Review of processes controlling arsenic retention and release in soils and sediments of Bengal basin and suitable iron based technologies for its removal

Author’s Accepted Manuscript Review of processes controlling Arsenic retention and release in soils and sediments of Bengal basin and suitable iron based technologies for its removal Tuhin Banerji, Komal Kalawapudi, Sudheer Salana, Ritesh Vijay www.elsevier.com/locate/gsd

PII: DOI: Reference:

S2352-801X(17)30223-0 https://doi.org/10.1016/j.gsd.2018.11.012 GSD177

To appear in: Groundwater for Sustainable Development Received date: 23 February 2018 Revised date: 18 September 2018 Accepted date: 22 November 2018 Cite this article as: Tuhin Banerji, Komal Kalawapudi, Sudheer Salana and Ritesh Vijay, Review of processes controlling Arsenic retention and release in soils and sediments of Bengal basin and suitable iron based technologies for its r e m o v a l , Groundwater for Sustainable Development, https://doi.org/10.1016/j.gsd.2018.11.012 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting galley proof before it is published in its final citable form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

Review of processes controlling Arsenic retention and release in soils and sediments of Bengal basin and suitable iron based technologies for its removal a

Tuhin Banerjia,b, Komal Kalawapudia*, Sudheer Salanaa, Ritesh Vijaya,b

Mumbai Zonal Centre, CSIR- National Environmental Engineering Research Institute, 89/B, Dr. A.B. Road, Worli, Mumbai- 400018, b Director’s Research Cell (DRC), CSIR-National Environmental Engineering Research Institute, Nehru Marg, Nagpur – 440020 *Corresponding author Phone: 022-24974607, Email: [email protected]

Abstract Arsenic in the soil environment has gained renewed interest because of the emerging cognizance that arsenic poisoning is a global concern. Groundwater in the Bengal Basin is significantly polluted by naturally occurring arsenic (As), a toxic metalloid, which adversely affects human health and among the countries facing As contamination problems, India and Bangladesh are the most affected. In soils and sediments, arsenic is often associated with Fe(III) (hydr)oxides and multiple processes/reactions govern its release into groundwater, including abiotic or biotically mediated oxidation-reduction and ligand exchange reactions. Reductive dissolution of arsenic-bearing Fe(III) (hydr)oxides and As(V) reduction to As(III) are the two main mechanisms controlling arsenic partitioning in soils, sediments and groundwater. Even though arsenic reduction is favourable over a wide range of conditions, Fe(III) reduction in nature is dependent on the biotic systems. This review reflects the current state of research for the understanding of arsenic in the soil environment with an emphasis on iron based technologies for its removal. It attempts to collate all the relevant literature such that it can be a useful resource for researchers or policy makers to help recognize and explore useful treatment options.

Graphical Abstract:

Keywords: Bengal basin, anoxic phase, arsenite oxidation, iron, arsenic removal Introduction The costs of ingestion of arsenic for human health can be extremely serious, and may extend from general malaise to death. Health effects of arsenic arise from its use as a poison, its use in medicine, its inclusion in various manufactured products, accidental exposure from industrial and even from naturally occurring arsenic in soil and water (Hughes et al., 2011; Mandal and Suzuki, 2002; Ng et al., 2003; Hughes, 2002; Aposhian and Aposhian, 2005; Basu et al., 2001; Saha et al., 1999, Kapaj et al., 2006). The term arsenicosis characterises the various clinical manifestations caused by chronic arsenic toxicity due to ingestion of arsenic over a long period, but, there is no standard definition of the symptoms of arsenic poisoning (Guha Mazumder, 2003). Acute arsenic poisoning can be fatal, but sudden demise is hardly

ever caused by environmental sources (Ravenscroft et al., 2009). The chronic poisoning impacts health by imposing dermatological effects, carcinogenic effects and systemic noncarcinogenic effects. Often, only the dermatological manifestations can be traced definitely to arsenic poisoning in individuals, the others being inferred from epidemiological studies (Ravenscroft et al., 2009). Bangladesh was ranked 12th worldwide on the United Nations Development Programme Human Poverty Index (UNDP, 1997). Unfortunately, the health effects of arsenic exposure increase poverty by decreasing productivity (Mathieu, 2008). For instance, arsenic-induced skin lesions, which primarily form on the hands and feet, are painful and make manual labour difficult. Chronic illnesses, such as cancers, can make work impossible and ultimately shorten lifespans. The net economic benefit of reducing drinking water arsenic concentrations to 50 ppb is $7/month/household in West Bengal, India (Roy, 2008). One main reason is that poor nutrition correlates with a likeliness to show symptoms of arsenic poisoning (Maharjan et al. 2007; Islam et al. 2004). In addition, the poor have less financial means to seek alternative water sources or receive medical treatment for symptoms of arsenic exposure. Thus, the poor suffer most from arsenic poisoning. Much research has been carried out to understand the precise cause of presence of arsenic in the groundwater (Bhattacharya et al. 1997; Ahmed et al. 2004; Ahmad et al. 2017; Ravenscroft et al. 2005; Kocar et al. 2008; Polizzotto et al. 2005; Polizzotto et al. 2008; Smedley and Kinniburgh, 2002, Nickson et al. 2000; Nickson et al. 1998; Harvey et al. 2002; McArthur et al. 2001; Zobrist et al. 2000, van Geen et al. 2004), but there is little clarity in the various approaches adopted or the causes elucidated. An attempt has been made here to review the present knowledge base and present it in a very concise manner so as to make the reader aware of the possible causes of arsenic presence in groundwater of poverty stricken areas of Bengal Basin. Also, various technologies (Mohan and Pittman, 2007; Sarkar et al. 2005; Misra et al. 2010; Sarkar et al. 2010; Sarkar et al. 2012; Padungthon et al. 2014; Padungthon et al. 2015; Biswas et al. 2012; von Brömssen et al. 2007; Hossain et al. 2014; Ahmad et al. 2017) exist for arsenic removal or for availing arsenic free drinking water but their suitability to all locations is not well defined (Hug et al. 2008; Sarkar et al. 2010). An attempt has been made here to

review available technologies for arsenic removal which would be suitable for the rural areas of the Bengal basin. Causes of Arsenic retention and release in Bengal basin soils and sediments Arsenic contamination in the environment is predominantly geologic (Bhattacharya et al. 1997; Ahmed et al. 2004) but sometimes increased by anthropogenic activity, including arsenic containing pesticide application and mining and smelting (Smedley and Kinniburgh, 2002). Several biogeochemical processes have been hypothesized so far to explain arsenic mobilization into environment, including arsenic pyrite oxidation due to fluctuation of the water table (Chowdhury et al., 2000), reductive dissolution of arsenic bearing Fe(III) (hydr)oxide minerals (Smedley and Kinniburgh, 2002; Nickson et al. 2000; Nickson et al. 1998; Harvey et al. 2002; McArthur et al. 2001; Zobrist et al. 2000) with concurrent arsenate reduction to arsenite, arsenic desorption at high pH > 8.5 (Darland and Inskeep, 1997), chemical weathering of silicate minerals (Neumann et al. 2010), ion exchange by phosphate (Darland and Inskeep, 1997), by carbonate (Apello et al. 2002), by silicic acid (Waltham and Eick, 2002), and reduction of Fe(III) (hydr)oxide minerals by organic carbon (Redman et al. 2002). The conceptual geochemical models are not yet well understood as arsenic adsorption occurs on many solid phases other than oxyhydroxides, like magnetite, siderite and apatite (Neumann et al., 2010; Swartz et al., 2004; Polizzotto et al. 2006). Fe(III) (hydr)oxide minerals are considered as the major arsenic bearing minerals in Bengal Basin as well as other basins of South and Southeast Asia (Ravenscroft et al. 2005; Kocar et al. 2008; Polizzotto et al. 2005; Ahmed et al. 2004; Bhattacharya et al. 1997; Polizzotto et al. 2008). In contrast, geological evidence of very little Fe(III) (hydr)oxide minerals deposition in Ganges Delta sediments was also explained (McArthur et al. 2004). The predominance of available experimental evidence supports the oxidation of natural organic matter (NOM) with associated reduction of arsenic bearing Fe minerals cause mobilization of arsenic and Fe in Holocene aquifers (Smedley and Kinniburgh, 2002, Nickson et al. 2000; Nickson et al. 1998; Harvey et al. 2002; McArthur et al. 2001; Zobrist et al. 2000, van Geen et al. 2004). Microbial organic matter decomposition and an anoxic environment are the two preconditions for reductive dissolution of arsenic in Holocene sediments (Hossain et al. 2014; Biswas et al. 2012). An anoxic environment has been generated naturally in Holocene

aquifers by sediment burial (Benner and Fendorf, 2010). Once the microbial O2 demand exceeds O2 diffusion rates under flooded conditions this leads to use of alternative electron acceptors, starting with nitrate and manganese oxides to arsenic (arsenate to arsenite) and Fe [Fe(III) to Fe(II)] reduction. Organic carbon and Arsenic cycling Bengal basin grey aquifer sediments are high in natural organic matter (NOM) content (Ravenscroft et al. 2005; Biswas et al. 2012). Solid phase arsenic content in aquifer sediments is strongly correlated with Fe and organic carbon (Meharg et al. 2006). Importance of NOM in arsenic retention and mobilization has been studied over the last decade. The source of NOM is still a point of controversy. Three major hypothesis have been postulated in literature so far, (i) that organic carbon leaching from nearby peat lenses (McArthur et al. 2004), (ii) that surfaced organic carbon from rice field, organic rich pond and river sediments is drawn down during groundwater recharge due to groundwater irrigation for agriculture in dry season (Harvey et al. 2002; Neumann et al. 2010), and (iii) co-deposition of organic carbon and arsenic (Meharg et al. 2006). Decomposition and reframing of biogenic matter from plants, animals and microbes leads to a complete mixture of varied organic molecules termed as NOM (Wang and Mulligan, 2006). The properties of NOM depend on qualities of the original material and the conditions and processes during its transformation. In water, NOM is found as dissolved molecules, colloids, and particles and NOM can be inter-converted among these forms by different chemical processes including dissolution/precipitation, sorption/desorption, aggregation/disaggregation (Perdue and Ritchie, 2004). Dissolved Organic Carbon (DOC) is the water soluble fraction of organic molecules that passes through a 0.45 μm filter. DOC includes molecules of different molecular weight and chemical structure, such as sugars, amino acids and refractory humic substances. The DOC concentration in natural water is highly variable depending on hydrological and biogeochemical conditions. Dissolved NOM comprises majorly of humic substances owing to its recalcitrant nature and functional groups with multifarious properties. Thus, these are important reactive species occurring in natural waters which significantly affects the biogeochemistry of metals and trace elements. A low concentration of DOC in groundwater (0.7 mg C/L) compared to river water (6.8 mg C/L) was also observed (Perdue and Ritchie, 2004).

Arsenic redox transformation and organic carbon

Reductive dissolution of Fe(III) (hydr)oxides and oxidation of sulphur containing minerals are the two major redox reaction based hypothesis causing Arsenic release from solid phase (Benner and Fendorf, 2010). However, research findings over the last two decades have transferred the focus to organic carbon driven biogeochemical reduction of Arsenic bearing oxyhydroxide minerals with concomitant dissolution of arsenite. Pollutant degradation and metal speciation are strongly influenced by redox reactions using DOC as the electron donor (Redman et al. 2002). Redox active functionalities of DOC are highly variable in its redox potential under different environmental conditions. DOC react by reductive dissolution of Fe(III) (hydr)oxides with coupled As release (Harvey et al. 2002; McArthur et al. 2004; Kumar et al. 2016) and, therefore, are a part of the reactions related to As speciation and transformation. The redox speciation of inorganic elements like Fe and Cr by DOC is found in some previous reports (Redman et al. 2002; Buschmann et al. 2005; Buschmann et al. 2006; Palmer et al. 2006). The microbial influence in redox reactions releasing As into environment is well documented (Kumar et al. 2016; Zobrist et al. 2006; van Geen et al. 2004; Rowland et al. 2006; McLean et al. 2006). The microbial processes require an electron donor, such as DOC (Charlet and Polya, 2006). Arsenic release shows positive correlations with DOC concentration in the aqueous environment for simulations of microbial action (Harvey et al. 2002; Rowland et al. 2007). The nature, availability and concentration of DOC strictly control arsenic mobilization by microbial redox transformation (Postma et al. 2007). In contrast, it was also reported that desorption of arsenic from sediments collected from 30 m depth is an abiotic chemical process (Polizzotto et al. 2006). Also the variation in colour of the sand sediments from grey to brown due to microbial action on the organic content gave reliable inference regarding the presence or absence of Arsenic in the groundwater (Biswas et al. 2012; Hossain et al. 2014; Ahmad et al. 2017).

Sorption/desorption of Arsenic and organic carbon

Organic carbon is considered to have strong influence on the adsorption/desorption of Arsenic on clay mineral and metal oxides surfaces causing Arsenic mobilization/transport in the environment (Redman et al. 2002; Bauer and Blodau et al. 2006; Sharma et al. 2011).

NOM functional groups are capable of sorption to Fe oxides both in their outer sphere and inner sphere complexes or surfaces of colloidal clay (Kumar et al. 2016; Biswas et al. 2012; von Brömssen et al. 2007). Humic anion sorption on mineral surfaces restrain arsenic sorption or induces arsenic desorption in the system resulting in arsenic mobilization (Redman et al. 2002; Grafe et al. 2002). Besides this process, arsenic association with colloidal particles has also been reported (Tadanier et al. 2005). Arsenic distribution in aqueous environment can be controlled by other complexation mechanisms including covalent bonds between Arsenic and humic molecules (Buschman et al. 2006) and ternary complexes between NOM, Fe and Arsenic (Redman et al. 2002; Sharma et al. 2011). Fe ions in aqueous solution bridging Arsenic with NOM has been recently studied by Sharma et al. (2010, 2011) who reported ~ 94 % arsenic associated with colloidal Fe-NOM and the rest ~6 % arsenic in the dissolved form.

Community vs Domestic Arsenic Removal Technologies

An arsenic removal plant (ARP) can be either community scale unit [eg. Amal filter (Sarkar et al. 2005)] or a domestic unit [eg. Kanchan filter (Ngai et al., 2007), Sono filter (Hussam and Munir, 2007), Naval Materials Research Laboratory-Defense Research and Development Organization (NMRL-DRDO) filter (Misra et al., 2010)]. The average total (direct+indirect) water intake in Bengal basin is approximately 23.12 L/family/day (Hossain et al. 2013). The community scale units are generally capable of catering to 200 families’ daily water requirements (Sarkar et al. 2010) whereas each domestic unit serves only one family. Thus to cater the same number of families, 200 domestic units would have to be installed. In developing countries, a major issue of installation of treatment plants and their operation and maintenance is the high cost. (Reisinger et al. 2005). Treatment plants installed by public authorities is an alternative which serves as a centralised facility but the distribution of treated water to each and every user remains the point of concern. Therefore, there is a rising interest in developing low cost treatment units like domestic filters, which can be installed at the users’ locations (Banerji and Chaudhari, 2017). While domestic scale units are initially easier to install compared to community scale systems, the latter provide a more sustainable solution with respect to robust and regenerable adsorbent media, collection and regeneration of exhausted media and ecologically safe containment of removed arsenic in addition to its

potential for economic growth (Sarkar et al. 2010; Banerji and Chaudhari, 2017). It is known that domestic units have drawbacks namely – bothersome maintenance, high costs, insufficient treatment rate, and/or reliance on materials unavailable in remote villages (Ngai et al. 2007). Thus regardless of the design/system of domestic arsenic removal, reliability and quality control is necessary and this can be attained by regular analysis of treated water for pH, arsenic , TDS, phosphate etc. Community scale units would require fewer samples for monitoring compared to domestic units, thus water quality monitoring is easier in community scale in comparison to domestic units (Banerji and Chaudhari, 2017). Also, it is comparatively more manageable to introduce modifications and innovations in well-head or community scale units for performance improvement (Sarkar et al. 2010). Domestic units using adsorbent media or coagulants will always generate arsenic-laden sludge or solids (Bordoloi et al. 2013; Qiao et al. 2012). Coordinating collection and safe handling and disposal of sludge or used media from individual households poses a level of intricacy and enforcement effort which is difficult to sustain in far-flung villages. The community scale systems reduce such management problems (Sarkar et al. 2012). Both capital and operating costs of the community scale systems are significantly lower than the total sum of equivalent two hundred or more domestic units (Sarkar et al. 2010; Banerji and Chaudhari, 2017). Because of these reasons, interviews with a spectrum of stakeholders from government officials to householders have shown that community water treatment systems were preferred to household/domestic systems (Johnston et al. 2014; Khan and Yang, 2014).

Technologies for Arsenic Removal

Arsenic concentrations in groundwater are greatly variable laterally on large (BGS, 2001) and small (van Geen et al. 2003) spatial scales but exhibit a consistent vertical profile, i.e. high arsenic (>50 ppb) rarely occurs at depths >150 m below ground surface (BGS, 2001; van Geen et al. 2003; Harvey et al. 2002; Ravenscroft et al. 2005). Installing deeper household wells (presently most are shallower than 100 m), has been suggested as an easy way out that does not require expensive and large infrastructure (Opar et al. 2007; Ahmed et al. 2006; Ahmad et al. 2017). Deep wells would initially provide low-arsenic groundwater and exhibit the least disease risk of domestic alternatives (Howard et al. 2006). Though this would provide low-arsenic water for the short term, arsenic is transported by groundwater and thus,

arsenic may migrate downward, eventually contaminating the deeper water leading to arsenic in drinking water (Michael and Voss, 2008; Burgess et al. 2010). It is also possible to achieve <10 ppb As(tot) in treated water by simply using the groundwater iron to precipitate along with arsenic and filter the precipitates through a sand filter (Voegelin et al. 2014; Nitzsche et al. 2015). But this approach would not work in Bangladesh and the Ganga Meghna Brambhaputa (GMB) plain as the average arsenic and phosphate concentrations found in these regions are quite high and the amount of iron required to remove arsenic to <10 ppb is not present in groundwater of this region (Hug et al. 2008). Researchers have evaluated arsenic removal by sand filters typically used for iron removal in Vietnam and they mentioned that Fe/As ratios of ≥50 or ≥250 were required to ensure arsenic removal to levels below 50 or 10 ppb, respectively (Berg et al. 2006). However, closer analysis of their data revealed that few of the sand filters could achieve less than 10 ppb for Fe/As ratio of around 10. Thus, it seems that in a continuous or semicontinuous flow system through granular media, possibly there are conditions under which arsenic removal by iron gets enhanced (Banerji and Chaudhari, 2017; Chaudhari et al. 2014). Whilst the best solution to the arsenic problem is to switch over to treated surface water which does not have any arsenic contamination but development and maintenance of surface water based drinking water system for the entire population (including household connections) is costly, time-consuming, requires more cumbersome maintenance and investment-intensive (Sarkar et al., 2010). Another option can be the demarkation of safe aquifer(s) that can be drilled for tubewell installation (Hossain et al. 2014; Biswas et al. 2012; von Brömssen et al. 2007; Ahmad et al. 2017). But this option also cannot be safely applied without adequate training and supervision. Hence, with all these complications, it is unlikely for developing countries like India and Bangladesh to switch over to surface water or demarcate safe groundwater aquifers in a short time. Thus, to meet the drinking water requirements till such a switch is possible, construction of arsenic removal plants (ARPs) is required. The established arsenic removal processes can broadly be classified into three process types; co-precipitation, membrane and adsorption. But these methods have inherent problems as outlined in Table 1. Thus new processes need to be developed which would address these issues.

Iron Based Arsenic Removal System

Iron has a high affinity for arsenic and, in most cases, arsenic in nature is present bound to iron minerals. Iron based arsenic removal technologies make use of the strong natural chemical association of arsenic with iron, removing arsenic by adsorption processes (Yuan et al. 2002; Sylvester et al. 2007; Badruzzaman et al. 2004), or co-precipitation using ferric or ferrous salts (Fuller et al. 1993; Meng et al. 2000; Roberts et al. 2004), with many systems reporting >90% arsenic removal from treated water. A range of inexpensive, iron based, water purification technologies have been developed to address the problem of arsenic contamination in groundwater. Many authors have examined the adsorption of arsenic by iron oxides (Ahmad et al. 2017; Smedley and Kinniburgh,2002;Mohan and Pittman,2007), highlighting the tendency of arsenic to strongly bind to hydrous ferric oxides (HFO) (as monodentate or bidentate inner sphere complexes), even at very low aqueous arsenic concentrations. The term HFO is quite loosely used in literature and it generally means any partially or fully hydrolysed precipitates of Fe(III) formed in the absence or presence of contaminant (Wilkie and Hering, 1996; Jang et al. 2008; Hering et al. 1997). A detailed review of the performance of many arsenic removal technologies, including those utilizing iron as an adsorbent or precipitant, is already in literature (Mohan and Pittman, 2007). Most iron-based treatment methods are more effective in removing As(V), rather than the more toxic As(III) species, as in aqueous phase with pH 8 and lowermost As(III) exist as neutral molecule H3AsO3, and so require oxidation as a pre-treatment (Sorlini and Gialdini, 2010). The urge to innovate and upgrade has led to designing robust, effective and economic devices for household use leading many researchers to examine the use of low-cost, local and/or waste materials such as iron-coated sand, cast iron filings, steel wool, minerals and amended blast furnace slag which may be used in filters for arsenic removal (Joshi and Chaudhuri, 1996; Leupin et al. 2005; Banerjee et al. 2008). Only groundwater iron is also being used for arsenic removal in domestic Vietnamese sand filters (Berg et al. 2006; Voegelin et al. 2014; Nitzsche et al. 2015). Many of the adsorbent based technologies, do however show significant potential for effective removal of arsenic from contaminated drinking water(Sarkar et al. 2005;

Misra et al. 2010; Sarkar et al. 2010; Sarkar et al. 2012;

Padungthon et al. 2014; Padungthon et al. 2015; Ahmad et al. 2017) but may need to be relooked at considering the specific requirements of rural areas (SenGupta et al. 2017; Hossain et al. 2014).

Effects of Phosphate, Silicate and NOM on Arsenic Removal by Iron Based Technologies

The arsenic contaminated areas in India are primarily agricultural and, as phosphatic fertilizer use is quite abundant, it is generally found in the

groundwater along with arsenic

(Kinniburgh et al. 2003). Phosphate and As(V) are structurally quite similar (Kish and Viola, 1999). The affinity of phosphate for sites on Fe(III) precipitates is similar to arsenate (Manning and Goldberg, 1996; Ryden et al., 1987; Gao and Mucci, 2001). The influence of phosphate on adsorption of As(III) and As(V) on ferrihydrite as a function of pH for P:As(tot) ratios of 1:1 and 10:1, which is within the range reported in Bangladesh (BGS, 2001), has been studied (Jain and Loeppert, 2000). They found that adsorption of both As(V) and As(III) decreased with increasing PO4-P concentration, and for As(V) this was significant over the entire pH range. They also report that neutral H3AsO30, dominant at pH >7, is better able to compete with negatively charged phosphate species. Other researchers have found similar effects (Hongshao and Stanforth, 2001; Elias et al. 2012; Biswas et al. 2014). It has been reported that phosphate is the main factor limiting arsenic removal in Fe(II) coprecipitation systems (Roberts et al. 2004). Further, presence of phosphate would cause requirement of more time for Fe(II) oxidation as phosphate can complex with ferrous iron (Pryor and Cohen, 1951). Phosphate adsorption kinetics often follows a complicated pattern, with a rapid initial adsorption followed by a slow reaction that does not reach equilibrium up to the end of the experiment (Torrent et al. 1992). During adsorption of phosphate and arsenate on goethite, it was found that at a given total anion concentration, the amount adsorbed was greater in the mixed ion system than in the single ion samples, implying that there were sites on the surface that were specific for each ion as well as some nonspecific sites on which both ions could adsorb (Hingston et al. 1971). As(III) and As(V) removal by adsorption on ferric phosphate has also been reported (Lenoble et al. 2005) but arsenic removal in their system was most probably occurring by chemical equilibrium reactions as phosphate was observed to get leached out. Silicate adversely affects arsenic removal by Fe(III) precipitates (Meng et al. 2000; Meng et al. 2002; Bang et al. 2005; Biswas et al. 2014). The iron oxide/hydroxide surface has a strong affinity for silicate (Sigg and Stumm, 1981), however silicate adsorption is weaker than As(V) adsorption (Swedlund and Webster, 1999). Silicic acid (H4SiO4) has been shown

to effectively compete with arsenic for adsorption sites (Davis et al., 2001) or may polymerise and coat the adsorption sites thereby reducing access to the adsorbate, hence reducing the adsorption capacity of iron oxides (Christl et al. 2012). As(III) removal is not affected by silicate during co-precipitation with Fe(II) but is negatively affected during coprecipitation with Fe(III) as during co-precipitation with Fe(II), As(III) was oxidized and this did not happen during co-precipitation with Fe(III) (Roberts et al. 2004). Co-precipitation with Fe (II) is similar to electrocoagulation with Fe electrodes ECFe (Banerji and Chaudhari, 2016). Similarly, silicate has very little effect on arsenic removal by ECFe as in this system As(III) is oxidized and hence it was not showing any significant effect up to 20 mg/L silicate (Wan et al. 2011; Banerji and Chaudhari,2016; ). Other researchers have reported high concentrations of silica polymerize to physically block access to sorption sites and not through competitive sorption (Zeng et al., 2008). It has also been reported that silica significantly slowed Fe(II) oxidation to Fe(III) in aerated water at pH 6.5 (Rushing et al. 2003). Natural Organic Matter (NOM) has been reported to cause reduction in arsenic removal efficiency by zero valent iron based processes (Giasuddin et al. 2007; Mak et al. 2009; Rao et al. 2009; Mak and Lo, 2011; Tanboonchuy et al. 2012 ). Arsenic removal by Fe was inhibited in the presence of humic acid (HA) (representative NOM compound) due to the formation of soluble Fe-humate in aqueous solutions which suppressed the formation of ferric hydroxide (Tang et al. 2014). More Fe was required for arsenic removal when HA concentrations were higher. When no more complexation with HA and dissolved Fe was possible, further Fe(II) dissolution produced Fe(III) (hydr)oxide which adsorbed arsenic (Giasuddin et al. 2007; Mak et al. 2009; Rao et al. 2009;). Oxidation of NOM can be achieved by Fe(IV) (Graham et al. 2010) which would reduce As(III) oxidation and thus may reduce arsenic removal.

Chemical Oxidation of As(III)

Adsorption capacity of Fe(III) hydr(oxide) for As(V) is much higher than As(III) (Hering et al., 1996), but significant fraction of As(III) occurs in natural ground water (Mukherjee and Bhattacharya, 2001), leading to reduced arsenic removal and thereby non-compliance of drinking water standards. Moreover, As(III) is more toxic than As(V), hence its removal is

of prime importance (Korte and Fernando, 1991). Henceforth, as an initial step, oxidation of As(III) to As(V) by chemical oxidants has been suggested (Ghurye and Clifford, 2004). As(III) can be oxidized by strong oxidants such as hydrogen peroxide, ozone, chlorine, potassium permanganate or ferrate (Kim and Nriagu, 2000; Lee et al. 2003; Bissen and Frimmel, 2003; Ghurye and Clifford, 2004; Sorlini and Gialdini, 2010; Lescano et al. 2011;). The possible oxidants along with their benefits and drawbacks are listed in Table 2.

Co-Oxidation of As(III) to As(V) during Fe(II) Oxidation

Oxidation of As(III) to As(V) was found during the corrosion of iron powder and filings under aerobic conditions within 4–7.5 days (Manning et al. 2002). Various studies report that arsenic is oxidized in parallel to the oxidation of dissolved Fe(II) by DO, most likely by reactive intermediates such as Fe(IV) (Hug et al. 2001; Hug and Leupin, 2003; Roberts et al. 2004). Studies show that As(III) is oxidised during Fe(0) corrosion on a time scale of several hours (Leupin and Hug, 2005). Overall reactions proposed by them include: Fe0 + 2H2O + ½ O2 → Fe2+ + H2O + 2OH-

(1)

Fe2+ + ¼ O2 + H2O → Fe3+ + ½ H2O + OH-

(2)

Fe3+ + 3H2O → Fe(OH)3 + 3H+

(3)

As3+ + intermediates (.OH, Fe4+) → As4+

(4)

As4+ + O2 → As5+ + .O2-

(5)

The Fe(II)-catalyzed As(III) oxidation mechanism has been described in detail elsewhere (Hug and Leupin, 2003). Many other researchers have postulated the formation of Fe(IV) and As(IV) as intermediates of oxidation of As(III) to As(V) during oxidation of Fe(II) to Fe(III) as detailed below. The sequence begins with a series of one-electron transfers in which molecular oxygen is reduced to superoxide and hydrogen peroxide as Fe(II) is oxidized to Fe(III) after Fe(II) is dissolved from the anode in ECFe or is added in systems with Fe(II) dosing, as outlined in Supporting Information T1. The reaction of Fe(II) and Fe(OH)+ with hydrogen peroxide

occurs producing the intermediates (INT) which quickly degrade to form a reactive oxidizing species,

Fe(IV).

The

nature

of

the

intermediates

has

been

suggested

to

be

[Fe(OH)(HO2)(H2O)4]+ and Fe(IV) thought to exist as [Fe(OH)3(H2O)4]+ (Bossman et al. 1998). These two species interconvert rapidly enough such that they are in equilibrium. At acidic pH, the hydroxyl radical is formed along with Fe(III) (Rigg et al. 1954). Since the pH of groundwater is typically higher (pH 6 – 8), Fe(IV) species is of greater importance as the oxidising radical in the arsenic removal systems using Fe based technologies (Hug and Leupin, 2003). The Fe(IV) species is less prone to scavenging by dissolved organic species than the hydroxyl radical (Hug and Leupin, 2003). Another sequence of one-electron transfers after the formation of the Fe(IV) species oxidizes As(III) to As(IV) as Fe(IV) is reduced to Fe(III), and then As(IV) is oxidized to As(V) as molecular oxygen is reduced to superoxide or might be more Fe(IV) oxidizes this As(IV) species to As(V) (as shown in Supporting Information T1). Alternatively, the Fe(IV) can also be reduced by Fe(II) to form Fe(III) as Fe(II) is oxidized to Fe(III). Most possible As(III) and Fe(II) oxidation reactions have been tabulated in Supporting Information T1. Oxidizing potential of Fe4+aq/Fe(III)aq ≈ 1.4 at pH 7 (Bossman et al. 1998 and the references therein). Addition of Fe(II) in multiple doses with sufficient time for Fe(II) oxidation leads to higher fractions of oxidized As(III) than one single addition of the same total amount of Fe(II) (Roberts et al. 2004). The reason is that Fe(II) itself consumes oxidizing species and outcompetes As(III) at the usually higher Fe(II) concentrations (Leupin and Hug, 2005). Thus, the effect of Fe(II) concentration on the overall process is difficult to predict quantitatively. The oxidation of As(III) is optimal with prolonged low concentrations of Fe(II), which is continuously oxidized by dissolved oxygen as in ECFe. During ECFe as Fe(II) is continuously added, the reactions forming Fe(IV) which led to As(III) oxidation are expected (Banerji and Chaudhari, 2016). Accordingly, As(III) oxidation during ECFe was negligible if Fe(II) oxidation was suppressed by the addition of 3 mM 2,2′-bipyridine and that in ECFe systems Fe(IV), and not OH•, is the most likely oxidant for As(III) as negligible effect of 2-propanol on As(III) oxidation was observed (Li et al. 2012). Adsorption vs Co-precipitation Literature dealing with arsenic removal using iron based technologies generally refers to two possible routes. They use either co-precipitation with Fe(III) precipitates or adsorption onto

pre-formed adsorbents. In co-precipitation, Fe(III) precipitates are formed in the presence of arsenic oxyanions leading to incomplete hydrolysis of Fe(III) i.e. formation of fractions of FeOH2+, Fe2(OH)24+ , Fe(OH)2+ , Fe(OH)30, and Fe(OH)4- and these hydrolysed Fe(III) forms then complexed with arsenic (Fuller et al. 1993; Stefánsson, 2007). During adsorption onto pre-formed adsorbent, iron (hydr)oxides are formed prior to the addition of arsenic. The adsorbent ageing period before the arsenic is added allows some time for aggregation of the adsorbent and crystallite growth to occur, both processes which tend to decrease the number of adsorption sites (Fuller et al.; 1993, Waychunas et al. 1993). If, however, a sorbing ion is present in solution during the hydrolysis and precipitation of Fe(III) (hydr)oxides (as in coprecipitation), then the ion may sorb to surface sites before aggregates are formed. This maximizes the number of available surface sites, increasing the sorptive capacity of the Fe(III) (hydr)oxides (Fuller et al.1993). In a situation where there is continuous in-situ formation of adsorbent in presence of the adsorbate a situation may arise when the adsorbate, due to less concentration, starts to adsorb and not co-precipitate (Fuller et al. 1993; Waychunas et al. 1993; Crawford et al. 1993a,b; Opiso et al. 2009; Wogelius, 2013; Chaudhari et al. 2014; Park et al. 2016). It is quite difficult to differentiate at which concentration adsorption predominates and at which concentration co-precipitation predominates (Chaudhari et al. 2014; Banerji and Chaudhari, 2016). Use of Zero Valent Iron for Arsenic Removal The use of zero valent iron (ZVI), i.e., Fe(0) to remove arsenic has been investigated by many groups (Farrell et al. 2001; Manning et al. 2002; Melitas et al. 2002; Nikolaidis et al. 2003; Bang et al. 2005; Katsoyiannis et al. 2008). The exposed surface area of ZVI plays a major role in both the removal kinetics and capacities. Corrosion of ZVI and subsequent oxidation of Fe(II) by dissolved oxygen produces Fe(III) precipitates which can complex with As(V) (Bang et al. 2005). Arsenic forms mono- and bidentate complexes with iron oxides produced from iron corrosion (Lackovic et al. 2000; Farrell et al. 2001; Kumar et al. 2014). The rate of Fe(III) (hydr)oxide formation from ZVI is directly related to the corrosion rate and higher corrosion rate leads to higher amount of Fe(III) (hydr)oxide which leads to more arsenic removal (Bang et al. 2005). Once Fe(II) is produced from ZVI the Fe(II) and As(III) oxidation and complexation/precipitation reactions

are the same as ECFe. Only difference between ECFe and ZVI is that Fe(II) dissolution in ECFe is controlled by current and is uncontrolled in ZVI. The arsenic removal rate of ZVI is dependent on dissolved oxygen concentration as well as pH, with the required contact time for removal increasing significantly as DO is decreased or pH is increased (Bang et al. 2005; Lackovic et al. 2000). It was reported that after 9 hours of reaction with 1 g/L ZVI filings in presence of DO, 99.8% removal was achieved for initial As(V) concentrations of 100 µg/L at pH 6, while only 55.5% and 2% were removed at pH 7 and 8 respectively and 99.8% of As(V) was removed at pH 6 in solution open to air compared to only 9% in pH 6 anoxic solution (Bang et al. 2005). Under anoxic conditions, iron oxide such as magnetite (Fe3O4) can be formed on the Fe(0) particle surface (Kumar et al. 2014). The reduction of As(III) and As(V) to As(0) by ZVI is thermodynamically favourable (Bang et al. 2005). However, it was also reported that no measurable reduction of As(V) to As(III) was observed (Farrell et al. 2001; Manning et al. 2002). Nanoscale (1–120 nm diameter) zero valent iron for rapid, first order As(III) and As(V) removal (kobs= 0.07–1.3 min-1) with innersphere surface complexation mechanism was synthesized and this rate was found to be about 1000 times faster than that of micron-sized iron (Su and Puls, 2001). It is important to note that the corrosion rate is very sensitive to surface area. In the limit of a low ratio of adsorbing species to adsorption sites, the arsenic removal rate is limited by mass transfer to adsorption sites, and therefore exhibits first order dependence on arsenic concentration (Farrell et al. 2001; Melitas et al. 2002). However, at high solution concentrations, removal becomes limited by the rate of adsorption site generation (related to the corrosion rate and Fe(III) (hydr)oxide formation rate) and is zero order in arsenic concentration (Farrell et al. 2001; Melitas et al. 2002). With 200 days of operation 18,000 mg arsenic [1:3 ratio of As(III):As(V)] was reported to be removed using 400 g ZVI filings – a capacity of 45 mg As/g Fe (Bang et al. 2005). However, breakthrough of arsenic began to occur after only 100 days, indicating that filings would have to be replaced long before the capacity was reached. Various researchers have also mentioned the possibility of oxidation of As(III) during arsenic removal by ZVI (Leupin et al. 2005; Katsoyiannis et al. 2008; Banerji and Chaudhari, 2016; Banerji and Chaudhari, 2017).

Arsenic Removal by Electrocoagulation with Iron Electrodes (ECFe)

Electrocoagulation using iron electrodes (ECFe) is an emerging technology which has shown lot of promise for arsenic removal. Arsenic removal efficiency increases in the sequence aluminium < titanium < iron when using monopolar electrodes (Kumar et al., 2004). It has been reported by some authors that As(III) oxidation is possible in ECFe (Kumar et al. 2004; Lakshmipathiraj et al. 2010; Wan et al. 2011; Banerji and Chaudhari, 2016) while others report on the contrary (Lakshmanan et al., 2008; Lakshmanan et al., 2010; Amrose et al. 2014). During ECFe, Fe(II) is dissolved from the anode when current is passed. This Fe(II) oxidizes to Fe(III) in the presence of DO. During formation of Fe(II) to Fe(III) some Fe(IV) might be forming, which is an oxidizing radicle and has been reported to be able to oxidize As(III) to As(IV), which is quickly converted to As(V) in the presence of DO (Hug and Leupin, 2003; Leupin and Hug, 2005; Li et al. 2012; Amrose et al. 2013; , Banerji and Chaudhari, 2016). Thus, oxidation of As(III) and its removal can be achieved in ECFe without the addition of any chemicals. This method promises to be the most efficient with respect to arsenic removal/unit weight of iron added. Conclusion Low income regions in Bengal basin face arsenic problems mostly caused by geogenic groundwater contamination. This review paper looked into the various theories of geogenic contamination of groundwater by arsenic. Reduction of Fe(III) (hydr)oxide

by

microorganisms followed by desorption of arsenic from the iron surface has been identified as the main mechanism of arsenic contamination in groundwater in the Bengal Basin. This review also sums up the iron based arsenic removal technologies currently being investigated for drinking waters, with a special focus on economically depressed regions. These regions need low-cost, effective technologies that can be readily available at the household or community level. Based on the simplicity and ease of availability of raw materials, a low– cost, effective, user friendly community-based filter for arsenic removal based on Fe may be developed which can be fitted with a hand-pump. The treatment technology must be applicable over the broad range of arsenic concentrations found in groundwater and should also be able to achieve As(III) oxidation. The materials for the treatment should be economical and easily obtainable in Indian villages, and/or suitable for reuse. Thus, ZVI and

ECFe are two technologies that are showing maximum promise for effective arsenic removal in the economically backward areas of India. Hence this review has summarized principles of conventional technologies currently described in the literature for arsenic removal that can be used in the developing countries, compared them with the ideal technology requirement and discussed their advantages and disadvantages along with a brief analysis of cause of arsenic presence in groundwater.

Table 1 Advantages and disadvantages of the technologies developed for arsenic removal from groundwater Process Ion Exchange resins

References Chen et al. (1999); Ghurye et al. (1999); Litter et al. (2010); Pirnie (1999)

Membrane processes

Elcik et al. (2013); Litter et al. (2010); Pirnie (1999)

Aluminium based removal processes (coagulation with alum, adsorbents or electrocoagulation with Al electrodes) Iron based adsorbents

Lin and Wu (2001); Singh and Pant (2004); Sarkar et al. (2005, 2010); Kumar et al., (2004); Balasubramanian et al. (2009); Pirnie (1999) Pierce and Moore (1982); Joshi and Chaudhuri (1996); Banerjee et al. (2008); Yuan et al. (2002); Sylvester et al. (2007); Badruzzaman et al. (2004) Mohan and Pittman (2007)

Other adsorbents

Remarks Efficiency of ion exchange process strongly affected by a number of factors such as, total dissolved solids, competing ions, redox potential, sulphate, resin type, etc. Ion exchange for arsenic removal is only applicable for low sulphate and low TDS source water. Very expensive. Narrow pH range for optimal arsenic removal More effective in removal of As(V) as compared to As(III). But oxidizing agents harm the membranes. High capital and running costs. High-tech operation and maintenance. Treated water may contain Al which is a neurotoxin. Less efficiency in arsenic removal than Fe. Factors such as arsenic oxidation state, pH and competing ions significantly affect arsenic removal. Problem with the use of adsorbents is that they have limited life and require

and references therein

Co-precipitation with Fe salts

Meng et al. (2000); Fuller et al. (1993); Hering et al. (1997); Qiao et al. (2012)

Zero Valent iron

Manning et al. (2002); Farrell et al. (2001); Melitas et al. (2002); Nikolaidis et al. (2003); Bang et al. (2005); Leupin et al. (2005) Electrocoagulation with Fe Kumar et al., (2004); Amrose electrodes et al. (2013); Ucar et al. (2013)

frequent regeneration and backwashing. Also most adsorbents (other than GFH – Banerjee et al., 2008 and alumina – Sarkar et al., 2010) have been tested upto lab scales only More effective in removal of As(V) as compared to As(III). Better efficiency than adsorption (w.r.t. amt of arsenic removed/wt of Fe added). Increased chemical costs and requirement of trained personnel for chemical addition. Large quantity of sludge production. Surface area exposed plays a major role and control of Fe(II) dissolution is quite difficult. High efficiency and As(III) oxidation is possible Reactions for arsenic oxidation and removal in ECFe similar to ZVI but as Fe(II) dissolution is controlled, much better efficiencies can be obtained.

Table 2 Comparison of oxidizing agents Oxidant Chlorine

Benefits - Low relative cost - Primary disinfection capability - Secondary disinfectant residual - Oxidizes arsenic in less than 1 min

Drawbacks References Formation of Sorlini and Gialdini, (2010) disinfection by- Bissen and Frimmel, (2003) products - Membrane fouling - Special handling and storage requirements

Hydrogen peroxide

- Oxidizes As(III) in <1 - High relative cost min in absence of - Membrane fouling interfering reductants and presence of iron (hydr)oxide - H2O2 alone takes large amount and long time for As(III) oxidation but reaction is fast in

Lescano et al., (2011) Hug and Leupin (2003)

presence of iron (hydr)oxide Permanganate Unreactive with membranes - No formation of disinfection by-products - Oxidizes arsenic in less than 1 min MnO2 media regenerant

Ozone

Ferrate

- No chemical storage or handling required - Primary disinfection capability - No chemical byproducts left in water - Oxidizes arsenic in less than 1 min in the absence of interfering reductants - No requirement for adsorbent as Fe(III) formed from ferrate adsorbs arsenic

- Formation of MnO2 particulates - Pink Water - Difficult to handle - High relative cost No primary disinfection capability - An additional oxidant may be required for secondary disinfection - An additional oxidant may be required for secondary disinfection - Sulphide and TOC interfere with conversion and increase the required contact time and ozone dose for oxidation

Sorlini and Gialdini, (2010) Driehaus et al. (1995)

Khuntia et al. (2014) Kim and Nriagu (2000)

- Large quantities of Leupin and Hug (2005) sludge produced Lee et al., (2003) - Not established and proved only upto lab scales

Highlights: 1. Groundwater in the Bengal Basin is significantly polluted by naturally occurring arsenic (As), a toxic metalloid, which is responsible for mass poisoning at huge extent. 2. The oxidation of natural organic matter (NOM) with associated reduction of arsenic bearing iron minerals cause mobilization of arsenic and iron in Holocene aquifers. 3. Reductive dissolution of arsenic-bearing Fe(III) (hydr)oxides and As(V) reduction to As(III) are the two main mechanisms controlling arsenic partitioning in soils, sediments and groundwater. 4. Iron-based technologies are very effective in removal of arsenic from groundwater and are more preferred as naturally arsenic is present in soil bound to iron oxides. 5. Community scale water treatment systems are more preferred than household/domestic systems due to its cost-effectiveness and ease of contamination monitoring.