Tributyltin (TBT) in freshwater and marine water
Toxicant default guideline values for protecting aquatic ecosystems
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).
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).
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.
ANZECC 1992. Water quality guidelines for fresh and marine waters. Australian and New Zealand Environment and Conservation Council, Australian Government Publishing Service, Canberra.
ANZECC & ARMCANZ 2000. Australian and New Zealand Guidelines for Fresh and Marine Water Quality, Australian and New Zealand Environment and Conservation Council and Agriculture and Resource Management Council of Australia and New Zealand, Canberra.
Batley GE, Fuhua C, Brockbank KL & Flegg KJ 1989. Accumulation of tributyltin by the Sydney rock oyster, Saccostrea commercialis. Australian Journal of Marine and Freshwater Research 40, 49-54.
Brooke LT, Call DJ, Poirier SH, Markee TP, Lindberg CA, McCauley DJ & Simonson PG 1986. Acute toxicity and chronic effects of bis(tributyltin)oxide to several species of freshwater organisms. Centre for Lake Superior Environmental Studies, University of Wisconsin-Superior, Wisconsin.
Bryan GW, Gibbs PE, Hummerstone LG & Burt GR 1987. The effects of tributyltin (TBT) accumulation on adult dog-whelks. Journal of the Marine Biological Association UK 67, 525-544.
Bryan GW, Gibbs PE, Hummerstone LG & Burt GR 1986. The decline of the gastropod Nucella lapillus around southwest England: Evidence of the effect of tributyltin from antifouling paints. Journal of the Marine Biological Association UK 66, 611-640.
CCME 1991. Canadian water quality guidelines, updates. Canadian Council of Ministers for the Environment, Ottawa.
CCREM 1987. Canadian water quality guidelines. Canadian Council of Resource and Environment Ministers, Ontario.
Ellis D & Pattisina LA 1990. Worldwide neogastropod imposex: A biological indication of global TBT contamination? Marine Pollution Bulletin 21, 248-253.
Evans SM, Hutton A, Kendall MA & Samosir AM 1991. Recovery in populations of dogwhelks Nucella lapillus (L) suffering from imposex. Marine Pollution Bulletin 22, 331-333.
Foale S 1993. An evaluation of the potential of gastropod imposex as a bioindicator of tributyltin pollution in Port Phillip Bay, Victoria. Marine Pollution Bulletin 26, 546–552.
Gibbs PE, Pascoe PL & Burt GR 1988. Sex change in the female dog-whelk, Nucella lapillus, induced by tributyltin from antifouling paints. Journal of the Marine Biological Association UK 68, 715-731.
Jarvinen A W & Ankley G T 1999. Linkage of effects to tissue residues: Development of a comprehensive database for aquatic organisms exposed to inorganic and organic chemicals. SETAC Technical Publication Series, SETAC Press, Pensacola FL.
Kannan K, Tanabe S, Iwata H & Tarsukawa R 1995. Butyltins in muscle and liver of fish collected from certain Asian and Oceanic countries. Environmental Pollution 58, 279–290.
King N, Miller M & de Mora S 1989. Tributyltin levels for seawater, sediment and selected marine species in coastal Northland and Auckland, New Zealand. New Zealand Journal of Marine and Freshwater Research 23, 287–294.
Nias DJ, McKillup SC & Edyvane KS 1993. Imposex in Lepsiella vinosa from southern Australia. Marine Pollution Bulletin 26, 380–384.
NSW EPA 2000. Analytical Chemistry Section, Table of Trigger Values 20 March 2000, LD33/11, Lidcombe, NSW.
OECD 1992a. Report of the OECD workshop on extrapolation of laboratory aquatic toxicity data to the real environment. OECD Environment Monographs No 59, Organisation for Economic Co-operation and Development, Paris.
Smith BS 1981. Male characteristics on female mud snails caused by antifouling bottom paints. Journal of Applied Toxicology 1, 22-25.
Smith PJ & McVeagh M 1991. Stewart C, de Mora SJ, Jones MRL & Miller MC 1992. Imposex in New Zealand neogastropods. Marine Pollution Bulletin 24, 104-209.
Stewart C, de Mora SJ, Jones MRL & Miller MC 1992. Imposex in New Zealand neogastropods. Marine Pollution Bulletin 24, 104-209.
Waite ME, Waldock MJ, Thain JE, Smith DJ & Milton SM 1991. Reductions in TBT concentrations in UK estuaries following legislation in 1986 and 1987. Marine Environmental Research 32, 89-111.
Wilson SP 1994. Tributyltin induced imposex in two species of intertidal gastropods. PhD Thesis, University of Technology Sydney, 349pp.
Wilson SP, Ahsanullah M & Thompson GB 1993. Imposex in neogastropods: An indicator of tributyltin contamination in eastern Australia. Marine Pollution Bulletin 26, 44-48.
Wong PTS, Chau YK, Kramar O & Bengert GA 1982. Structure-toxicity relationship of tin compounds on algae. Canadian Journal of Fisheries and Aquatic Science 39, 483-488.
Zabel TF, Seager J & Oakley SD 1988. Proposed environmental quality standards for list II substances in water: Organotins. Water Research Centre, Medmenham, UK. ESSL TR 255.