TBT and Imposex

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This article describes the use of Tributyltin (TBT) in aquatic antifouling paints, its behaviour in the marine environment and one of its powerful negative effects in non-target species - the phenomenon of imposex in marine gastropods - which ultimately led to the global prohibition of TBT-based antifouling systems.


Introduction

Tributyltin (TBT) was widely used from the 1960s onward as a highly effective biocide in antifouling paints for ship and boat hulls. Its use caused widespread contamination of coastal waters and sediments and severe effects on non-target organisms. The best-known effect is imposex in marine gastropods, in which females develop male sexual characteristics after exposure to very low TBT concentrations.

TBT-containing antifouling systems are since 2008 globally prohibited on ships under the International Maritime Organization's Anti-Fouling Systems Convention. Nevertheless, TBT remains an environmental concern because it persists for long periods in contaminated sediments and hotspots still occur in harbors, shipyards and other areas with a history of intensive antifouling use. Imposex therefore remains a useful biological indicator of residual TBT contamination. Imposex is the development of male sexual characteristics, especially a penis and vas deferens (sperm duct), in female neogastropods. In severe cases the vas deferens can obstruct the female reproductive tract, causing sterility and population decline.

The need for antifouling

The problem of hull fouling

Any submersed rigid structure can work as substrate and be colonized by several marine organisms. It is estimated that there are over 4000 marine fouling species [1]. In the case of vessels, the degree of fouling of the hull depends on the time of submersion, the time the vessel is immobilized or its speed, but mainly on the features of the marine environment. Without an antifouling protection, the fouling can under favorable conditions reach up to 150 kg per square meter, in less than 6 months[2]. This phenomenon leads to an increase in the weight of the vessel and the drag resistance of the hull surface, which directly affect the speed, maneuverability and the fuel consumption (increasing up to 40%), leading to more frequent maintenance operations, higher costs and higher emissions of polluting gases [3]. Additionally, the hulls can work as vectors of translocation of organisms from one place to another, increasing the risks of introducing non-native, invasive species [4].

Fouling on the hull of a small boat

Antifouling methods and TBT

The problem of fouling in vessels was recognized since the beginning of navigation. The ancient Phoenicians and Carthaginians were thought to have used copper sheathing and the Greeks and Romans both used lead sheathing on their ships’ hulls [5]. More recent methods included the use of paints containing organic compounds of lead, arsenic, mercury and halogens (e.g. DDT) and copper oxide [3]. The latter is still widely used. The first antifouling paints using organic compounds of tin started appearing in the second half of the 20th century and quickly dominated the markets during the following decades.

Sources and behavior of TBT in aquatic systems

Historically, antifouling systems were the major direct source of TBT to the marine environment. A TBT-based paint can be composed up to 3% of tin and a large commercial vessel can release more than 200g of TBT to the aquatic environment in only 3 days of permanence in a port [6]. Additionally, dry-docks and boatyards can also be relevant sources of antifouling paints (and other pollutants), where old paint removal and repaint procedures take place. Most of the residues end up in the surrounding environment.

When released into the water, TBT can be degraded into less harmful forms by microorganisms and ultraviolet radiation. However, due to its high affinity to particles it will be easily transported to the sediments, where its concentration is typically higher than in the water. In sediments, organotin compounds are exceptionally stable and the concentration can remain high for a long time even after the sources have ceased [7]. In the water, TBT can remain for a few days or months but in the sediments its half-life can extend for several months, years or even decades [3].

Dry-docks and boatyards: Lack of proper containment during antifouling paint removal can result in deleterious substances being released into the aquatic environment.

Effects of TBT on marine organisms

The case of the Bay of Arcachon (France)

During the period when TBT was being widely used as antifouling, the production of oysters in the Bay of Arcachon (France) almost collapsed. This coastal area is simultaneously a place of production of this shellfish and an area of intense recreational boating [8]. Although the knowledge of TBT was very limited at the time, the French authorities restricted the use of the compound in antifouling paints in the region, in a rare example of precautionary principle[9]. Later studies established that TBT was responsible for reproductive failure and abnormal shell development in the oysters.

Female gastropod Stramonita brasiliensis with imposex. Photo credit Barreira and Castro (2020[10]).

Imposex in marine snails

Also in the beginning of the 70’s certain reproductive abnormalities in other molluscs were discovered, which were later proved to result from exposure to TBT. In certain species of gastropods with separate sexes, the females presented a penis and/or vas deferens. The term 'imposex' was given as 'a superimposition of male features in females' and was first described in dog whelk (Nucella lapillus) [11]. Soon it was clear that this was a generalized phenomenon – not only all the populations of dog whelk analyzed in southwest England were affected but worldwide the same phenomenon was reported and for different species of snails, particularly in areas of intense maritime traffic. By 2007, imposex and intersex (a similiar phenomenon) had been described in over 150 species of marine snails[12]. More developed stages of imposex can lead to the sterilization and premature death of the females, affecting the entire population.

TBT acts as an endocrine disruptor in gastropods. The endocrine systems of molluscs differ substantially from those of vertebrates, and the detailed mechanism of imposex is therefore still not completely resolved. Several mechanisms have been proposed for TBT-induced imposex. Modulation of the retinoid X receptor (RXR) is one likely mechanism, and interaction of TBT with the heterodimeric nuclear receptor pair PPAR/RXR may be another.[13]

TBT can act at extremely low concentrations: a few nanograms per liter is enough to trigger imposex in marine snails[11] - the equivalent of 1 g of salt dissolved in a square pool of 100 m side and 100m depth. These concentrations were close to the detection limits of routine chemical analyses when imposex monitoring was first developed.

Effects in other species

TBT can also cause toxic and endocrine effects in other marine organisms, including molluscs, crustaceans, fish and mammals. The sensitivity and mechanisms of toxicity differ considerably among species, and some effects remain incompletely understood. It is known that top predators from marine ecosystems can accumulate significant amounts of pollutants. TBT is not an exception and has already been detected in cetaceans and seals, sharks and tunas [14]. Some studies suggest harmful effects on the immune and neurological systems and embryos in mammals [15] and described toxicity to plankton, algae, fish and seabirds[8]. However, imposex in sensitive gastropods remains its most distinctive and widely used biological effect indicator.

Imposex as an indicator of TBT contamination

Imposex is one of the strongest historical examples of a biological-effect biomarker directly linked to a specific contaminant. Some species of snails have been used as bioindicators to evaluate and compare the degree of TBT contamination in aquatic environments. They are suitable species since:

  • the severity of imposex correlates with TBT exposure and with TBT concentrations in tissues and the surrounding environment[11]
  • the imposex is triggered by extremely low concentrations – close to the level of detection of measuring instruments when imposex monitoring was first developed
  • marine snails are common in certain habitats and have restricted mobility, so responses reflect local contamination

Imposex monitoring has documented both contamination gradients and subsequent recovery after regulation.

Regulation and recovery

The severe ecological effects of TBT led progressively to national and international restrictions. France prohibited the use of TBT-based antifouling paints on vessels smaller than 25 m in 1982, and several other countries introduced similar restrictions during the 1980s and 1990s. In 2001, the International Maritime Organization adopted the International Convention on the Control of Harmful Anti-fouling Systems on Ships (AFS Convention). The Convention entered into force in 2008 and prohibits the use of organotin compounds acting as biocides in antifouling systems.

Following these restrictions, TBT concentrations and the incidence and severity of imposex declined in many coastal areas, and recovery of affected gastropod populations has been reported. However, TBT persists in contaminated sediments, and local hotspots remain near harbors, shipyards and other areas with a history of intensive antifouling use.[16][17]


Related articles

Antifouling paints
TBT and intersex in periwinkles
Endocrine disruption by marine pollutants
Biomonitoring of pollution impacts in the marine environment
Biomarkers for assessing marine pollution effects
Coastal pollution and impacts
Biomarker
Bioindicator
Endocrine disrupting compounds
Endocrine system
Environmental risk assessment of marine activities
Coastal Wiki articles related to ecotoxicology


References

  1. ↑ Almeida, E., Diamantino, T. and Sousa, O. 2007. Marine Paints: the particular case of antifouling paints. Progress in Organic Coatings, 59: 2-20
  2. ↑ Bray, S. 2006. Tributyltin pollution on a global scale. An overview of relevant and recent research: impacts and issues. Langston, W.J. (Ed.)
  3. ↑ 3.0 3.1 3.2 Omae, I. 2003. Organotin antifouling paints and their alternatives. Applied Organometallic Chemistry 17: 81-105
  4. ↑ Champ, M. 2000. A review of organotin regulatory strategies, pending actions, related costs and benefits. The Science of the Total Environment 258: 21-71
  5. ↑ Callow, M.E. and Callow, J.A. 2002. Marine biofouling: a sticky problem. The Biologist 49: 10-14
  6. ↑ Batley, G. 1996. The distribution and fate of tributyltin in the marine environment. In Tributyltin: case study of an environmental contaminant. de Mora, S. (ed). Cambridge University Press: London, U.K. p. 139-166
  7. ↑ Langston, W.J., Bryan, G.W., Burt, G.R. and Gibbs, P.E. 1990. Assessing the impact of tin and TBT in estuaries and coastal regions. Functional Ecology 4: 433-443
  8. ↑ 8.0 8.1 Terlizzi, A., Fraschetti, S., Gianguzza, P., Faimali, M. and Boero F. 2001. Environmental impact of antifouling technologies: state of the art and perspectives. Aquatic Conservation: Marine and Freshwater Ecosystems 11: 311-317
  9. ↑ Ruiz, J.M., Bachelet, G., Caumette, P. and Donard, O.F. 1996. Three decades of tributyltin in the coastal environment with emphasis on Arcachon Bay, France. Environmental Pollution 93: 195-203
  10. ↑ Barreira, C.A.R. and Castro, I.B. 2020. Compostos triorganoestanicos. In: Contaminantes orgânicos em ambientes aquáticos, Imprensa Universitária da Universidade Federal do Ceará
  11. ↑ 11.0 11.1 11.2 Gibbs, P.E. and Bryan, G.W. 1994. Biomonitoring of tributyltin (TBT) pollution using the imposex response of neogastropods molluscs. In: Biomonitoring of Coastal Waters and Estuaries. Kramer, K.J. (Ed.) 1994. CRC Press Inc. Boca Raton, p: 205-226
  12. ↑ Sousa, A., Matsudaira, C., Takahashi, S., Tanabe, S. and Barroso, C. 2007. Integrative assessment of organotin contamination in a southern European estuarine system (Ria de Aveiro, NW Portugal): Tracking temporal trends in order to evaluate the effectiveness of the EU ban. Marine Pollution Bulletin 54: 1645-1653
  13. ↑ Metcalfe, C.D., Bayen, S., Desrosiers, M., Munoz, Z., Sauvé, S. and Yargeau, V. 2022. An introduction to the sources, fate, occurrence and effects of endocrine disrupting chemicals released into the environment. Environmental Research 207, 112658
  14. ↑ Tanabe, S., Prudente, M., Mizuno, T., Hasegawa, J., Iwata, H. and Miyazaki, N. 1998. Butyltin contamination in marine mammals from North Pacific and Asian coastal waters. Environmental Science & Technology 32: 193-198
  15. ↑ Berge, J., Brevik, E., Bjorge, A., Folsvik, N., Gabrielsen, G. and Wolkers, H. 2004. Organotins in marine mammals and seabirds from Norwegian territory. Journal of Environmental Monitoring 6: 108-112
  16. ↑ Beyer, J., Song, Y., Tollefsen, K.E., Berge, J.A., Tveiten, L., Helland, A., Øxnevad, S. and Schøyen, M. 2022. The ecotoxicology of marine tributyltin (TBT) hotspots: A review. Marine Environmental Research 179: 105689.
  17. ↑ Unger, M.A., Harvey, E., Southworth, M., Mann, R., Marquardt, A. and Otto, N. 2025. Decreases in TBT concentrations in southern Chesapeake Bay: Evidence for long-term TBT degradation and reduced imposex in the invasive gastropod, Rapana venosa. Marine Pollution Bulletin 212: 117524.


The main author of this article is Veiga, Joana M
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Citation: Veiga, Joana M (2026): TBT and Imposex. Available from http://www.coastalwiki.org/wiki/TBT_and_Imposex [accessed on 7-09-2026]
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