Review Arsenic 2015

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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

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. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2534.1.1. Oxidation and ltration. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2534.1.2. Photochemical oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2534.1.3. Photocatalytic oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2544.1.4. Biological oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 254

4.1.5. In-situ oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2544.2. Phytoremediation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2554.3. Coagulationocculation. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 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

Contents lists available atScienceDirect

journal homepage: www.elsevier.com/locate/ecoenv

Ecotoxicology and Environmental Safety

http://dx.doi.org/10.1016/j.ecoenv.2014.10.0090147-6513/&2014 Elsevier Inc. All rights reserved.

n Corresponding authors.E-mail addresses: [email protected](V.P. Singh), [email protected](S.M. Prasad).

Ecotoxicology and Environmental Safety 112 (2015) 247270
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4.6.4. Indigenous lters 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 microltration. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.9.2. As removal using ultraltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.9.3. As removal using Nanoltrations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.9.4. As removal using reverse osmosis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 262

4.10. As removal by advanced hybrid and integrated technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2624.10.1. 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 in

South-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 occurring

trace 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 (AsIII) 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 As III 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)3AsCH2COOH], arseno-choline [AsC], arsenosugars [AsS], arsenolipids etc. (Tangahu et al.,2011). In general, inorganic arsenicals are more toxic than organic

ones (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 reach

upto 16 mg As/kg sh oil (Sele et al., 2012). Recent ndings

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;Kile

et 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 L1. However, the levels of Asin unpolluted surface water and groundwater vary typically from110g L1. Groundwater concentrations of As is reported to bevery large range from less than 0.55000 mg L1 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 g L1 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 g L1 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 for

humans, which causes a life-threatening problem for millions of

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13,900mg L1) 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 L1.

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 and

2.08 mg L1

, has been transported from the abandoned CaribooGold Quartz mine tailings of that area, and gets accumulated inhigh concentration (upto 1104mg g1) 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 stratication 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 the

Moira Lake, Ontario, Canada with an average concentration of47mg L1 in surface water, compared to 22mg L1 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 stratication in lake water also causes the release of Asinto the water column from bottom sediments due to depletion ofO2levels 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 L1 (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 g L1

(Ng, 2005), and its concentrations in Atlantic and deep Pacicwaters are between 1.01.8mg L1 (Cullen and Reimer, 1989),3.1g L1 in marine waters of the Pacic coast near Nakaminato

(Ibaraki, Japan) (Ishikawa et al., 1987), and 1.11.6mg L1 in coastal

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 efuents and geothermal water. The physi-cal mixing of the fresh and seawater masses and salinity may in-uence 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 kg1. 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 areidentied 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 kg1 (Stalov et al., 2010) but the backgroundconcentration can signicantly differ depending on soil conditions.Arsenic contamination in soil in different parts of the world issummarized inTable 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 in uenced by theamount of food in the diet. The amount of arsenic in various food

stuffs and its safe limit is summarized inTable 4.

Table 3

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

Country Region Soil As concentration in mg/kg References

Bangladesh Noakhali 3.626 mg/kg (Meghna River) Nriagu et al., 2007Brazil Minas Gerais (Southeastern Brazil) 200860 mg/kg Bundschuh et al., 2012Chile Esquia Up to 489 mg/kg (Ro Caritaya region) Bundschuh et al., 2012;Nriagu et al., 2007India Uttar Pradesh 16417 mg/kg (Central India) 5.4015.43 ppm (Uttar Pradesh) Das et al., 2013;Srivastava and Sharma, 2013Mexico Lagunera 22152675 mg/g (Highly polluted area) Nriagu et al., 2007Poland Lower Silesia, (Southwestern Poland) Up to 18,100 mg/kg (Highly polluted area) Karczewska et al., 2007Spain Duero Cenozoic Basin 23 mg/kg (Mean) Gmez et al., 2006Turkey Simav plain (Kutahya) Up to 660 mg/kg (Highly polluted area) Gunduz et al., 2010United Kingdom Cornwall 217 mg/kg (Bioaccessible) Palumbo-Roe et al., 2005USA Tulare lake average 280 mg/kg (Hawaii) Nriagu et al., 2007



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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 birds

raised for organic certication. It is estimated that roxarsone isadded to poultry feed at the rate of 22.745.4 g per ton, or 0.00250.005 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 world

which can be used as organic fertilizers on agricultural lands.

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 0.42119.0 mg kg1. Therefore, there is a growing public concern aboutthe arsenic residues accumulation in pig manure. Silbergeld andNachman (2008)also observed arsenicals in pig manure that may

likely 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; rst, 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 signicant 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 kmin WestBengal has been identied as As contaminated zone among which38,861 kmarea has been identied 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;Chatterjee

et al., 2010;Samal et al., 2011) which is the potential human health

Table 4

Arsenic concentrations in various food stuffs and its recommended limit.

Food stuff(s) Arsenic concentra-tion (mg As kg1)

Recommended limit ofarsenic in food (FAO)

Reference

Rice 153.1 1 mg kg1 EFSA, 2014Wheat 22.0 do EFSA, 2014Oat 27.3 do EFSA, 2014Corn 49.3 do EFSA, 2014

Vegetables 903900 do Das et al.,2004

Pulses 1300 do Santra et al.,2013

Chicken meat 286 do Islam et al.,2013

Fish 3000 do Lin and Liao,2008

Fig.1. Schematic diagram is showing transfer of arsenic from soil and water to human beings through food chains. Intake of arsenic by human beings causes several diseases.

R. Singh et al. / Ecotoxicology and Environmental Safety 112 (2015) 247270 251


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4.1. Arsenic revomal by oxidation techniques

These techniques use various processes and are discussedbelow.

4.1.1. Oxidation andltration

The main purpose of oxidation is to convert the soluble AsIII toAsV, 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 ltration refer to the processes which are designedto remove the naturally occurring iron and manganese from water.The process involves the oxidation of the soluble forms of iron andmanganese to their insoluble forms and then removal by ltration.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 ooded paddy elds, 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 afnity to mineral surfaces, while

arsenate 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)3HOCl-AsO43Cl4H (1)

ClO2: As(OH)3

2ClO2

H2O-

AsO43

2ClO2

5H

(2)O3: As(OH)3O3-AsO4

3O23H

(3)

H2O2: As(OH)3H2O2-AsO433HH2O (4)

NH2Cl: As(OH)3NH2ClH2O-AsO43NH4

Cl3H (5)

MnO4: 3As(OH)32MnO4

--3AsO4

3-2MnO27H

H2O (6)

FeO42: 3As(OH)32FeO4

2H2O-3AsO4

32Fe(OH)3

5H (7)

When used for oxidation of AsIII to AsV, the reaction is very fastfor Cl2, O3and MnO4

compared to the H2O2and NH2Cl (Lee et al.,2003;Ghurye et al., 2004;Dodd et al., 2006). By using air and pureoxygen, about 5457% 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 andH2O2are sluggish in reacting with As

III while Cl2and O3react 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 ClO2must be kept low. For example, in the United States,

dosages ranging from 1 to 1.4 mg L1

are used mainly for the

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

III

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 As III oxidation with Cl2, O3, andFeO4

2 . However, specic 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 is

better to use ozonation rather than chlorination because O3reactsslowly 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 As III byMnO2 coated PEEC-WC nanostructured capsules and demon-

strated that when water contains a low concentration of As, theyhave a higher efciency than conventional oxidation methods(Criscuoli et al., 2012).

4.1.2. Photochemical oxidation

The most widely tested chemical oxidant in presence of natu-rally occurring iron in theeld 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: (FeOH2) 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 0.52.5, 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 system was also found to beuseful under natural water conditions (Hug et al., 2001). An oxi-dation of AsIII solution containing 0.065 mg L1 FeII and FeIII using90 W m2 UV-A light removed more than 90% of the 500 g L1

total As in 23 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) in water 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 Fe

III andH2O2 increased oxidation efciency but humic acid did not showany inuence on the oxidation. In order to achieve effective oxi-dation of AsIII, intense UV light source was used with the potassium

peroxydisulphate (KPS) system (Neppolian et al., 2008).



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4.1.3. Photocatalytic oxidation

The efcient 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 TiO2was investigated (Duttaet al., 2004,2005;Miller et al., 2011). The PCO of AsIII in suspen-sions with low TiO2loadings followed by subsequent adsorption ofAsV onto TiO2 surfaces in slightly acidic media reduced As con-centrations below the WHO drinking water limit of 10 g L1 in

water (initial [As]66.7 M) (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 inu-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 g1

TICB and 4925 mg AsVg1] followed by UV radiation compared tothe solution that was not exposed to UV light [2198 mg AsIII g1

TICB and 2050 mg AsVg1]. A nanocrystalline Al2O3and TiO2 im-pregnated chitosan for As removal was prepared byYamani et al.(2012)and their study proposed a mechanism where AsIII is photo-oxidized to AsV by TiO

2, then adsorbed by Al

2O3. When a very low

amount of TiO2 is present, the TiO2/UV system has an inefcientremoval 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, uoride, phosphate and humicacid (Guan et al., 2012). The PCO increases in the presence of NOMbut at higher concentration of NOM (215 mg L1), the adsorptionof AsVon the surface of TiO2decreases (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 modiesthe 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 oxidation

It 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 As

contaminated 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 MnIVand 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 lter medium, offering afavorable environment for the adsorption of As because As in theform of AsIII cannot be efciently 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 L1. The kinetics ofbacterial AsIII oxidation and subsequent removal of AsVby sorptiononto biogenic manganese oxides during ground water treatmentwas studied byKatsoyiannis et al. (2004). Their ndings 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 efciently 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 articial ground water with high As and iron concentra-tion through a mixture of 1.5 g iron llings and 34 g quartz sand

in 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 eld bySen 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 L1) 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. Saaleld 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 articial groundwater enriched with sul-phate (0.0810 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 oxidation

This 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 of

tube well water.



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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 hyperaccumulating

large 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 14427526 mg kg1 As infronds from contaminated soils, and up to 27,000 mg kg1 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-afnity chelating molecules present in the cytoplasm of the plant.For example, an arsenate activated glutaredoxin from the fern P.vittataL. 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 efciency 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 bene t 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, phytoltration, 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 is

possible to phytoremediate the As contaminated substrates. Phy-toltration involves several steps, (i) the selection of the mostpromising plants capable of removing the contaminant from waterand 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 bodies

using 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 (Spyrogiraspp.) (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 sh scales, coconut ber, dried roots

of 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. Coagulationocculation

In As removal processes, coagulation and occulation areamong the most common method ever employed. The addition ofa coagulant followed by the formation of a oc 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 occulated precipitate. The nalproducts are larger particles or oc, which settle under the inu-ence of gravity or ltered 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 ocs 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 ltration. 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 efciencies 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 and

Katsoyiannis, 2002;Han et al., 2002) and anionic (Zouboulis and



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Katsoyiannis, 2002) polymers for increasing As removal are lim-ited in numbers.Pallier et al. (2010)used kaolinite and FeCl3as acoagulant/occulent and they observed over 90% and 77% re-movals of AsVand AsIII, respectively, using 9.2 mg L1 of Fe3 . 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 L1

. Further their study suggests that thealuminum species regulates the removal of As and therefore, theefciency of As removal can be improved by adjusting the pH. Thepresence of sulfates signicantly 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/occulation 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. TiCl4successfully achieved high organicmatter removal to the same extent as Al and Fe salts and the re-sulting ocs are with better settleability. The most signicant ad-vantage of using TiCl

4as 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 satised the requirement of the World HealthOrganization's (WHO) guidelines (0.515mg L1) for drinkingwater (Ravenscroft et al., 2009). Therefore, TiCl4is expected to be apromising alternative coagulant for conventional Al and Fe salts.

4.4. Electrocoagulation (EC)

It is an alternative process to CF (coagulation/occulation). In-stead of adding a chemical reagent as ferric chloride, metallic ca-tions are directly generated in the efuent to be treated by ap-plying a current between iron electrodes to dissolve soluble an-odes. In EC, electrolytic oxidation of a sacricial 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 oc. 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 ocs 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 eld trials of two 100 L batch reactors inWest Bengal.

The ECAR was found to lower initial As concentrations as high

as 3000 g L1

to below the WHO-MCL of 10 g L1

, and easily

reached below 5 g L1. Strong oxidants produces during theFenton-type reactions were found to oxidize As III 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 sludge

production from 100 L reactor eld 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 efciency,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 (Gimnez 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 Ca2 competing for the adsorption sites (Giles et al.,2011,Zhu et al., 2013). There are some common adsorption tech-niques used for the efcient As removal from water are discussedbelow.

4.6.1. Activated alumina

Activated alumina (AA) is a physical/chemical process by whichions 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 rst 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)3at high temperature.It is a porous, granular material having typical diameter of 0.30.6 mm and a high surface area for good sorption properties. TheAA is used in packed beds to remove contaminants such as As,uoride, 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 lled, 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 56 whereasAsIII is best adsorbed at pH 78(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 signicant 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 of

and replaced with fresh media. Regeneration is achieved through a



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sequence of rinsing with regenerant, ushing 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 ow through the column resulting in high As con-taminated caustic wastewater. Onsite regeneration of AA mediatypically produces 3747 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 H2SO4rinse. It hasbeen reported that at pH range 2.811.5, 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 L1 (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-dened 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 g1 (Ma-cherey-Nagel, Germany) is 15.9 mg g1 (Lin and Wu, 2001). On theother hand, mesoporous materials have attracted much attentionin the eld of chemical engineering, optics, electromagnetics,biomedicine, industrial catalysis and adsorption, environmentalprotection, and fabrication of novel nano-object materials due tothe high surface area, well-dened and adjustable pore diameterof 250 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 Asremoval. Removal has been attributed to ion exchange, specicadsorption to surface hydroxyl groups or coprecipitation. Cur-rently, IBS products available in the market are granular ferrichydroxide (GFH), iron coated sand, modied 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 and

Gagnon, 2009;Burton et al., 2009). On the basis of chemistry ofthe remediation process, iron based technologies can be dividedinto two overlapping groups; rst 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 sufcient surface area (Smedley andKinniburgh, 2002). The mechanism of AsV adsorption on GFH wasstudied in detail byGuan et al. (2008), and they proposed that at7.4 pH, bidentate binuclear complexes with GFH are formed asevidenced by an average FeAsV bond distance of 3.32 by EXAFSanalyses. The impacts of temperature on adsorption kinetics and

equilibrium capacities for AsIII

and AsV

on GFH have been reported

byBanerjee 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 AsVand AsIII by GFH wasa spontaneous endothermic process.

Polymorphs of iron (III) hydroxy-oxide mineral [FeO(OH)] suchas goethite (a-FeOOH), akaganite (b-FeOOH) and lepidocrocite (c-FeOOH) have been described as good adsorbents for AsV. Pure

goethite was synthesized byMohapatra et al. (2007)with differentdopant cations (Cu, Ni and Co) for adsorption of AsV. With330 m2 g1 surface area, akaganite exhibited adsorption capacityas high as 120 mg g1 (Deliyanni et al., 2003). The adsorptioncapacity of akaganite (b-FeOOH) can be enhanced by isomorphicsubstitution of Fe3 by Zr4 on its structure (Sun et al., 2013). AtpH 7.0, the maximum adsorption capacity for AsV by akaganitewas 60 mg g1. Adsorption capacities for lepidocrocite have beenreported 25.17 mg g-1 with specic area of 103.9 m2 g1 (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 specic area and narrowpore size distribution. The uniqueness of this FeOOH polymorph is

that 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; Gimnez et al., 2007; Jnsson 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 found

applicable 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 t to use for the groundwater treatment. Under acidicpH conditions, 9699% 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 L1 solution of both species. However, inpresence of phosphate in the solution, the efciency of the ad-sorption process suffered. Less than 60% As uptake was achievedfrom the natural groundwater containing more than 5 mg L1

phosphate and 1.13 mg L1 of As. This is a practical problem to befaced in the eld 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 AsVper gram of catalyst (Luo et al., 2012) while in thepresence of Fe3O4/graphene composite, a high adsorption capacity(180.3 mg g1) for AsVwas achieved (Mishra and Ramaprabhu, 2012).Composites of iron oxides and TiO2 or Al2O3 presented adsorptioncapacity for AsV of 7.8 and 54.55 mg g1, respectively(D'Arcy et al.,2011; Basu et al., 2012)). On the other hand, natural iron oxidesshowed very low adsorption capacity (0.020.4 mg g1) because of itslow specic surface area(Zhang et al., 2004;Guo et al. 2007a,2007b;

Gimnez et al., 2010).



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13/24

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-bazitephillipsite, clinoptilolite, and volcanic glass. The As removalefciency for chabazitephillipsite was 6080% and for clin-optilolite-bearing rocks was 4060%. Removal of As is inunced bythree key factors rst, mineralogy of the zeolites occurring in the

volcanic 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 lters are

manufactured including Sono 3-Kolshi Filter, Granet Home-madeFilter, Chari Filter, Adarsha Filter, Sha Filter, and Bijoypur Clay/Processed Cellulose lters. The Garnet home-made lter containsrelatively inert materials like brick chips and sand as lteringmedia without added any chemical to the system. Air oxidationand adsorption on iron-rich brick chips and ocs of naturallypresent iron in ground water could be the possible reason for Asremoval from ground water. The study demonstrates that three-pitcher lters are an effective option as a short-term measure forAs removal. However, the three-pitcher lters that are not effec-tive option for long duration. The Chari lter also uses brick chipsand inert aggregates in different Charis as lter media. The Shaand Adarshs lters use clay material as lter media in the form ofcandle. Although the Sha lter was reported to have better Asremoval capacity but clogging of lter media is still a problem.Khair (2000)found that Bijoypur clay and treated cellulose werealso able to adsorb As from water. These units/lters 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). LaterNgai et al. (2006),Ahammedand Davra (2011)and Noubactep et al. (2012)made certain mod-ications to BSF design and lter 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 signicant pollutants removal for safe drinkingwater production. Kanchan arsenic lter (KAF), a saturated sandlter was designed byNgai 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 ltrations (Noubactep, 2010).

Metal oxides/hydroxides coated adsorbents such as iron oxidecoated sand, manganese-coated sand, Fe3 impregnated activatedcarbon, siderite-coated quartz and hematite coated quartz havebeen extensively used to improve the lter efciency (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 lter systems for As removal. Many emergent ad-sorbents being used in lter-based treatments are not sufcient toremove total As (AsIIIAsV) 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)andSheng et al. (2014)experimentally proved HBC tobe useful for the pollutants removal.Sheng et al. (2014)conrmed

that iron-amended HBC efciently removed AsV

(961.5mg g1

) 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 adsorbents

In last years, a wide variety of adsorbent systems have beendeveloped for the removal of As. Activated carbon (Huang and Fu,

1984), y ash (Diamadopoulos et al., 1993), aluminum-loaded corallimestone (Ohki et al., 1996), modied y 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), modied 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 efcient 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 g1). 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 g1 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 g1) was observed onactivated carbon byRajakovic (1992).Rajakovic (1992)found thatcarbon pretreated with Ag or Cu2 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 the

removal of As from drinking water. The maximum adsorption ca-pacities were 5.96 mg As g1 for AsIII and 5.15 mg As g1 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 ower-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 inuences 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 adsorbent

materials; consequently affect the adsorption performance in



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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 (Soa 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 signicantly 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), FeZr binary oxides (Ren et al., 2011), CeTioxide (Deng et al., 2010), Zr (IV)-loaded ligand exchange ber(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 CeO2and ZrO2might 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 afnity (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

, HAsO42 and AsO4

3) are effectively ex-changed with the anionic charged functional group of the resin,thus produces efuents 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:

HCrO44CrO4

24ClO44SeO4

24SO4

24NO3

4Br4

(HPO42, HAsO4

2, SeO32, CO3

2)4CN 4NO24Cl4

(H2PO4, H2AsO4

, HCO3)4OH4CH3COO

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 ushed with HCl so asto implant labile Cl on the surface of the resin, which later easilyexchanged with As. The AsVcan 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)andDonia et al.(2011) have reported As removal by using strong base anion ex-change resins. The efciency of ion exchange process is improved

by 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 efciency 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 inuent. Otherfactors include the afnity 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 signicant 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 be

produced.Hll (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 efuent is greater than the As levelin the inuent. 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, Fe III, 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 ne-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 thecleaning agent, several transport mechanisms (electroosmosis,electromigration and electrophoresis) and electrochemical reac-tions (electrolysis and electrodeposition) are induced (Acar andAlshawabkeh, 1993). When a direct electrical eld is applied acrossa wet mass of contaminated soil, the migration of non-ionic poreuids by electro-osmosis and the ionic migration of dissolved ions

take place towards the electrodes. This technique has certain ad-vantages over the conventional methods: (1) it is efcient in lowpermeability soil, which is difcult 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 ow 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-2HO2(g)2e

(8)

2H2O2e-2OH-H2(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 the

pH near the cathode also increases. Electro-kinetic remediationtechniques demonstrated 8595% As removal efciency, from low-permeability soils such as clay, peat, kaolinite, high-purity nequartz, Na and sand montmorillonite mixtures, as well as fromargillaceous sand (Yeung et al., 1997). However, kaolinite showedmore than 90% removal efciencies 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



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a catholyte chamber, showed the dominant mechanism for theremoval of diesel and compared them to those of electroosmoticow.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)andWong 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 L1

and in some cases to below 10 mg L1. It produces large residualvolumes and is more expensive than other As treatment technol-ogies. Researchers have explored various types of pressure drivenmembrane such as microltration (MF), ultra-ltration (UF), nano-

ltration (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 ow or solu-tion diffusion (Shih, 2005;Choong et al., 2007).

4.9.1. As removal using microltration

The MF is a low pressure driven membrane process for separ-ating colloidal and suspended particles in the range 0.110m.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 occulation 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 occulants and studied the removalrate, which was dependent on the adsorption of As on to the Fe III-complex and their subsequent removal from the solution. Theirresults showed that by the combination of occulation and MFtechnique, the As removal efciency 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 containingocs 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 4.010.0, negativelycharged AsV anions got effectively adsorbed by forming surfacecomplex while AsIII removal was poor because in the pH range4.010.0, 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-gulationmicroltration process. The size of As containing parti-cles increases via coagulationocculation 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 g L1 in presence of uoride and iron contaminant to a As

content of 8.7 g L1

. The EC experiment, consisted of a bath with

four aluminium sheets of 0.15 m0.05 m0.002 m, were continuedupto 45 minutes with a current density of 625 A m2.

4.9.2. As removal using ultraltration

Ultraltration (UF) is a size exclusion-based low pressure-dri-ven membrane separation process having pore sizes in the rangefrom 10 to 1000 and is capable of retaining species in the mo-lecular weight ranging from 300 to 5,00,000 Da. The rejection of

As by charged membrane explored the in

uence of co-occurringdivalent ions and natural organic matter (NOM). In presence ofdivalent cations such as Ca2 , Mg2 , AsV rejection reduced almostupto zero. This reduction in AsV rejection probably due to theformation of ion pairs between counter ions and the xed 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

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Review Arsenic 2015