LL1-13 hindsight check: Tributyltin (TBT) antifoulants (Santillo, Johnston and Langston), Late lessons from early warnings (EEA, 2001), Ch. 13, pp. 135–148#

Checked 25 September 2026. The check covers what happened between 2001 and September 2026. Page numbers refer to the 2001 report.

Overview#

This chapter was written just before the global treaty it describes. That makes it unusually easy to test: most of its forecasts (the treaty, the end of shipping inputs, slow recovery, sediment legacy, the problem of substitutes) had clear outcomes within 5 to 20 years. On the whole the chapter has aged well. Its causal story and its forecasts about shipping have been borne out more clearly than it could show in 2001. The weak points are the mechanism it names for imposex, the evidence behind the Japan example, and the sweep of its closing verdict that no action on TBT was precautionary.

Vindicated or strengthened - The global regime came, and it worked where it applied. - The IMO convention was adopted on 5 October 2001, as the chapter expected, and entered into force on 17 September 2008. - By December 2025 it had 100 parties, covering 95.77% of world merchant tonnage. - In the OSPAR area (north-east Atlantic), the share of monitoring sites with imposex above the environmental assessment criterion (EAC, the level set to protect sensitive species) fell from 81% (2008) to about 21% (2020). Imposex declined significantly in all nine subregions with enough data. - In Norway, no imposex was found at any monitoring station in 2017. - Shipping was the main remaining source (pp. 139–140). Along Norway’s coast, TBT and imposex near shipping lanes were significantly higher before the 2008 all-vessel ban than after it (Schøyen et al., 2019). The same pattern appears in English estuaries used by commercial shipping (Langston et al., 2015) and in southern North Sea sediments (OSPAR, 2022). This is strong support for the chapter’s central causal claim about seagoing vessels. - “Partial recovery, not solved” (pp. 138–139). Dogwhelks recolonised shores where they had died out, but hotspots near ports, dockyards and sediment sinks have persisted into the 2020s. That matches the chapter’s reading. - Persistence and legacy (p. 141). The chapter’s “several years” was, if anything, an understatement: - sediment half-times reach 33 years in Southampton Water (Langston et al., 2015); - southern North Sea sediments still averaged about 3.6 times the OSPAR quality standard in 2022 (OSPAR, 2022); - in 2018 TBT still caused more than 650 EU water bodies to fail good chemical status (EEA, 2018); - the treaty did not settle who pays for dredging contaminated harbour sediments (Champ, 2003). - Regrettable substitution (p. 141). - Irgarol 1051 (cybutryne), the chapter’s example, was restricted nationally (Denmark 2000; UK 2001), not approved by the EU (2016), and banned by the IMO from antifouling systems applied to ships from 2023. The 2013 volume devoted a chapter to it: “Booster biocide antifoulants: is history repeating itself?” - Copper from antifouling paint has since risen in Arcachon Bay, the very place where the TBT story started (Araújo, 2025). - Continuing organotin inputs from other uses (p. 142). The EU restricted organotins in consumer articles from 2010 to 2015. OSPAR’s 2020 target of ceasing releases was not met. A 2026 review reports global organotin production now far above late-1990s levels (Pang et al., 2026).

Wrong, superseded or overstated - Mechanism (pp. 135, 141). The chapter attributes imposex to interference with “steroid hormone metabolism”, the leading hypothesis in the 1990s. Since 2004 the leading explanation has been that organotins activate the retinoid X receptor (RXR), a nuclear receptor (Nishikawa et al., 2004; Horiguchi, 2023). The vertebrate-steroid picture of mollusc endocrinology has since been questioned more generally (Panagiotidis et al., 2026). The chapter itself flagged that mechanistic knowledge was “likely to continue to develop” (p. 141). - Japan (p. 142). The example is weaker than the chapter makes it. - Japan’s controls were not a “national ban on all applications”. Japan’s environment ministry describes a 1990 prohibition of one compound (TBTO), restriction of the other TBT compounds, and voluntary industry controls (Japan MOE, 2010). - The ministry does attribute continuing loading chiefly to foreign vessels, which supports the chapter. - But TBT was still at 1.0–2.8 ng/l in Seto Inland Sea surface waters in the 2020s, 15 years after the global ban (Harino et al., 2023). And imposex did not decline in Hong Kong between 2004 and 2015 (Ho et al., 2016). Global restrictions were necessary but not sufficient. - “None of the actions … precautionary” (p. 142). This is true for the UK’s 1987 action, and substantially true for the global treaty. But the chapter’s own account shows France acting “on the best information available” (p. 136). The IMO extension to large ships was agreed while the chapter itself described the far-field evidence as under “intense debate” (p. 139). Later evidence vindicated that decision (Claim 3). - Marine mammals and human health (p. 140). The immune-effect hypothesis for wild marine mammals was never established as a population-level cause. In Danish harbour porpoises, organotins declined and were not among the contaminants above effect thresholds; PCBs and DDT were (Dietz et al., 2026). In Europe, human intakes from seafood were later estimated to be below the tolerable daily intake (TDI) that EFSA (the EU food-safety agency) set in 2004. That TDI was itself based on immunotoxicity (SCHER, 2006; Guérin et al., 2007).

What the chapter could not see - Harm moved to new endpoints. At nanomolar concentrations TBT activates RXR and PPARγ, another nuclear receptor, in vertebrates. It was characterised as an “obesogen” (a chemical that promotes fat-cell formation) in 2005–06 (Grün et al., 2006). SCHER (2006), the EU scientific committee on health and environmental risks, judged that this required no change to the TDI. - Phenyltins took over in some regions. In South China, phenyltin levels in dolphins kept rising over 2003–2021 while butyltins fell (Zhang et al., 2022). Phenyltins are related organotins that the chapter mentions only in passing. - Choice of indicator species now shapes compliance verdicts. OSPAR notes that assessment status depends on whether the rock-shore dogwhelk or the sediment-dwelling netted whelk is monitored (OSPAR, 2022; Ruiz et al., 2017).

Weight for the lens. The chapter’s strongest lessons have been reinforced by the later record: - thresholds derived from the obvious high-dose endpoint fail; - early controls hit the tractable source rather than the largest one; - transboundary mobile sources need an institution with matching reach; - persistence creates legacy costs that fall on others; - substitution can reproduce the problem.

Its weakest element is the closing verdict. It holds descriptively, but the chapter’s own evidence (France 1982, the far-field decision) cuts against it. Standpoint also matters. Two authors worked for Greenpeace. The third, Langston, is also lead author of one of the key recovery studies used here (Langston et al., 2015). I have tried to rely on independent monitoring (OSPAR, Cefas, Norwegian, Irish, Dutch and Iberian programmes) wherever possible.

2013 update. The 2013 report (Late lessons II) does not update this case in Annex 3; its updated cases are fisheries, benzene, asbestos, PCBs, halocarbons, DES, antimicrobials, MTBE and growth hormones. Annex 2 summarises the case. Table A2.1 lists the first early warning as the 1976–81 collapse of the Arcachon oyster beds, effective action as 1982–1987 (the French, UK and north-east Atlantic bans) and the 2008 global ban, and “5–30” years of “substantial inaction” (EEA, 2013, pp. 702, 714). Chapter 12 of the 2013 volume (Price and Readman) carries the story forward to booster biocides (Claim 7). I read these from the EEA’s web copies, not the local extracts.

Claim 1. The planned global regime: IMO Resolution A.895(21) set 2003 and 2008 deadlines; the convention was expected in 2001; after 2008 “widespread inputs from shipping should, at least, be a thing of the past” (pp. 140–142; Table 13.1, p. 143)#

Original claim. - Deadlines in the draft convention were 2003 for ending application of organotin paints and 2008 for their presence on hulls. They were “adopted under IMO Assembly Resolution A.895(21) in November 1999, and formal adoption of the convention is expected in 2001” (pp. 140–141). - The IMO convention was a “global, legally binding treaty” whose significance “should not be underestimated” (p. 142). - “Complete phase-out of organotin paints from the global shipping fleet by 2008 will mark the closure of an important chapter … widespread inputs from shipping should, at least, be a thing of the past” (p. 142).

What happened since. - Adoption and entry into force. - The AFS Convention (the International Convention on the Control of Harmful Anti-fouling Systems on Ships) was adopted on 5 October 2001. It entered into force on 17 September 2008 (IMO). - Entry into force therefore came after both target dates. - The EU brought the dates into its own law through Regulation (EC) No 782/2003: - no application of TBT to EU-flagged ships from 2003; - ships with active TBT coatings barred from EU ports from 1 January 2008 (EEA, 2013, Table 12.3, p. 272; Langston et al., 2015). - Langston et al. (2015) state that “elsewhere, sealing or removal was permissible until 2013”, to allow for the five-year life of a coating applied in 2008. I did not check this reading against the treaty text. - Coverage. Equatorial Guinea became the 100th party in December 2025. Parties “now cover 95.77% of world merchant shipping by tonnage” (IMO, 17 December 2025). - Mechanism for new substances. The convention’s procedure for adding other harmful substances was used once, for cybutryne (Irgarol 1051): amendment MEPC.331(76), adopted 17 June 2021, in force 1 January 2023 (IMO). - Trade controls. Tributyltin compounds were added to Annex III of the Rotterdam Convention on prior informed consent in trade: as pesticides at COP-4 (October 2008; in force 1 February 2009), and as industrial chemicals by decision RC-8/5 (2017) (IISD ENB, 2008; InforMEA). - Did shipping inputs end? Largely, where the ban applied. - OSPAR’s imposex indicator (published 30 June 2022): - sites above the EAC fell from 81% (2008) to about 21–22% (2020); - sites at near-background levels rose from 0.5% to about 22%; - “Of the nine OSPAR subregions with enough data, imposex levels have decreased significantly in all subregions.” - The OSPAR hazardous substances assessment (2023): the 2008 ban “seems to have worked”, with imposex “below the EAC in open waters”. - Norway: no imposex at any station in 2017 (Schøyen et al., 2019). - Similar recoveries have been reported from: - New Zealand ports: imposex “largely absent” from Tauranga Harbour by 2015–17 (Jones and Ross, 2018); - Western Australia: “virtually complete recovery” by 2019 (Wells and Gagnon, 2020); - South Korea: significant decline after its total ban, though still enough exposure to cause imposex (Kim et al., 2017). - Not entirely a thing of the past. - OSPAR (2022) expects inputs to continue “from countries not in compliance with the ban, from disused vessels or installations” and from redistributed sediments. - It also notes that TBT “is still manufactured in the USA and is available on the market”. The source is Uc-Peraza et al. (2022), who found banned TBT paints on sale in the Caribbean, Central America and beyond in 2021, with a possibly “even more global distribution than previously thought”. - In Portugal a butyltin degradation index below 1 at about 90% of sites in 2014 pointed to fresh TBT inputs. The authors suggest illegal use and release from sediments (Laranjeiro et al., 2018). - Where the convention was not ratified or enforced, contamination persisted: - Cabo Verde, a non-party: no decline between 2012 and 2019 (Gomes et al., 2021); - Chile: imposex at 11 of 15 sites in 2016, attributed to lack of regulation (Batista et al., 2016).

Verdict: held up. The convention came when expected, and the phase-out largely achieved what the chapter forecast in the regions that applied it. Two qualifications: - legal force arrived later than the target dates everywhere except the EU; - “a thing of the past” is true for widespread inputs, not for localised, illegal or non-party use.

Weight. This strongly supports digest insight 7: a mobile, transboundary source needed an institution with matching reach. The record adds a nuance. A regional actor with port-state leverage (the EU refusing entry to TBT-coated ships from 2008) bridged the gap while the global treaty waited for ratifications. Recovery has tracked ratification and enforcement rather than the treaty text alone.

Claim 2. Small-vessel controls (France 1982, UK 1987, EU 1991) substantially reduced inputs and allowed partial recovery, but did not solve the problem (pp. 136, 138–139)#

Original claim. - Retail restrictions caused “a major shift away from the use of TBT paint on leisure craft” and “partial recovery of mollusc populations” (p. 138). - But recovery at Killybegs (the chapter misprints it as “Ballybegs”) was “slower than expected”. Sullom Voe and Yell Sound recovered “against a continuing high background of imposex”. Butyltins in the southern Gulf of St Lawrence “remained remarkably high eight years after retail restrictions”. French dissolved TBT had “stabilised” above effect levels (pp. 138–139). - The authors dispute the view that the 1980s controls “solved” the problem (p. 138).

What happened since. - Recovery was real and continued. - Dogwhelks recolonised shores where they had died out: the Isle of Cumbrae, north-east England, Shetland and south-west England (Birchenough et al., 2002). - Seven of eleven mainland sites in central southern England were colonised by 2007–08 (Bray et al., 2011). - At Plymouth Hoe, where dogwhelks were eliminated by 1985, a breeding population with a size spectrum like that of 1967 had re-established by 2013. By 2012–13 there was “no trace of imposex” at Bude, Widemouth or Renney (Langston et al., 2015). - Hotspots persisted where shipping and sediments kept supplying TBT. - Near Falmouth, imposex was still elevated in 2012–13 (mean VDSI, a standard imposex index, 2.8). Previously recorded populations in the Fal Estuary were “still extinct” (Langston et al., 2015). - In Southampton Water, the 1987 rule worked in the small-boat Hamble and Itchen. In the commercially used Test, “the original assumption … that TBT concentrations would be diluted sufficiently … was not upheld”. Clams at Cracknore were absent for two decades (Langston et al., 2015). - Poole Harbour met the 2 ng/l water standard at most sites by 2009. - UK status in 2024 (Cefas, UK Marine Strategy assessment, data 2004–2021, dogwhelks only): - “All stations are either showing a decreasing trend or stable level of imposex”; - 62% of Greater North Sea and 83% of Celtic Seas stations were below the EAC; - the Eastern Channel was above the EAC; - at Sullom Voe, none of 20 sites exceeded the EAC in the last two surveys. - Ireland. In 2010–11, sterile females were still found at 14 of 63 sites, and 75% of sites met OSPAR’s ecological quality objective. There was “a significant decline … following 2005” (Wilson et al., 2015). I found no separate post-2001 series for Killybegs. - France and Iberia. - Langston et al. (2015), citing Alzieu (2000), report “continuing low-level effects” in French oysters two decades after 1982. - Galicia met the objective in rock-shore dogwhelks by 2009 but showed “no further improvement thereafter”. The netted whelk lagged until 2011–2015 (Ruiz et al., 2017, 2018). - Portugal’s median VDSI fell from 3.96 to 0.78 in dogwhelks and from 3.39 to 0.29 in netted whelks over 15 years. Hotspots shifted to fishing ports and marinas (Laranjeiro et al., 2018). - Netherlands. Imposex had gone from the Eastern Scheldt by 2013 and from the North Sea coast by 2016. The dogwhelk returned to the Western Scheldt in 2014 (CLO indicator, version 10, 22 February 2022). Coastal populations were then hit by beach nourishment (sand replenishment) and dike works, an unrelated pressure. - Gulf of St Lawrence. I found no post-ban trend series. A 2006 survey, taken before the global ban, found butyltins at up to 489 ng Sn/g in benthic organisms even where sediments were only lightly contaminated (Michaud and Pelletier, 2006). - Timing. Much of the sustained improvement came after the whole-fleet measures of 2003–2008: Portugal after 2003, Ireland after 2005, Norway after 2008. The chapter was therefore right that small-boat controls alone would not finish the job.

Verdict: strengthened. The chapter’s middle position has been confirmed at both ends: substantial but partial recovery under small-boat controls, persistent hotspots, and no “problem solved”. Recovery accelerated once all vessels were covered.

Weight. Supports digest insights 6 (early controls hit the tractable segment) and 8 (partial success recast as “solved”). Langston et al.’s title, “A problem solved?”, answers the 2001 debate with a qualified no. One caution: Langston was a co-author of the chapter. The independent OSPAR and Cefas assessments point the same way.

Claim 3. Seagoing vessels are “potentially the most significant single contributor”; about 68 t TBT a year enters the North Sea from shipping; offshore imposex correlates with shipping density, though its significance was disputed (pp. 139–140)#

Original claim. - Large ships had been exempted on the argument that they spent most of their time at sea (p. 139). - Davies et al. (1998) estimated 68 t TBT a year entering the North Sea from shipping, mainly from leaching from hulls (pp. 139–140). - Imposex in common whelks from the open North Sea, the Strait of Malacca and Galicia correlated with shipping intensity (p. 139). - Nicholson and Evans (1997) questioned the significance of “mild” imposex “against a background of overexploitation by shell fisheries” (p. 140).

What happened since. - The source attribution has been tested by the ban, and it passed. - Norway: “Significantly higher levels of TBT and imposex were measured in coastal areas close to shipping lanes along most of the coast prior to 2008 than afterwards … In 2017, no sign of imposex was found” (Schøyen et al., 2019). Small-boat controls had been recommended across the north-east Atlantic since 1987 (PARCOM, p. 138). The persistence of imposex near shipping lanes until the all-vessel ban, and its disappearance afterwards, points to large ships as the remaining source. I did not check the date of Norway’s own small-boat rules. - Southern North Sea sediments: TBT has fallen by about 10% a year since the ban. OSPAR expects exceedances to persist for about 12 more years “along the major shipping lanes and big ports” (OSPAR, 2022). - North Sea brown shrimp: organotin levels fell about tenfold after the ban, coinciding with a recovery of the shrimp stock. The link to the stock is correlational (Verhaegen et al., 2012). - OSPAR (2023) reports imposex “below the EAC in open waters”. - A refinement. Ten Hallers-Tjabbes et al. (2003) found that the shipping–imposex relation in the North Sea depends on hydrography. Whelks at stratified deep-water stations in the Skagerrak had low organotin burdens despite heavy traffic, because dissolved organotins do not easily cross a pycnocline (the density layer between stratified water masses). Shipping density matters, but it does not act alone. - The rival explanation for population decline remains live. - The Dutch government’s environmental indicator reported “no recovery” of the common whelk despite falling TBT. It attributed this to both residual TBT and beam-trawl fishing (CLO, 2013). - The whelk was later dropped from the indicator altogether because bottom trawling confounded TBT-related trends (CLO, version 10, 2022). - A historical study of the Dutch Wadden Sea whelk fishery (1946–1970) found signs of overfishing and dredge damage before TBT arrived (de Vooys and van der Meer, 2010). - So TBT from shipping explains the offshore imposex, but the offshore population declines of whelks may owe as much to fishing. - The 68 t estimate. I found no later re-estimate of North Sea shipping inputs.

Verdict: strengthened on the source: the natural experiment of the large-ship ban confirmed shipping as the dominant remaining source. Contested (unresolved) on the population significance of offshore imposex, where fishing remains a confounder, as the chapter’s critics argued.

Weight. This strongly supports digest insight 6 (the segment exempted on an exposure assumption, “they spend their time at sea”, turned out to be the largest source). Grade it strong. When citing the offshore evidence, separate exposure (well established) from population effects (confounded).

Claim 4. Dose–response: sterilisation of female dogwhelks at 3–5 ng/l, almost all females affected at 10 ng/l; the 1985 20 ng/l target was inadequate and was replaced by 2 ng/l; imposex results from interference with “steroid hormone metabolism” (pp. 135, 137–138, 141)#

Original claim. - “Sterilisation of female dogwhelks occurs at TBT concentrations as low as 3–5 ng/l … with almost all females affected once concentrations reach 10 ng/l” (Gibbs et al., 1988; p. 137). - The 20 ng/l target was set on the assumption, from acute toxicity tests, that effects needed µg/l concentrations. It was replaced by a 2 ng/l standard (pp. 137–138, 141). - Imposex results “from interference with steroid hormone metabolism” (p. 135). Had “the endocrine-mediated mechanisms” been identified earlier, low-dose effects would have been clear; knowledge is “likely to continue to develop” (p. 141).

What happened since. - Standards moved further down. The EU water-quality standards directive set, for TBT (as the cation): - an annual-average standard of 0.0002 µg/l (0.2 ng/l); - a maximum allowable concentration of 0.0015 µg/l (1.5 ng/l); - classification as a priority hazardous substance (Directive 2008/105/EC, as amended by 2013/39/EU, Annex I).

That is ten times below the UK’s 2 ng/l and 100 times below the 1985 target. OSPAR notes that gastropods “are not the most sensitive species, as some fish larvae are hampered in their development at even lower TBT concentrations” (OSPAR, 2023). I found no post-2001 work contradicting the Gibbs dose–response. - The mechanism was revised. - Nishikawa et al. (2004) showed that organotins bind the retinoid X receptor with high affinity, and that injecting its natural ligand (9-cis retinoic acid) induces imposex in the rock shell. - Independent groups later confirmed organotin activation of RXR in dogwhelks and other species: Urushitani et al. (2018) with Portuguese co-authors; Fonseca et al. (2020); Giulianelli et al. (2020), who also implicate PPARγ. - Horiguchi (2023), whose group proposed the RXR hypothesis, summarises the field (in Japanese; quotations are my translation). The aromatase-inhibition (steroid) hypothesis was “most favoured” through the 1980s–2000s. His group could support none of the four older hypotheses, and RXR activation “is now regarded as the most likely” mechanism. - A 2026 systematic review found the vertebrate-steroid framing of mollusc endocrinology poorly supported. For example, 95% of studies measuring hormones in molluscs did not test receptor binding (Panagiotidis et al., 2026). - The RXR case rests heavily on one Japanese group’s work but has been replicated in Portugal, Argentina and elsewhere. I treat it as the mainstream view, not a settled consensus.

Verdict: partly held up. - Strengthened: the dose–response numbers, and the judgement that the 20 ng/l target was inadequate. Later standards went ten times lower again. - Superseded: the attribution to “steroid hormone metabolism”, now displaced by nuclear-receptor (RXR) activation. The chapter’s hedge (p. 141) anticipated revision.

Weight. Digest insight 1 (thresholds set from the obvious high-dose endpoint can be wrong by ten times or more) is strongly reinforced: the error ran to about 100 times. The mechanism revision adds a lesson the chapter could not state. A highly specific indicator can establish causation robustly even while the mechanism is misidentified. The misidentification then hides where else the chemical acts (see Claim 8 on vertebrate effects through the same receptor).

Claim 5. Persistence: half-lives of days in eutrophic surface waters but up to several years in nutrient-poor waters and sediments, especially anaerobic ones; a legacy that will delay recovery and burden dredging authorities (p. 141)#

Original claim. - “Half-life estimates in the order of days for eutrophic surface waters must be contrasted with up to several years for residues in nutrient-poor waters … and marine sediments …, especially anaerobic sediments” (p. 141). - This “raises the prospect of delayed recovery” and “represents a substantial legacy to be borne by authorities responsible for dredging operations”, which is “only now receiving formal recognition” in the dumping conventions (p. 141).

What happened since. - Longer than “several years” in places. - Langston et al. (2015) measured sediment half-times of up to 8 years in Poole and up to 33 years in Southampton, “longest near commercial shipping”. They observed that timescales in estuarine muds are “broadly consistent with” earlier projections of more than ten years for anoxic sediments. - A dredging programme in Southampton Water in the late 1990s “coincided with a reversal in the decline of TBT levels in water”. Approval was given in 2013 to dredge a further 23 million tonnes next to one marine Special Area of Conservation, with disposal near another. - Regional status. - Southern North Sea sediments: “no monitoring station is yet significantly below the EQS”; the mean was about 3.6 times the standard and up to 30 times at some stations; at least a decade more of exceedance is expected (OSPAR, 2022). - OSPAR adds that TBT “is only released slowly, though faster release occurs during dredging operations”. Harbour hotspots “typically close to shipyards and busy harbours” remain (OSPAR, 2022). - EEA (2018): TBT caused failure of good chemical status in 659 water bodies in 14 Member States. TBT is classed as a “uPBT” (a ubiquitous persistent, bioaccumulative and toxic substance) “owing to the difficulty in remediating contaminated areas”. “There is little that can be done to remediate water bodies failing for this substance”, other than removal or burial of sediments. - Who pays. Champ (2003) noted that the 2001 treaty added a requirement for safe removal and disposal of old coatings. But “the liability for the future dredging and disposal costs of TBT-contaminated port and harbor sediments has not been addressed.” - Hotspots and disposal sites. - Belgian ports (2023): TBT at 237–546 times the predicted no-effect concentration (Vanavermaete et al., 2023). - Portuguese shelf: high TBT around disposal sites for harbour dredge spoil; Portuguese dredging rules set no action limits for butyltins (Mil-Homens et al., 2023). - A French Mediterranean marina: sediment organotins rose after dredging (Montigny et al., 2025). - Swedish researchers are testing chemical and electrochemical treatment of TBT-contaminated dredged sediment (Norén et al., 2022). - In the UK, dredged material is screened against Cefas action levels for TBT (0.1 and 1 mg/kg) (Langston et al., 2015). - Monitoring. TBT in sediment and its biological effects became mandatory elements of OSPAR’s coordinated monitoring programme (OSPAR, 2022).

Verdict: strengthened. Persistence in anoxic sediments proved longer than the chapter’s upper estimate. The legacy fell, as predicted, on port and dredging authorities, and the treaty left its costs unassigned.

Weight. Digest insight 10 (persistence turns a use decision into a long legacy borne by others) is strongly reinforced. Grade it strong. The case now also shows a second-order effect: routine maintenance of the infrastructure (dredging) re-mobilises the legacy.

Claim 6. Only universal, global restrictions can address the totality of the problem; evidence is continued contamination in Japan despite “a national ban on all applications of TBT for marine antifouling” (p. 142)#

Original claim. “The continuation of the TBT problem in coastal waters of Japan …, despite a national ban on all applications of TBT for marine antifouling, clearly illustrates the transboundary nature of the problem. It would seem that universal, global restrictions are the only way to address the totality of the TBT problem” (p. 142).

What happened since. - What Japan actually did. Japan’s environment ministry (March 2010) describes three steps: - tributyltin oxide (TBTO) was designated a Class I specified chemical substance in January 1990, prohibiting manufacture, use and import; - other TBT compounds were designated Class II in September 1990, requiring notification, “with the quantities the production or import restricted accordingly”; - these followed voluntary industry controls.

The 2013 EEA chapter refers to “a national ban in the late 1990s” (EEA, 2013, p. 267). Either way, “a national ban on all applications” overstates the legal position during the period the chapter’s evidence covers. - The ministry’s own reading supports the chapter’s inference. - TBT in fish and shellfish fell from 1985, but the decline “has slowed down since 1998”. - Sediment contamination was highest in metropolitan bays. Given that TBT paints “has already been prohibited in Japan”, “a chief ongoing loading source may be due to the navigation of foreign vessels” (Japan MOE, 2010). - After the global ban, the picture in Asia was mixed. - Seto Inland Sea (published 2023): TBT at 1.0–2.8 ng/l in surface water (higher in bottom water) and 2.0–28 ng/g in sediment. Triphenyltin in sediment reached 2,700 ng/g. Organotins “have still been detected … from closed sea areas” (Harino et al., 2023). Those water levels fall within the low-ng/l range the chapter associates with imposex induction (p. 136), though below the 3–5 ng/l sterilisation range. They are roughly 5–14 times the EU annual-average standard, allowing for possible differences in how concentrations are expressed. - Hong Kong, 2004–2015: imposex incidence stayed at 100%, and “imposex levels and tissue concentrations … did not decline”. The authors conclude that the global convention “alone may be inadequate” (Ho et al., 2016). - South Korea: a significant decline after its total ban (Kim et al., 2017). - Across Asian seas, phenyltins now exceed butyltins in many areas, “possibly due to regional differences in policies or industries” (Pang et al., 2026). - I found no national post-2008 imposex series for Japan. - Elsewhere, recovery tracked ratification and enforcement (Claim 1): strong in Norway, New Zealand, Australia and Korea; absent in non-party Cabo Verde and in Chile.

Verdict: partly held up. - The transboundary argument is supported, including by the Japanese ministry itself, and global action proved necessary. - But “only” global restriction is not enough. Legacy sediments, non-antifouling uses (including phenyltins), illegal supply and non-parties sustain contamination. - The Japan example rests on an overstated description of Japan’s ban.

Weight. Digest insight 7 remains strong for the jurisdictional point. Treat the Japan illustration as moderate: it supports the inference only in the ministry’s more cautious form.

Claim 7. Alternatives: copper with booster biocides such as Irgarol 1051 “in some ways mirror early findings with respect to TBT”; biocide-free “non-stick” coatings are “perhaps the most promising contemporary solution”, though their use on large vessels was “under evaluation” (p. 141)#

Original claim. - Irgarol 1051’s “widespread occurrence … in certain estuaries” and “direct effects on plant growth in the field … in some ways mirror early findings with respect to TBT” (p. 141). - Using such examples to argue against substituting TBT is “a somewhat negative argument” (p. 141). - Non-stick coatings retain “a significant market share for leisure craft”. Their “viability and performance on larger vessels remain under evaluation, although applicability for fast-moving craft has been demonstrated” (p. 141). - Options “which do not rely on the release of hazardous substances … should, perhaps, be viewed most favourably” (p. 142).

What happened since. - Irgarol (cybutryne) followed the path the chapter implied. - Denmark banned Irgarol and diuron on pleasure craft in 2000. In 2001 the UK limited small-boat antifoulants to three substances, removing Irgarol and diuron. Irgarol concentrations then fell, which the authors attribute to control at the manufacturer level (Cresswell et al., 2006, as summarised in EEA, 2013, pp. 271–273). - Bermuda banned Irgarol and diuron paints in 2005 to protect corals (EEA, 2013, Box 12.2). - Cybutryne was listed as an EU priority substance with a 2.5 ng/l annual-average standard (Directive 2013/39/EU). - The EU did not approve it for antifouling use, citing “unacceptable risks” (Commission Implementing Decision (EU) 2016/107, 27 January 2016). Germany’s environment agency reported that it could not be used after 27 January 2017 (UBA, 20 April 2016). - The IMO banned its application to ships from 1 January 2023 (MEPC.331(76)). - The 2013 EEA volume devoted a chapter to the question, “Booster biocide antifoulants: is history repeating itself?”. It concluded that booster biocides “followed a very similar pattern to the evolution and demise of organotin”. It also noted that they lack TBT’s endocrine action but have “more broad-spectrum impacts” on primary producers (EEA, 2013, pp. 265, 273). - Copper remained the default, with growing evidence of its costs. - OSPAR (2022): in European waters TBT “has almost universally been replaced by paints containing other tin substances and/or copper”. It cites evidence that such paints lead to exceedance of quality standards. - In Arcachon Bay, copper in oysters rose over four decades, and copper isotopes “confirm that antifouling paints are the major local anthropogenic source” (Araújo, 2025). Copper transferred through the food chain lowered the condition of oyster spat in experiments (Akcha et al., 2022). - A copper-free “environmentally safe” product (tralopyril) showed “the highest toxicity and environmental risk” in a 2026 comparison of leisure-boat coatings (Lagerström et al., 2026). - Biocide-free coatings: proven, but not dominant on large ships. - A 2026 field comparison ranked a biocide-free silicone foul-release coating highest for sustainability on leisure boats, “combining strong antifouling efficacy with minimal acute toxicity” (Lagerström et al., 2026). - EEA (2018) reports that non-biocidal alternatives became established for leisure craft within the previous decade. - For commercial ships I found no authoritative figure for fleet share. A 2020 news feature in ACS Central Science reports that “almost every commercial and recreational vessel” carries antifouling paint, most of it containing copper and zinc. Fouling-release coatings are “readily commercially available” but “have not been widely adopted”. Reasons given: switching costs, and the fact that they do not work when ships sit idle (Fore, 2020; a news feature, not a primary source). - Market-research reports give higher numbers but are not independent. I have not used them. - Silicone coatings raise their own, lesser questions about leachates (Feng et al., 2012).

Verdict: held up. - On Irgarol it is strengthened: the “mirror” was borne out and ended in national, EU and global bans. - On biocide-free coatings it is partly held up: they are viable and the best-ranked environmentally, but copper-biocide paints still dominate large-ship use 25 years on.

Weight. Digest insight 11 (imperfect alternatives are used to resist substitution; framing the question around the function, not the chemical, helps) is reinforced as a pattern of regrettable substitution. The chemical-for-chemical route produced a second problem (Irgarol) and a third that is still growing (copper). As the chapter predicted, none proved “as damaging as TBT” (p. 142). Grade it moderate to strong. The chapter’s normative preference for non-release options is supported by later comparative testing, but the market has not followed.

Claim 8. Butyltins accumulate in marine mammals (up to 10 ppm in porpoise liver) and “might” impair immune function; human seafood intake could exceed immunotoxicity-based TDIs for some products (Belfroid et al., 2000), though another assessment found US intakes well below concern (Cardwell et al., 1999) (p. 140)#

Original claim. - “Although uncertainty remains, evidence suggests that accumulation of butyltins in top predators might adversely effect the immune system.” This rests on correlations in stranded bottlenose dolphins and diseased Californian sea otters (p. 140). - On human intake, the chapter sets Belfroid et al. (2000) against Cardwell et al. (1999) (p. 140).

What happened since. - Accumulation confirmed, then declining where controls applied. - California sea otters: total butyltins “decreased significantly … since the 1990s”, with estimated half-lives of about three years. Concentrations in a few otters that died of infectious disease were “close to or above the threshold levels for adverse health effects” (Murata et al., 2008). - Danish harbour porpoises: TBT and total butyltins declined between 1987–90 and 2010–19. Only PCBs, DDTs and mercury were above effect thresholds. The authors conclude that PCBs and DDTs “played a major role” in the Baltic porpoise decline (Dietz et al., 2026). - Immune effects remain plausible but unproven in the wild. - In vitro, dibutyltin and TBT suppressed phagocytosis, natural-killer and T-cell functions of harbour seal immune cells at 50–200 nM (Frouin et al., 2008). - I found no study establishing organotins as a cause of marine-mammal die-offs. - New endpoints and new compounds. - Pearl River Estuary humpback dolphins (2003–2021): butyltins fell after the ban, but phenyltins “continued to increase”. Tissue-relevant doses activated dolphin PPARγ and altered fatty-acid profiles (Zhang et al., 2022). - In laboratory mammals TBT promotes fat-cell formation through RXR and PPARγ at low doses (Grün et al., 2006). SCHER (2006) judged this “a previously uncharacterized type of toxic effect”, occurring at doses similar to those causing immunotoxicity, and so not requiring a TDI change. - Human intake. - EFSA’s 2004 opinion set a group TDI for TBT, dibutyltin, triphenyltin and dioctyltin based on immunotoxicity. It corresponds to 0.1 µg Sn/kg body weight per day (SCHER, 2006), or 0.25 µg/kg as bis(tributyltin) oxide (Eguchi et al., 2010). I could not open the EFSA opinion itself. - So Belfroid’s endpoint became the regulatory basis. But later exposure estimates found intakes below the TDI: - Norway: median intake 0.007 µg Sn/kg bw per day (mean 0.033); high consumers 0.015 (median) to 0.070 (mean), that is, up to about 70% of the TDI (EFSA’s estimates as reported in SCHER, 2006); - Finland: “a limited fraction” of the TDI in a market-basket study (SCHER, 2006); - France: exposure for high seafood consumers under 47% of the TDI; “Nobody would exceed this limit” (Guérin et al., 2007); - Mexico’s Yucatán: below the TDI (Uc-Peraza et al., 2022). - SCHER (2006) cautioned that consumers of fish from contaminated areas could have much higher intakes.

Verdict: partly held up. Accumulation in top predators is confirmed. The chapter’s “might” on immune effects remains a might for wild mammals, and other contaminants (PCBs) proved the stronger explanation for at least one population. On human exposure, the Cardwell reading prevailed for typical European diets, while the Belfroid endpoint became the basis of the standard.

Weight. Use this section for digest insight 2 (predicted behaviour fails elsewhere: accumulation in top predators was “simply not envisaged”, p. 141). That part is strong. Treat the immune-harm claims as suggestive only. The more durable later lesson is that the same molecular target (RXR) links the invertebrate effect to metabolic effects in vertebrates, an expansion of harm the chapter could not foresee.

Claim 9. Organotin inputs will continue through consumer-product uses; whether regulation will draw on the TBT lessons “remains to be seen” (p. 142); OSPAR’s target of ceasing all organotin releases by 2020 (Table 13.1, p. 143)#

Original claim. - “Organotin inputs will continue, of course, through their use as additives in a wide range of consumer products. Whether efforts to address these emerging challenges will draw on the lessons of the past remains to be seen” (p. 142). - Table 13.1 records OSPAR’s 1998 aim of “cessation of all releases of organotins to marine environment … in 2020” (p. 143).

What happened since. - The EU restricted organotins in consumer articles. Commission Decision 2009/425/EC (28 May 2009) set limits of 0.1% by weight of tin: - tri-substituted organotins (TBT, TPT) in articles, from 1 July 2010; - dibutyltin in mixtures and articles for the public, from 1 January 2012, with derogations for sealants, PVC and coatings until 1 January 2015; - dioctyltin in listed articles (textiles, gloves, footwear, childcare articles, nappies and others), from 1 January 2012.

Its first recital begins with the antifouling history. The rationale is immunotoxicity “via the thymus gland”, acting “in a cumulative way”, with “risk to human health, particularly for children”. To that extent regulation did draw on the TBT record. - Inputs continued. - EEA (2018): TBT also came from wood preservatives, silicone sealants, roof sheeting and textiles. “The remaining production and use of TBT continues to result in emissions from industry and UWWTPs [urban waste-water treatment plants].” - OSPAR (2022) lists wastewater treatment plants and landfills as potential sources. - A 2026 review: “The annual production of organotin has far exceeded that at the end of the last century.” Other organotins (mono- and dibutyltins, methyltins, tetrabutyltins) “may become or have become main organotin pollutants” (Pang et al., 2026). - OSPAR’s 2020 cessation target was not met. The 2023 Quality Status Report says OSPAR “has moved towards the 2020 cessation target” and the objective “has been partially fulfilled”. It describes progress towards cessation as “considerably slower” because of re-release from sediments. TBT specifically remains above sediment standards in the southern North Sea and above the EAC for imposex in two subregions (OSPAR, 2022, 2023). - Global instruments. TBT compounds were listed under the Rotterdam Convention (2008 and 2017; Claim 1). I found no listing under the Stockholm Convention on persistent organic pollutants, but I did not check this exhaustively.

Verdict: held up. Organotin inputs did continue from non-antifouling uses, and the 2020 cessation target was missed. “Remains to be seen” has a partly positive answer: the EU’s consumer-product restriction and the group TDI explicitly carry the TBT record forward.

Weight. This supports digest insight 10 (a long legacy), and adds a point the chapter only gestures at. Banning a substance in its most visible use leaves other uses of the same chemical family, and total production can grow while the headline problem shrinks.

Claim 10. Interpretive: none of the actions on TBT up to 2001 were precautionary; all followed extensive documentation of harm, including the “fundamentally retrospective” IMO convention (p. 142); weighed against France acting in 1982 “on the best information available” (p. 136)#

Original claim. - “It would be difficult to argue … that any of the actions to address TBT to date have been precautionary, resulting as they have from extensive documentation of ecological impacts” (p. 142). - The IMO convention “has come only after the consequences of continued use have been well documented … it too is fundamentally retrospective in action” (p. 142). - The same chapter says France acted “on the best information available linking the oyster collapse to the presence of TBT” before reliable environmental measurements existed (p. 136).

What happened since. - Later assessments agree that the response was mainly reactive. - The 2013 EEA volume records “5–30” years of “substantial inaction” for TBT (Annex 2, Table A2.1, p. 702). - Its booster-biocide chapter lists, among the lessons of TBT, that “imposex observed in 1970 in predatory gastropods … was considered acceptable” until the oyster crop failed (EEA, 2013, p. 267). - A 2026 review of seven decades of ecotoxicology uses TBT as one of its main cases. It concludes that unpredicted sensitivities, exposure pathways and modes of action “have typically resulted in reactive rather than preventive regulatory responses” (Sumpter and Margiotta-Casaluci, 2026). - But the later record complicates “fundamentally retrospective” for the global treaty. - When the IMO moved (1996–1999), the evidence for effects far from coasts was, in the chapter’s own words, “the subject of intense debate” (p. 139). - Champ (2000), writing just before the treaty, records a contested case: advocates cited offshore TBT, spreading imposex and bioaccumulation, while opponents pressed cost and alternatives arguments. - The ban’s results (Claim 3) showed that the contested offshore link was real. For that part of the problem, the IMO acted before the evidence was settled. That is closer to precaution than the chapter allows. - The coastal harm the treaty addressed was, as the chapter says, already well documented. - France 1982. France acted within about six years of the oyster failures and before detailed environmental data existed (p. 136). The chapter’s own wording fits “acting on reasonable grounds for concern”, which is closer to precaution. What made it possible was commercial loss, not the earlier ecological warning (the loss of the oyster drill). - The same pattern repeated for the substitute. - Irgarol was detected at high levels in 1993. - National restrictions followed in 2000–2001, the EU non-approval in 2016 and the IMO ban in 2023. - Each step followed documentation of exposure and effects (EEA, 2013; Claim 7). Bermuda, acting within about ten years to protect corals, is the faster exception.

Verdict: partly held up. “Mostly retrospective” is well supported and now echoed in later reviews. “None … precautionary” is too sweeping. France in 1982, and the IMO’s extension to large ships in the face of contested far-field evidence (later vindicated), both carry precautionary elements. As the digest notes, the verdict is partly definitional: if strong evidence of harm at the time of action counts as disqualifying, retrospection follows by construction.

Weight. Digest insight 9 (action that waits for conclusive evidence is retrospective by construction) stays strong descriptively. Use the IMO case as a mixed example: retrospective for coastal harm, anticipatory for offshore harm. The case also supports a narrower claim. Commercially salient harm (oysters) triggered the first, fastest action. Harm to species nobody valued commercially needed more documentation (digest insight 3, still moderate).

Minor factual checks#

• “Ballybegs” (p. 139) is a misprint for Killybegs, Co. Donegal. The cited paper is Minchin et al. (1997), “Biological indicators used to map organotin contamination from a fishing port, Killybegs, Ireland”.
• Japan’s “national ban on all applications” (p. 142). Japan’s own account is a 1990 prohibition of TBTO, restriction of other TBT compounds, and voluntary industry controls (Japan MOE, 2010). The 2013 EEA volume dates a national ban to “the late 1990s” (p. 267).
• EU small-boat ban date. The 2001 chapter and Table 13.1 give 1991. The 2013 volume says the European Community restricted TBT in 1989 (pp. 265, 267). The difference probably reflects adoption versus application of the Community directive. I did not verify the directive text.
• The 68 t North Sea estimate (pp. 139–140) comes from Davies et al. (1998). I found no later re-estimate.

Implications for the section’s transferable insights#

Stated in technology-neutral terms and keyed to the digest’s list.

New lessons from hindsight, not drawn in the chapter (technology-neutral): - A correct causal attribution can coexist with a wrong mechanism. The wrong mechanism can hide where else an agent acts. The receptor behind the invertebrate effect later linked the agent to metabolic effects in vertebrates. - Restricting the most visible use of a chemical family does not reduce total use. Production of related compounds grew, and related compounds became the main residue in some regions. - A global agreement’s real reach is set by ratification, enforcement and illicit supply, not its text. Legacy reservoirs keep releasing long after inputs stop. - Routine maintenance can re-release a legacy. Dredging needed to keep ports open releases stored TBT. - The choice of monitoring indicator can change the verdict on whether a target has been met. - Recovery can bring its own surprises. In New Zealand ports, the authors suggest that falling antifouling toxicity may have contributed to changes in community structure and the arrival of non-indigenous species (Jones and Ross, 2018). This is speculative.

Method and access notes#

• Search. I used web search until the session’s search budget ran out partway through the check. After that I worked from primary documents retrieved directly:
• Europe PMC and Crossref metadata and abstracts;
• EUR-Lex texts (Decision 2009/425/EC, Directive 2008/105/EC as amended, Implementing Decision 2016/107);
• OSPAR assessment pages;
• EEA reports (2013 Late lessons II chapter 12, Annexes 2–3 and Summary; Report 18/2018);
• the Japanese environment ministry report;
• the SCHER opinion;
• the Dutch CLO indicator pages;
• the IMO and UBA web pages;
• the Langston et al. (2015) author manuscript (Plymouth repository).
• Not opened or not verified.
• The EFSA 2004 opinion (blocked, HTTP 403): its TDI is taken from SCHER (2006) and Eguchi et al. (2010).
• The text of Regulation (EC) No 782/2003: its dates are taken from the 2013 EEA volume’s Table 12.3 and Langston et al. (2015).
• The resolution MEPC.331(76) PDF: dates are taken from the IMO convention page.
• Grün et al. (2006): cited via SCHER (2006).
• No national post-2008 imposex series for Japan; no post-2001 series for Killybegs or the Gulf of St Lawrence; no re-estimate of the 68 t figure; no authoritative fleet-share figure for foul-release coatings.
• The Stockholm Convention status was not checked exhaustively.
• Standpoint.
• Langston, a chapter author, is lead author of Langston et al. (2015). I have cross-checked its conclusions against OSPAR, Cefas and non-UK monitoring.
• The RXR hypothesis originated with Horiguchi’s group, whose 2023 summary I cite. Independent confirmation comes from Portuguese, Argentine and other groups.
• The CLO indicator is an official Dutch government compendium.
• Scratch files. Downloads and text extracts are in the session scratchpad (ll113/). None were written into the project.

Sources#

Treaties, regulation and official assessments 1. IMO. International Convention on the Control of Harmful Anti-fouling Systems on Ships (AFS) – convention page (adoption 5 Oct 2001; entry into force 17 Sep 2008; cybutryne amendment in force 1 Jan 2023). https://www.imo.org/en/about/conventions/pages/international-convention-on-the-control-of-harmful-anti-fouling-systems-on-ships-(afs).aspx (accessed 25 Sep 2026) 2. IMO. “Two international treaties to protect the ocean reach 100 accessions” (17 Dec 2025). https://www.imo.org/en/mediacentre/pages/whatsnew-2409.aspx 3. IMO. Resolution MEPC.331(76), adopted 17 June 2021 (not opened; dates from source 1). https://wwwcdn.imo.org/localresources/en/KnowledgeCentre/IndexofIMOResolutions/MEPCDocuments/MEPC.331(76).pdf 4. Regulation (EC) No 782/2003 on the prohibition of organotin compounds on ships (14 Apr 2003) (not opened; as summarised in source 16 and source 26). https://eur-lex.europa.eu/eli/reg/2003/782/oj 5. Commission Decision 2009/425/EC of 28 May 2009 amending Directive 76/769/EEC as regards organostannic compounds (OJ L 138, 4.6.2009). https://eur-lex.europa.eu/eli/dec/2009/425/oj 6. Directive 2008/105/EC on environmental quality standards, consolidated with Directive 2013/39/EU (Annex I, entries 30 and 40). https://eur-lex.europa.eu/eli/dir/2008/105/oj 7. Commission Implementing Decision (EU) 2016/107 of 27 Jan 2016 not approving cybutryne for product-type 21 (OJ L 21, 28.1.2016). https://eur-lex.europa.eu/eli/dec_impl/2016/107/oj 8. Umweltbundesamt. “EU Commission declines to authorise antifouling ingredient”, press release 16/2016 (20 Apr 2016). https://www.umweltbundesamt.de/en/press/pressinformation/eu-commission-declines-to-authorise-antifouling 9. IISD Earth Negotiations Bulletin. Rotterdam Convention COP-4 summary report (27–31 Oct 2008). https://enb.iisd.org/events/4th-meeting-conference-parties-cop4-rotterdam-convention-prior-informed-consent-procedure-5 10. InforMEA. Rotterdam Convention decision RC-8/5, “Listing of tributyltin compounds in Annex III” (2017). https://www.informea.org/en/listing-tributyltin-compounds-annex-iii-rotterdam-convention 11. OSPAR. QSR 2023 indicator assessment: Status and Trends in the Levels of Imposex in Marine Gastropods (TBT in Shellfish) (30 June 2022). https://oap.ospar.org/en/ospar-assessments/quality-status-reports/qsr-2023/indicator-assessments/tbt-shellfish/ 12. OSPAR. QSR 2023 indicator assessment: Status and Trends of Organotin in Sediments in the Southern North Sea (30 June 2022). https://oap.ospar.org/en/ospar-assessments/quality-status-reports/qsr-2023/indicator-assessments/status-trends-organotin-sed/ 13. OSPAR. QSR 2023 Hazardous Substances Thematic Assessment (2023). https://oap.ospar.org/en/ospar-assessments/quality-status-reports/qsr-2023/thematic-assessments/hazardous-substances/ 14. Cefas. Marine Online Assessment Tool: Imposex (UK Marine Strategy 2024 assessment, data 2004–2021). https://moat.cefas.co.uk/pressures-from-human-activities/contaminants/imposex/ (accessed 25 Sep 2026) 15. EEA. Chemicals in European waters: knowledge developments, EEA Report No 18/2018, section 3.8.1 and Box 3.7 (pp. 47–49). https://www.eea.europa.eu/en/analysis/publications/chemicals-in-european-water 16. EEA. Late lessons from early warnings: science, precaution, innovation, EEA Report No 1/2013. Chapter 12, Price, A. R. G., Readman, J. W., “Booster biocide antifoulants: is history repeating itself?” (pp. 265–278); Annex 2 (Table A2.1, p. 702; TBT summary, p. 714); Annex 3 (from p. 717); Summary. https://www.eea.europa.eu/en/analysis/publications/late-lessons-2 17. Compendium voor de Leefomgeving (CLO, Netherlands). Indicator “Purperslak en wulk en aangroeiwerende verven”: version 5 (5 Feb 2008), https://www.clo.nl/indicatoren/nl110405-purperslak-en-wulk-en-aangroeiwerende-verven ; 1956–2012 version (17 Dec 2013), https://www.clo.nl/indicatoren/nl110408-purperslak-en-wulk-en-aangroeiwerende-verven-1956-2012 ; version 10 (22 Feb 2022), https://www.clo.nl/indicatoren/nl1104-purperslak-en-wulk-en-aangroeiwerende-verven 18. Ministry of the Environment, Japan. Present status of marine pollution in the sea around Japan (PDF dated 29 Mar 2010), section 3.1.1(4) “Butyltin compounds”. https://www.env.go.jp/water/kaiyo/monitoring/status_report/en-2.pdf 19. SCHER (EU Scientific Committee on Health and Environmental Risks). Opinion on organotin compounds (30 Nov 2006). https://ec.europa.eu/health/ph_risk/committees/04_scher/docs/scher_o_047.pdf 20. EFSA CONTAM Panel. Opinion on the health risks to consumers associated with exposure to organotins in foodstuffs. EFSA Journal 2004;102 (not opened; access blocked). https://doi.org/10.2903/j.efsa.2004.102

Recovery, monitoring and sources

21. Langston, W. J., Pope, N. D., Davey, M. et al. Recovery from TBT pollution in English Channel environments: a problem solved? Mar Pollut Bull 2015;95:551–564. https://doi.org/10.1016/j.marpolbul.2014.12.011 ; author manuscript https://plymsea.ac.uk/id/eprint/6535/
22. Birchenough, A. C., Evans, S. M., Moss, C., Welch, R. Re-colonisation and recovery of populations of dogwhelks Nucella lapillus on shores formerly subject to severe TBT contamination. Mar Pollut Bull 2002. https://doi.org/10.1016/s0025-326x(01)00308-3
23. Bray, S., McVean, E. C., Nelson, A. et al. The regional recovery of Nucella lapillus populations from marine pollution, facilitated by man-made structures. J Mar Biol Assoc UK 2012;92:1585–1594 (online 2011). https://doi.org/10.1017/s0025315411001317
24. Schøyen, M., Green, N. W., Hjermann, D. Ø. et al. Levels and trends of TBT and imposex in dogwhelk along the Norwegian coastline from 1991 to 2017. Mar Environ Res 2019. https://doi.org/10.1016/j.marenvres.2018.11.011
25. Wilson, J. G., Minchin, D., McHugh, B. et al. Declines in TBT contamination in Irish coastal waters 1987–2011, using the dogwhelk as a biological indicator. Mar Pollut Bull 2015. https://doi.org/10.1016/j.marpolbul.2015.08.038
26. Laranjeiro, F., Sánchez-Marín, P., Oliveira, I. B. et al. Fifteen years of imposex and tributyltin pollution monitoring along the Portuguese coast. Environ Pollut 2018. https://doi.org/10.1016/j.envpol.2017.09.056
27. Galante-Oliveira, S. et al. Nucella lapillus imposex levels after legislation prohibiting TBT antifoulants: temporal trends from 2003 to 2008 along the Portuguese coast. J Environ Monit 2011. https://doi.org/10.1039/c0em00140f
28. Ruiz, J. M., Carro, B., Albaina, N. et al. Bi-species imposex monitoring in Galicia (NW Spain) shows contrasting achievement of the OSPAR Ecological Quality Objective for TBT. Mar Pollut Bull 2017. https://doi.org/10.1016/j.marpolbul.2016.10.058
29. Ruiz, J. M., Carro, B., Albaina, N. et al. Extended imposex monitoring in N Atlantic Spain confirms punctual attainment of European environmental objectives for TBT. Mar Pollut Bull 2018. https://doi.org/10.1016/j.marpolbul.2017.11.048
30. ten Hallers-Tjabbes, C. C., Wegener, J. W., van Hattum, B. et al. Imposex and organotin concentrations in Buccinum undatum and Neptunea antiqua from the North Sea: relationship to shipping density and hydrographical conditions. Mar Environ Res 2003. https://doi.org/10.1016/s0141-1136(02)00217-9
31. de Vooys, C. G. N., van der Meer, J. The whelk (Buccinum undatum) in the western Dutch Wadden Sea in the period 1946–1970: assessment of population characteristics and fishery impact. J Sea Res 2010;63:11–16. https://doi.org/10.1016/j.seares.2009.08.005
32. Verhaegen, Y., Monteyne, E., Neudecker, T. et al. Organotins in North Sea brown shrimp after implementation of the TBT ban. Chemosphere 2012. https://doi.org/10.1016/j.chemosphere.2011.11.028
33. Michaud, M. H., Pelletier, E. Sources and fate of butyltins in the St. Lawrence Estuary ecosystem. Chemosphere 2006. https://doi.org/10.1016/j.chemosphere.2005.12.002
34. Jones, M. R. L., Ross, P. M. Recovery of the New Zealand muricid dogwhelk Haustrum scobina from TBT-induced imposex. Mar Pollut Bull 2018. https://doi.org/10.1016/j.marpolbul.2017.11.034
35. Wells, F. E., Gagnon, M. M. A quarter century of recovery of the whelk Thais orbita from tributyltin pollution off Perth, Western Australia. Mar Pollut Bull 2020. https://doi.org/10.1016/j.marpolbul.2020.111408
36. Kim, N. S., Hong, S. H., Shin, K. H., Shim, W. J. Imposex in Reishia clavigera as an indicator to assess recovery of TBT pollution after a total ban in South Korea. Arch Environ Contam Toxicol 2017. https://doi.org/10.1007/s00244-017-0369-x
37. Ho, K. K., Zhou, G. J., Xu, E. G. et al. Long-term spatio-temporal trends of organotin contaminations in the marine environment of Hong Kong. PLoS One 2016;11:e0155632. https://doi.org/10.1371/journal.pone.0155632
38. Batista, R. M., Castro, I. B., Fillmann, G. Imposex and butyltin contamination still evident in Chile after TBT global ban. Sci Total Environ 2016. https://doi.org/10.1016/j.scitotenv.2016.05.039
39. Gomes, D. M., Galante-Oliveira, S., Almeida, C. et al. Temporal evolution of imposex and butyltin contamination in Gemophos viverratus from São Vicente (Cabo Verde). Mar Pollut Bull 2021. https://doi.org/10.1016/j.marpolbul.2021.112633
40. Uc-Peraza, R. G., Castro, Í. B., Fillmann, G. An absurd scenario in 2021: banned TBT-based antifouling products still available on the market. Sci Total Environ 2022;805:150377. https://doi.org/10.1016/j.scitotenv.2021.150377
41. Harino, H., Ohji, M., Kono, K. et al. Current status of antifouling biocides contamination in the Seto Inland Sea, Japan. Arch Environ Contam Toxicol 2023. https://doi.org/10.1007/s00244-023-01036-8
42. Eguchi, S., Harino, H., Yamamoto, Y. Assessment of antifouling biocides contaminations in Maizuru Bay, Japan. Arch Environ Contam Toxicol 2010. https://doi.org/10.1007/s00244-009-9394-8
43. Minchin, D., Bauer, B., Oehlmann, J., Schulte-Oehlmann, U., Duggan, C. B. Biological indicators used to map organotin contamination from a fishing port, Killybegs, Ireland. Mar Pollut Bull 1997. https://doi.org/10.1016/s0025-326x(96)00108-7

Sediments and dredging

44. Champ, M. A. Economic and environmental impacts on ports and harbors from the convention to ban harmful marine anti-fouling systems. Mar Pollut Bull 2003. https://doi.org/10.1016/s0025-326x(03)00106-1
45. Champ, M. A. A review of organotin regulatory strategies, pending actions, related costs and benefits. Sci Total Environ 2000. https://doi.org/10.1016/s0048-9697(00)00506-4
46. Vanavermaete, D., Hostens, K., Everaert, G. et al. Assessing the risk of booster biocides for the marine environment: a case study at the Belgian part of the North Sea. Mar Pollut Bull 2023. https://doi.org/10.1016/j.marpolbul.2023.115774
47. Mil-Homens, M., Almeida, C. M. R., Dias, S. et al. Spatial distribution and temporal trends of butyltin compounds in short sediment cores of the SW Portuguese Shelf. Sci Total Environ 2023. https://doi.org/10.1016/j.scitotenv.2023.165872
48. Montigny, C., Chouba, C., Domeau, A. et al. Assessing pollution in sediment and water before, during and after sediment dredging in a Mediterranean harbor. J Environ Manage 2025. https://doi.org/10.1016/j.jenvman.2025.126182
49. Norén, A., Lointier, C., Modin, O. et al. Removal of organotin compounds and metals from Swedish marine sediment using Fenton’s reagent and electrochemical treatment. Environ Sci Pollut Res 2022. https://doi.org/10.1007/s11356-021-17554-8

Mechanism and wider effects

50. Nishikawa, J., Mamiya, S., Kanayama, T. et al. Involvement of the retinoid X receptor in the development of imposex caused by organotins in gastropods. Environ Sci Technol 2004. https://doi.org/10.1021/es049593u
51. Urushitani, H., Katsu, Y., Kagechika, H. et al. Characterization and comparison of transcriptional activities of the retinoid X receptors by various organotin compounds in three prosobranch gastropods. Aquat Toxicol 2018. https://doi.org/10.1016/j.aquatox.2018.03.029
52. Giulianelli, S., Primost, M. A., Lanari, C., Bigatti, G. RXR expression in marine gastropods with different sensitivity to imposex development. Sci Rep 2020. https://doi.org/10.1038/s41598-020-66402-1
53. Fonseca, E., Ruivo, R., Borges, D. et al. Of retinoids and organotins: the evolution of the retinoid X receptor in Metazoa. Biomolecules 2020;10:594. https://doi.org/10.3390/biom10040594
54. Horiguchi, T. Imposex in gastropods and its induction mechanism (symposium paper, in Japanese). Nippon Suisan Gakkaishi 2023;89(2). https://www.jstage.jst.go.jp/article/suisan/89/2/89_WA3008-1/_pdf
55. Panagiotidis, K., Miller, T. H., Martin, O. V., Baynes, A. Mapping molluscan endocrinology: a systematic and critical appraisal. Biol Rev 2026. https://doi.org/10.1002/brv.70112
56. Grün, F., Watanabe, H., Zamanian, Z. et al. Endocrine-disrupting organotin compounds are potent inducers of adipogenesis in vertebrates. Mol Endocrinol 2006 (cited via SCHER 2006; not read). https://doi.org/10.1210/me.2005-0367
57. Pang, J., Chen, H., Huang, S., Zhang, Y., Lin, K. Both legacy and emerging: organotin pollutants in marine environments. Mar Pollut Bull 2026. https://doi.org/10.1016/j.marpolbul.2026.119599
58. Sumpter, J. P., Margiotta-Casaluci, L. Ecotoxicology in the context of biodiversity loss: lessons from seven decades of chemical impacts and paths forward. Environ Sci Technol 2026. https://doi.org/10.1021/acs.est.6c00982

Marine mammals and human exposure

59. Murata, S., Takahashi, S., Agusa, T. et al. Contamination status and accumulation profiles of organotins in sea otters found dead along the coasts of California, Washington, Alaska and Kamchatka. Mar Pollut Bull 2008. https://doi.org/10.1016/j.marpolbul.2008.01.019
60. Dietz, R., Berenguer López, G., Bossi, R. et al. Decreases in contaminant loads and risk of adverse effects in harbour porpoises from Inner Danish waters over three decades. Chemosphere 2026. https://doi.org/10.1016/j.chemosphere.2026.144850
61. Zhang, X., Yu, R., Xie, Y., Yu, R. Q., Wu, Y. Organotins remain a serious threat to the Indo-Pacific humpback dolphins in the Pearl River Estuary. Environ Sci Technol 2022. https://doi.org/10.1021/acs.est.2c02780
62. Frouin, H., Lebeuf, M., Saint-Louis, R. et al. Toxic effects of tributyltin and its metabolites on harbour seal immune cells in vitro. Aquat Toxicol 2008;90:243–251. https://doi.org/10.1016/j.aquatox.2008.09.005
63. Guérin, T., Sirot, V., Volatier, J. L., Leblanc, J. C. Organotin levels in seafood and its implications for health risk in high-seafood consumers. Sci Total Environ 2007. https://doi.org/10.1016/j.scitotenv.2007.08.027
64. Uc-Peraza, R. G., Delgado-Blas, V. H., Rendón-von Osten, J. et al. Organotin contamination in seafood from the Yucatán Peninsula, Mexico: is there a potential risk for the health of consumers? Chemosphere 2022. https://doi.org/10.1016/j.chemosphere.2022.136178

Alternatives

65. Araújo, D. F. Copper isotope ratios in oysters from the French Mussel Watch program confirm antifouling paints as a major source of copper in Arcachon Bay. Mar Pollut Bull 2025. https://doi.org/10.1016/j.marpolbul.2025.118503
66. Akcha, F., Coquillé, N., Sussarellu, R. et al. Trophic transfer of copper decreases the condition index in Crassostrea gigas spat. Sci Total Environ 2022. https://doi.org/10.1016/j.scitotenv.2022.153841
67. Lagerström, M., Le Bihanic, F., Veensalu, L. et al. Sustainability ranking of antifouling coatings for leisure boats – balancing efficacy and environmental impact. J Hazard Mater 2026. https://doi.org/10.1016/j.jhazmat.2026.141600
68. Feng, D., Rittschof, D., Orihuela, B. et al. The effects of model polysiloxane and fouling-release coatings on embryonic development of a sea urchin and a fish. Aquat Toxicol 2012;110–111:162–169. https://doi.org/10.1016/j.aquatox.2012.01.005
69. Fore, M. Seeking nontoxic coatings to keep ship hulls clean. ACS Cent Sci 2020 (news feature; secondary). https://doi.org/10.1021/acscentsci.0c01350