Human health implications, risk assessment and ...

Science of the Total Environment 601–602 (2017) 756–769

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Review

Human health implications, risk assessment and remediation of As-contaminated water: A critical review Muhammad Bilal Shakoor a,⁎, Rab Nawaz b, Fida Hussain c, Maimoona Raza d,e, Shafaqat Ali a, Muhammad Rizwan a, Sang-Eun Oh c, Sajjad Ahmad f a

Department of Environmental Sciences and Engineering, Government College University, Faisalabad, Allama Iqbal Road, 38000 Faisalabad, Pakistan Department of Environmental Sciences, The University of Lahore, Lahore, Pakistan Department of Biological Environment, Kangwon National University, 200-701 Chuncheon, Kangwon-do, South Korea d Department of Geology, Kangwon National University, 24341 Chuncheon, South Korea e National Water Quality Laboratory, Pakistan Council of Research in Water Resources, Islamabad, Pakistan f Department of Environmental Sciences, COMSATS Institute of Information Technology (CIIT), Vehari Campus, Pakistan b c

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Arsenic is recognized as a Class-A human carcinogen. • Groundwater As contamination has affected over 200 million people worldwide. • This paper reviews current knowledge regarding As in the environment. • A critical assessment of remediation of contaminated water is presented.

a r t i c l e

i n f o

Article history: Received 24 April 2017 Received in revised form 19 May 2017 Accepted 24 May 2017 Available online xxxx Editor: D. Barcelo Keywords: Arsenic Biomarkers Groundwater Metabolism Toxicity

a b s t r a c t Arsenic (As) is a naturally occurring metalloid and Class-A human carcinogen. Exposure to As via direct intake of As-contaminated water or ingestion of As-contaminated edible crops is considered a life threatening problem around the globe. Arsenic-laced drinking water has affected the lives of over 200 million people in 105 countries worldwide. Limited data are available on various health risk assessment models/frameworks used to predict carcinogenic and non-carcinogenic health effects caused by As-contaminated water. Therefore, this discussion highlights the need for future research focusing on human health risk assessment of individual As species (both organic and inorganic) present in As-contaminated water. Various conventional and latest technologies for remediation of As-contaminated water are also reviewed along with a discussion of the fate of As-loaded waste and sludge. © 2017 Elsevier B.V. All rights reserved.

⁎ Corresponding author. E-mail address: [email protected] (M.B. Shakoor).

http://dx.doi.org/10.1016/j.scitotenv.2017.05.223 0048-9697/© 2017 Elsevier B.V. All rights reserved.

M.B. Shakoor et al. / Science of the Total Environment 601–602 (2017) 756–769

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Contents 1. 2.

Introduction . . . . . . . . . . . . . . . . . . . . Sources of arsenic in the environment . . . . . . . . 2.1. Natural sources . . . . . . . . . . . . . . . 2.2. Anthropogenic sources . . . . . . . . . . . . 3. Arsenic distribution in the environment. . . . . . . . 3.1. Arsenic in groundwater/drinking water . . . . 3.2. Arsenic in surface water . . . . . . . . . . . 3.3. Arsenic in marine water/saline water . . . . . 3.4. Arsenic in soil . . . . . . . . . . . . . . . . 3.5. Arsenic in food. . . . . . . . . . . . . . . . 3.6. Arsenic in the atmosphere . . . . . . . . . . 4. Arsenic exposure and bioavailability . . . . . . . . . 5. Metabolism and biomarkers of arsenic exposure . . . . 6. Effects of arsenic on human health . . . . . . . . . . 7. Health risk assessment of arsenic. . . . . . . . . . . 7.1. Chronic daily intake . . . . . . . . . . . . . 7.1.1. Oral exposure . . . . . . . . . . . . 7.1.2. Dermal exposure . . . . . . . . . . 7.2. Hazard quotient (HQ) and hazard indices (HI) . 7.3. Carcinogenic risk (CR) and cancer indices (CI) . 8. Remediation techniques for arsenic removal from water 8.1. Precipitation. . . . . . . . . . . . . . . . . 8.2. Coagulation/flocculation . . . . . . . . . . . 8.3. Ion exchange . . . . . . . . . . . . . . . . 8.4. Membrane filtration . . . . . . . . . . . . . 8.5. Floatation . . . . . . . . . . . . . . . . . . 8.6. Phytoremediation . . . . . . . . . . . . . . 8.7. Sorption. . . . . . . . . . . . . . . . . . . 9. Ultimate fate of arsenic-laden sludge and wastes. . . . 10. Conclusions and future research . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . Appendix A. Supplementary data . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . .

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1. Introduction Arsenic is a naturally existing metalloid and a carcinogenic element. The background natural concentration of As in the soils is ~5–10 mg kg−1 (Smedley and Kinniburgh, 2002; Basu et al., 2014); however, its distribution is not homogenous in the Earth's crust and is concentrated in certain geologic settings (Ravenscroft et al., 2009; Rahman and Hasegawa, 2011; Shah, 2014). Arsenic concentration in groundwater ranges from 0.5 to 5000 μg L−1 and contamination from natural sources has been reported for N105 countries (Ravenscroft et al., 2009; Kippler et al., 2016). The major sources of As in groundwater are minerals such as realgar (AsS), arsenopyrite (FeAsS) and orpiment (As2O3). Overall, As sources are attributed to both natural processes, such as oxidative/reductive dissolution of As containing compounds sorbed onto pyrite minerals and anthropogenic activities such as use of pesticides, irrigation with Ascontaminated water, semi-conductor manufacturer, phosphate fertilizers, mining and smelting activities, burning of coal, and timber preservatives (Shakoor et al., 2015; Shakoor et al., 2016). Arsenic occurs in four oxidation states, i.e. arsenite (As(III)), arsenate (As(V)), elemental As (As0), and arsine (As(III)) (Sharma et al., 2014a). The solubility and mobility of As largely depends on pH and redox conditions, and is present in both organic and inorganic forms in the groundwater. Arsenite predominates in a reduced environment (deep groundwater) at a high pH value, while As(V) species prevail in an oxidized environment (shallow groundwater) at low pH (Abdallah and Gagnon, 2009; Bundschuh et al., 2012; Jackson et al., 2012). In aqueous environments, As(V) exists as weak triprotic acid and it has different acid dissociation constants (pK) such as 2.20 and 6.97– 11.53. At natural pH range (4–8) such as in groundwater, As(V) species 2− are present as H2AsO− 4 and HAsO4 . Arsenate species prevail as an

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amphiprotic material at pH 2.24–11.50, and thus it accepts or donates protons (H+) depending upon the variation in the pH of the aqueous environment (Mandal and Suzuki, 2002; Naidu et al., 2006a; Markley and Herbert, 2009). On the other hand, As(III) occurs as a hydroxo-acid and has three dissociation constants (9.22, 12.13 and 13.40). Since the protons (H+) present on the hydroxyl groups (OH) of As(OH)3 do not possess a doubly bonded oxygen neighbor, As(III) is considered to be a very weak acid and exists as a neutral or non-ionic species (H3AsO3) in most of the groundwater conditions. Arsenite also predominates as both the amphiprotic and the polyprotic acid at pH N 9.22 in aqueous environment (Naidu et al., 2006a). Organic As is also present in the environment as monomethylarsonic acid {CH3AsO(OH)2; MMA}, dimethylarsinic acid {(CH3)2AsOOH; DMA}, trimethylarsine oxide {(CH3)3AsO; TMAO}, arsenosugars (AsS), arsenobetaine (AsB), arsenocholine (AsC) and arsenolipids etc. (Singh et al., 2015). Methylated As compounds occur in a small amount (Gan et al., 2014); however, they may also be present as a major component in the soil. Both the MMA and DMA (cacodylic acid) were used extensively as herbicides and pesticides in the past – DMA was also applied as a cotton defoliant (Diwakar et al., 2015). Arsenic contamination of groundwater is a global public health and environmental risk for over 200 million people in the world (Rasool et al., 2015; Mishra et al., 2016). Groundwater in most of the countries also acts as a pathway for As to enter into the food chain via drinking, cooking, bathing, and irrigation of food crops with As-contaminated water (Abid et al., 2016). Thus, intake of As-contaminated water and food is a major pathway for As exposure to humans. Long-term exposure to As is reported to cause skin lesions, neurotoxicity, diabetes, cardiovascular diseases, and various kinds of cancers such as skin, liver, bladder, and kidney (Hsu et al., 2016; Rasheed et al., 2017).

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In general, inorganic As is 100 times more toxic compared to organic As and almost all the trivalent and pentavalent species of As are significantly absorbed into the body and transported through the blood stream to the tissues (Basu et al., 2014; Rasheed et al., 2016). Arsenic metabolism depends on oxidation/reduction processes causing the inter-conversion of As species (trivalent and pentavalent), and methylation of As(III) to produce methylated As species (Kim et al., 2011; Yadav et al., 2014). The main reason behind more toxicity of As(III) than As(V) is possibly the interaction of As(III) in enzymatic reactions in which it binds with sulfhydryl (−SH) or hydroxyl (−OH) functional groups (Naujokas et al., 2013). Previous data have indicated that trivalent methylated As species are significantly more toxic (60 times) as compared to the inorganic pentavalent As species, however, the relative toxicity of individual arsenicals, such as MMA(III) or DMA(III) is still not fully understood (Basu et al., 2014). Past studies also showed that the methylation of inorganic As species decreases the relative toxicity, but data are contradictory (Naidu et al., 2006a; Mishra et al., 2016). Hence, uncertainties still exist about the potential health risks and resulting toxicity of specific As species in the human body. Health risk assessment of exposure due to consumption of As-contaminated water and food can be determined using health risk models (USEPA, 2005; Shakoor et al., 2015; Rasheed et al., 2017). Here we review As health risk assessment models/frameworks. We also discuss the distribution and potential health risks of As species in groundwater from multiple exposure pathways and the impact of remediation techniques.

(Arain et al., 2009). Most of the common As minerals are ores or byproducts of such minerals. Arsenic is commonly found in deposits of silver, zinc, copper, cadmium, gold, mercury, tin, uranium, iron, cobalt, lead, nickel, selenium, phosphorus, sulfur, antimony, bismuth, tellurium, tungsten, molybdenum, and platinum (Zhang et al., 2017). The most common As-bearing minerals are arsenolite (As4O6), claudetite (As2O3), AsS, FeAsS, pentoxide (As2O5), and scorodite (FeAsO4.2H2O) and among all these minerals, the most abundant and common mineral is FeAsS (Sarkar and Paul, 2016; Singh et al., 2015). 2.2. Anthropogenic sources There are a number of uses of As in different countries of the world (Fig. 1). Major anthropogenic sources of As are pesticides, herbicides, paints, cosmetics, dyes, mining operations, smelters, wood treatments, cattle dips, electronic manufacturing, vitamin supplements, poultry and swine feed additives, pharmaceuticals, cigarettes and mining operations and processing of wastes (Smedley and Kinniburgh, 2002; Basu et al., 2014). Anthropogenic sources account for As emissions of almost 30,000 tons per year into the atmosphere (Basu et al., 2014). Coal combustion and copper smelting contribute almost 60% of anthropogenic As contamination of the environment (Basu et al., 2014). Land is also directly contaminated with As from dumping of sludge, slag and wastewater discharges from refineries and smelters (Cullen and Reimer, 1989; Andreae and Andreae, 1989; Amonoo-Neizer and Amekor, 1993; Berg et al., 2001; Ahsan et al., 2006; Clewell et al., 2007; Brammer and Ravenscroft, 2009; Bhattacharya et al., 2010).

2. Sources of arsenic in the environment 3. Arsenic distribution in the environment 2.1. Natural sources 3.1. Arsenic in groundwater/drinking water Arsenic is a major component of over 200 minerals, and it does not occur frequently in its pure arsenical form. Important minerals of As include elemental As, sulfides, arsenates, arsenites, arsenides and oxides

Arsenic is mobilized from the natural sources to groundwater and has been reported at concentrations up to 5000 μg L−1 in groundwater

Fig. 1. Arsenic distribution in the environment and its transfer pathways to humans.


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Table 1 Concentration of arsenic in groundwater of the world. Area (countries)

Concentration (μg L−1)

Source

National standard (μg L−1)

References

Cordoba, Argentina Bangladesh Calcutta, India Chile Hanoi, Vietnam Hungary Inner Mongolia, China Lagunera region, Mexico Mekong River floodplain, Cambodia Nakhon Si Thammarat Province, Thailand Nepal Ronpibool, Thailand West Bengal, India Western USA Xinjiang, PR China Japan Australia Taiwan Central Punjab, Pakistan Southern Punjab, Pakistan Jamshoro, Pakistan Tharparkar, Pakistan

100–3800 b10 to N1000 b50–23,080 470–770 1–3050 1–174 1–2400 8–624 1–1340 1.25–5114 8–2660 1–5000 3–3700 1–48,000 0.05–850 1–293 1–5000 0.15–3590 N1–1900 1–201 1–106 1–2580

– Wells Sediments – Sediments Deep groundwater Drinking water; bores Well waters Groundwater Shallow (alluvial) groundwater, mining Drinking water Mining activities Sediments Drinking water Well water Groundwater Minerals, groundwater Groundwater Groundwater Groundwater Mining activities, groundwater Mining activities, groundwater

10 50 50 10 10 50 50 50 50 50 50 50 50 10 50 10 10 10 50 50 50 50

(Nicolli et al., 1989) (Chakraborti and Roy, 1997) (Mandal et al., 1996) (Naidu et al., 2006b) (Berg et al., 2008) (Sancha et al., 2008) (Smedley and Kinniburgh, 2002) (Mahimairaja et al., 2005) (Buschmann et al., 2007) (Williams et al., 1996) (Naidu et al., 2006b) (Naidu et al., 2006b) (Mandal, 1996) (Welch and Lico, 1998) (Yinlong, 2001) (Naidu et al., 2006b) (Naidu et al., 2006b) (Naidu et al., 2006b) (Sultana et al., 2014) (Shakoor et al., 2015) (Baig et al., 2009) (Brahman et al., 2013)

as given in Table 1 and Table S1; Supplementary information and World Health Organization (WHO) has recommended 10 μg L−1 of As in drinking water as a safe limit and this was reduced from 50 μg L−1 in 1993, therefore in many countries the reduced (10 μg L−1) limit of As in drinking water is being implemented (Ravenscroft et al., 2009; Pio et al., 2015). Arsenic contamination of groundwater has been reported in various countries including Bangladesh, China, India, Chile, Pakistan, Mexico, Taiwan, Poland, Argentina, Canada, New Zealand, Hungary, Japan, and the United States (Naidu et al., 2006b; Naujokas et al., 2013). In West Bengal, India (Chakraborti et al., 2010) 59 districts, in Bangladesh majority of population and many large areas of Punjab and Sindh in Pakistan (Farooqi et al., 2007; Shakoor et al., 2015) rely on As-contaminated groundwater for drinking and irrigation purposes (Abedin et al., 2002). Keeping in view the WHO guideline for As in drinking water, global population of N100 million is at risk of As poisoning and over 45 million people facing As contamination b50 μg L−1 (maximum permissible limit in these countries) live in third world countries (Queirolo et al., 2000; Prieto-GarcÃ-A et al., 2005; Roberge et al., 2009; Singh et al., 2015). Although the main source of As exposure to humans is drinking water, food is another source. Some edible crops like vegetables (peas, potato, cabbage, onion, carrots, chili) and rice were reported to contain high contents of inorganic As grown on soils with elevated As concentration in the soil and irrigated with As-contaminated groundwater (Maest et al., 1992; Molla et al., 2004; Rahman and Hasegawa, 2011; Bhattacharya et al., 2012).

contamination up to 13,900 μg L−1 in water of Mole River (Ashley and Lottermoser, 1999). Concentrations of As in lake waters are typically lower than the As concentrations reported in river waters worldwide (Table S1; Supplementary information). Research has shown that As levels in lake waters around Canada and British Columbia were varied between 0.2 and 2.08 μg L−1 and As was released from the mine tailing of Gold Quartz in that region and increased up to 1104 μg g−1 in deep sediments of different lakes catchments (Yusof et al., 1994; Azcue and Nriagu, 1995; Smedley et al., 1996; Wilkie and Hering, 1998; Smedley and Kinniburgh, 2001; Smedley and Kinniburgh, 2002). Geothermal sources of As directly and mining facilities indirectly have enhanced As levels in lake waters (Peterson and Carpenter, 1983; Navarro et al., 1993; Nriagu et al., 2007; Palmieri et al., 2009; Basu et al., 2014). In recent years, the main As affected areas are identified in great deltas and in the sideways of major rivers originating from the Himalayas notably the Bengal delta which has been described as worst As-exposed area where N 88% of 45 million population are at risk of contact to As concentrations up to 50 μg L− 1 (Li et al., 2003; Ravenscroft et al., 2009; Singh et al., 2015). Other As-affected regions are river deltas and basins in South Asia specially the Red river delta (Berg et al., 2008; Muhammad et al., 2010), the Mekong river delta (Buschmann et al., 2008; Piyawat Saipan, 2009) and the river basins of Salween; Indus, Ganges, Chenab, Brahmaputra, Chindwin-Irrawady (Thakur et al., 2010; Sarkar and Paul, 2016). 3.3. Arsenic in marine water/saline water

3.2. Arsenic in surface water The concentration of As varies from 0.15–0.45 μg L−1 in freshwaters including rivers, lakes and streams depending on catchment geochemistry, source and availability (Singh et al., 2015). (Table S1; Supplementary information). The main natural causes of As contamination in river waters are geothermal inputs, groundwater contamination and evaporation (Del Razo et al., 1990; Chowdhury et al., 1997; Chakraborti et al., 2001; Choprapwon and Porapakkham, 2001; Gómez et al., 2006; Díaz et al., 2008; Chetia et al., 2011; Das et al., 2013). A study on river water of Lao River (Northern Chile) indicates the As contamination up to 21,000 μg L−1 due to natural processes of geothermal inputs and evaporation (Cáceres et al., 1992). Some anthropogenic activities like mining can also facilitate the presence of high levels of As in rivers. A case study from New South Wales Australia shows the As

Arsenic concentration in sea water is usually b 2 μg L−1 and in major oceanic water such as Atlantic and deep Pacific waters, it ranges 1– 1.8 μg L− 1 (Ng, 2005; Sloth and Julshamn, 2008; Seow et al., 2012), 3.1 μg L− 1 in marine waters of the Pacific coast (Japan) and 1.1– 1.6 μg L− 1 in coastal waters of southern parts of Australia (Maher, 1985) (Table S1; Supplementary information). The concentration of As is more uniform in estuarine waters compared to open marine waters. In estuarine waters, As concentration could be affected by the effluents from industrial activities and mine tailing processes and geothermal water intrusion (Lerda and Prosperi, 1996; Mora et al., 2001; Munoz et al., 2002; Palumbo-Roe et al., 2005). The physical mixing of the fresh water with seawater masses and salinity may influence the concentration of dissolved As in estuarine water (Hasegawa, 1996; Guo et al., 2001; Gurzau and Gurzau, 2001; Harvey et al., 2002; Gunduz et al.,

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2010). A linear rise in total As contents, reaching from 0.13 μg L−1 in freshwaters to 1.8 μg L−1 in marine waters, also with increase in the salinity, has been documented in Krka Estuary, Yugoslavia (Seyler and Martin, 1991). 3.4. Arsenic in soil Background concentrations of As differ greatly relying on the type and structure of soil. Generally, the baseline levels of As in the soils range from 5 to 10 mg kg−1 and the average concentration of As in European topsoil is found to be 7 mg kg− 1 (Julshamn et al., 2004; Karczewska et al., 2007; Stafilov et al., 2010). Higher As levels are found in peats and bog soils (average 13 mg kg−1). Acid sulfate soils also contain relatively elevated concentrations of As (Smedley and Kinniburgh, 2002; Karagas et al., 2015; Rasheed et al., 2017) The United States-Environmental Protection Agency (US-EPA) has recommended the safe limit of As in soil as 24 mg kg−1 but this limit has been exceeded in many parts of the world as shown in Table S1; Supplementary information (Bundschuh et al., 2011; Mondal et al., 2013; Hossain et al., 2017). 3.5. Arsenic in food Rice is a significant source of As exposure to humans in the recent past (Table S2; Supplementary information). Rice is one of the staple foods for a large number of the world's population, because rice is an enriched source of many nutrients like carbohydrates, vitamin B6, thiamin, and some other essential trace elements like zinc, magnesium, iron and copper (Howard et al., 1988; Kurttio et al., 1998; Kondo et al., 1999; Hasegawa et al., 2009; Sun et al., 2014; Tsuji et al., 2014). The United States Department of Agriculture (USDA) examined that rice production in 2016/2017 would be 481.5 million metric tons while the rice production in 2014 was 472.2 million tons globally a 2% increase (Cooke et al., 2016). Most of the rice producing countries have human health risks to high concentrations of As in their ground waters and soils (Rahman and Hasegawa, 2011). Rice can accumulate lethal concentrations of toxic elements specifically As and due to its daily use as a staple food, As can accumulate in the human body and pose a serious risk (Shraim, 2014). Further, some other cereal crops such as corn, oat, wheat, and buckwheat are also potential sources of As toxicity to humans (Table S2; Supplementary information). Many vegetables, dairy, and meat products are also reported as a good source of As accumulation in humans (Robinson et al., 1995; Schoof et al., 1999; U.S Food and Drug Administration, 2009; Wang et al., 2013). The amount of food taken in the diet may influence the amount of daily As ingestion via food by humans (Fig. 1). Arsenic concentrations in different food items and its safe concentrations are given in Table S2; Supplementary information. Arsenic is commonly used in poultry in the form of roxarsone (an additive in feed for conventionally raised boilers) (Li et al., 2016). Roxarsone is applied to control parasites of protozoan (coccidians) and to increase weight gain, but feeding As in any form to laying hens is prohibited (Farooq et al., 2016a, 2016b). The birds raised for organic certification are prohibited for feeding As as per organic regulations. It was reported that the roxarsone was added at a rate of 22.7 to 45.4 g ton− 1 (0.0025–0.005%) to the poultry feed (Mandela, 2015). Mostly the roxarsone ingested by birds is excreted from the body in stable form (Nigra et al., 2017), and on average, each broiler excretes around 150 mg of roxasone during 42 days of growth period (Menahem et al., 2016). Furthermore, the fate of roxarsone and the products after transformation {(As(V), As(III), 3-amino-4hydroxyphenylarsonic acid, monomethylarsonate, dimethylarsinate and 4-hydroxyphenylarsonic acid)} were investigated in the chicken manure (Rosal et al., 2005). The results showed that roxasone and its transformed products were responsible for the presence of As in the chicken.

3.6. Arsenic in the atmosphere Although the atmospheric concentration of As is very low, volcanic activities, smelting and mining operations, fossil fuel combustion and other industrial activities significantly increase concentration in the atmosphere (Fig. 1). Arsenic concentration has been found about 0.001– 0.00001, 0.003–0.18 and N1 μg m−3 in unpolluted areas, urban areas and industrial plants, respectively (WHO, 2011). Much of the As present in the atmosphere is in the form of particulate and total As deposition has been reported between b1–1000 μg m−2 year−1 based on dry and wet deposition, and proximity to the source of emission (Basu et al., 2014). Arsenic as dry deposition (29–55%) in mid-Atlantic coast was found to be 38–266 μg m−2 year−1 (Wang and Mulligan, 2006). 4. Arsenic exposure and bioavailability Previous research describes that As uptake and accumulation in different plant species vary widely (Sharma et al., 2014b; Rasheed et al., 2017) depending on many factors such as concentration of As in soil, properties of soil, microbial activities, pH, redox potential, type of plant species, and their water requirements (Gan et al., 2014; Yunus et al., 2016). Arsenic enters in the food chain when herbivores eat As-contaminated feedstock/fodders or drink As-contaminated water from water supplies (Petrick et al., 2000; Pérez-Carrera and Fernández-Cirelli, 2005). In humans, the major food sources through which As enters into the body have been reported to be fish, fruit, crops (rice, cereals), poultry, meat and milk (Table S2; Supplementary information). The WHO permissible limit for As in food is 1 mg kg−1 that has been exceeded by various types of foods (Table S2; Supplementary information). The As concentration in different food types has been reported as 1.9 mg kg−1 (cereals), 22.4 mg kg−1 (fruits and fruit juices), 13 mg kg−1 (vegetables) and 42.6 mg kg−1 (animal products and Rice) (Nookabkaew et al., 2013; Rasheed et al., 2016). Rice crop is an excellent scavenger of As and accumulates as much as ten times more As compared to other cereal crops that might be due to the growth in flooded conditions (Chen et al., 2010; Khan et al., 2010; Bakhat et al., 2017). Hence, As exposure is very likely to be higher for individuals who eat large quantities of rice daily and for infants, who are fed upon rice-based baby food as solid meal. The toxicity and mobility of As in foods depends upon bioaccessibility and the chemical form (Juskelis et al., 2013). Apart from water, both organic and inorganic As species have been thoroughly studied in a wide variety of foods ranging from rice to milk and fish (Carey et al., 2010; Norton et al., 2013). Arsenic speciation and its relative toxicity should be thoroughly investigated in water in order to understand Asbioavailability to food crops from irrigation with As-contaminated water (Erban et al., 2011). 5. Metabolism and biomarkers of arsenic exposure Arsenic is a member of group 15 in the periodic table, and occurs with phosphorous and nitrogen that are essential nutrients for humans. The chemistry and molecular structure of As(V) is analogous to phosphate, but still As is not considered as essential element for humans (WHO, 2011). Although the literature shows that human metabolism can adapt to As exposure up to a certain extent, but to our knowledge not even a single study has been reported which indicates beneficial effects of As in humans as it has in the form of roxarsone for poultry (Li et al., 2016).Water soluble inorganic arsenicals (80–90% of single dose) are quickly and heavily sorbed from the gastrointestinal tract and substantially bind to DNA or protein molecules in humans (Zheng et al., 2014). The estimated biological half-life of inorganic As in humans is around 4 days and the major route of excretion is urine, while small concentrations are eliminated through hair, sweat, skin, breast milk and faeces (Sharma et al., 2014b). Metabolism of As in the human body depends on the inter-conversion of its species i.e. As(III) and As(V) and


characterized by oxidation/reduction and biological methylation of As species. Arsenate concentration that is absorbed into human gastrointestinal tract is about 40–100% of total As (Saha et al., 1999). Arsenite can bind with thiol (–SH) group of proteins (Shah, 2014), however, the absorption rate into the body is less than other soluble forms of inorganic As in water (Anjos et al., 2012). Although, the detailed processes taking part in the metabolic pathways of organic and inorganic As species are yet to be explored, an overall summary of As metabolism has been illustrated in Eq. (1) (Rasheed et al., 2016) and Fig. 2. AsðIIIÞ

oxidation=reduction

MMAðIIIÞ

→

methylation

→

AsðVÞ

DMAðVÞ

methylation

→

MMAðVÞ

possible reduction

→

reduction

→

ð1Þ

DMAðIIIÞ

To understand the role of As in metabolism of humans, certain biomarkers have been studied as indicators. Biomarkers are defined as “the quantifiable changes in biochemical, physiological or behavioural states within cells, tissues or whole individuals because of external stressors” (Timbrell, 2001). These are classified as markers of exposure and effect/susceptibility (National Academy of Science) and gives information about the fate and metabolic reactions of As within the body of humans. To enhance our understating on the fate and metabolism of As in the human body, As concentration in nails, hairs, blood and urine have been analyzed (Table 2). It was demonstrated that As accumulation in nails and hair was due to the strong binding of As with proteins (keratin) (National Research Council, 1998). Arsenic analysis in nails and hair is used to confirm As exposure (intake of drinking water, skin exposure) and related accumulation within the human body (Table 2). The maximum concentrations of As reported in human hair, nails and urine are 1500 mg kg−1, 5406 mg kg−1 and 1000–6200 μg L−1, respectively (Lindberg et al., 2006; Concha et al., 2010), while in blood As was found to be lowest as 1–14.3 μg L−1. Only limited data are available on As speciation in nails and hair as compared to urine probably because of the complexity involved in sample preparation to eliminate impurities adsorbed on the surface of materials (Mandal, 1996). Urinary As metabolites are used to correlate exposure of As with intake rates, mechanism involved in As methylation, bioaccumulation of As into the human body and excretion potential (Smith et al., 2003). The carcinogenic and noncarcinogenic health effects of As can also be determined from As present in urine (Fig. 2, Table 2). Previous studies on urinary metabolites (listed in Table 2) showed that major part of ingested As in humans is methylated followed by significant excretion in the form of DMA (79–85%) and in small concentration as MMA (5–6%) and as inorganic As (8–16%) (Christian et al., 2006). Despite a lot of studies available on urinary As metabolites, the processes controlling As uptake and excretion from

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various foods are still unclear, and also there is a need to unravel how exposures from different sources lead to toxic health effects in humans (Christian et al., 2006).

6. Effects of arsenic on human health Arsenic toxicity and resulting health effects have been associated with drinking of As-contaminated water containing As N 10 μg L− 1 (Basu et al., 2014). Various kinds of human health effects originating from exposure to As in the world have been illustrated in Table 3. Long term exposure to As in drinking water or chronic toxicity cause lung, skin and kidney cancers, pigmentation changes, neurological disorders, skin thickening (hyperkeratosis), loss of appetite, muscular weakness, and nausea (Table 3) (Rahman et al., 2009). However, acute poisoning causes oesophageal and abdominal pain, vomiting, and diarrhea (Fig. 1). Elevated levels of As in drinking water can also cause an increase in miscarriages and spontaneous abortions (Smedley and Kinniburgh, 2002).

7. Health risk assessment of arsenic Assessing the human health risk of As by comparing the concentration of As in water with permissible limit of As is not a reliable and comprehensive approach. The risk to an individual human from intake of Ascontaminated water depends upon many factors such as water intake, exposure time, exposure duration, exposure frequency, body weight and type of population (Mahimairaja et al., 2005; Farooqi et al., 2007; Caylak, 2012). Furthermore, the permissible limit of As in drinking is set for total As only, but As is present mainly as inorganic species in groundwater thus this limit does not differentiate the toxicity of inorganic and organic As species in water. Health risk assessment models/frameworks estimate potential health effects on humans caused by exposure to As-contaminated water and are important initial steps in the development of remedial and management strategies to protect people from toxicity. Table S3; Supplementary information summarized a variety of risk assessment models or approaches that have been used to assess human health risk for As in water (Mondal and Polya, 2008; Mondal et al., 2010; Liang et al., 2013). Among these models/frameworks US-EPA health risk assessment model equations seems to be more promising and comprehensive approach for health risk assessment of As from intake of As-contaminated water (Table S3; Supplementary information) (USEPA, 2005), and therefore discussed here.

Fig. 2. Sources of arsenic exposure, metabolism and excretion through urine in human body (adopted from Navas-Acien et al., 2009).

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Table 2 Summary of human studies measuring biological arsenic in hair, nail, blood and urine. Biomarker type

Arsenic level

Unit

Analysis

References

Hair

0.10–4.57 0.018–1 0.01–57.21 0.0059–0.0644 0.20 to 6.50 20–1500 5.52 Total arsenic: 0.07–4.61, As(III): 0.21–2.64, As(V): 0.08–1.54, DMA(III): 0.02–0.13, MMA(V): 0.02–0.2 0.61–27.89 0.19 7.8 0.10 to 7.95 Finger nail: 0.03–28.47 Toenail: 0.10–21.89 Total arsenic: 5406, As(III): 11,477, As(V): 2899, DMA(V): 84, MMA(V): 73 0.02 to 2.11 21.7 Total arsenic: 1.47–7.39, As(III): 0.95–2.76, As(V): 0.27–1.31, MMA(III): 0.09–0.21, DMA(III): 0.11–0.38, DMA(V): 0.04–0.09 3.29–8.82 (exposed cancer patients) 14.3 1.31–10.37 (new borne blood) 1.0–18.3 Exposed: 6.6 Unexposed: 5.0 117 ± 8.3 As(III): 16.8, As(V): 1.8, MMA: 1.8, DMA: 88.6 N3.5 (total arsenic + MMA + DMA) Inorganic arsenic: 11–509.4, MMA: 55–2192.5, DMA:6.8–687.4 260 As(III): b1–22.6, MMA(V): b1–20.3, DMA(V): 17.7–86, As(V): b1–35.1 172 11.1–54.5 Total arsenic: 13–440, inorganic As + MMA + DMA: 9–405

mg mg mg mg mg mg mg mg

kg−1 kg−1 kg−1 kg−1 kg−1 kg−1 kg−1 kg−1

Total As Total As Total As Total As Total As Total As Total As Speciation based analysis

(Aldroobi et al., 2013) (Normandin et al., 2013) (Phan et al., 2011) (Essumang, 2009) (Gault et al., 2008) (Concha et al., 2010) (Hinwood et al., 2003) (Mandal et al., 2003)

mg mg mg mg mg

kg−1 kg−1 kg−1 kg−1 kg−1

Total As Total As Total As Total As Total As

(Rahman et al., 2005) (Cottingham et al., 2013) (Cui et al., 2013) (Gault et al., 2008) (Phan et al., 2011)

mg kg−1

Speciation based analysis

(Button et al. 2009)

mg kg−1 mg kg−1 mg kg−1

Total As Total As Speciation based analysis

(Michaud et al., 2004) (Hinwood et al., 2003) (Mandal et al., 2003)

Total arsenic Total arsenic Total arsenic Total arsenic

(Wadhwa et al., 2011) (Hall et al., 2006) (Intarasunanont et al., 2012) (Vahter et al., 1995) (Neamtiu et al., 2015)

Total arsenic Total arsenic Speciation based analysis Speciation based analysis Total arsenic Speciation based analysis Total arsenic Total arsenic Speciation based analysis

(Liu et al., 2013) (Hata et al., 2012) (Fillol et al., 2010) (Loffredo et al., 2003) (Asante et al., 2008) (Agusa et al., 2006) (Hall et al., 2006) (Maharjan et al., 2005) (Harrington et al., 1978)

Nails

Blood

Urine

a

μg μg μg μg

L−1 L−1 L−1 L−1

μg g−1 of creatininea μg L−1 μg L−1 μg L−1 μg L−1 mg g−1 creatinine μg L−1 μg g−1 of creatininea μg L−1

Urinary arsenic reference value: 28 μg mmol−1 creatinine.

The US-EPA equations calculate the carcinogenic and non-carcinogenic risk to humans. Health risk from As-contaminated water and food can be determined which are the two major exposure routes of As in humans (Shakoor et al., 2015). From food, health risk assessment can be determined in terms of estimated daily intake (EDI) by evaluating the exposure dose of As due to the consumption of food crops. However, health risk assessment based on consumption of As-contaminated water is highly significant due to the direct, intense and continuous human exposure to As. Human health risk assessment can be computed by estimating chronic daily intake (CDI), hazard quotient (HQ), and carcinogenic risk (CR) (Kavcar et al., 2009; Nguyen et al., 2009; Rahman et al., 2015).

7.1. Chronic daily intake Chronic daily intake for As via oral and dermal exposure to As-contaminated water is calculated by using Eqs. (2)–(9) as discussed below. 7.1.1. Oral exposure For this purpose, CDI (oral) of As due to oral intake of As-contaminated drinking water is calculated by the Eq. (2): (USEPA, 2005)

CDI ðoralÞ ¼

C IR ED EF BW AT

ð2Þ

Table 3 Effects of arsenic toxicity on human health. Health effects

Study type

Country

References

Peripheral neuropathy Bladder cancer Skin lesions Lung, bladder cancer Neurobehavioral function Cough, bronchitis Hyperpigmentation, hyperkeratosis and skin tumours Prominent transverse white lines in the fingernails and toenails called Mee's lines Diarrhea and stomach issues Cardiovascular disease Restrictive or obstructive lungs diseases, and bronchitis Increased frequency of miscarriages Respiratory disease Skin lesion, lung disease

Cohort Case-control Cross-sectional Case-control Cross-sectional Case-control Cross-sectional population survey Case-control

India USA Chile Taiwan Taiwan Bangladesh India USA

(Mukherjee and Bhattacharya, 2001) (Steinmaus et al., 2003) (Smith et al., 2000) (Morales et al., 2000) (Tsai et al., 2003) (Milton et al., 2005) (Mazumder et al., 1992) (Fincher and Koerker, 1987)

Cross-sectional Case-control retrospective assessment of exposure Cross-sectional Prospective cohort Cross section survey Cross section survey

Australia

(Poklis and Saady, 1990) (Pinto et al., 1977) (Mazumder et al., 2000) (Rahman et al., 2009) (Arain et al., 2009) (Shakoor et al., 2015)

USA Bangladesh Pakistan Pakistan


where C is the concentration of As in water (mg L−1), IR is the water ingestion rate (L day−1), ED is exposure duration, EF is exposure frequency (365 days year−1), BW is a body weight (Sultana et al., 2014) and AT is average life time (24,455 days). 7.1.2. Dermal exposure The CDI (dermal) for dermal exposure to As via drinking water is calculated by Eq. (3): (Li and Zhang, 2010). CDI ðdermalÞ ¼

LAs SA Kp ET EF ED Fc1 Fc2 BW day

ð3Þ

where CDI (dermal) is the chronic daily As exposure dose through dermal contact of water (mg−1 kg −1 day−1), LAS is the level of As in drinking water (μg L−1), SA is drinking water exposed skin area (cm2), Kp is coefficient for dermal permeability (cm h−1), ET is exposure time during bathing and shower (min day− 1), EF is exposure frequency (days year−1), ED is exposure duration (year), BW is body weight (kg), day = time (days), Fc1 is the factor for conversion from μg to mg (0.001), and Fc2 is the factor for conversion of unit (L 1000 cm− 3) (0.001). According to USEPA database Kp value is 1.10−3 cm h−1 for As (USEPA, 2005).

Hazard quotient (oral and dermal) and HI are computed by the Eqs. (4), (5) (USEPA, 2005). ðoralÞ

¼

CDI ðoralÞ RfD ðoralÞ

ð4Þ

where RfD (oral) represents oral reference dose (0.0003 mg kg−1 day−1) for As calculated by USEPA (2005). HQ

ðdermalÞ

¼

CDI ðdermalÞ RfDðdermalÞ

ð5Þ

where RfD (dermal) is dermal reference dose (0.000190 mg kg−1 day−1) for As according to USEPA (2005). The HI is also calculated by using Eq. (5). If HQ or HI is N1, it represents a non-carcinogenic As risk for human health (USEPA, 2005). HI ¼ HQ

ðoralÞ

þ HQ

ðdermalÞ

ð6Þ

7.3. Carcinogenic risk (CR) and cancer indices (CI) Carcinogenic risks for oral and dermal exposure to As are calculated by using following Eqs. (7), (8). CR ðoralÞ ¼ CDI ðoralÞ SF

ð7Þ

where CSF (oral) is the cancer slope factor for oral exposure to As which is 1.5 mg kg−1 day−1, according to USEPA (2005) CR ðdermalÞ ¼ CDI ðdermalÞ SF

ð8Þ

where CSF (dermal) is cancer slope factor (3.66 mg kg−1 day−1) for dermal exposure to As (USEPA, 2005). To estimate the total potential cancer risks of As by oral and dermal exposure, CI is calculated (Eq. (9)). A CI value N 10−6 indicates potential CR (De Miguel et al., 2007). CI ¼ CR ðoralÞ þ CR ðdermalÞ

toxicity and bioavailability in water (Rasheed et al., 2016; Waqas et al., 2017). It is therefore critical to analyze As species absorbed from water, food or soil, metabolized in liver, bladder and kidneys, concentrated in hairs and nails and eventually excreted by urine or faeces. To our knowledge, a few studies have been conducted to evaluate the speciated As contents in humans. Normandin et al. (2013) reported that As(III), As(V), DMA(V) and MMA(V) ranged 0.03–7.38, 0.03–13.3, 0.32–7.38 and 0.03–31.5 μg L−1 in human urine (Table 2). In another study, Mandal et al. (2003) analyzed total As (0.07–4.61 mg kg− 1), As(III) (0.21–2.64 mg kg−1), As(V) (0.08–1.54 mg kg−1), DMA(V) (0.02–0.13 mg kg− 1) and MMA(V) (0.02–0.2 mg kg− 1) in human hairs and nails. The literature is also limited on the health risks caused by exposure to different As species separately such as As(III), MMA(III), DMA(III) or As(V), MMA(V), DMA(V) from a range of sources such as groundwater (Mahimairaja et al., 2005; Basu et al., 2014; Rasheed et al., 2016). Hence there is a need to study speciation of As in groundwater, plants, and various parts of humans such as hair, nails, blood and urine which might be helpful in assessing health risks caused by different As species. 8. Remediation techniques for arsenic removal from water 8.1. Precipitation

7.2. Hazard quotient (HQ) and hazard indices (HI)

HQ

763

ð9Þ

In assessment of human health risk, speciation and bioavailability of As are crucial factors as different As species have varying levels of

Precipitation is a conventional method for As and other heavy metals removal from contaminated water and this process is extensively employed due to its simplicity and cost-effectiveness (Luqman et al., 2013). The chemicals interact with As to make insoluble precipitates of hydroxide, sulfide, and carbonates which could be removed from water through filtration or sedimentation process (Matisoff et al., 1982; McLaren and Kim, 1995; Meacher et al., 2002; Mohan and Pittman, 2007; Yang et al., 2013). The main disadvantage of chemical process is that it is generally suitable to treat contaminated water having high concentration of As and is less effective when the concentration is low (Chiban et al., 2012). 8.2. Coagulation/flocculation One of the widely used treatment method for As removal involves coagulation followed by flocculation. Coagulation is a process in which colloidal particles are destabilized by neutralizing those opposite forces which keep them separate. The cation based coagulants offer positive electric charges in order to neutralize the negatively charged colloids, thus the particles combine to make larger particles and this is achieved by rapid mixing of solution which spreads the coagulant agent completely (Cui et al., 2014). On the other hand, in flocculation various polymers are used to make bridges between the flocs and these polymers help to bind large clumps or agglomerates of the colloids. A negatively charged flocculent will bind cationic particles via sorption and results in destabilization either by forming a bridge or neutralization of opposite charges (Luqman et al., 2013). Iron salts such as ferric chloride and ferric sulfate are most successful among various types of coagulants studied in the past for As removal from water (Hu et al., 2012). The major drawback of the coagulation/flocculation process is the large amount of sludge produced having substantial As concentration. 8.3. Ion exchange Ion exchange is a process which removes As and other metals from contaminated water. It is a promising and widely used treatment method for As removal. Complete removal and recovery of subjected metal(loid) and minimum/no production of toxic sludge are some of the important benefits of this method compared to other conventional treatment methods. Natural or synthetic ion-exchange resins are used in this process which has the explicit potential to exchange anion with

764


the As in contaminated water. In past years, use of natural zeolites (clinoptilolite, montmorillonite, kaolinite) has also been encouraged due to their high sorption potential, metal(loid) selectivity, low-cost and biodegradable nature (Li et al., 2011; Wu et al., 2014). Although this method is well-established and promising but ion-exchange resins must be rejuvenated consistently to maintain and achieve complete removal of As and other metals which could increase the cost of the whole system along with toxic sludge generation (Chiban et al., 2012; Shakoor et al., 2016).

significantly influenced by the occurrence of other competing ions like phosphate, sulfate, carbonates and biocarbonates calcium, magnesium, sodium competing binding sites (Shakoor et al., 2016). Recently, some studies have investigated the use of novel materials like alumina (Han et al., 2013), clays (Anjum et al., 2011), iron oxides (Sun et al., 2013), activated carbon (Zhang et al., 2008), zeolites (Swarnkar and Tomar, 2012), agricultural/food industry biowastes (Abid et al., 2016. Shakoor et al., 2016) etc. as sorbents for As removal from water.

8.4. Membrane filtration

9. Ultimate fate of arsenic-laden sludge and wastes

In the membrane filtration process, synthetic membranes are used which allow certain dissolved compounds to pass through but retains some others. Membranes can be employed as efficient filters to remove As from contaminated water due their ability of selective permeability. High pressure is needed for water to go through the membrane from a concentrated solution to dilute solution, thus due to this reason membrane filtration is known as a pressure driven process (Basu et al., 2014; Shakoor et al., 2016). Recently, membrane filtration technology has been upgraded as electro-ultrafiltration which showed high potential in eliminating As from contaminated water (Mostafazadeh et al., 2016).

All the above methods produce sludge/wastes containing high concentrations of As that could be a risk to environmental health, hence for safe disposal of the toxic wastes or sludge produced, proper handling and care are required. There are a number of methods used to manage As-loaded wastes worldwide. For example, in West Bengal, India, people use pit holes to dump the sludge (produced from domestic As treatment systems) far away from food crops and children's playing sites (Sarkar et al., 2010). Whereas, in commercial and municipal As removal systems, As-loaded sludge/wastes results due to reaction with the sorbents thereby slowly escalate head loss in the column and decrease the flow rate. Once every day, it is necessary to backwash the column and collects As-loaded material from the top of the filter. Sorbents used to treat As-contaminated water are regenerated sometimes in the central regeneration system and following the treatment, spent regenerates produce As-loaded sludge (Sarkar, 2006). Chemically, both forms of wastes are similar as both contain significant amount of As. Most often, environmental laws dealing with the safe disposal of Asloaded wastes/sludge in third world countries either do not exist or they are not enforceable. Therefore, removal of As-loaded wastes/ sludge having a negative ecological impact and human health risk is equally important as the elimination of As is to supply safe and Asfree drinking water. In recent years, As-loaded sludge and wastes are disposed of regularly in landfills in developing countries like India and Bangladesh, although, several recent studies have publicized that As leaching is stimulated or increased in a landfill or in a poorly managed waste burial site (Delemos et al., 2006). Both pH and redox potential distinctively regulate speciation of As in landfills that in turn control leachability of As (Stumm and Morgan, 2012). In an oxidized environment As leaching is minimized while in a reduced environment As is mobilized in a landfill environment (Sarkar et al., 2008). In some developing nations, for example, in India, As wastes are dumped openly in waste sites or the treatment residuals are stored inside aerated sand filters. The coarse sand filter is designed with the facility of air venting pipes in such a way that its inside possesses high ventilation. The solids present in the sludge stick inside the sand filter whereas the solution infiltrates through the sand bed to soil beneath (Rahman et al., 2015). Some other treatment technologies such as compaction, incineration, composting, pyrolysis and direct disposal could be exploited for As-loaded wastes or sludge. However, a detailed systematic research is required to test the applicability of these technologies for safe disposal of As-loaded wastes or sludge and to avoid secondary pollution of the environment (Sas-Nowosielska et al., 2004).

8.5. Floatation It is another separation method for As removal which involves hetero-phase system (suspensions and liquid emulsions detached from dispersed phase). Floatation is achieved by using air bubbles and resulting flocks are gathered and eliminated at the top (Yenial et al., 2014). The major types of floatation process used commonly for elimination of As from contaminated water are (i) dissolved air flotation, (ii) ion flotation and (iii) precipitation flotation. 8.6. Phytoremediation Phytoremediation is an ecofriendly method which uses plants and microbes for remediation of metal(loid)-contaminated soils and water (Adrees et al., 2015; Rizwan et al., 2016, 2017; Tauqeer et al., 2016). Although there are various types of phytoremediation but in the recent past, phytofiltration has emerged as a very environmental-friendly and cheap option for remediation of As from contaminated water streams. Phytofiltration is based on several steps, (a) selection of the most efficient plants which can remove contaminant from water by absorbing in roots, and (b) relocation of screened plants to a built wetland in which As removal from water takes place (Farooq et al., 2016a, 2016b). The plant with greater As removal potential have been discovered in the last few years, i.e. P. vittata (Ma et al., 2001). Furthermore, aquatic macrophytes including Lemna minor, Eichhornia crassipes and Lessonia nigrescens were mostly considered suitable for immobilization of As in contaminated water streams specially surface water i.e. river, lake and stream, etc. (Buschmann et al., 2007). 8.7. Sorption In sorption process, different solid materials are used to remove As and other metals either from gaseous or liquid solutions. Sorption has following advantages due to which it is getting more attention: (i) usually it does not need additional chemicals and large volume (ii) it is easier to arrange as a point of entry or point of use (POE/POU) As removal procedure, and (iii) it avoids the production of harmful by-products (Zhang et al., 2008) and more importantly it is cost effective. Arsenic removal by sorption depends on the pH and speciation of As, thus pH b 7 shows better As(V) removals as compared to the As(III) while pH 7–9 is highly suitable for maximum As(III) removal from water (Kamsonlian et al., 2012; Abid et al., 2016). The As sorption rate from water is

10. Conclusions and future research Arsenic occurs in different chemical forms in groundwater including inorganic (As(III), As(V)), and organic forms (MMA, DMA). Exposure to As is largely contributed to direct intake of drinking water in humans or via consumption of edible crops irrigated with As-contaminated water. Groundwater used for drinking purpose or cooking of food items and eating rice (As-laced) as a staple food are the main pathways for As entry into the people of As-contaminated regions. There could be a greater risk of As uptake in crops grown on soils irrigated with Ascontaminated groundwater. The chance of As exposure is significantly


higher for the people, especially children and women in less developed low-income communities having insufficient food and lack of access to safe drinking water. Availability of As-free safe and healthy drinking water that is also used for cooking purpose, and rice with minimum or no As must be the focus of mitigation plans to protect people from lethal As intakes in rural communities of less developed countries. Furthermore, previous research work has mainly focused on human exposure to the total As by groundwater, comparatively a little data is available on the role of individual As species and this is an existing research gap which needs to be addressed in future research. Human health risk assessment models/frameworks that cover all sources of As exposure are rarely reported, although As sources and exposure are extensively explored. Among all the health risk assessment models/frameworks, US-EPA models are considered to be more reliable compared to others, hence these models must be explored further to determine carcinogenic and non-carcinogenic risk caused by individual As species in groundwater. Moreover, current treatment technologies for As-contaminated water have a number of disadvantages, and the wastes or sludge generated can be a potential source of secondary pollution. Thus, for better protection of our environment from As, new hybrid technologies are needed accompanied by safe disposal options for As-loaded wastes/ sludge. Acknowledgements Authors acknowledge the financial support from Government College University, Faisalabad, Pakistan. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2017.05.223. References Abdallah, E.A., Gagnon, G.A., 2009. Arsenic removal from groundwater through iron oxyhydroxide coated waste products. Can. J. Civ. Eng. 36 (5), 881–888 (A paper submitted to the Journal of Environmental Engineering and Science). Abedin, M.J., Cotter-Howells, J., Meharg, A.A., 2002. Arsenic uptake and accumulation in rice (Oryza sativa L.) irrigated with contaminated water. Plant Soil. 240 (2), 311–319. Abid, M., Niazi, N.K., Bibi, I., Farooqi, A., Ok, Y.S., Kunhikrishnan, A., Ali, F., Ali, S., Igalavithana, A.D., Arshad, M., 2016. Arsenic (V) biosorption by charred orange peel in aqueous environments. Int. J. Phytoremediation 18 (5), 442–449. Adrees, M., Ali, S., Rizwan, M., Rehman, M.Z., Ibrahim, M., Abbas, F., Farid, M., Qayyum, M.K., Irshad, M.K., 2015. Mechanisms of silicon-mediated alleviation of heavy metal toxicity in plants: a review. Ecotoxicol. Environ. Saf. 119, 186–197. Agusa, T., Kunito, T., Fujihara, J., Kubota, R., Minh, T.B., Trang, P.T.K., Iwata, H., Subramanian, A., Viet, P.H., Tanabe, S., 2006. Contamination by arsenic and other trace elements in tube-well water and its risk assessment to humans in Hanoi, Vietnam. Environ. Pollut. 139 (1), 95–106. Ahsan, H., Chen, Y., Parvez, F., Argos, M., Hussain, A.I., Momotaj, H., Levy, D., Van Geen, A., Howe, G., Graziano, J., 2006. Health effects of arsenic longitudinal study (HEALS): description of a multidisciplinary epidemiologic investigation. J. Expo. Sci. Environ. Epidemiol. 16 (2), 191–205. Aldroobi, K.S.A., Shukri, A., Bauk, S., Munem, E.M.A., Abuarra, A.M., 2013. Determination of arsenic and mercury level in scalp hair from a selected population in Penang, Malaysia using XRF technique. Radiat. Phys. Chem. 91, 9–14. Amonoo-Neizer, E.H., Amekor, E., 1993. Determination of total arsenic in environmental samples from Kumasi and Obuasi, Ghana. Environ. Health Perspect. 101 (1), 46–49. Andreae, M., Andreae, T., 1989. Dissolved arsenic species in the Schelde estuary and watershed, Belgium. Estuar. Coast. Shelf Sci. 29 (5), 421–433. Anjos, V.E., Anjos, V.E., Machado, E. d. C., Grassi, M.T., 2012. Biogeochemical behavior of arsenic species at ParanaguÃ¡ estuarine complex, southern Brazil. Aquat. Geoche. 18, 407–420. Anjum, A., Lokeswari, P., Kaur, M., Datta, M., 2011. Removal of As (III) from aqueous solutions using montmorillonite. J. Anal. Sci. Methods Inst. 1, 25. Arain, M.B., Kazi, T.G., Baig, J.A., Jamali, M.K., Afridi, H.I., Shah, A.Q., Jalbani, N., Sarfraz, R.A., 2009. Determination of arsenic levels in lake water, sediment, and foodstuff from selected area of Sindh, Pakistan: estimation of daily dietary intake. Food Chem. Toxicol. 47 (1), 242–248. Asante, K., Agusa, T., Kubota, R., Subramanian, A., Ansa-Asare, O., Biney, C., Tanabe, S., 2008. Evaluation of urinary arsenic as an indicator of exposure to residents of Tarkwa, Ghana. West Afr J. App. Ecol. 12. Ashley, P., Lottermoser, B., 1999. Arsenic contamination at the mole river mine, northern New South Wales. Aust. J. Earth Sci. 46 (6), 861–874.

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