Tellurium

Endpoint summary

Description of key information

Aquatic biota have always been exposed to varying levels of essential and non-essential metals. Over time, they learned to cope with varying exposure levels by developing strategies that allow the regulation of internal concentration levels (e.g., active uptake/elimination, storage, sequestering and detoxification) (George et al, 1980; mason and Nott, 1981; Rainbow et al, 1980; Simkiss, 1981; White and Rainbow, 1982; Rainbow, 1988; Viarengo, 1989; Depledge and Rainbow, 1990). It can thus be concluded that the BCF model concept is not valid for metals for several reasons:

• Fundamental physicochemical differences between inorganic substances (metals), and organic substances for which the BCF-model was developed
• Natural occurrence of metals in the environment, leading to active regulation and homeostatic control
• Physiological uptake mechanisms that exhibit saturable kinetics
• Active elimination strategies
• Storage of detoxified metal forms by aquatic organisms (e.g., binding to metallothionein-like proteins (Noel-Lambot et al, 1980; Roesjadi, 1980; Langston and Zhou, 1986; Hylland et al, 1994). The use of granules as a storage mechanism can lead to extremely high tissue concentrations (and BCF), but still unrelated to adverse impacts.

It is therefore not possible to determine a relevant bioaccumulation factor for essential and non-essential metals, including tellurium. With regard to tellurium it should be noted that - despite the fact that this element probably has no purpose in living organisms – it belongs together with Po and Se to a chemically relatively homogenous group. All of these elements can follow the same metabolic pathways, such as biomethylation (Hussain et al., 1995; Kim et al., 2000; Chasteen and Bentley, 2003). Therefore the physiological mechanisms that invalidate the applicability of the classic BCF-model for essential metals will also be relevant for tellurium.

Additional information

Bioaccumulation is an intrinsic property of a substance, and when used in the context of a hazard/effects assessment, it should remain independent of exposure concentrations (i.e., a constant value within the range of evaluated exposure conditions) (McGeer et al, 2003). A BCF/BAF-value, for instance, is a critical part of the PBT-assessment in order to understand a potential environmental hazard. For metals, adverse effects are only expected where the uptake exceeds the elimination in excess of a threshold concentrations at the specific site of action. 

The typical BCF/BAF model which describes the relationship between bioaccumulation and potential adverse effects, is originally developed/validated for a fairly limited number of neutral, lipophilic, synthetic organic substances with narcosis as the mode of toxic action. 

The applicability of such a model for natural occurring substances such as metals (essential and non-essential) has been questioned (Franke et al, 1994; Chapman et al, 1996; Barron et al, 1990; Chapman, 1996; Franke, 1996; Adams et al, 2000; Chapman et al, 1999) as it does not take into account complex internal processes such as active uptake/elimination, internal sequestration and storage.

 

Using bioaccumulation in the context of hazard identification is only possible when the BCF is independent from the exposure concentration. This type of intrinsic property typically translates into a linear relationship between external exposure concentration and internal concentration (constant over a wide range of exposure conditions). For neutral organic substances the main uptake process is simple passive diffusion across the lipid bilayer of a biological membrane (McKim, 1994). In essence, the BCF model is a hydrophobicity model (Barron, 1992), which is supported by some findings on an inverse relationship between BCF and water solubility (Kenaga, 1980; Banerjee et al, 1980; Chiou et al, 1977; Clayton et al, 1977). It has been found that some of the properties – like lipophilicity and K_ow- that correlate with bioaccumulation for organic substances, are irrelevant for metals and are unrelated to accumulation (OECD, 2001; Newman, 1995; McKim, 1994; Newman and Jagoe, 1994; Langston and Bryan, 1984). Trace metals, both essential and non-essential, are always present in the environment and can be found in all aquatic biota (Cowgill, 1976; Shearer, 1984). Uptake of ionic substances such as dissolved metal ions (incl. tellurium) occurs via physiological mechanisms that exhibit saturation kinetics as it depends on the availability of binding receptor sites and competing processes at these sites. Therefore a constant uptake rate regardless of the exposure concentration is not considered relevant for metals. For example, at background-levels the calculated BCFs can be as high as 300000, but such values are meaningless within the context of hazard/effects assessment.

 

Virtually no measured BCF-values have been reported in open literature for tellurium. Waska et al (2008) analyzed Te in different tissues of the squid Todarodes pacificus, with concentrations of 0.9 ± 0.6 ng/g in muscle, 2.4 ± 4.1 ng/g in stomach, 0.9 ± 0.1 ng/g in gills and 3.4 ± 2.5 ng/g in hepatopancreas. These results indicate a moderate Te bioaccumulation, with values in the range of 5800-22000. It should be noted, however, that these concentration factors are based on previously published seawater reference values. The data did not allow to assess a potential concentration-dependency of the calculated BCF.

 

As stated, there are a number of physicological processes that invalidate the applicability of the classic BCF-model for metals. Literature data have been identified which show that several of these processes are relevant for tellurium.

Boriova et al (2014), for instance, demonstrated that tellurium can undergo methylation by some organisms, resulting in a biovolatilization of this element. The rate and amount of biovolatilized Te in the filamentous fungus Scopulariopsis brevicaulis also appeared to be concentration-dependent, which suggest that this organism actively regulates its internal concentration levels. This finding confirmed the conclusions of earlier studies by e.g. Gharieb et al (1999) who reported Te-methylation and volatilization by the fungi Fusarium sp.

Gharieb et al (1999) also reported a detoxification mechanism that was based on the reduction of tellurite to elemental tellurium. This process was demonstrated by electron microscopy which revealed the deposition of black granules outside the hyphae and intracellularly, apparently occupying some of the vacuoles of the mycelium. Energy-dispersive X- ray microanalysis of the granules confirmed that these particles were composed of elemental tellurium. Reduction to elemental tellurium was also found in Escherichia coli where it was deposited in or around cells (Taylor et al., 1988; Lloyd-Jones et al., 1994), and was also seen in the yeast Schizosaccharomyces pombe (Smith, 1974). Finally, Ramadan et al (1989) showed that Te was readily incorporated into amino-acids by fungi in place of Se, thus showing that the mechanisms which regulate Se-levels in organisms could also be applicable on Te.

 

Transfer of Te from one trophic level to another was investigated by Nolan (1991) who exposed marine algae (the centric diatom Thalassiosira pseudonana and the green alga Dunaliella tertiolecta) as well as two crustacean species (the brine shrimp Artemia salina and the Monaco shrimp Lysmata seticaudata) to Te radiotracer.

Nolan et al (1991) found strong indications that Te was accumulated and metabolized by the algae via rapid adsorption to the cell surface, followed by a gradual translocation across the cell membrane to the interior. The rapid loss of the radiotracer from the Artemia coincided with the voiding of the gut contents. This finding indicates that no true assimilation occurred and that any adsorption of radiotracer to the crustacean exoskeleton was weak and completely reversible upon transfer to clean water. The lack of assimilation of the radiotracer by both Artemia and Lysmata is indicative of an extremely low transfer factor through different trophic levels and is in accordance with the non-nutrient type profile of tellurium in ocean waters (Lee & Edmond, 1985). Thus filter-feeders and zooplankton grazers may well be responsible, by virtue of the quantities of particulate material processed, for a rapid removal of tellurium from surface waters via the production of faecal pellets.