Tributyltin (TBT) in freshwater and marine water

​Toxicant default guideline values for protecting aquatic ecosystems

October 2000

Extracted from Section 8.3.7 ‘Detailed descriptions of chemicals’ of the ANZECC & ARMCANZ (2000) guidelines.

The default guideline values (previously known as ‘trigger values’) and associated information in this technical brief should be used in accordance with the detailed guidance provided in the Australian and New Zealand Guidelines for Fresh and Marine Water Quality.

Description of chemical

Although inorganic tin is generally considered non-toxic, the attachment of alkyl or aryl groups to the tin atom greatly increases the toxicity (CCME 1991 Appendix X). Tributyltin (TBT) is the most common of a group of organotin compounds, which have widespread usage in marine antifouling points and for wood preservation. The general structure of tri-organotins is R4Sn, and they are generally associated with an anion such as acetate, chloride and fluoride. Tributyltin oxide TBTO (CAS 56-35-9) has the structure Bu3Sn-O-SnBu3, formula C24H54OSn2 and molecular weight 596.1. Their solubility in freshwater varies from 6 to 256 mg/L (Zabel et al. 1988). The solubilities for TBTO were 19.4 mg/L in distilled water and 1.4 mg/L in salt water. TBTO is usually applied as a slow-release coating to boats. The current analytical practical quantitation limit (PQL) for TBT is 0.001 µg/L as [Sn] in both fresh and marine water (NSW EPA 2000).

Due to its high toxicity to marine bivalves, the use of TBT has been restricted in Australia and New Zealand, usually to use on larger boats > 25 m in length (Wilson et al. 1993). The major issues of concern were deformities in oysters (Batley et al. 1989) and induction of imposex in gastropods (Smith & McVeagh 1991, Stewart et al. 1992, Wilson et al. 1993). The phenomenon of imposex, the formation of male sex characteristics in female gastropods, is a worldwide issue (Ellis & Pattisina 1990). It was linked to organotins in the 1980s (Smith 1981) and was identified as being responsible for a marked decline in populations of Nucella lapillus in southwest England (Bryan et al. 1986). There has been evidence of recovery in TBT levels and degrees of imposex since restrictions have been placed on TBT internationally (Evans et al. 1991, Waite et al. 1991, Wilson et al. 1993). ANZECC resolved in 1990 that antifoulants should not release more than 5 µg TBT/cm2/day; however, this rate may be reduced (ANZECC 1992).

Environmental fate

Tributyltin is strongly bound to sediments (CCME 1991) but subsequent remobilisation by biota is possible (CCME 1991). A sediment-water partition coefficient of 3288 has been reported (Zabel et al. 1988). Biodegradation occurs by sequential dealkylation to di- and mono-butyltins and eventually elemental tin. These degradation products have much lower toxicity. Photolylic half-lives for the butyltins of 18 days and > 89 days have been reported, depending on conditions (CCREM 1987). Possible concentration in the surface microlayer may aid photolytic breakdown. Half-lives may be several months at lower temperatures (CCME 1991).

TBT is bioaccumulated and has been found to accumulate in tissues of molluscs (Zabel et al. 1988). Gibbs et al. (1988) reported a maximum bioaccumulation factor of 250,000 in the snail Nucella lapillus after 54 days exposure at concentrations of 1 to 2 ng/L tributyltin (as Sn).

Aquatic toxicology

CCME (1991 Appendix X) summarised the toxicity data for tributyltin for freshwater animals. The acute 96-hour LC50 to freshwater fish ranged from 2.6 µg/L to 13 µg/L. In a study of the chronic toxicity of TBT, Brooke et al. (1986) found that Daphnia magna exposed to 0.2 µg/L of TBT for 21 days showed a significant reduction in the number of surviving young. They estimated that the lowest observed effect concentration (LOEC) for TBT for growth of fathead minnows Pimephales promelas was 0.08 µg/L.

Screened toxicity figures for TBT below are given as µg/L of tin (Sn). This has limited the data that could be considered, as some papers did not report the type of measurement.

Freshwater fish: 48 to 96-hour LC50, two species, 4.8 to 8.4 µg/L.

Freshwater crustaceans: D. magna 48 to 96-hour EC50, 2.2 to 6.6 µg/L.

Freshwater algae: 96-hour LC50, photosynthesis, 10 µg/L.

Zabel et al. (1988) and CCME (1991 Appendix X) have reviewed the effects of TBT on marine organisms. Significant reductions in growth of larval inland silverside Menidia beryllina were noted at 0.93 µg/L. There is an extensive dataset for marine invertebrates, with 96-hour LC50 values ranging from 0.42 µg/L for Acanthomysis sculpta to 19.5 µg/L for Palaeomonetes pugio. Chronic LOECs as low as 0.023 µg/L have been reported (CCME 1991). The dog-whelk Nucella lapillus exhibited a high percentage of imposex at 0.019 µg/L (Bryan et al. 1986).

Data used for trigger value calculations are as follows:

Marine fish: three species, 48 to 96-hour LC50, 2.1 to 400 µg/L. Chronic NOEC (30-day mortality), 0.6 µg/L.

Marine crustaceans: five species, 48 to 96-hour LC50, 0.13 to 63 µg/L. The amphipod Rhepoxynius abronius, was least sensitive and the calanoid copepod Acartia tonsa was most sensitive. Most figures were 2.6 µg/L. NOEC (6-day mortality) of 0.004 µg/L for A. tonsa gave an acute-to-chronic ratio (ACR) of 100 (the LOEC was 0.01 µg/L).

Marine molluscs: 48 to 96-hour LC50, 2 species 84 to 717 µg/L. A 48-hour EC50 (devel.) of 0.5 µg/L for Isognomon californicum did not satisfy screening requirements. Chronic NOEC on three species: Crassostrea virginica (66-day growth), 0.13 µg/L; Mytilus edulis (33 to 66-day growth), 0.002 to 0.8 µg/L (48 to 96-hour LC50 0.92 to 120 µg/L); Scrobicularia plana (23 to 30-day mortality, 0.05 to 1 µg/L).

Algae: one species, red algae Porphyria yezoensis, 96-hour EC50 (population growth), 27 to 33 µg/L and two species diatom 92 to 96-hour EC50 (growth). 0.13 to 376 µg/L (geometric means of 0.44 and 1.36 µg/L).

Field and mesocosm studies

Much of the difficulty of interpreting the results of field studies with TBT arises from the fact that chemical measurements in the water column do not often give a good indication of exposure, given its propensity for adsorption to sediments or collecting in the surface microlayer. None of these studies satisfied OECD (1992a) requirements but they did provide corroborative information for levels derived from laboratory tests.

Wong et al. (1982) studied the effect of trialkyltins on natural phytoplankton communities in freshwater lakes and TBTO was found to be toxic at concentrations as low as 0.1 µg/L.

Laboratory tests have shown that very low concentrations of TBT can cause reduced growth and thickening of shells (Zabel et al. 1988). The placing of two small boats treated with TBT in a pristine estuary in New South Wales caused significant deformities in shells of oysters (Saccostrea commercialis), which were correlated with tissue TBT levels up to 27 µg/L Sn/kg in oysters over 300 m from the boats.

Numerous field studies have demonstrated the occurrence of imposex in field populations of marine intertidal gastropods (e.g. Bryan et al. 1987), including those in New Zealand and Australia (King et al. 1989, Smith & McVeagh 1991, Stewart et al. 1992, Nias et al. 1993, Foale 1993, Wilson et al. 1993, Kannan et al. 1995). A high percentage of imposex in the European dog-whelk Nucella lapillus, was associated with TBT levels of 0.019 µg/L (Bryan et al. 1986, 1987). Gibbs et al. (1988) suggested that 3 to 5 ng Sn/L has been linked with development of imposex. Nias et al. (1993) demonstrated that TBT concentrations of 0.5 µg/L caused 65% incidence of imposex in Lepsiella vinosa from southern Australia, and 0.1 µg/L caused 41% incidence, compared to 7% in controls. Some effects were even induced at 0.01 µg/L. Wilson (1994) demonstrated that concentrations of TBT around 0.5 µg/L induced significant levels of imposex in Thais orbita within one month and in Morula marginalba within 2 to 4 months.

Jarvinen and Ankley (1999) report data on tissue residues and effects for TBT for around four freshwater species and nine marine species. It is not possible to summarise the data here but readers are referred to that publication for more information. TBT concentrations of 3 to 5 ng/L, corresponding to tissue residues of 0.18 mg/kg, caused a reduction in reproduction for Nucella lapillus. No effect was noted at 1 to 2 ng/L (Gibbs et al. 1988). Effect levels for other species were generally higher.


A freshwater low reliability trigger value of 0.002 µg/L​ (2 ng/L) expressed as [Sn] was derived using an assessment factor (AF) of 1000. This should only be used as an interim indicative working level. A marine high reliability trigger value of 0.006 µg/L (6 ng/L) expressed as [Sn] was derived using the statistical distribution method with 95% protection. This is considered sufficiently protective in slightly to moderately disturbed ecosystems.


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