to its analogue S, it is widely distributed in environment as a major and minor constituent of most of the sulfide ores (Cooper et al. 1970) or as selenides of nickel (Ni), copper (Cu), silver (Ag), lead (Pb), and Mercury (Hg). Uranium ore contain highest (~600 μg/g) of Se content (Ralston et al. 2009). Rocks contain around 40% of the Se of the total of Earth crust (Wang and Gao 2001), values reported for igneous rocks (0.35 μg/g) (Fordyce 2005), sedimentary rocks (0.0881 μg/g) (Tamari et al. 1990), shales (0.24–277 μg/g) (Lakin and Davison 1967), phosphatic rocks (1.4–178 μg/g) (Robbins and Carter 1970), coal (1–5 μg/g) (Cooper et al. 1970), limestone (0.03–0.08 μg/g) (Fordyce 2005), and sandstone (0–112 μg/g) (Lakin and Davison 1967). Leaching from these Se‐rich sources can elevate the Se concentration in environment up to 1200 μg/g (Paikaray 2016). Rosenfeld and Beath (1964) have compiled the data of Se concentrations in rocks and seleniferous soils. These seleniferous rocks are the major source of Se in soil, ground water, and atmosphere.
The distribution of seleniferous rocks and deposits in geologic and topographic map indicate the areas of Se‐rich soil and water. Soil and water at distant places can become seleniferous when soluble and/or suspended Se species are imported to such areas through surface water. In the environment, concentration of Se varies from place to place; this uneven distribution of Se is mainly governed by the processes including weathering, interaction of rocks and water, and microbial activities. These processes control the transportation of Se from rocks to soil, water, and air. In soil the amount of Se is primarily influenced by parent material and possible leaching from rocks during soil formation. The confined environmental surroundings such as properties of soil, aeration, organic matter, pH, and microbial activity are also the responsible factors for Se distribution in soil. The average range and mean global concentration of Se in soil is 0.01–2 and 0.4 μg/g, respectively (Dungan et al. 2002). Further, the reactivity and bioavailability of Se, in addition to total concentration, also depends on different chemical forms available in soil and water. Here bioavailability means the part of substance that becomes soluble and accessible for adsorption by means of membrane (Reeder et al. 2006; Ruby et al. 1996) of living organisms.
The toxic, tolerable, and deficient areas of Se level exist alongside and for these different Se levels local environmental conditions are responsible. Depending upon the sampling done for the available Se concentration in vegetation and plants grown in the soil, seleniferous soil has been categorized as toxic, moderate, and low level of Se. Soil that provides an adequate amount of Se to make toxic plants is referred to as toxic seleniferous soil. Contrary to this, the soil may have high Se level as exhibited by toxic Se soil but provide less Se to the plants, known as non‐toxic seleniferous soil. From deficient to most‐seleniferous soil, the concentration of Se reported is 0.01 and 1200 μg/g (Fleming 1980; Jacobs 1989; Neal 1995). Many countries have elevated level of Se including USA (Presser 1994), India (Dhillon and Dhillon 2003), Ireland (Seby et al. 1997), and China (Wang and Gao 2001). Central region and Great Plains of North USA and Prairie region of Canada (Ihnat 1989) is formed from Cretaceous shale (2 μg/g) and exhibit relatively high concentration. In Australia, Ireland, and various other countries with toxic Se level, shales are the parent material (Johnson 1975). Florida, South Carolina, and Tennessee are ranging lower (0.8–9 μg/g) in Se due to phosphatic rocks of that region (Rader and Hill 1935). A low level of Se has been documented in Finland and New Zealand. Se content in Hawaiian and Japanese volcanic sulfur ranged from 1026 to 2000 and 67–206 μg/g, respectively (Lakin and Davison 1967). Many parts of Africa were recognized with low Se; however, in Asia both high and low Se concentrations have been reported (National Research Council 1983). Most of the parts of the world are characterized as moderate to low bioavailability as compared to high Se soil content. Among most studied Se‐contaminated water bodies, Kesterson Reservoir of San Joaquin Valley, California USA is one of them. The main Se source is Se‐rich marine sedimentary rocks (mean values = 8.9 μg/g) of the coastal range, which raise the Se content to the reservoir by weathering and other beneath mechanisms (Milne 1998; Presser and Piper 1998). Human and industrial activities are also responsible for the discharge of Se in rivers and lakes. Dhillon and Dhillon (2003) have compiled a comprehensive review on seleniferous soils.
Water applied for irrigation purpose to soil can mobilize the soluble Se species from upper soil surface and create a hydraulic gradient, and as a result Se is discharged to surface water. Apart from natural Se sources, this agriculture drain water and other anthropogenic activities can also discharge Se to water bodies and is available to aquatic life. The aquatic bodies have also played an important role in the transportation of Se from one place to other. The estimated value for the fluxes of Se is ~14 000 tons per year that are relocated from continents to oceans along aquatic pathway (Nriagu 1989, 1991). The mode of Se transfer is either via selenium‐bearing sediments (~85%) or through soluble species (Cutter and Bruland 1984). Leaching from soil to groundwater is another segment of hydrological transport of selenium species. However, contribution through leaching processes is not very clear because of the heterogeneity of the interfaces between water‐rocks, ‐soils and ‐sediment (Dhillon et al. 2008). Ihnat (1989) has reported that the average residence period of selenium in the deep ocean is ~1100 years, almost of the order of residence time of water. The soluble selenium species from industrial waste move downward through infiltration and elevate the Se level in ground water during monsoon season (Kumar and Riyazuddin 2011). Se concentration in water is ranged between 1 and 10 μg/l; due to the emission of geogenic and/or anthropogenic Se, this may exceed 100 μg/l occasionally in outstanding cases. In industrial effluents Se concentration is in 10–100 μg/l range, under rare conditions it can exceeds up to 1000 μg/l. Typical Se concentrations in ambient water are <1 μg/l in the dearth of direct Se sources. The existing Se concentration of marine waters is 0.02–0.04 μg/l. However, for fresh water the background Se concentration is considered to be comparable to marine water. Ghosh et al. (2008) reported that the elevated concentration of Se in soil (1100 μg/l) and water (3 μg/g) in localities of Mumbai and Punjab of India is due to the alkaline pH (pH 8–9.2) of soil in industrially polluted areas. The concentration of Se is high in the saturated water aquifers which have fluctuations in their water table. Se concentration in aquifers also depends on mixing ratio of discharge stream into receiving water. In addition, the probable interaction of discharge to Se removing sediments is relatively more important than its dispersion in aqueous phase. The water quality criterion for Se in water is primarily determined by ecotoxicological considerations for aquatic living organism, e.g. predatory fish and waterfowl, which are at the top of aquatic food chains. Countries like the US and Canada provide regulatory guidelines that include 5 and 1 μg/l (with varying guidelines for individual regions up to 100 μg/l), respectively (US Environmental Protection Agency 2008). The relationship between Se concentration in water and environmental toxicology is not fully understood because of the existing complexity of the biogeochemical Se cycle. However, several efforts are in progress to establish water quality measures on considering different ecosystems that are subjected to emission of Se in particular areas specific for sites.
In addition to this, adsorption and desorption of elements, precipitation of minerals, and incineration of municipal wastes (Plant et al. 2004) have also contributed to the insertion of Se into atmosphere. Organometallic compounds of Se are introduced partly to the atmosphere by chemical or microbial redox reactions and to soil and water by metabolic uptake and release by animals and plants (McNeal and Balistrieri 1989). Consequently, it enters into the food chain through crops, plant, and aquatic lives (Paikaray 2016). Frost (1967) has reported that sea water, earth crust, animals, and plants contain 0.004, 0.09, 1–20, and 0.02–4000 μg/g, respectively, which indicates that plants and animals have ability to concentrate Se from earth crust. It again enters into environment through the decomposition of these species and has excreted from human body (~50–80% through urine) to environment. For this in‐and‐out pathway of Se into environment, several cycles have been proposed including geological cycling of Se where animals and plants had a role, proposed by Moxon et al. (1939) and Lakin and Davidson (1967). Shrift (1964) and Frost (1973) have proposed a biological cycle of Se in which involvement of reduction–oxidation (redox) reactions of Se by plants, bacteria, and fungi have been incorporated. Allaway et al. (1967) and Olson (1967) have reviewed the cycling of low and high levels, respectively, of Se in soils, plants, and animals.
Se forms more than 170 solid compounds, whereas very limited compounds