Arsenic occurs naturally as an element, ranks as the 20th most occurring trace element in the earth's crust (NRC 1977) and is widely distributed in the environment. The ultimate source of arsenic on the Earth's surface is igneous activity (Nriagu 1994). Arsenic is widely spread in the upper crust of the Earth, although mainly at very low concentrations, with arsenic concentrations in soil ranging from 0.1 to more than 1,000 ppm (mg kg-1). In atmospheric dust, the range is 503,400 ppm. In seawater, the average arsenic level may be 2.6 ppb and in fresh water about 0.4 ppb. Arsenic at significant levels is all around us (Mukhopadhyay et al. 2002). The global arsenic geocycle elucidates how arsenic enters into the soil, sediment, water, and food chain of living organisms (Fig. 12.2). The major source of As contamination is from naturally existing minerals; however, anthropogenic activities have also contributed extensively (Nordstrom 2002). A range of As compounds, both organic and inorganic, are introduced into the environment through geological (geo-genic) and anthropogenic (human activities) sources. Small amounts of As also enter the soil and water through various biological sources (biogenic) that are rich in As. Although the anthropogenic source of As contamination is increasingly becoming important, it should be pointed out that the recent episode of extensive As contamination of groundwaters in Bangladesh and the Indian state of West Bengal is of geological origin, transported by rivers from sedimentary rocks in the Himalayas over tens of thousands of years, rather than anthropogenic.
Arsenic exists in several oxidation states (—3, 0, +3, and +5), enabling it to mobilize under various environmental conditions and hinders many remediation technologies from efficiently removing it from water. Under oxidizing conditions, As(V) is the dominant form at lower pH while As(III) becomes dominant at higher pH. However, the uncharged form of As(III) [As(OH)3] becomes dominant under reducing environments, which is more toxic and difficult to remove (Smedley and Kinniburgh 2002). Its association with some non-weathering-resistant mineral deposits (e.g., sulfide minerals) has contributed to its release in large amounts into the environment (Murdoch and Clair 1986).
Arsenic is also being introduced into the environment through various anthropogenic activities. These sources release As compounds that differ greatly in chemical nature (speciation) and bioavailability. Major sources of As discharged onto land originate from commercial wastes (40%), coal ash (22%), mining industry (16%), and the atmospheric fallout from the steel industry (13%) (Eisler 2004; Nriagu and Pacyna 1988). Arsenic trioxide (As2O3) is used extensively in the manufacturing of ceramic and glass, electronics, pigments and antifouling agents, cosmetics, fireworks, and Cu-based alloys (Leonard 1991). Arsenic is also widely used for wood preservation in conjunction with Cu and chromium (Cr), i.e., copper-chromium-arsenate (CCA). The use of sodium arsenite (NaAsO2) to control aquatic weeds has contaminated small fish ponds and lakes in several parts of the United States with As (Adriano 2001). Arsenic contamination in soil was also reported due to the arsenical pesticides used in sheep and cattle dips to control ticks, fleas, and lice (McBride et al. 1998; McLaren et al. 1998). A study of 11 dip sites in New South Wales indicated considerable surface soil (0-10 cm) contamination with As (37-3,542 mg kg—1) and significant movement of As (57-2,282 mg kg—1) down the soil profile at 20-40 cm depth (McLaren et al. 1998). Continuous application of phosphatic fertilizers that contain trace levels of As also results in As contamination of soil (Peng et al. 2011), thereby reaching the food chain through plant uptake (McLaughlin et al. 1996). Similarly, in New Zealand, timber treatment effluent is considered to be the major source of As contamination in aquatic and terrestrial environments (Bolan and Thiyagarajan 2001). Because As is widely distributed in the sulfide ores of Pb, Zn, Au, and Cu, it is released during their mining and smelting processes. The flue gases and particulate from smelters can contaminate nearby ecosystems downwind from the operation with a range of toxic metal(loid)s, including As (Adriano 2001). Coal combustion not only releases gaseous As into the atmosphere, but also generates fly and bottom ash containing varied amounts of As. Disposal of these materials often leads to As contamination of soil and water (Beretka and Nelson 1994).
Arsenic is present in many pesticides, herbicides, and fertilizers. Arsenic may accumulate in agricultural soils due to the agricultural practices such as the applications of As-containing pesticides and herbicides, pig manure, and phosphorous fertilizers, and it has raised more concerns about the risk of As to the environment and human health (Chirenje et al. 2003; Brouwere et al. 2004; Li and Chen 2005). Industries that manufacture As-containing pesticides and herbicides release As-laden liquid and solid wastes that, upon disposal, are likely to contaminate soil and water bodies. For example, indiscriminate discharge of industrial effluents from the manufacturing of Paris Green pesticide [Cu(CH3COO)2-3Cu(AsO2)2] resulted in the contamination of soil and groundwater in residential area of Calcutta, India (Chatterjee and Mukherjee 1999). The use of horticultural pesticides, lead arsenate (PbAsO4), calcium arsenate (CaAsO4), magnesium arsenate (MgAsO4), zinc arsenate (ZnAsO4), zinc arsenite [Zn(AsO2)2], and Paris Green in orchards has contributed to soil As contamination in many parts of the world (Merry et al. 1983; Peryea and Creger 1994). Soil contamination due to the use of organoarsenical herbicides such as monosodium methanearsonate (MSMA) and disodium methanearsonate (DSMA) was also reported (Merry et al. 1983; Peryea and Creger 1994). The accumulated As in agricultural soils can distribute among different soil components, such as organic matter, iron (Fe) and manganese (Mn) oxides, carbonates and sulfides, and such distribution could affect its mobility, bioavailability, and toxicity (Cummings et al. 1999; Islam et al. 2004; deLemos et al. 2006). The distribution and redistribution process of As in soils can be influenced by microbial activities, because microbes could mediate the transformation of As and As adsorbents (Bentley and Chasteen 2002; Mukhopadhyay et al. 2002; Oremland and Stolz 2003, 2005; Islam et al. 2004). It is reported that microbially mediated As release to the groundwater for drinking has threatened the health of millions of people in Bangladesh, West Bengal, and some regions of China (Smith et al. 2000; Shen et al. 2005).
The adverse effects of As in groundwater used for irrigation water on crops and aquatic ecosystems is also of major concern. In addition to potential human health impacts caused by ingestion of food containing As, the potential for reduced crop yield due to its buildup in the soil is an active area of research. The fate of As in agricultural soils is often less well studied compared to groundwater, and in general has been studied in the context of As uptake by different plants (Huq et al. 2001; Das et al. 2004; Al Rmalli et al. 2005; Correll et al. 2006). Crop quality and the effect of As on crop quality and yield is becoming a major worldwide concern, particularly for rice which forms the staple for many South Asian countries where groundwater is widely used for irrigation (Meharg and Rahman 2003). In a recent study it was reported that irrigation has increased in Bangladesh since 1970, while since 1980, the area under groundwater irrigation for the cultivation of Boro (winter) rice has increased by almost an order of magnitude (Harvey et al. 2005). Based on available information on the distribution of As concentration in groundwater (BGS and DPHE British Geological Survey and Department of Public Health Engineering 2001) and the area under shallow tubewell irrigation (BADC 2005), Saha (2006) estimated that approximately 1,000 Mg of As is cycled with irrigation water during the dry season of each year. Rice yield has been reported to decrease by 10% at a concentration of 25 mg kg-1 As in soil (Xiong et al. 1987). Table 12.1 shows how As might accumulate in soil over time at different concentrations in irrigation water, assuming an annual water application of 1,000 mm. A soil irrigated with 1,000 mm of water containing 100 ppb As receives 1 kg ha-1 As per annum. The limited evidence at present suggests that the safe limit of soil As for rice lies somewhere in the range 25-50 mg kg-1 (Saha and Ali 2007; Duxbury and Panaullah 2007).
Table 12.1 Potential effect of arsenic concentrations in irrigation water on soils with time (from Brammer and Ravenscroft 2009)
Years of Arsenic in irrigation water (ppb)
Arsenic added to soil
50 14 28 70 140 280
In Table 12.1, the soil arsenic concentration values between 25-50 and >50mgkg-1 As have been shown when, in principle, these soil concentrations might be reached (Brammer and Ravenscroft 2009). Actual soil loading rates will vary with the amount of irrigation water applied, As concentrations in the water, and losses due to volatilization, leaching, and crop removal. Not all the As delivered by tube wells actually reaches the fields irrigated. In most As-affected areas of South and Southeast Asia, groundwater is rich in iron (e.g. DPHE 1999; Gurung et al. 2005; Postma et al. 2007). That iron is oxidized when the water is exposed to the air and is then precipitated as iron hydroxides which adsorb As. Hossain (2005) reported that topsoil arsenic concentrations at the Faridpur site, which had been irrigated for about 20 years, ranged from 61 mg kg-1 in the field nearest the wellhead to 11 mg kg-1 in a field near the far side of the command area. Williams et al. (2006), who measured As contents of 37 vegetables, pulses, and spices commonly grown in Bangladesh, found levels were highest in radish leaves (0.79 mg kg-1), arum stolons, spinach, and cucumber, and lowest (0.2 mg kg-1) in most fruits, vegetables, and spices. Roychowdhury et al. (2002) also found great differences between 30 crops and food items from 34 As-affected households in West Bengal, India, inter alia reporting a significant difference between potato skins (0.526 mg kg-1) and potato flesh (0.00728 mg kg-1). A greenhouse study by Abedin et al. (2002) revealed reduced yield of a local variety of rice (BR-11) irrigated with water having As concentrations in the range of 0.2-8 mg L-1. The accumulation of As in rice field soils and its introduction into the food chain through uptake by the rice plant is of major concern (Duxbury et al. 2003).
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