BEHAVIOUR OF INORGANIC ARSENIC

Identification of the aqueous-solid-phase-interactions between aquifer sediments and the surrounding groundwater are the key to predict As mobility in aquifer systems. These interactions are strongly influenced by the activity of autochthonous microbes that interfere with the hydrochemical composition of groundwater and the mineralogical inventory of the aquifer sediments. According to the prevailing conditions, sediments can act both, as source and as sink for As. Important processes are complexation,

redox-have the most important influence on the mobility of As. Surface adsorbed As, which is either weakly adsorbed (electrostatic attraction) or strongly bound (ligand exchange), is easily accessible to interactions with dissolved compounds and microbes. Both, As(III) and As(V), have high binding affinities for Fe-(oxyhydr)oxides (e.g., goethite) and form strong bidentate complexes via ligand exchange (DIXIT & HERING 2003, MÜLLER et al.

2010, ONA-NGUEMA et al. 2005). Especially amorphous hydrous Fe-oxides and poorly crystalline Fe-(oxyhydr)Fe-oxides (e.g., ferrihydrite) have large surface areas, resulting in a chemical reactivity that is far out of proportion to their abundance (BORCH et al. 2010). Iron-(oxyhydr)oxides often occur in sediments as alteration products in form of partial coatings around mineral grains and act as important sinks for many trace elements including As (EICHE at al. 2010, GUO et al. 2007). To a lesser extent, As in sediments is associated with Mn- and Al-(oxyhydr)oxides, clay minerals, sulphates, calcium carbonates and organic acids (O’DAY 2006).

It is very difficult to estimate the sorption behaviour of dissolved As in a certain system, which is primarily influenced by pH and redox state of the solution and the presence of adsorbing mineral phases (DIXIT & HERING 2003, GOH & LIM 2004). Despite the complex nature of a multi-component system like natural aquifers, previous studies that examined the sorption behaviour of As used strongly simplified experimental setups (MOHAN &

PITTMAN 2007). Additionally, the similar sorption behaviour of PO4

and other anions plays an important role regarding competitive ion exchange and adsorption (POSTMA et al. 2007). In case of a tropical soil rich in Fe-(oxyhydr)oxides, As exchange potentials declined in order of PO43-

>> CO3

2-> SO42- ≈ Cl^- (GOH & LIM 2005). Arsenic retention through adsorption generally depends on respective flow velocities, available binding sites and adsorption partners, as well as concentrations of competing ions and solutes in groundwater.

2.1.4 TOXICITY

Arsenic has influenced human history for thousands of years according to its extremely high toxic potential and can therefore be entitled as the king of poison. Arsenic uptake causes acute as well as chronic intoxications, even at very low doses. Acute arsenic poisoning mostly manifests after accidents with pesticides or homicidal intensions, while chronic intoxication is mainly derived from oral ingestion of arsenic-enriched drinking water (MELIKER & NRIAGU 2007). Resulting effects of chronic poisoning are complex and depend on the prevailing chemical form, whereas both in-organic species As(V) and As(III) are much more reactive than methylated organic forms (HOPENHAYN 2006, WHO 2003). Inorganic As is supposed to act genotoxic, carcinogenic and teratogenic (WHO 2003). Due to its similarity to phosphate, As(V) can interact with up to 200 enzymes, most of them being part of the adenosine-tri-phosphate (ATP) synthesis pathway or the DNA synthesis and repair system (ABERNATHY et al. 1999, ISLAM 2008). Reduced inorganic As(III) is considered even more toxic to human organism, which results from its high affinity for reactive thiol groups of enzymes (KNOWLES & BENSON 1983). The principal organ of As meta-bolism is the liver, where inorganic As is methylated to dimethylarsinic acid (DMAA) and monomethylarsonic acid (MMAA), before it is excreted via urine. The half-life of inorganic As compounds in the human body is 2-40 days after resorption, but a continuously uptake results in enduring enrich-ment in liver, kidneys, heart, lungs and ectodermic tissues (POMROY et al.

1980). Chronic exposure to increased concentrations of inorganic As, especially As(III), is known to entail severe diseases and its carcinogenic character promotes an increased appearance of cancer (skin, lung, bladder and liver) in affected populations, which was observed in various case studies in Bangladesh, Taiwan and China (e.g., CHEN et al. 1992, KAPAJ et al. 2006, SMITH et al. 2000). Another characteristic expression is arsenicosis, a collective term for skin lesions like keratosis, hyperkeratosis and pigmentary abnormalities of the extremities (AHMAD et al. 1997).

One of the most famous examples for an endemic occurrence of arsenicosis is an area in the southwest of Taiwan, where local villagers had

changed their drinking water source from surface water to arsenic-enriched artesian groundwater in the 1920’s. This undiscovered exposure soon caused symptoms of chronic As intoxications, which were first described in the 1950’s as “black foot disease” (IPCS 2001). Long-term cohort studies in affected villages of Vietnam, China and Bangladesh imply that chronic As uptake may also trigger foetal loss and infant death, development of diabetes mellitus, cardio-vascular disease and eventually neurotoxic effects and inhibition of children’s mental development (ALAM et al. 2002, ARGOS et al. 2010, FUJINO et al. 2006, LIN et al. 2004, RAHMAN et al. 1998 &

2007, WASSERMANN et al. 2004). A profound overview over health effects related to As can be found in the work of NRIAGU (1994). To the present day, no proper therapy for arsenicosis exists, which is why mitigation strategies are the only available possibility to avoid diseases related to chronic As uptake. It is further very problematic to assess a proper threshold value for drinking water and food, since no dose-response relationship exists according to the carcinogenic character of inorganic As.

The WHO released a provisional guideline value for total As in drinking water of 10 µg L^-1, based on the level that can be achieved through practical treatment methods (WHO 2011). In India, there is actually a legal limit of 50 µg L^-1 for the total As concentration in drinking water in force (Indian Standard Specifications for drinking water IS 10500, reaffirmed 1993).

2.1.5 WORLDWIDE OCCURRENCE AND THE HUMAN HEALTH