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Antimony metal and antimony containing compounds will dissolve and generate antimony ions. The environmental fate section will therefore discuss the fate of antimony in general. Antimony, being an elemental substance, cannot be degraded, but may be transformed between different phases, chemical species, and oxidation states.

The speciation and physico-chemical state of antimony are important for its behaviour in the environment and availability to biota. For example, antimony incorporated in mineral lattices is inert and unlikely to be bioavailable. Most analytical methods for antimony do not distinguish between the various forms of antimony. While the total amount of antimony may be known, the nature of the antimony compounds, the importance of adsorption, etc. are not. This information, which is critical in determining the availability of antimony, is likely to be site-specific.

There are uncertainties surrounding the thermodynamic data for Sb compounds and, as a consequence, the Eh-pH diagrams differ between different sources. Earlier diagrams suggest that antimony is immobile under oxidizing conditions, occurring as solid oxides, but more recent diagrams show that in oxidizing conditions, Sb(OH)₆^-is the most important species, confirming the relatively high mobility of Sb under oxidizing conditions.

The conclusions that are drawn in the EU RAR for antimony trioxide on the fate of antimony in water are:

 i) In natural waters antimony exists almost exclusively in the dissolved phase in the two valency states + 3 and + 5. Both Sb(III) and Sb(V) ions hydrolyse easily, and Sb(III) is present as the neutral species Sb(OH)₃, and Sb(V) as the anion, Sb(OH)₆^-.

ii)  According to thermodynamic calculations, antimony should almost exclusively be present as Sb(V) in oxic systems and as Sb(III) in anoxic systems, when they reach equilibrium. Even though the dominant species in oxic waters is Sb(V), Sb(III) has been detected in concentrations much above what is predicted, and the reverse is true for Sb(V) in anoxic systems. It is clear from speciation measurements of well oxygenated antimony solutions that antimony oxidation processes in freshwaters are slow, with half-lives usually being in the order of months in laboratory solutions. Whilst there is evidence to suggest that factors such as UV light, dissolved organic carbon (DOC), and iron and manganese oxyhydroxide precipitates can all increase the rate of oxidation of Sb(III) to Sb(V), thermodynamic equilibrium is unlikely to be approached in natural waters. This is confirmed by the available monitoring data which include speciation measurements of Sb in the field. Equilibrium predictions of the chemical speciation of antimony in natural waters must therefore be treated with caution.

 iii)  Reports exist on both conservative behaviour (i. e. the concentration only changes with dilution or evaporation), and a behaviour corresponding to a mildly scavenged element with surface (atmospheric) input.

 iv) In addition to the inorganic forms of antimony, there also exist methylated forms of trivalent and pentavalent antimony.

 v)  Interactions between the antimony species (anionic Sb(OH)₆^-or the neutral Sb(OH)₃) present in natural waters and the predominantly negatively charged natural organic matter may occur, but firm conclusions cannot currently be drawn on its overall importance in Sb speciation.

 vi) Transformation/dissolution studies indicate that diantimony trioxide solubility is dependent on the pH conditions, being more soluble at pH 8 than at pH 6. The maximum observed dissolution was 0.18 mg l^-1from a 1 mg l^-1loading after 28 days. Redox and salinity conditions may also affect the rate of dissolution. Antimony metal and NaSb(OH)₆showed greater solubility in transformation/dissolution tests, although 1 mg l^-1loadings resulted in concentrations of up to only 0.6 mg l^-1after 28 days.

As no information is available about the relative toxicities of the different forms of antimony, and speciation analysis of environmental samples is also extremely uncommon, the PEC is expressed in terms of the dissolved antimony concentration.

The conclusions that are drawn in the EU RAR for antimony trioxide on the fate of antimony in sediment are:

 i) The adsorption of antimony in oxic sediments has been correlated with the presence of iron-, manganese-, and aluminium oxides.

 ii)  The decrease in bioavailable antimony in water by oxic sediments is not a permanent decrease, as the adsorption on the hydrous oxides is dependent on both pH and oxic condition (which may change). This is largely due to the redox cycling of iron (oxy) hydroxide precipitates with which antimony may be associated. In addition, antimony may become bioavailable to organisms inhabiting the sediment through ingestion of the sediment.

 iii) In anoxic systems, and in the presence of sulphur, antimony forms soluble or insoluble stibnite, SbS₂^-and Sb₂S₃(s), respectively, depending on pH. This may result in a larger decrease in bioavailable antimony, when compared to the oxic part of the sediment.

The conclusions that are drawn in the EU RAR for antimony trioxide on the fate of antimony in soil are:

i) The sorption and precipitation of Ca[Sb(OH)₆]₂seems to be a more important process overall in the fate of antimony in the environment than the dissolution processes of Sb₂O₃.

 ii) The solubility of antimony compounds depends on the soil conditions (Eh/pH) and the time given to dissolve. The results of transformation/dissolution tests indicate that some antimony substances may be more soluble at higher pH (e. g. pH 8.5) than at low pH (e. g. pH 6).

 iii) The most important soil characteristic to influence the mobility of antimony in soil (and sediments), appears to be the presence of hydrous oxides of iron, manganese, and aluminium, to which antimony may adsorb. In addition, these hydrous oxides seem to oxidise dissolved trivalent antimonite (Sb(OH)₃) to the pentavalent antimonate (Sb(OH)₆^-).

 iv) The largest effect of pH on sorption seems to be around pH 3 - 4, with decreasing sorption at higher pH values. The effect of pH as such is probably less important when compared to the effect of the hydrous oxides. The effect of pH on antimony mobility seems to be via the influence of hydrous oxides on the valence of antimony and the solubility of the antimony compound, and also via the increasing negative charge of the soil at increasing pH (and hence, weaker sorption of the negatively charged Sb(OH)₆^-).

 v) Due to the anionic character of the dissolved species (Sb(OH)₆^-), antimony is expected to have a low affinity for organic carbon. However, there are results that indicate that the sorption of Sb(V) by humic acid in acid soils with high proportions of organic matter may be more important than previously suspected, although the strong Sb(V) scavenging potential of Fe(OH)₃probably results in a diminished role of organic matter binding in soils with high amounts of amorphous hydroxides.

 vi) The cationic exchange reactions, which are the main sorption reactions on clay minerals, are not expected to be important for the anionic antimony.

 vii)  Initial differences in sorption depending on the type of antimony compound diminish with time.

 viii) The influence of the concentration of added antimony on sorption appears to be small.

 ix)  A higher Sb porewater concentration can be achieved in transformation studies when using Sb₂O₃, when compared to SbCl₃. The limiting factor appears to be precipitation of Ca[Sb(OH)₆]₂.

The conclusions that are drawn in the EU RAR for antimony trioxide on the fate of antimony in air are:

 i) Anthropogenic activities may result in long-range transport of antimony far from its source.

 ii) Combustion/incineration processes transform antimony compounds to diantimony trioxide regardless of the pre-incinerated form of antimony.

 iii) There are indications that diantimony trioxide may dissolve in the atmosphere and that the trivalent form will oxidise to the pentavalent form.