of cellulose acetate RO membrane,AsV removal efficiency have been achieved of above 90% with theRO system operated at high-pressure around 400 psi (Cliffordet al., 1986; Fox, 1989) while, AsIII removal efficiency was lessthan 70%. The operational parameters for As removal by ROmembranes was investigated by Akin et al. (2011) and found thatalthough the feed water concentration had no effect on the re-jection rate, the rejection of As was largely affected by the pH ofthe feed water and operating pressure. Yoon et al. (2009) studiedthe effect of charged membranes on the As removal and theirstudy concluded that there was a considerable rejection of theoxy-anionic AsV while below pH 10 AsIII removal was low due tothe existence of uncharged AsIII species in solution. Thus, thepresence of AsIII plays an important role for removal of As. Ac-cording to the Walker et al. (2008) when AsIII was the dominant Asspecies in the groundwater, the As removal efficiency decreased to

less than 50%. Therefore, where AsIII is the dominant species ROseems to be ineffective for As contaminated natural water espe-cially for reducing conditions.

4.10. As removal by advanced hybrid and integrated technologies

4.10.1. As removal using membrane distillationPresently, As remediation is taking place with better under-

standing of complex chemical processes due to the recent advance-ment in science and technology. Among the different kinds ofmembrane distillation (MD), direct contact membrane distillation(DCMD) is the most simple, economical and efficient techniquewhere the hot feed and the cold permeate are directly separated bythe membrane. The MD is a non-isothermal membrane separationprocess that employs a microporous hydrophobic membrane withpore size ranging from 0.01 μm to 1 μm. The main requirements forMD process are that the membrane should not be wetted and onlyvapor and non-condensable gases should be present within its pores.Based on these requirements hydrophobic, microporous membranesmade of polytetrafluoroethylene (PTFE), polyethylene (PE), poly-propylene (PP), and polyvinylidenefluoride (PVDF) are now com-mercially available. Recently, there have been studies to remove Asusing DCMD (Islam, 2004; Manna et al., 2010; Yarlagadda et al.,2011). In one such study, Manna et al. (2010) had performed ex-periments by using three different types of hydrophobic membranesfor solar driven MD and studied the removal of AsIII and AsV fromgroundwater. Their study demonstrated that nearly 100% of the Asremoval from the contaminated groundwater with high flux of49.80 kg m�2 h�1. Recently, AsV and AsIII removal studied by Quet al. (2009) using self-made polyvinylidene fluoride (PVDF) mem-brane in a DCMD unit and the experimental results indicated that thepermeate AsIII and AsV were under the maximum contaminant limit(10 μg L�1) until the feed AsIII and AsV concentration achieved 40 and2000 mg L�1, respectively. Macedonio and Drioli (2008) tested theperformance of an integrated RO and MD system. This integratedprocess (RO followed by MD) removes total As fromwater and leavesless environmental hazardous waste on comparing to oxidation fol-lowed by the RO and thus is very promising for contaminated waterpurification.

4.10.2. As removal using forward osmosisForward osmosis (FO) is another membrane process that has

been used for treatment of industrial waste, concentrate landfillleachate, liquid food in food industry and desalination of seawater(Cath et al., 2006). Water is filtered due to the osmotic pressuredifference, which is the driving force for FO separation. Cath et al.(2006) had determined several advantages of the use of FO: (i) it isa simple equipment set-up (ii) there is low or no requirement ofhydraulic pressure, (iii) the membrane have a high rejection, and(iv) the probability of lower membrane fouling than other con-ventional membrane processes. Recently, Jin et al. (2012) in-vestigated the removal of As using FO and also studied the influ-ence of membrane orientation and membrane fouling. The As fromsolution was successfully removed by the FO and the removal/rejection was better when the surface active membrane side facedthe feed water compared to the draw water due to more severeconcentrative internal concentration polarization. In addition, theyalso found that for FO, the AsIII rejection from solution was greaterthan that of RO; the rejection increased by increasing water flux.Further, they concluded that the lower rejection by RO was due tothe deformation of the fabric supported membrane followed byexpansion of membrane pores by hydraulic pressure, which en-abled As and other solute particles to pass through the membrane.Although in DCMD and FO, the waste water with an increased Asconcentration after separation may be re-treated, the As con-centration significantly increases in the feed water. As a result, the


risk for membrane damage increases along with the cost oftreatment. Therefore, this concentrate then can be treated withZVI under an inert atmosphere so as to reduce the soluble As tocommercially valuable metallic As (Bang et al., 2005a; Sun et al.,2011).

4.11. Disposal of As laden sludges and wastes

In all the above mentioned methods, sludge is generated as aresult of precipitation of Fe or Al hydroxides depending on theamount of coagulant present naturally and/or added. This sludgecontains high concentrations of As, therefore, for disposal of thissludge care should be taken. In a recent study, Sarkar et al. (2010)addressed to dig holes and bury the sludge away from edible cropsand children's playing area is one of the possible ways of disposingof the sludge from household removal units, whereas the sludgefrom community based removal units can be stored in aeratedcoarse sand filters.

5. Conclusion and future perspective

South-East, South-West and North-East USA, Inner Mongolia(China), South-West Taiwan coastal regions, Sonora (Mexico),Pamplonian Plain (Argentina), West Bengal (India), Northern Chile,and Bangladesh are highly As polluted centres. Therefore, treat-ment of As contaminated water and soil could be the only effectiveoption to minimize health hazard. To achieve this, various tech-niques are being used. However, the adopted technologies havesome drawbacks and their by-products can be a further potentialsource for secondary As pollution. Therefore, for better clean up ofthe environment, new technologies with the options of new hy-brid technologies are needed to challenge the menace of As.Among various techniques, membrane technologies, electro-re-mediation and phytoremediation are seemed to be more suitablefor removal of As fromwater and soil. However, further research isneeded to develop more efficient techniques without drawbacksfor As free life.

Acknowledgments

Authors are thankful to the University Grants Commission, NewDelhi for providing financial assistance to carry out the work.

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