ARSENIC DISTRIBUTION IN SOIL: Generally, As concentrations in uncontaminated soils seldom exceed10 mg kg - 1. However, anthropogenic sources of As have elevated the background concentration of As in soils. For example, in areas near As mineral deposits, As levels in soils may reach up to 9300 mg kg - 1.
Depending on the nature of the geogenic and anthropogenic sources, As concentration in soils can range from <1 to 250,000 mg kg -1. However, there is a large fluctuation among countries due to variation in soil parent material; for example, calcareous soils can be expected to have higher levels of As than non-calcareous soils. Although the dominant source of As in soils is geological, additional inputs may also be derived locally from industrial sources, such as smelting and fossil fuel combustion products and agricultural sources, namely pesticides and phosphatic fertilizers.
In soils, As forms a variety of inorganic and organic compounds. Arsenic forms solid precipitates with Fe, aluminum (Al), calcium (Ca), magnesium (Mg), and nickel (Ni). A number of studies involving solid phase speciation have shown that As is prevalent mostly in the oxalate fractions associated with amorphous and crystalline Fe and Al oxides, indicating the strong affinity of As for these soil components. The soluble As concentration in soil is largely determined by redox conditions, pH, biological activity, and adsorption reactions. The adsorption and mobility of As in soil are affected more strongly by the presence of H2PO2 ion than any other anions. Arsenic is subject to both chemical and biological transformations in soils, resulting in the formation of various species.
DISTRIBUTION IN AQUATIC EVIRONMENT
Arsenic in an aquatic environment is distributed in both the aqueous solution and sediments. Elevated concentrations of As in natural waters are usually associated with As-rich sedimentary rocks of marine origin, weathered volcanic rocks, fossil fuels, geothermal areas, mineral deposits, mining wastes, agricultural use, and irrigation practices. Uncontaminated waters usually contain less than 0.001-g As liter-1.
In contaminated areas, however, high levels of As have been reported in water bodies. It should be noted that considerable variation in As concentration exists within the same geological area as reported by different researchers. As discussed earlier, one of the principal causes of high As concentrations in subsurface waters is the reductive dissolution of hydrous Fe oxides and/or the release of adsorbed As proposed that the release was primarily due to reduction (and dissolution) of “ferric arsenates” instead of changes in the As speciation. The high As in groundwater can be associated with reducing conditions, resulting from the presence of dissolved organic carbon, particularly in alluvial and delta environments. While the exact mechanisms responsible for this remain uncertain, it is possible that both reductive dissolution and desorption of As from oxides and clay play an important role in elevating As concentration.
A significant proportion of As in aquatic environment is derived from the sediments, and the relative distribution of As in water and sediments depends mainly on the nature and amounts of sediments. The As-rich sediments act as a buffer in maintaining the As concentration in water bodies, thereby controlling the dynamics and bioavailability of As in the aquatic environment.
CHEMICAL FORM AND SPECIATION
Speciation of metalloids can be achieved by both analytical processes and on the basis of theoretical consideration. The analytical processes involved in the speciation of metalloids in soils can be grouped into solid phase speciation and solution-phase speciation. In view of the limitations of many of the analytical procedures used in speciation, often species distribution is predicted using a number of speciation models that are based on theoretical chemical (thermodynamic) concepts. Although the fundamental thermodynamic principles that drive these models are based on scientific facts, problems arise when these principles are applied to complex natural matrixes.
A large number of sequential extraction schemes have been used for soils, generally attempting to identify metalloids held in any of the following fractions: soluble, exchangeable, sulfide/carbonate bound, organically bound, oxides bound, and residual or lattice mineral bound. The bioavailability of metalloids in soils has been examined using the physiologically based in vitro chemical fractionation schemes that include the physiologically based extraction test (PBET), potentially bioavailable sequential extraction (PBASE), and gastrointestinal (GI) test. These innovative tests predict the bioavailability of metalloids in soil/sediments when ingested by animals and humans.
A vast number of analytical techniques are available for solution-phase characterization and quantification of metalloids. These include electro-analytical techniques, cation/anion-exchange resins and chemical adsorbents to fractionate ionic and nonionic forms, ultrafiltration, dialysis, and gel permeation techniques for molecular size fractionation, spectroscopic techniques measuring the oxidation state of elements, X-ray techniques to measure trace element distribution, and chromatographic techniques to measure the phase distribution of metalloids. Arsenic speciation is determined by both biotic and abiotic variables.
Arsenic speciation is important not only for understanding the biogeochemical cycling of As in different ecosystems and mechanisms of As accumulation and detoxification, but also for designing safe disposal options of As-rich biomass. In soil, As occurs both as inorganic [As(III) and As(V)] and as organic forms. Trivalent As can exist as arsenous oxide (As2O3), arsenious acid(HAsO2), arsenite (H2AsO-3, HAsO2-3 , AsO3-3 ) ions, arsenic trichloride(AsCl3), arsenic sulfide (AsS3), and arsine (AsH3). Pentavalent As commonly occurs as arsenic pentoxide (As2O5), orthoarsenic acid (H3AsO4), metaarsenic acid (HAsO3), and arsenate (H2AsO-4, HAsO2-4, AsO3-4 ) ions.
The presence of different forms of organic As, such as mono-methylarsonic acid[MMA, CH3AsO(OH)2], dimethylarsenic acid [DMA, (CH3)2AsO(OH)],trimethylarsine oxide [(CH3)3AsO], methylarsine (CH3AsH2), dimethylarsine[(CH3)2AsH] and trimethylarsine [TMA, (CH3)3As], has also been observed in contaminated soil and water. The most common forms of As in the environment are the inorganic oxy-ions of As(III) and As(V). Arsenite [As(III)] is more toxic and relatively mobile in contaminated soils, whereas arsenate [As(V)] is relatively less toxic. Both As(III) and As(V) compounds are highly soluble in water and may change valency states depending on the pH and redox conditions. Results of a literature search on the speciation of As in environmental and biological samples are presented in Table 1.
In contaminated soils, generally As(V) predominates over As(III), whereas in waters, the relative proportion of these two species varies depending on a number of factors, including As sources, redox potential, pH, and microbial activity and studied the influence of redox potential and pH on As speciation and solubility in a contaminated soil. They observed that alterations in the oxidation state of As, as influenced by redox potential and pH, greatly affected its solubility in soil. A toxic redox levels (500–200 mV), As solubility was low and the major part (65–98%) of the As in soil solution was present as As(V).
At alkaline pH, the reduction of As(V) to As(III)released substantial proportions of As into solution. Under moderately reducing conditions (0–100 mV), As solubility was controlled by the dissolution of Fe-oxyhydroxides. At an anoxic redox level of -200 mV, soluble As increased 13-fold as compared to an toxic redox level of 500 mV. The apparent slow kinetics of the As(V) to As(III) transformation and the high concentrations of manganese (Mn) present indicate that, under reducing conditions, As solubility could be controlled by the Mn3(AsO4)2 phase.