Márcia Alexandra the Course of TBT Pollution in Miranda Souto the World During the Last Decade

Márcia Alexandra The course of TBT pollution in Miranda Souto the world during the last decade

Evolução da poluição por TBT no mundo durante a última década

DECLARAÇÃO

Declaro que este relatório é integralmente da minha autoria, estando devidamente referenciadas as fontes e obras consultadas, bem como identificadas de modo claro as citações dessas obras. Não contém, por isso, qualquer tipo de plágio quer de textos publicados, qualquer que seja o meio dessa publicação, incluindo meios eletrónicos, quer de trabalhos académicos.

Márcia Alexandra The course of TBT pollution in Miranda Souto the world during the last decade

Evolução da poluição por TBT no mundo durante a última década

Dissertação apresentada à Universidade de Aveiro para cumprimento dos requisitos necessários à obtenção do grau de Mestre em Toxicologia e Ecotoxicologia, realizada sob orientação científica do Doutor Carlos Miguez Barroso, Professor Auxiliar do Departamento de Biologia da Universidade de Aveiro.

O júri

Presidente Professor Doutor Amadeu Mortágua Velho da Maia Soares Professor Catedrático do Departamento de Biologia da Universidade de Aveiro

Arguente Doutora Ana Catarina Almeida Sousa Estagiária de Pós-Doutoramento da Universidade da Beira Interior

Orientador Carlos Miguel Miguez Barroso Professor Auxiliar do Departamento de Biologia da Universidade de Aveiro

Agradecimentos A Deus, pela força e persistência que me deu durante a realização desta tese.

Ao apoio e a força dados pela minha família para a realização desta tese.

Á Doutora Susana Galante-Oliveira, por toda a aprendizagem científica, paciência e pelo apoio que me deu nos momentos mais difíceis ao longo deste percurso. Ao Sr. Prof. Doutor Carlos Miguel Miguez Barroso pela sua orientação científica.

Agradeço todo o apoio dado pelos meus amigos.

“ Não deixe que a saudade sufoque, que a rotina acomode, que o medo impeça de tentar. Desconfie do destino e acredite em você. Gaste mais horas realizando que sonhando, fazendo que planejando, vivendo que esperando…” (Luiz Fernando Veríssimo).

Acknowledgements To God, by the strength and persistence that gave me during the making of this thesis.

To support and strength given by my family for the realization of this thesis. To Susana Galante-Oliveira, by all the scientific learning, patience and by the support that gave me in the most difficult moments along this route.

To Prof. Dr. Carlos Miguel Miguez Barroso for the scientific guidance. .

I appreciate all the support given by my friends.

"Do not let that missing suffocates, that the routine accommodates, that fear hinders you from trying. Mistrusts destiny and believes you. Spend more hours realizing than dreaming, doing than planning, living than expecting ... " (Luiz Fernando Veríssimo).

palavras-chave Organoestânicos; Tributilestanho; TBT; imposex; biomarcador; bioindicador; biomonitorização da poluição por TBT; Organização Marítima Internacional (OMI); Convenção AFS

resumo Os organoestânicos (OTs) são compostos organometálicos. Apesar das múltiplas aplicações dos OTs, a notoriedade destes compostos, no que diz respeito a fenómenos de poluição, é devida ao tributilestanho (TBT), um potente biocida usado em tintas antivegetativas desde a década de 60 para prevenir a bioincrustação nas superfícies imersas, nomeadamente nos cascos de embarcações. Apesar da extrema eficácia destas tintas no combate à bioincrustação são também extremamente tóxicas para organismos não-alvo pelo que várias medidas legislativas restringindo o seu uso foram implementadas em vários países. O TBT provoca uma enorme variedade de efeitos nefastos em espécies não- alvo (sub-letais / letais) de diferentes grupos taxonómicos (de bactérias a mamíferos). A primeira evidência de efeitos nefastos induzidos pelo TBT em espécies não alvo surgiu na década de 70 em ostras da espécie Crassostrea gigas na Baía de Arcachon, França. As suas conchas sofreram um espessamento com a consequente diminuição do volume da parte comestível, diminuindo o seu valor comercial. Na mesma década verificou-se a ocorrência de características sexuais masculinas em fêmeas de gastrópodes prosobrânquios (formação de um pénis e/ou desenvolvimento de um vaso deferente), por exposição ao TBT. Esse fenómeno foi designado de "imposex" por Smith em 1971. O imposexo é o melhor exemplo conhecido de disrupção endócrina provocada por um poluente sendo utilizado como biomarcador para monitorizar a poluição ambiental por TBT.

resumo (cont.) Nesta tese fez-se um levantamento da metodologia utilizada na monitorização da poluição por TBT a nível mundial durante a última década (2003-2013), nomeadamente, o número de artigos publicados por ano/continente e país sobre as espécies bioindicadoras do imposexo mais utilizadas na biomonitorização da poluição por TBT nesta década e em função da sua localização geográfica. Neste trabalho avaliou-se, também, a eficácia da Convenção AFS na redução da poluição por TBT a nível mundial, através da evolução temporal (2003-2013) dos níveis de imposexo por espécie bioindicadora e zona geográfica, e avaliou-se a evolução temporal (2003-2013) da concentração de OTs no biota, água e sedimentos por zona geográfica. Verificou-se um aumento do número de publicações de monitorização da poluição ambiental por TBT em 2008, coincidente com o momento de entrada em vigor da Convenção AFS; um decréscimo gradual do número de publicações de 2008 a 2013, após a entrada em vigor da Convenção AFS; a maioria dos estudos de monitorização no período 2003-2013 foi realizada na Europa, seguindo-se os continentes: America, Ásia, África e Oceania; neste período foram usadas 96 espécies bioindicadoras de gastrópodes (imposex), sendo que as 5 mais utilizadas foram: Nucella lapillus, Hexaplex trunculus, Nassarius reticulatus, Stramonita haemastoma e Thais clavigera. Neste trabalho é possível verificar a utilização de uma maior diversidade de espécies bioindicadoras na América e na Ásia em relação à Europa, África e Oceania são os continentes que apresentam menor número de espécies bioindicadoras. Independentemente da região geográfica onde foram realizados os estudos, e apesar da grande diversidade de espécies utilizadas, os bioindicadores foram todos caenogastropodes (subclasse Caenogastropoda) e mais frequentemente das ordens Neogastropoda e Littorinimorpha, sendo a maioria da família Muricidae. Neste trabalho também é possível verificar descida dos níveis de imposexo nas diferentes espécies bioindicadoras, utilizadas por continente e país, e descida das concentrações de OTs no biota, água e sedimentos, por continente e país, desde a entrada em vigor da convenção AFS em Setembro de 2008.

keywords Organotins; Tributyltin; TBT; imposex; biomarker; bioindicator; TBT pollution biomonitoring; International Maritime Organization (IMO); AFS Convention

abstract Organotins (OTs) are organometallic compounds. Despite of the multiple applications of the OTs, the notoriety of these compounds is due to tributyltin (TBT), a potent biocide used in antifouling paints since the 60s to prevent biofouling on submerged surfaces, including the hulls of ships. Despite of the extreme effectiveness of these paints in combating biofouling, they are also extremely toxic to non-target organisms, therefore several legislative measures restricting their use were implemented in several countries worldwide. TBT causes a huge variety of adverse effects on non-target species (sublethal / lethal) of different taxonomic groups (from bacteria to mammals). The first evidence of adverse effects induced by TBT on non-target species emerged in the 70s in oysters of the species Crassostrea gigas in the Bay of Arcachon, France. Their shells have suffered thickening with consequent decrease in the volume of edible portion, reducing its commercial value. In the same decade was verified the occurrence of male sexual characteristics in female prosobranch gastropods (formation of a penis and / or development of a vas deferens), by exposure to TBT. This phenomenon was coined as "imposex" by Smith in 1971. The imposex is the best known example of endocrine disruption caused by a pollutant being used as a biomarker to monitoring the environmental pollution by TBT.

abstract (cont.) The gastropods are used globally for monitoring environmental pollution by TBT. In the present work we studied how gastropods have been used to track the course of TBT pollution worldwide during the last decade (2003-2013), and the main results obtained. We searched for the number of published articles by year / continent and country regarding the bioindicator species most used in biomonitoring TBT pollution in function of their geographical location; we evaluated the efficacy of the AFS Convention in reducing TBT pollution worldwide; we evaluated the temporal evolution (2003-2013) of the levels of imposex by bioindicator species and geographical area, as well as the concentration of OTs in the biota, water and sediments. There was noticed an increase of the number of publications for monitoring environmental pollution by TBT in 2008, coincident with the time of entry into force of the AFS Convention and a gradual decrease of the number of publications from 2008 to 2013, after the entry into force of this Convention; most of the monitoring studies developed in the period 2003-2013 were conducted in Europe, followed by the continents: America, Asia, Africa e Oceania; in this period were used 96 bioindicator species of imposex, being that the 5 more used were: Nucella lapillus, Hexaplex trunculus, Nassarius reticulatus, Stramonita haemastoma and Thais clavigera. A greater diversity of bioindicators was used in America and Asia compared to Europe, whist Africa and Oceania were the continents with smaller number of bioindicator species used. Regardless of the geographic region, and despite the wide diversity of species used, the bioindicators were all caenogastropods (Subclass Caenogastropoda) and most frequently of Order Neogastropoda and Littorinimorpha, being the majority from family Muricidae. There was a decrease of the levels of imposex in the different bioindicators used by continent and country, as well as a decline of the concentrations of OTs in biota, water and sediments, by continent and country, since the entry into force of the AFS Convention in September 2008.

Table of Contents

Chapter 1: General Introduction ...... 21 1.1 Organotin (OT) compounds ...... 22 1.1.1 Characterization and physicochemical properties ...... 23 1.1.2 Production and applications ...... 27 1.1.2.1 Stabilizers of PVC and CPVC ...... 30 1.1.2.2 Biocides in Agriculture ...... 32 1.1.2.3 Wood preservation...... 33 1.1.2.4 Antitumor drugs in medicine ...... 33 1.1.2.5 Biocides in antifouling (AF) paints ...... 34 1.2 Adverse effects on living organisms ...... 40 1.3 The imposex phenomenon ...... 46 1.3.1 Possible mechanisms ...... 48 1.3.2 Imposex as a biomarker of TBT pollution ...... 56 1.3.2.1 Vantages and disadvantages ...... 58 1.3.2.2 Indicator species ...... 60 1.3.2.3 Biomonitoring guidelines ...... 64 1.4 Restrictive measures on the use of Organotins (OTs) ...... 68 References ...... 74 Chapter 2: The dissertation ...... 102 2.1 Thesis aims and organization ...... 103 2.2 Methods for the data obtaining and processing ...... 104 Chapter 3: Evolution of the worldwide pollution by TBT in the last decade...... 106 3.1 Global trend of TBT pollution monitoring in the last decade ...... 108 3.1.1 General results on the bibliographic search ...... 108 3.1.1.1 Number of studies published between 2003 and 2013 ...... 108 3.1.1.2 Worldwide distribution of the studies published ...... 109 3.1.2 Species used as bioindicators of imposex from 2003 to 2013 ...... 113 3.1.2.1 By taxonomic group...... 119 3.1.2.2 By geographic region ...... 122 3.2 TBT pollution evolution from 2003 to 2013 ...... 126

3.2.1 Imposex levels in gastropod females between 2003 and 2013 ...... 126 3.2.2 OTs concentrations in biota, sediment and water between 2003 and 2013 ...... 132 References ...... 142 Chapter 4: Final considerations ...... 166 4.1 General conclusions ...... 167 References ...... 171

Acronyms %I………………………………. Percentage of Imposex Affected Females %S……………………………… Percentage of Sterile Females AAS……………………………. Atomic Absorption Spectroscopy AF...... Antifouling AF paints ...... Antifouling paints AFS Convention ...... International Convention on the Control of Harmful Antifouling Systems on Ships AV...... Antivegetatives BDI…………………………….. Butyltins degradation index BTs...... Butyltins CEMP ...... Co-ordinated Environmental Monitoring Program COPR…………………………. Control of Pesticides Regulations CPVC…………………………. Chlorinated PVC CYP19 ...... Cytochrome P450-dependent aromatase DBT ...... Dibutyltin EAC……………………………. Ecotoxicological / Environmental Assessment Criteria EC...... European Community EcoQ………………………….. Ecological Quality EcoQO……………………….. Ecological Quality Objective EDCs ...... Endocrine Disruptors EU ...... European Union FEPA …………………………. Food and Environment Protection Act FPL ...... Females Penis Length FSUs…………………………… Floating storage units FPSOs………………………… Floating production storage and offtake units IMO ...... International Maritime Organization JAMP…………………………. Joint Assessment and Monitoring Program MBT ...... Monobutyltin MEPC ...... Marine Environment Protection Committee

MPL...... Males Penis Length OSPAR ...... Oslo and Paris Convention OT ...... Organotin OTs...... Organotins OTs paints ...... Organotins based antifouling paints PCI ...... Penis Classification Index PL ...... Penis Length PMF ...... Penis Morphogenic Factor PVC ...... Polyvinyl Chloride RPLI ...... Relative Penis Length Index RPSI ...... Relative Penis Size Index TBT ...... Tributyltin TPT ...... Triphenyltin VDS ...... Vas Deference Sequence VDSI ...... Vas Deferens Sequence Index

List of Tables

Table 1.1: Industrial applications of different organotin species indicated by the respective general formula. R: organic substituent group linked to the tin (Sn) central atom (e.g. methyl, ethyl, butyl, propyl, octyl, phenyl); X: functional anionic ligand (e.g. chloride, fluoride, oxide, hydroxide, carboxylate, thiolate). Adapted from Sousa et al. (2013).………………………………………….29

Table 1.2: Summary of ecotoxicological effects of TBT indicated by taxonomic group. ……………………………………………………………………………………………………………………………………………………42

Table 1.3: Coloured scheme presented by CEMP for the TBT, showing specific biological effects assessment classes in dog-whelks and other gastropods. In this scheme, green means that the OSPAR EcoQO on imposex is met. However, it should be taken into account that the EcoQO only applies to the species in the white columns. Extracted from Galante-Olivera (2010). …………………………………………………………………………………………………………………………………………………….67

Table 3.1: Bioindicator species of imposex used in studies published from 2003 to 2013 in Europe……………………………………………………………………………………………………………………………………….114

Table 3.2: Bioindicator species of imposex used in studies published from 2003 to 2013 in America……………………………………………………………………………………………………………………………………..115

Table 3.3: Bioindicator species of imposex used in studies published from 2003 to 2013 in Asia……………………………………………………………………………………………………………………………………………117

Table 3.4: Bioindicator species of imposex used in studies on the topic under analysis in Africa from 2003 to 2013…………………………………………………………………………………………………………………….118

Table 3.5: Bioindicator species of imposex used in studies published from 2003 to 2013 in Oceania……………………………………………………………………………………………………………………………………..119

Table 3.6: Number different species (Ns) used as bioindicators of imposex levels for TBT pollution monitoring between 2003 and 2013 in Europe, indicated per country, Order and Family………………………………………………………………………………………………………………………………………..122

Table 3.7: Number different species (Ns) used as bioindicators of imposex levels for TBT pollution monitoring between 2003 and 2013 in America, indicated per country, respective Orders and Families……………………………………………………………………………………………………………………………………..123

Table 3.8: Number of different species (Ns) of bioindicators of imposex levels used in TBT pollution biomonitoring studies published between 2003 and 2013 in Asia, indicated per country, respective Orders and Families…………………………………………………………………………………………………124

Table 3.9: Number different species (Ns) of bioindicators of imposex levels for TBT pollution biomonitoring between 2003 and 2013 in Africa, indicated per country, respective Orders and Families……………………………………………………………………………………………………………………………………..125

Table 3.10: Number different species (Ns) of bioindicators of imposex levels used in TBT pollution biomonitoring studies published between 2003 and 2013 in Oceania, indicated per country, respective Orders and Families………………………………………………………………………………………………….125

Table 3.11: Evolution of the imposex levels in gastropod females reported in Europe from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of several imposex indices (%I, VDSI, FPL, RPLI, RPSI and %S) indicated by colored arrows (↓, decreasing trend; ↔, levels maintenance; and ↑, increasing trend) are presented by the sampling year ascending order……………………………………………………………………………………………127

Table 3.12: Evolution of the imposex levels in gastropod females reported in North America from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of several imposex indices (%I, VDSI, FPL, RPLI, RPSI and %S) indicated by colored arrows (↓, decreasing trend; ↔, levels maintenance; and ↑, increasing trend) are presented by the sampling year ascending order………………………………………………………………………128

Table 3.13: Evolution of the imposex levels in gastropod females reported in South America from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of several imposex indices (%I, VDSI, FPL, RPLI, RPSI and %S) indicated by colored arrows (↓, decreasing trend; ↔, levels maintenance; and ↑, increasing trend) are presented by the sampling year ascending order………………………………………………………………………………………………………………………………………….129

Table 3.14: Evolution of the imposex levels in gastropod females reported in Asia from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of several imposex indices (%I, VDSI, FPL, RPLI, RPSI and %S) indicated by colored arrows (↓, decreasing trend; ↔, levels maintenance; and ↑, increasing trend) are presented by the sampling year ascending order……………………………………………………………………………………………130

Table 3.15: Evolution of the imposex levels in gastropod females reported in Africa from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of several imposex indices (%I, VDSI, FPL, RPLI, RPSI and %S) indicated by colored arrows (↓, decreasing trend; ↔, levels maintenance; and ↑, increasing trend) are presented by the sampling year ascending order……………………………………………………………………………………………130

16 colored arrows (↓, decreasing trend; ↔, levels maintenance; and ↑, increasing trend) are presented by the sampling year ascending order………………………………………………………………………131

Table 3.17: Evolution of OTs concentrations in gastropod females reported in Europe from 2003 to 2013. The sampling year, the bioindicator used, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend), are presented by the sampling year ascending order. The Butyltin Degradation Index (calculated by MBT+DBT/TBT) is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)………………………………..133

Table 3.18: Evolution of the OTs concentrations in sediments, reported in Europe from 2003 to 2013. The sampling year, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)………134

Table 3.19: Evolution of OTs concentrations in water samples in Europe between 2003 and 2013. The sampling year, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)………134

Table 3.20: Evolution of OTs concentrations in gastropod females tissues reported in North America from 2003 to 2013. The sampling year, the bioindicator used, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend), are presented by the sampling year ascending order………………………………………………………………………………………………….135

Table 3.21: Evolution of the OTs concentrations in biota reported in South America from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)………………………………………………………………………………………………………136

Table 3.22: Evolution of the OTs concentrations in sediments reported in South America from 2003 to 2013. The sampling year, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)………………………………………………………………………………………………………………………………137

Table 3.23: Evolution of the OTs concentrations in biota reported in Asia from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The

BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)………………………………………………………………………………………………………………………………138

Table 3.24: Evolution of the OTs concentrations in sediments reported in Asia from 2003 to 2013. The sampling year, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)……….139

Table 3.25: Evolution of OTs concentrations in water samples in Asia between 2003 and 2013. The sampling year, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)……….140

Table 3.26: Evolution of the OTs concentrations in biota reported in Africa from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of TBT, DBT and MBT concentrations (indicated by colored arrows: ↓, decreasing trend; ↔, maintenance; and ↑, increasing trend) are presented by the sampling year ascending order. The BDI is also indicated by colored arrows meaning: ↓, BDI>1 (old input of TBT) and ↑, BDI<1 (recent input of TBT)………………………………………………………………………………………………………………………………141

List of Figures

Figure 1.1: Tributyltin (TBT) molecule, characterized by a central atom of tin (Sn⁺⁴) covalently bound to three organic substitutes (butyl groups, n-C4H9). Adapted from Galante-Oliveira (2010). ……………………………………………………………………………………………………………………………………………………26

Figure 1.2: Bioincrustation phenomenon: settlement and growth of a wide variety of marine organisms perennial or temporary in structures immersed in water such as: buoys (A and B), ship hulls (C and E), and oil platforms (D). Pictures A, B, C and D adapted from http://m.gizmodo.uol.com.br/e-assim-que-voce-vive-fundo-oceano. Picture E adapted from Galante-Oliveira (2010). ……………………………………………………………………………………………………………35

Figure 1.3: Biocide dispersal mode: “contact leaching” paints, adapted from Bennet (1996). ……………………………………………………………………………………………………………………………………………………37

Figure 1.4: Mode of action of the ablative paints in soluble matrix with the biocide being released by diffusion, adapted from Galante-Oliveira (2010). ……………………………………………………………………………………………………………………………………………………38

Figure 1.5: Mode action of self polishing copolymer paints with TBT as biocide (i.e. copolimerised with the resin), adapted from Bennet (1996). ……………………………………………………………………………………………………………………………………………………39

Figure 1.6: Distribution of organotin compounds in the aquatic environment from different pathways. Extracted from Hoch (2001). ……………………………………………………………………………………………………………………………………………………40

Figure 1.7: Pacific oyster Crassostrea gigas. (A) normal oyster shell and (B) oyster shell showing intense “chambering” as a result of TBT exposure. Extracted from Sousa (2009). ……………………………………………………………………………………………………………………………………………………45

Figure 1.8: Nucella lapillus specimen exhibiting imposex; (P) penis, (Vd) vas deferens, (Vb) blockage of vaginal opening and (ne) neoplasmic like growth . Extracted from Hagger et al. (2005). ……………………………………………………………………………………………………………………………………………………47

Figure 1.9: TBT interferes in the steroid biosynthesis inducing imposex by inhibiting cytochrome P- 450 dependent aromatase. Extracted from Galante-Oliveira (2010) (Adapted from Bettin et al., 1996). …………………………………………………………………………………………………………………………………………50

Figure 1.10: Model for imposex induction based both on changes in peptide hormones and steroid hormones. According to this model, TBT acts in the nervous system, while steroids act on the positive feedback loop which maintains the Accessory Sex Organs (ASO). (+) imposex induction; (-) imposex inhibition. Model adapted from Oberdörster and McClellan-Green (2002). …………………………………………………………………………………………………………………………………………………….52

Figure 1.11: Speculative mechanism of imposex induction from an interaction between organotins (TBT and TPT) and RXR in the limited tissue area behind the right tentacle (i.e., the penis-forming area) in female gastropods. Adapted from Horiguchi et al. (2007). ……………………………………………………………………………………………………………………………………………………54

Figure 1.12: Model proposed for the mechanism for imposex induction by TBT. TBT: Tributyltin; AXAT: Acyl CoA-acyltransferase; RXR: Retinoic X Receptor; RXRRE: RXR Response Elements. Model dapted from Sternberg et al. (2009). ……………………………………………………………………………………………………………………………………………………55

Figure 1.13: Nucella lapillus VDS csore scheme. Abbreviations: a, anus; b, blister; gp, genital papilla; n, nodule; p, penis; v, vulva; vd, vas deferens; A, either blisterslike protuberances around the papilla; B, nodules of hyperplasic tissue. Adapted from Gibbs et al. (1987). ……………………………………………………………………………………………………………………………………………………62

Figure 1.14: Geographical distribution of Nucella lapillus in the world. Extracted from Tyler- Walters (2008).……………………………………………………………………………………………………………………………63

Figure 3.1: Number of publications on TBT pollution environmental monitoring using imposex as a biomarker from 2003 to 2013……………………………………………………………………………………………………108

Figure 3.2: Percentage of articles on the topic “TBT environmental pollution monitoring” by continent, during the period from 2003 to 2013……………………………………………………………………….109

Figure 3.3: Pie charts of the percentage of scientific publications on TBT pollution environmental monitoring using imposex as biomarker, published between 2003 and 2013, by country for each continent: Europe, America, Asia, Africa and Oceania……………………………………………………………….112

Figure 3.4: Bioindicators of imposex used worldwide for TBT pollution monitoring between 2003 and 2013……………………………………………………………………………………………………………………………………113

Figure 3.5: Number of species used as bioindicators of imposex, in studies published in Europe from 2003 to 2013, by Family…………………………………………………………………………………………………….119

Figure 3.6: Number of species used as bioindicators of imposex, in studies published in America from 2003 to 2013, by Family……………………………………………………………………………………………………120

Figure 3.7: Number of species used as bioindicators of imposex, in studies published in Asia from 2003 to 2013, by Family…………………………………………………………………………………………………………….120

Figure 3.8: Number of species used as bioindicators of imposex, in studies published in Africa from 2003 to 2013, by Family…………………………………………………………………………………………………….121

Figure 3.9: Number of species used as bioindicators of imposex, in studies published in Oceania from 2003 to 2013, by Family…………………………………………………………………………………………………….121

Chapter 1: General Introduction

Table of Contents

1.1 Organotin (OT) compounds

Organotins (OTs) are organometallic compounds well-known as environmental pollutants. Most OTs are of anthropogenic origin, except methyltins that can also be produced by biomethylation, a process naturally accomplished by bacteria (Guard et al., 1981; Hallas et al., 1982). Their industrial production started in 1940s, when the plastic

22 industry, especially the one involved in the production of PVC, began to expand (Sousa et al., 2013). Several applications of OTs are recognized, from their use as biocides, polyvinyl chloride stabilizers and industrial catalysts for the manufacture of silicone and polyurethane foams (Sousa et al., 2013). Although, their adverse effects on some living organisms and most aquatic ecosystems have led to their global restriction in 2008 (IMO, 2001), the phase out process and its evolution in the last decade (from 2003 to 2013), in terms of pollution monitoring and global status, is the main target of the present thesis.

1.1.1 Characterization and physicochemical properties

Organotin compounds (OTs) are characterized by a central atom of tin (Sn) that is covalently bound to one or more organic chains (methyl, ethyl, butyl, propyl, octyl or phenyl) and to another functional group, such as chloride, oxide, hydroxide, among others (Hoch, 2001). Chemically, they are represented by the general formula RSnX, in which R represents an organic (alkyl or aryl) group, and X an inorganic or organic ligand as, for instance, chloride, fluoride, oxide, hydroxide, carboxylate or thiolate (WHO, 1990; Hoch, 2001; Sekizawa et al., 2003). According to the number of organic groups R, OTs can be classified into four distinct classes: monoorganotins (RSnX₃), diorganotins (R₂SnX₂), triorganotins (R₃SnX) and tetraorganotins (R₄Sn). These compounds chemical and physical properties vary significantly, being not only dependent on the number and nature of the R groups, but also on the type of the other functional ligand X. Their solubility in water, for instance, tends to decrease with both the increase in the number and length of the organic chains, but the nature of the ligand can also play an important role (WHO, 1990 and Hoch, 2001). According International Programme on Chemical Safety (IPCS, 1990) the Sn-C bonds are stable in the presence of water or atmospheric oxygen and at temperatures up to 200°C, therefore the thermal decomposition of OTs is not significance. According to the same program the number of Sn-C bonds and the length of the alkyl chains have effects on the chemical and physical properties of OTs. Some of

23 these properties facilitate OTs occurence in the aquatic environment, namely their low solubility in water. Noteworthy that methylated OTs are more soluble (and have lower boiling points) than other OT species (as for example Tributyltin i.e., TBT), due to their smaller organic functional groups (Cui et al., 2011).

The low solubility of the majority of OTs in water depends of the pH, temperature, and ionic strength (Graceli et al., 2013). In aqueous solution, the OT⁺ ion is in equilibrium with the OT-OH and OT-Cl forms. For instance, in the marine environment, under normal conditions of pH and salinity, TBT is mainly in the form TBT-OH and its bioavailability increases when it is this form at pH> 8 (Fent, 1996; Alzieu, 1998). The half-life (T½) of TBT in water varies from days to weeks and depends on various environmental parameters including pH and temperature, but also turbidity and light (Alzieu, 1996; Fent, 1996). The lipophilic character, associated with a high partition coefficient (Kd), also facilitates these compounds adsorption to the particulate matter in suspension (IPCS, 1990; Langston et Pope, 1995; Sekizawa, 1999 and Hoch, 2001). The Kd is defined as the ratio between the concentration of a given OT species, e.g. TBT, in the sediment (µgKg⁻¹) and water (µgl⁻¹) i.e., is a relative measure of the affinity of the OT to the particulate phase. For instance, according to Gadd (2000), approximately 95% of the TBT present in the water column is adsorbed to the material in suspension, including plankton, while the remainder is largely associated to organic matter dissolved and to the organic and inorganic ligands. In fact, these compounds high affinity to organic matter (Langston and Pope, 1995; Minchin et al., 1997; Hoch, 2001), and their long half-life period in sediments (several years; Fent and Hunn, 1991; Sekizawa, 1999) may cause an increase in their persistence in the environment. This is of great relevance for TBT, which persistence in the aquatic environment is dependent on a large number of complex physical-chemical parameters such as its high affinity to particulate matter and slow release into the water column through resuspension (Unger et al., 1988; Page et al., 1996; Ruiz et al., 1998, 2005; Amourox et al., 2000; Axiak et al., 2000). Thus, as OTs exhibit hydrophobic, lipophilic and ionic properties, these properties also contribute to their toxicity for the living organism. Due to their lipophilicity, these compounds easily cross cell membranes and accumulate in the adipose tissue (Borges,

2012). Due to their ionic properties, they can bind to proteins such as glutathione and α- keratins (Kannan and Falandysz, 1997 and Appel, 2004). Furthermore, beyond these compounds bioaccumulation, they also suffer biomagnification along food chains (Tanabe, 1999; Fernadez, 2001; Fent, 2003; Godoi et al., 2003; Ortiz et al., 2005; Strand and Jacobsen, 2005; Murai et al., 2008). The bioaccumulation is the uptake of organic compounds in the food chains from of water or food. Many toxic organic chemicals attain concentrations in biota several orders of magnitude greater than their aqueous concentrations, and therefore, bioaccumulation poses a serious threat to both the biota of surface waters and the humans that feed on these surface-water species (Smith, 1988).

OTs methylation also increases their toxicity. In anoxic sediments, OTs can suffer methylation by anaerobic sulfate-reducing bacteria, leading to the formation of Sn compounds, more toxic and volatile, in the aquatic environment (Hoch, 2001). In general, inorganic Sn is considered to be non-toxic, whereas the trisubstituted forms have maximum toxicological activity (Hoch 2001; Sekizawa et al., 2003), followed by the diorganotins, monoorganotins and finally the tetraorganotins (Ebdon et al., 1998).

According to Omae (2003), Carvalho (2010) and Santelli (2010), OTs are easily degraded by UV light and microorganisms in the aquatic environment. In summary, the factors that can influence OTs degradation into inorganic Sn are: temperature, light intensity, pH, salinity, and the nature and density of microbial communities (Hoch, 2001; Carvalho and Santelli 2010). These compounds degradation was proved to be accelerated at higher temperature, higher salinity and under light exposure, according to Batley (1996), Fent (1996), Meador (2000) and Burton et al. (2006). Although of the biological degradation mechanism (from microorganisms) yet not be understood is considered the most important factor in the process of degradation of organotin compounds (Gadd, 2000).

As above mentioned, OTs trisubstituted forms have the maximum toxicological activity (Hoch, 2001; Sekizawa et al., 2003). The most toxic OTs are the triorganotins TBT (see Fig. 1.1) and triphenyltin (TPT).

Figure 1.1: Tributyltin (TBT) molecule, characterized by a central atom of tin (Sn⁺⁴) covalently bound to three organic substitutes (butyl groups, n-C4H9). Adapted from Galante-Oliveira (2010).

The TBT and TPT compounds have biocidal properties (Bennett, 1996). According Bennett in 1996 these trisubstituted compounds were applied in the preservation of wood, textiles and paper as disinfectants, fungicides, insecticides and antibiotics, preventing degradation. The TBT and TPT compounds were widely used as active ingredients in antifouling (AF) paints to prevent the biofouling on the surfaces immersed, mainly on the hulls of ships, and, in this way, constituted TBT compounds’ main source to the aquatic environment until to date (de Mora, 1996; Yebra et al., 2004). The TBT was more used as biocide in AF paints that the TPT, the major application of TPT was as fungicide in agriculture (Fent, 1996). Is reported by the authors Omae (2003), Carvalho (2010) and Santelli (2010) that TBT and to a lesser extent TPT, leaches directly from the antifouling paints into the water column where it is degraded into less toxic compounds by light (UV-irradiation) or microorganisms through a series of progressive removal according to the following scheme: R₃Sn⁺→ R₂SnX²⁺→ RSnX³⁺→Sn(IV). This stepwise dealkylation or dearylation ultimately forms a non-toxic element (inorganic tin).

TBT degrades to dibutyltin (DBT) and monobutyltin (MBT), and TPT degrades to diphenyltin (DPT) and monophenyltin (MPT), by a stepwise dealkylation or dearylation that ultimately forms a non-toxic element (inorganic Sn). Thus, TBT and TPT degradation occurs by a progressive loss of the butyl and phenyl groups attached to the tin atom

(Omae, 2003). The final degradation product of TBT and TPT is inorganic tin (Sn⁴⁺), which is stable and considered non-toxic. Because of their organic chains, TBT and TPT are hydrophobic, a feature that directly depends on the alkyl or aryl bond of the tin atom (Graceli et al., 2013). These compounds have low solubility in water due to three organic groups attached to the tin atom. The increase of TBT and TPT toxicity is influenced by the number and nature of the organic groups attached to the tin atom and may be related to this insolubility in water because their hydrophobicity is the main chemical characteristic responsible for their bioaccumulatin (Fent, 2003). According to several authors (Omae, 2003; Carvalho, 2010; Santelli, 2010 and Sousa et al., 2013), TBT and TPT are easily degraded by UV light and microorganisms in the aquatic environment.

Despite of the fact that TBT can be degraded into less toxic forms: their derivatives (DBT and MBT) and inorganic Sn in the water column under aerobic conditions, its low aqueous solubility and strong affinity to particulate matter and suspended particles favors its deposition in sediments (Sousa et al., 2013). The degradation processes of TBT are very slow in sediments, under anaerobic conditions, with reported time of half life (T½) of several years (see Fent, 2006, and references therein) and the T½ can vary between 1.9 and 3.8 years in the deep sediments (Batley, 1996). Therefore, sediments are considered TBT ultimate sink (Hoch, 2001 and Omae, 2006) from which it may become bioavailable again through resuspension and diffusion into the water column (Hoch, 2001; Díez et al., 2002 and Burton et al., 2005). Due to the application of TBT for decades, its release into the aquatic environment reached an unprecedented scale for a manmade chemical.

1.1.2 Production and applications

The existence of organotins (OTs) is known since 1853 (Graceli et al., 2013). Excluding methyltins, which can also be produced naturally by bacteria (Guard et al., 1981 and

Hallas et al., 1982), all organic forms of tin are of anthropogenic origin (Sousa et al., 2013). First OTs were synthesized in the mid-nineteen century by Lowing, in 1852, and Frakland in 1853 (Omae, 2003; de Carvalho and Santelli, 2010), this later using the reaction of iodide of ethylzinc and metallic tin, producing diiodide of diethyltin (Rochow, 1966 and Nicholson, 1989). Afterwards, in 1859, Buckton obtained tetraethyltin by reacting diethylzinc and tin tetrachloride (Davies, 1997). Tetraethyltin was latter on obtained when Letts and Collie, in 1886, did react ethyl iodide with a tin sprayed metallic zinc alloy (Dias, 2005). In 1900, Grignard published the synthesis of halides, i.e., organomagnesium compounds, in solution, getting monoorganotins, diorganotins and triorganotins. These compounds were more sensitive to air exposure than the ones synthesized by Frankland (Pope and Peachey, 1903). Popp and Peachey (1903), prepared and characterized a large number of simple compounds, by reacting tetralkitin and triphenyltin with Grignard reagents, route that has become standard to obtain of aryl- and alkyltin compounds. Krause and Von Grosse’s (1937), in “Organometallische Chemie”, described examples of the synthesis of tetralkyltin and tetraryltin compounds, and organotin di-halideorganotin, hydrides, carboxylates, hydroxides, oxides, alkoxides, phenoxides, compounds RnSn(II), ditin (R₃SnSnR₃) e oligotin (R₂Sn).

The industrial production of OTs started during the 1940s, when the plastic industry, especially the one involved in the production of polyvinyl chloride (PVC) began to expand (Sousa et al., 2013). The increased production and use of PVC brought consequences for ecosystems because its disposal led to an accumulation of mono- and dialkylated OT derivatives in the environment and to possible long-term effects on man and biota (Quevauviler et al., 1991). Even so, many applications of OTs were then discovered, reason why they belong to the most widely used organometallic compounds family, with an estimated annual production of about 60,000 tons (Mala, 2008) until 2008.

The first practical application of OTs was reported in 1925, with the registration of a patent, in which they were indicated as “anti moths agents”, although they have never

28 been used for this purpose (Luijten, 1972; Godoi et al., 2003). Nonetheless, they later came into extensive use in several industrial sectors with multiple applications as revised by Sousa et al. (2013) (see Table 1.1).

Applications General formula References Stabilizers in PVC films RSnX₃ WHO (1990) Glass treatment processes ATSDR (2005) Stabilizers in plastics industry as PVC and CPVC R₂SnX₂ WHO (1990) (Stabilization against decomposition by heat and light) Rodolfo et al. (2002) Scintillation detectors for c- and X-rays Davies (1997) Catalysts in the production of polyurethane foams Hoch (2001) and in room temperature vulcanization of silicones Glass treatment processes as precursors for SnO₂ film de Carvalho and Santelli (2010) Anti-inflammatory and cancer treatments drugs Valla and Bakola- Christianopoulou (2007) Poultry farming (Dewormer) Antizar-Ladislao (2008) Water-proofing agents for cellulosic materials (e.g., cotton textiles, paper and wood) Flame retardants for wool fabrics Binder in water-based varnishes Rodent-repellents R₃SnX OMS (1980) Biocides in antifouling paint formulation WHO (1990) Fishing nets Bright & Ellis (1990) Drinking water treatment Sadiki et al. (1996); Sadiki & Williams (1996) Agricultural pesticides and fungicides Luijten (1972); Poller (1970) and Hoch (2001) Surface disinfectants (RPA, 2005) Insecticides and fungicides in wood preservation Acaricides in vineyards Miticides in citrus fruits Pesticides for ornamental plants Biocides in construction materials and household items Biocides in: allergic pillows, insoles for shoes, cycling shorts padding, sprays for athlete’s foot treatment Insecticides and antifeedants in textiles Disinfectants and biocides for cooling systems in power stations, pulp and paper mills, textile mills, breweries, tanneries Laundry sanitizers Sousa et al. (2013)

Continued of table 1.1 Ballistic additives for solidrocket engine fuels R₄Sn WHO (1990)

Intermediates in the preparation of other organotin de Carvalho and compounds Santelli (2010) Oil stabilizers

OTs were applied as additives in PVC manufacturing process (Hoch, 2001), with the mono-and disubstituted being used as PVC and chlorinated PVC (CPVC) stabilizers (WHO, 1990; Rodolfo et al., 2002; Sousa et al., 2013); as catalysts in the production of polyurethane foams and silicones (Hoch, 2001); as pesticides and fungicides in agriculture (Poller, 1970 and Luijten, 1972); as biocides in construction materials and household items (RPA, 2005); as surface disinfectants (RPA, 2005); laundry sanitizers (Sousa et al., 2013); rodent-repellents (OMS, 1980; Sousa et al., 2013); in scintillation detectors for C- and X-rays (Davies, 1997); as ballistic additives for solid-rocket engine fuels (WHO, 1990); as ionophores in liquid membrane ion-selective electrodes and pharmaceuticals, e.g., anti-inflammatory and cancer treatments drugs (WHO, 1990; Allsopp et al., 2001; Valla and Bakola-Christianopoulou, 2007 and Mala, 2008); some were used as constituent of the material for food packaging and preservation of textiles, wood and paper (Dias, 2005). Trisubstituted OTs were used as biocides to remove particulate matter on drinking water sources (Sadiki et al., 1996; Sadiki and Williams, 1996) and in the antifouling (AF) paints (WHO, 1990; Sekizawa et al., 2003). In fact, the outstanding biocidal properties of trialkyltins (particularly tributyltin, TBT, and triphenyltin, TPT) were discovered only after the 1950s, and since then these compounds were broadly used as active ingredient in antifouling paints (WHO 1990; Sekizawa et al., 2003) to prevent the biofouling process on water submerged stuctures.

1.1.2.1 Stabilizers of PVC and CPVC

In 1940s, the plastic industry, mainly the involved in the production of polyvinyl chloride (PVC), used organotin compounds (OTs) as additives in the manufacturing

30 process to prevent discoloration and embrittlement under the influence of light and heat (Hoch, 2001). Mono-and disubstituted OTs were used as PVCand polyvinyl chloride (CPVC) stabilizers (Hoch, 2001; Dias, 2005 and Sousa et al., 2013) as PVC and its copolymers undergo modifications during degradation reactions. These alterations imply the loss of their properties, both during processing and their final use, and occur namely by heat, oxidizing agents, andultraviolet or infrared radiation (Hoch, 2001; Dias, 2005; Sousa et al., 2013). PVC processing occurs at temperatures between 180-200°C but it is also within these temperatures that its decomposition occurs (Hoch, 2001). This processing follows in the presence of oxygen, making it unstable, being necessary the use of additives as thermal, ultraviolet and antioxidant stabilizers (Rodolfo et al., 2002). PVC can also be degraded by the prolonged exposure to light due to the loss of HCl from the polymer (Hoch, 2001). Accordingly, Hoch (2001) reports that, the results from such exposure, are the product discolouration and embrittlement. To prevent the degradation of PVC by exposure to light, certain OTs (mainly mono- and dialkylated derivatives) are added to at a levels of 5-20 gKg⁻¹ of PVC (Blunden and Chapman, 1986; Lawson, 1986). OTs are present in PVC that is used in applications that vain from tubes and rigid profiles for use in the construction up toys and flexible laminates for packaging of blood and plasma (Rodolfo and Mei, 2007). The PVC is the second thermoplastic most consumed in all the world, with a global demand of resin upper at 33 million of tonnes in 2006, being the global capacity of production of PVC resin estimated at around 36 million tons / year (Rodolfo and Mei, 2007). According to Hoch (2001), OTs-stabilized PVC has numerous applications including in window frames, coating materials, packaging materials, foils, piping of potable water, wastewater and drainage water. Noteworthy the application of dioctyl(maleate)tin used as additive of PVC for packagings and shrink films, with the advantage of allowing the direct contact with food (Rodolfo Jr. et al., 2002) and pharmaceuticals (Rodolfo and Mei, 2007). Regarding CPVC, a variant of PVC, it is widely used for high-temperature water distribution systems (Hoch, 2001).

Trisubstituted OTs (e.g. TBT and TPT) are rarely applied since they are less effective as PVC stabilizers and have greater toxicity (Luijten, 1972; Rodolfo Jr. et al., 2002).

1.1.2.2 Biocides in Agriculture

The use in agriculture is the second largest application of organotins (Poller, 1970 and Luijten, 1971). The first compound used in this field was triphenyltin (TPT) acetate, and it was observed that triaryltin derivatives are less phytotoxic than the trialkyltin ones

(Poller, 1970 and Luijten, 1971). In general, triorganotins of the type R3SnX have high fungicide power and may cause complete growth inhibition of the fungus Botrytis allii, Penicilium italicum, Aspergillus niger and Rhizopus nigricans (Van Der Kerk and Luijten, 1956). TPT, for example, were used in the control of Phytophthora infestan in potatoes, and of Cercospora beticola in beets (Poller, 1970 and Luijten, 1972). From 1960, the products derived from TPT acetate entered the market under the trademark Brestan through the German chemical industry Hoescht (currently designated Aventis, S.A.), while TPT hydroxide was sold under the brandname Du-ter, produced by Philips Dufan. The later was extensively applied in agriculture for the control of pathogenic fungi on potatoes, sugar beets, celery, carrots, anions, and rice, and was also used to prevent tropical plant diseases in peanuts, pecan, coffee, and cocoa (Champ, 1996; Seligman, 1996 and Dias, 2005). In 1967, Dow Chemical produced and marketed the miticide Plictran, based on tricycloexyltin hydroxide, for the control of mites on apples, pears, and citrus fruits (Hoch, 2001). More recently, the German chemical company Bayer AG marketed the acaricide Peropal, based on tricycloexyltin-1, 2, 4-triazole (Crowe, 1984 and Sanyagina, 1993), still in use (Dias, 2005; Lima et al., 2005). The great concern on the application of OTs as pesticides in agriculture arises from the well-known significant portion of these chemicals in the environment, already proved to be derived from their direct input into the soil, water, and air, by spraying, leaching, and run off (Hoch, 2001).

1.1.2.3 Wood preservation

OTs are also used in wood preservation. They were applied as fungicides in the timber treatment to control insects, fungi and bacteria that break the cellulose and other complex substances that constitute the wood (Hoch, 2001). The first investigations into wood preservation based on OTs as ingredients began during the late 1950s by the Osmose Inc. (New York, USA). In 1960, this company marketed the wood preservative Oz, containing tributyltin (TBT) named, in analogy to the “Tin Man” in “The Wizard of Oz” (Bennett, 1996). Trisubstituted OTs as TBT oxide, TBT naphthenate and TBT phosphate, were extensively used as fungicides in the wood preservation, applied as a 1-3 wt % solution in an organic solvent (Hoch, 2001). These OTs were applied by one of several methods such as spraying, dipping, brushing and double vacuum impregnation (in specially designed impregnation chambers). This latter method is described by Hoch (2001) as the most effective treatment and was the more often used in the timber industry. According to this method, after solvent evaporation, the biocide remained safely within the wood structure, due to their low vapor pressures (Hoch, 2001). According to Hoch (2001) the impregnation chambers are closed systems and they not constitute a significant source of environmental pollution. There are care timber treatment facilities, but can occur liberation of TBT into the freshwater environment due to seepage, accidential spills and effluents (Hoch, 2001). Due to the reported high toxicity of TBT to non-target species, these compounds were progressively substituted by others (e.g. synthetic pyrethroids; Hoch, 2001).

1.1.2.4 Antitumor drugs in medicine

Many tin compounds, especially organometallics, exhibit biological action. The first information about OT compounds, precisely the diorganotin compounds, with antitumor activities was published in 1980 (Crowe, 1984). In the period between 1980 and 1990, more than 2000 diorganotins were tested for verification of their effects on biological

33 systems (Crowe, 1984). It was confirmed that this activity is related with the stability of the diorganotin compounds and with the length of the Sn-N bond, it was observed that, for compounds with bond length Sn-N less than 239 ηm, the biological effects are related with the dissociation of the ligand nitrogen as part of the mechanism of action (Crowe, 1984). More recently, studies on the effects of trialkytins in Ca²⁺ in the mobilization of the cells PC3 in human prostate cancer have been explored (Jan et al., 2002). It was found that the tin compounds showed effective against the lymphocytic leukemia, being the diorganotin derivatives more active than the triorganotin derivatives. It was also observed a greater biological activity related to the presence of aromatic nitrogen and phenyl- or ethyl- groups bonded to tin (Terra, 1997).

There are countless organometallic compounds that feature antitumor activity, in human cancer cells. One example is the cis-platine complex [Pt(NH₃)₂Cl₂] (Pellerito, 2002). Currently, the interest associate them with biologically active organic compounds, especially in oncology, with the goal of evaluate possible synergism effects. Thus, two forms can be recognized in this field. Firstly, the preparation of new compounds, associating the two classes mentioned above and its chemical study under various aspects. In second place is the biological assay and clinical, well as the planning of news substances from of objectives of pharmacological action (Filgueiras, 1998).

1.1.2.5 Biocides in antifouling (AF) paints

Antifouling (AF) paints are used to prevent (or at least reduce) the biofouling (or bioincrustation) phenomenon (see Fig. 1.2), characterized by the colonization of surfaces submerged in water by bacteria, diatoms, algae, hydroids, barnacles, mussels, and other marine invertebrates.

This phenomenon implies very high costs to the shipping industry since it accelerates the ships surfaces corrosion (Hoch, 2001 and Sousa, 2013) but also because it increases ship hulls’ roughness, increasing the frictional force between submerged surfaces and water with the consequent decreased speed and increased fuel consumption. It was estimated that during only six months at the sea without an AF system, a ship may accumulate about 150 kg m⁻² of fouling (i.e. an weight gain of about 6,000 tons for a large vessel with an immersed surface of 40.000 m², e.g. an Oil-tanker; IMO, 1999), which increases fuel consumption by about 40% to maintain the same travel speed (IMO, 1999). Thus, this justifies the need for the application of an AF system that prevent the establishment and growth of marine organisms on submerged surfaces. Furthermore, according to Omae (2003) and Abel (2000), the use of AF paints is of utmost importance since they not only reduce the fuel consumption, which in turn reduces the emissions of carbon dioxide and of sulfur to the atmosphere, but also because it reduces the risk of introduction of new species (exotic species) in ecosystems.

Were initially used as antifouling, products as quicklime, brimstone, rosin, pitch, tar and bitumen to protect ships hulls (IMO, 1999; Yebra et al., 2004). Some of these antifouling continued to be used but over an extra wood sheath applied to the hull surface (Galante-Oliveira, 2010), allowing easier repairs, by the simple replacement of that last layer, in dry dock and at regular intervals.

Were subsequently tested the materials: iron, zinc, lead and copper, but the later performed very well in protecting the hull from biofouling once, these materials in contact with water produces a kind of a “poisonous film” mainly of copper oxychloride (Kegley et al., 2008). This film is slightly soluble and it is therefore gradually washed away, leaving no way in which marine organism can attach itself to the vessel submerged surface (Galante-Oliveira, 2010).

Until the 60’s the copper continued to be used, but in the form oxide, in the protection the hull from biofouling. From this time and until recent times the control of biofouling has been assumed by organotin compounds. The tin organic derivatives have great biocidal capacity and this was recognited from a systematic study conducted in the 50’s by the International Council of Survey Paints (Institute of Organic Chemistry in Utrecht; van der Kerk and Luitjen, 1954 in Hoch, 2001). Thus, this study triggered these compounds application in the control of the biofouling phenomenon, particularly the TBT (Crompton, 1997), widely popular since the 60s until their global legal restriction in 2008 (Hoch, 2001; Godoi et al., 2003; Omae, 2003a; IMO, 2009). The commercialization of AF paints started in the 70’s. In this decade, due to durability and biocidal efficiency of the TBT, most of the maritime shipping vessels had their hulls coated with TBT-based AF paints (Yebra et al., 2004).

The use of AF paints was of extreme effectiveness in the reduction of the bioencrustation due to TPT and TBT compounds, particularly the TBT that was the most widely used biocide in the world (Bennet, 1996). TBT was considered the most potent and toxic biocide worldwide, due to its effectiveness and long duration, which prevented

36 costly dry dock operations and thought that TBT is easily degraded by UV light and microorganisms in the aquatic environment (IMO 1999; Omae 2003; Sousa et al., 2013). The antifouling paints consist of an array that contains ingredients biocides and pigments. The mode of action these paints is characterized by the release of small quantities of biocide to the water, forming a layer which repels fouling organisms. According to the authors Bennet (1996) and Hoch (2001) there are different types of paints that are classified depending on their mode of action.

According to Mora (1996) the first TBT-based AF systems were known as “free association” paints in which a high amount of biocide was dispersed in a matrix that could be insoluble. There are two distinct subtypes depending on the type of wherein the matrix is TBT (Bennet, 1996): A) paints of leaching from contact and B) ablative paints in soluble matrix.

A) Paints of leaching from contact

The initial rate of release of the biocide was high and controlled through microchannels from the matrix. The toxic components leached exponentially from the paint with time (Bennet, 1996) (Figure 1.3). According Hoch (2001) one reason for the diminishing of the leaching rate was that the microchannels in the paint surface might be clogged up and this inhibits the further release of the toxic component, the biocide. The paint dyes had maximum durability and two years vessels could not be repainted without removing the previous layer ink implying an additional source of pollution from construction sites (Mora, 1996).

Figure 1.3: Biocide dispersal mode: “contact leaching” paints, adapted from Bennet (1996).

B) Ablative paints in soluble matrix

The ablative paints incorporated the biocide in a weakly soluble matrix from which it was released by diffusion (Figure 1.4). Periodically the surface layer (depleted of the biocide) was separated from the surface allowing a new layer in contact with water and restarting the biocidal action. These paints have an erratic effectiveness and durability of several years (Bennet, 1996).

Figure 1.4: Mode of action of the ablative paints in soluble matrix with the biocide being released by diffusion, adapted from Galante-Oliveira (2010).

B.1) Self Polishing Copolymer:

The self polishing copolymer was introduced in 1974 and was the most commonly used and most effective type of AF system in the 1980s (Hoch, 2001). Surfaces, immersed in water, treated with TBT-based copolymer paints, reached a constant TBT leaching rate of 1.6 µg (Sn) cm⁻² per day (Hoch, 2001). The paint dye was insoluble in water and did not

38 penetrate the matrix in which TBT was copolymerised, delaying its release to the aquatic environment (Figure 1.5). The TBT release from the paint surface was caused by a chemical reaction with sea water by hydrolysis of the bond between the resin and TBT so that TBTO (tributyltin oxide) was released slowly at a speed controlled during the paint lifetime (Sousa, 2004). The hydrolysis promoted the erosion (or polishing) of the surface releasing the biocide and increasing the performance of the paint (Bennet, 1996). These paints reached a maximum durability of 60 months (IMO, 1999), not requiring removal just repainting. Since its appearance, in 1978, it occupied a prominent place in the shipbuilding industry that, in 1991, about 80% of the 4000 t tonnage boats used this AF tecnology with TBT as biocide (IMO, 1999).

Figure 1.5: Mode action of self polishing copolymer paints with TBT as biocide (i.e. copolimerised with the resin), adapted from Bennet (1996).

In summary, the most important application of OTs, in quantitative terms, was undoubtedly as PVC stabilizers (mainly dibutyltins, dioctyltins and monosubstituted organotins): about 80% of the annually produced OTs were applied in the PVC industry (Hoch, 2001) and 20% as biocides and pesticides (EVISA, 2009). However, the use of TBT as a potent biocide in AF paints gained special relevance due to its extreme toxicity towards non-target species and the consequent adverse effects in aquatic ecosystems reported all around the world.

1.2 Adverse effects on living organisms

Organotin compounds (OTs) have been classified by several authors as endocrine disruptors, immunotoxicants, carcinogens and obesogens (Bettin et al., 1996; Grün et al., 2006; Carfi’ et al., 2008; Newbold et al., 2009; Penza et al., 2011; Tonk et al., 2011; Ravanan et al., 2011). Their proved high toxicity in combination with their widespread use made them a matter of global concern. The direct input of some of these compounds in the aquatic ecosystems by different routes and their high persistence cause damaging effects on different trophic levels (see Figure 1.6).

Figure 1.6: Distribution of organotin compounds in the aquatic environment from different pathways. Extracted from Hoch (2001).

The possible routes or sources of TBT pollution to the aquatic environment are: boating activity e.g. commercial harbours (deep sea ports), marinas, ferry terminals, shipyards and/or dry dock facilities, naval facilities, moorings, and fishing harbours (artisanal fishing); shipping activity e.g. shipping lanes; and agriculture (Titley-O’Neal et al., 2011). The TBT is persistent in the aquatic environment due to their lipophile character. The TBT is bioaccumulated in the aquatic invertebrates, mainly in the molluscs (bivalves and

40 gastropods) and crustaceans (dacapods) (Hoch, 2001), being these taxonomic groups important seafood resources and are ecologically dominant in many habitats. Crustaceans and fish bioaccumulated much lower amounts of TBT because according to Laughlin (1996) their possession of efficient enzymatic machanisms that degrade TBT in the body. The higher trophic aquatic organisms also bioaccumulate TBT, this bioaccumulation proceeds through either uptake from solution alone or of a combination with diet ingestion. Some studies showed that marine mammals and birds also bioaccumulate high levels of butyltins in various tissues and organs (Iwata et al., 1995; Kannan et al., 1998 a, b). In regions with high shipping, harbours and shipyards there is a direct emission of TBT from antifouling paints into the water and accumulation this biocide in the sediment (Hoch, 2001), which gives rise to contamination of water and sediment of marinas, lakes and coastal areas. According to Hoch (2001) the TBT and TPT compounds applicated in antifouling paints have accumulated in the sediments over the last decades and may to be remobilized of the sediments. The TBT levels in sediments are related to the organic content because TBT adsorbs to organic material (Jacobsen and Asmund, 2000), but there seems not to be an obvious single factor responsible for the varying concentrations of TBT in the different harbour sediments (Mallon and Manga, 2007). Factors as shipping traffic intensities, distance to shipyard and quays, water renewal, bottom current regime and sedimentation are together responsible for the highly variable concentrations of TBT in sediments (Berge et al., 1997; Green et al., 2001).

TBT is extremely toxic for the representatives of different taxonomic groups, from bacteria to mammals, including humans. The high toxicity of TBT has resulted in numerous and widespread adverse biological effects to a wide range of organisms of the different taxonomic groups (see table 1.2). It is already proved to be a potent agonistic ligand of vertebrate nuclear receptors, retinoid X receptors (RXR) and peroxisome proliferator activated receptor-gamma (PPARy) (Graceli et al., 2013). The physiological consequences of these nuclear receptors activation have been demonstrated namely in experimental models of adipogenesis (Grün et al., 2006). Thus, TBT alters metabolic and

41 lipid homeostasis parameters, induces the differentiation of adipocytes in vitro and increases adipose mass in vivo (Grün et al., 2006; Tontonoz et al., 2008). Rodents exposure to OTs induce these developmental and reproductive toxic effects as well as alteration of metabolic homeostasis through its action as an obesogen (Omura et al., 2001; Ogata et al., 2001; Grün et al., 2006). According to Graceli et al. (2013), the adverse effects occuring in rodents have raised concern about OTs potential health risk to humans. OTs cause endocrine-disrupting effects in mammals, humans (Keithly, 1999) and rodents (Grote et al., 2004; Grote et al., 2006), as a consequence of the consumption of contaminated seafood. According to Dorneles et al. (2008), human exposure to OTs may result from dietary sources, such as seafood but also from consumption of contaminated drinking water. In vitro exposure to triorganotins, especially TBT or TPT, revealed a decrease in DNA and protein synthesis in human choriocarcinoma cell lines (Nakanishi et al., 2002) while the exposure of human cells to TPT inhibited human aromatase (Lo et al., 2003) and other steroidogenic enzymes, also affecting sexual development in rodents (Omura et al., 2001; Ogata et al., 2001; Grote et al., 2004). Therefore, OTs have many complex effects on the endocrine systems (in both genders) that can induce morphological changes in these target organs. Despite of OTs effects have been the focus of numerous investigations, the extension of OTs effects on mammals or invertebrates and/or their abnormal effects on hormonal modulation to their reproductive and metabolic functions is still not thoroughly known.

Table 1.2: Summary of ecotoxicological effects of TBT indicated by taxonomic group. Group Effect Important references Bacteria Toxic for bacteria (gram-positive more sensitive); ICPS, 1990 WHO (1990) Mendo et al., 2003 Inhibits growth and metabolism; Wuertz et al. (1991) Affects respiration; Gadd (2000) Reduces productivity; Mendo et al. (2003) Inhibits solute transport; Martins et al. (2005) Inhibits macromolecules biosynthesis and Cruz et al. (2012) transhydrogenase; Phytoplankton Reduction of marine microalgae growth; Beaumont and Newman (1986) WHO (1990) Petersen and Gustavson

Continued of table 1.2 (1998) Reduces respiration and photosynthetic activity; Bry an and Gibbs (1991) Alters photosynthetic pigment content; Fargasová (1996) Sidharthan et al., 2002 Induces drastic changes in biochemical Fargasová and Kizlink (1996) composition; Reduces primary productivity; Fargasová (1997) Induces changes in the community structure; Nudelman et al. (1998) Petersen and Gustavson (1998) White et al. (1999) Gadd (2000) Sidharthan et al. (2002) EC, 2005 Plants Impairment of macroalgae spores motility; ICPS, 1990 WHO (1990) Decreases growth and transpiration rates; Jensen et al. (2004) Reduction of several marine angiosperms growth; Trapp et al. (2004) Reduces photosynthetic activity; Caratozzolo et al., 2007 Stress induction, by bioaccumulation, in terrestrial Azenha et al., 2008 plants; Lespes et al., 2009 Carvalho et al. (2010) Crustaceans Reproductive performance reduction; ICPS, 1990 WHO (1990) Neonate survival decrease; Kusk and Petersen (1997) Inhibits larvae developmental ratios; Waldock et al., 1999 Juveniles growth rate decrease; Dahllöf et al., 2001 Community structure changes; Takeuchi et al., 2001 EC, 2005 Aono and Takeuchi, 2008

Molluscs Abnormal shell growth; Alzieu et al., 1981 Langston (1990) Inhibits egg development; Bryan and Gibbs (1991) Females virilisation (Imposex / Intersex); Bryan et al., 1986 Hagger et al. (2006) Induces sterility; Bauer et al., 1995 Park et al. (2012) Increased mortality; Page et al., 1996 DNA damage; M atthiessen and Gibbs, 1998 Induces heat-shock proteins; Barroso et al. (2002) Alters sex ratio; Sousa et al. (2005) Teratogenic effects – embryos malformations; Gabbianelli et al., 2006 Reduces survival ship of hatchlings; Leung et al. (2 007) Induces DNA damage; Sousa et al. (2008) Decreases growth ; Horiguchi (20 09) Community structure alterations; Rank, 2009 Fish Growth inhibition; WHO (1990) Induces masculinization; Fent (1996) Sperm abnormalities induction and reduced fecundity; Induce liver vacuolation; McAllister and Kime, 2003 Teratogenic effects – larvae malformations; Shimasaki et al., 2003

Continued of table 1.2 Induces sperm abnormalities; McAllister and Kime (2003) Hyperplasia of the hematopoietic tissue; EC, 2005 Zuo et al. (2012) Disruption of intracellular energy production; Zheng et al., 2005 Inhibits ovarian development; Zheng et al. (2005) Induces embryo abnormalities; Zhang et al. (2007) ATPase and ion-pump activities; Zhang et al., 2008 Neurotoxic through the modulation of the Zhang et al., 2009 glutamate Zuo et al., 2009 signalling pathway; Thymus atrophy and thymocytes apoptosis; Cytochrome P450 system inhibition; Zhang et al. (2013) Induces DNA damage; Induces severe damage in the thyroid gland; Induces lipotoxic; Mammals Induces reproductive anomalies, including reduced WHO (1990) spermatogenesis; Behavioural changes in rats; Foetal gonad morphology alterations by changes in Fent, 1996 gene expression profiles; Ema et al., 1997 Teratogenic effects – embryos malformations; Whalen et al., 2002 Inhibition of basal and calmodulin-dependent Ca²⁺- Grün et al., 2006 ATPase activity in rat brain synaptic membrane Aluoch et al., 2006 preparations; Inhibition of mitochondrial oxidative Ohtaki et al., 2007 phosphorylation or ATP synthesis; Kishta et al., 2007 Inhibition of natural killer cell cytotoxic function; Yonezawa et al., 2007 Induces neurobehavioral alterations; Suppression of osteoclastogenesis trough the Antizar-Ladislao (2008) retinoic acid; Induces immunological disorders receptor (RAR) Antizar-Ladislao (2008) pathway; Chen et al., 2008 Induces cardiovascular toxicity; Suppresses osteoclastogenesis; Mala (2008) Adipose tissue differentiation rise – obesity Yang et al., 2009 induction; Disturbance of the Ca⁺ homeostasis in human Grün and Blumberg, 2009 neutrophils; Human lymphocytes inhibition;

All these adverse effects of TBT have been described throughout the time. Although the first harmfull effect of the trisubstituted OTs on non-target organisms was recorded in 1970s, in oysters from a bay located in the French Atlantic coast with abundant oysters farming facilities, the Arcachon Bay (Alzieu et al., 1980; Alzieu, 1986). The Pacific oyster, Crassostrea gigas, revealed abnormal shell “thickening”, also known as “chambering” or “balling”, a phenomenon of irregular calcification of the shell (see Figure 1.7) by the

44 hypersecretion of interlayer gel, overlap of a calcitic layer and disappearance of the gel, resulting in the formation of chambers that gave the shell a different shap “ball shape”.

Figure 1.7: Pacific oyster Crassostrea gigas. (A) Normal oyster shell and (B) oyster shell showing intense “chambering” as a result of TBT exposure. Extracted from Sousa (2009).

Oyster Crassostrea gigas production in the Arcachon area was severely affected in the 1970’s and early 1980’s due to severe growth problems as shell calcification anomalies that were followed by a complete lack of reproduction (see review by Alzieu 2000). Such problems were associated with the existence of numerous marinas on the bay of Arcachon, where docked vessels were releasing high amounts of TBT from their AF paints into the aquatic environment (Sousa et al., 2013).

Almost simultaneously, another harmful effect of TBT to non-target organisms was also described. In 1970, Blaber recorded the appearance of a penis-like organ in females of the gastropod Nucella lapillus collected in Plymouth Sound, United Kingdom. The severity of appearance of a penis in females of the Nucella lapillus was higher inside harbors and lower further away, but the causative agent was at that time unidentified. One year later, Smith (1971) also noticed the development of similar male sexual characters, such as a penis and vas deferens, onto females of Nassarius obsoletus collected in United States. Smith coined this endocrine disruption syndrome imposex, but at that time the cause was still unknown. Later, the cause-effect relationship between this phenomenon and exposure to TBT was disclosed (Smith, 1981; Bryan et al., 1986; Gibbs and Bryan, 1987).

According to Evans (1999) and references therein, the imposex on gastropods is only one of the negative effects of TBT, others are mortality and hormonal imbalance in dolphins, crabs, lobsters, oysters, invertebrate larvae, seagrasses and algae. Even so, the imposex in female gastropods has been regarded as the most complete example of endocrine disruption caused by an environmental pollutant (Vos et al., 2000).

1.3 The imposex phenomenon

TBT acts as an endocrine disruptor in female gonochoristic gastropods (Matthiessen and Gibbs, 1998), and there are three different types of female masculinization (FM) that are known to occur in female gastropods in response to TBT exposure. The three types of female masculinization in gastropods are: the well-known pseudohermaphroditism (Jenner, 1979; Fioroni et al., 1990) or imposex (Smith, 1971), intersex (Oehlmann et al., 1994) and ovo-testis (Gibbs et al., 1988). The third type of female masculinization, ovo-testis, is inconspicuous and has only been described by histological analysis (Gibbs et al., 1988; Oehlmann et al., 1996; Horiguchi et al., 2000; Horiguchi et al., 2004; Sloan and Gagnon 2004; Horiguchi et al., 2005).

Imposex is a morphological phenomenon defined as the superimposition of male sexual characters, such as vas deferens and/or penis, onto female’s reproductive tract (Figure 1.8) and is the most evident known example of endocrine disruption in invertebrates (Matthiessen and Gibbs, 1998).

Figure 1.8: Nucella lapillus specimen exhibiting imposex; (P) penis, (Vd) vas deferens, (Vb) blockage of vaginal opening and (ne) neoplasmic like growth. Extracted from Hagger et al. (2005).

Imposex in gonochoric prosobranch species is the biological effect in response to TBT exposure more often reported in the literature and is described as being dose-dependent. Thus, the intensity of the phenomenon expression was directly correlated with the concentration of TBT to which the animal was exposed and in extreme cases it can lead to the complete functional sterilization and death of the affected specimens (Schulte- Oehlmann et al., 1997). According to Bryan and his team (1986), imposex can result in total reproduction failure and is believed to account for the extinction of dogwhelks, Nucella lapillus, in severely contaminated areas by TBT. The apparent unique susceptibility of prosobranch gastropods to TBT-induced imposex is, according to Sternberg and colleges (2009), probably a result from the difficulty in identifying this condition in hermaphroditic species.

The imposex phenomenon has been reported in more than 200 gastropod species (Shi et al., 2005). This phenomenon has been used worldwide as the specific biomarker of TBT pollution, namely in Europe; North and South America; Asia; Africa; Australia; and even in such remote locations as the Arctic and Antarctic (see e.g., Gibson and Wilson 2003; Shi et al., 2005; Strand et al., 2006; Negri and Marshall 2009; OSPAR, 2010; Tallmon, 2012; Horiguchi, 2012).

Even so, controversial evidence from museum samples collected prior to the 1960s when there was an increased use of TBT as a biocide in AF paints, and the occurrence of ‘‘natural masculinization’’ in some species demonstrates that imposex does occur without exposure to OTs (Titley-O’Neal et al., 2011). The occurrence of pseudo-imposex, or ‘‘natural imposex’’ and the ability of steroids (Bettin et al., 1996; Oehlmann et al., 1996; Oberdörster and McClellan 2000; Oberdorster and McClellan 2002; Santos et al., 2005) and other contaminants such as Aroclor 1260 (Garaventa et al., 2008) and nonylphenol (Evans et al., 2000) to induce imposex justifies the inability of researchers to confirm the actual mechanism of female masculinization in gastropods to date. Despite of the remaining uncertainties, ecotoxicology still reports imposex as one of the few examples of a specific response of organisms to a particular compound (Garaventa et al., 2006).

1.3.1 Possible mechanisms

Several hypotheses have been proposed to explain the imposex induction in female gastropods and three possible hypotheses have been identified: the neuroendocrine, the steroid and the retinoic. This section shows an overview of each mechanism. But these hypotheses are not yet fully clarified and there continues to be much debate on these different hypotheses.

Steroid pathway

According Nakanishi (2008) some OTs are known as endocrine-disrupting chemicals that modulate steroid hormone biosynthesis. The steroid pathway hypothesis is related to the mechanism of imposex induction in female gastropods regarding the modulation of testosterone. Gooding in 2002 indicated that the conversion rate of free testosterone to fatty acid esters in Ilyanassa obsoleta was decreasing as a result of TBT exposure, which increased free testosterone levels. The idea that TBT was toxic to the esterification of testosterone in I. obsoleta was confirmed by Gooding and his team in 2003 when TBT inhibited I. obsoleta ability to produce or retain testosterone as a fatty acid ester. TBT has

48 the ability to disrupt testosterone esterification, which results in an increase in free testosterone levels, thus explaining why elevated testosterone levels have been reported in a number of species affected by imposex (Gooding and LeBlanc, 2001).

This mechanism was also formulated by other authors, after evidences that imposex affected females have consistently exhibited elevated free testosterone titres when compared with the non affected ones by imposex (Spooner et al., 1991; Bettin et al., 1996; Barroso et al., 2002b; Barroso et al., 2005; Santos et al., 2005). According with the author Castro and his time (2007) two different pathways for explain such disruption can be recognized: (i) interference with steroid biosynthesis or (ii) interference with steroid excretion also known as Inhibition of testosterone excretion.

(i) Interference with steroid biosynthesis:

The biosynthesis of sex steroids from cholesterol requires trafficking between mitochondria and smooth endoplasmic reticulum, and involves many enzymatic steps. According to Nakanishi (2008) most of these steps use cytochrome P450 (CYP) haem- containing enzymes.

Data from a study using Nucella lapillus and Nassarius reticulatus show that TBT interferes in the steroid biosynthesis, suggesting that TBT induces imposex by inhibiting cytochrome P-450 dependent aromatase (CYP-19). Spooner and co-authors (1991) suggested that TBT inhibits the cytochrome P450 dependent of the said aromatase, responsible for the aromatization of androgens (androstenedione and testosterone) to estrogens (estrone and 17 β- estradiol), resulting high levels of testosterone. Later, Bettin et al. (1996) confirmed in N. lapillus and N. reticulatus that aromatase catalyses the aromatization of androgens to estrogens, and the hypothesis suggests that TBT-induced imposex is a result of increased androgen levels and not directly from TBT or OTs themselves (Figure 1.9).

Figure 1.9: TBT interferes in the steroid biosynthesis inducing imposex by inhibiting cytochrome P- 450 dependent aromatase. Extracted from Galante-Oliveira (2010) (Adapted from Bettin et al., 1996).

Was proven theory of the TBT-induced imposex as being the result of aromatase inhibition when a P-450 aromatase inhibitor, at 0.30 mg l⁻¹ induced imposex in N. lapillus (Santos et al., 2005).

(ii) Interference with steroid excretion or Inhibition of testosterone excretion

According the hypothesis interference with steroid excretion, the induction of imposex in female gastropods is due to inhibition of testosterone excretion. This hypothesis also explains the imbalance of the androgen / oestrogen ratio in imposex- affected females. It was raised by Ronis and Mason (1996) considering the possibility of an inhibition of the testosterone excretion. These authors demonstrated that TBT decreased the amount of testosterone sulphur-conjugates in Littorina littorea leading to an increase in the free testosterone levels in tissues. Results from an in vivo experiment revealed that TBT inhibited the sulfur conjugation of testosterone and phase I metabolites, which resulted in an excess of androgens in L. littorea tissues (Titley-O’Neal at al., 2011).

However, very high TBT concentrations were tested, during a time of 42h and no reference was made to the endogenous testosterone levels. The authors, Gooding and LeBlanc (2001) also showed that the metabolism of testosterone to polar sulphate conjugates is negligible in the imposex susceptible mud snail Ilyanassa obsoleta and that these organisms conjugate testosterone primarily to non-polar fatty acid esters (Galante-Oliveira 2010). Gooding and LeBlanc also demonstrated that TBT decreases the esterification of testosterone to fatty acids in Ilyanassa obsoleta leading to an increase in free testosterone, inducing so the imposex phenomenon.

Neuroendocrine pathway

The neuroendocrine hypothesis was proposed by Féral and Le Gall (1982, 1983) using the sting winkle, Ocenebra erinacea. In 1982, Féral and Le Gall conducted a series of in vitro experiments and suggested that TBT induced the penis morphogenetic activity in the pedal ganglia of normal females and that TBT inhibits the release of a neuroendocrine factor (Penis Regression Factor- PRF). Nevertheless those authors did not find any effects of TBT on the formation of Penis Morphogenic Factor (PMF), a molecule expressed in all prosobranch snails’ females and males (Oehlmann et al., 2007). Revitalization of the neuropeptide theory was supported by data from Oberdörster and McClellan-Green (2000, 2002) and Oberdörster et al. (2005) using Ilyanassa obsoleta, and a model for imposex induction based both on changes in peptide hormones and steroid hormones was proposed (see Figure 1.10).

The species I. obsoleta was exposed to a set of neuropeptides (APGWamide, conopressin, LSSFVRIamide and FMRFamide) and only APGWamide, which is normally co- localised in the right pedal ganglia, induced imposex (Oberdörster, 2000). From these results, APGWamide was proposed as PMF, which controls the development of secondary male sexual characters from the right pedal ganglia.

Titley-O’Neal at al. (2011) reported that, in a subsequent study, I. obsoleta was exposed to TBT, APGWamide, estradiol (E2), 3-methylcholanthrene (3MC), a highly carcinogenic polycyclic aromatic hydrocarbon and mixtures of TBT and 3MC. Exposure to

3MC or E2 alone did not induce imposex, but co-exposure of 3MC+ TBT and E2 + TBT blocked the TBT-induced imposex. On the other hand, and according to Oberdörster and

McClellan-Green (2002), when APGWamide was mixed in a cocktail with 3MC or E2, none of the mixtures was able to significantly block APGWamide-induced imposex. These data support the hypothesis that APGWamide-induced imposex occurs by neuropeptide action.

Imposex induction by testosterone is different from the induced by TBT, and is possible that testosterone interferes with a downstream signaling event to induce imposex. This fact is explained by Oberdörster et al. (2005) that provided further evidence supporting the neuropeptide functioning as the PMF hypothesis, from two fronts. At first, APGWamide administered exogenously increased male penis length confirming that APGWamide can function as PMF. Secondly, I. obsoleta females showing TBT-induced imposex had APGWamide levels comparable to control males (Oberdörster et al., 2005).

Even so, the neuropeptide hypothesis was also evaluated in other species and APGWamide failed to promote imposex in at least 2 species: (i) Bolinus brandaris (Santos et al., 2006) and (ii) N. lapillus (Castro et al., 2007).

Retinoic pathway:

The last proposed, and most accepted, hypothesis of imposex induction involves the activation of the nuclear receptor “retinoid X receptor” (RXR) by TBT. The participation of RXR on imposex induction was first mentioned by Nishikawa et al. (2004) that, through laboratory experiments with Thais clavigera and 9-cis Retinoic acid (9CRA) bound to the human RXR (hRXR) in females with high affinity in similar manner to OTs. Since the ligand binding in T. clavigera and vertebrate RXR, was similar for both OTs and 9CRA, it was suggested that RXR plays an important role in imposex induction via growth and differentiation of the male genital tract in female gastropods.

Subsequently, a speculative theory of imposex induction in T. clavigera via RXR was proposed by Horiguchi et al. (2007) referring the activation of a signalling cascade which would be dependent on the RXR activation / inhibition (see Figure 1.11). The imposex response in 9-cis RA treatment was similar to the one obtained in TBT treatment. These results were further confirmed for the same species through a series of laboratory experiments and also through the assessment of RXR gene expression and protein content in wild animals (Horiguchi et al., 2007a).

The authors Horiguchi et al. (2007) described a significantly higher RXR gene expression in the penises of males and of imposex-affected females than in the normal females penis-forming area (the region behind the right tentacle) and suggested that, for the differentiation and/or growth of the penis and vas deferens in T. clavigera, an interaction between OTs and RXR in this limited tissue area is primarily important.

The involvement of RXR on imposex induction has also been confirmed in N. lapillus (Castro et al., 2007). The retinoid signalling has been increasingly implicated in the regulation of male reproductive differentiation and development (Sternberg et al., 2008) and this stills the most accepted theory of imposex induction.

The Retinoic pathway hypothesis was then reinforced by a study published by Sternberg et al. in 2008 proposing that the early retinoid signalling via RXR may stimulate

54 male recrudescence, whereas later signaling may stimulate female recrudescence and, if so, TBT may induce the development of male sex characteristics in females (i.e. imposex) by initiating RXR signalling prematurely in females during the temporal gap of normal male recrudescence. The same working team has recently reviewed what is currently known on environmental and endocrine signals that regulate sexual differentiation / development in gastropods, as well as the known effects of TBT on relevant endocrine parameters in molluscs (Sternberg et al., 2009), and used the information to propose a rational model for the mechanism by which TBT causes imposex (see Figure 1.12).

According to them, TBT activates the RXR signalling pathway to initiate the transcription of genes necessary for male reproductive organs development. Activation of this pathway by TBT can be directly, by binding to and activating RXR or indirectly, by inhibiting acyl CoA acyltransferase, resulting in an increase in endogenous retinoids and testosterone levels (Sternberg et al., 2009). With increasing of the endogenous retinoids levels, RXR is then activated by endogenous free retinoid that, in turn, stimulates gene

55 transcription through interaction with RXR response elements. Finnaly, exogenously- administrated testosterone competitively inhibits retinoid esterification resulting in elevated endogenous free retinoids, capable of activate the RXR signalling pathway, leading to male reproductive organs development.

In any case, there are still several issues to clarify to achieve, for sure, the complete knowledge on the complex mechanism by which TBT-induced imposex develops. For now, the certainty that this phenomenon occurs in a TBT dose-dependent manner (Oehlmann et al., 1998) ensures its usefulness as a biomarker of TBT environmental pollution (Matthiessen and Gibbs, 1998).

1.3.2 Imposex as a biomarker of TBT pollution

Environmental monitoring of organotin compounds in coastal areas has traditionally utilized the analysis of physical / chemical parameters; biological variables were just occasionally considered, namely the assessment of benthic fauna changes in sediments monitoring (Lam and Gray, 2003; Galante-Oliveira, 2010). Was in the 70s that started to make pollution monitoring and this had as target the analysis of several ecosystem compartments as sediment, water and biota (Kramer, 1994). The term biomonitoring signify monitoring of biota and this has been intensively developed, not only regarding the optimization of methods for the detection of chemicals in organisms, but also defining new parameters and developing new methods to assess the biological effects of many contaminants, namely by the identification and validation of biomarkers (Galante-Oliveira, 2010).

As can be documented from the multiplicity of definitions of term “biomarker” or “biomarker response” found in the literature, this term has aroused much interest and not only amongst the scientific community. According Decaprio (2006) in toxicology, a

56 biomarker is usually defined as “quantifiable biochemical, physiological or histological measures that relate in a dose- or time-dependent manner the degree of dysfunction that a contaminant has produced”. The use of the term “biomarker” is usually restricted to cellular, molecular, biochemical or physiological changes that are measured in cells, body fluids, tissues or organs within an organism (Galante-Oliveira, 2010). The “bioindicators”, to contrary of biomarkers, are changes that occur at the individual, population and higher levels of the biological hierarchy (Schlenk, 1999; Lam and Gray, 2003). There is evidence that one of the functions of biomarkers is provide early warning signals of biological effects, and that it is generally believed that molecular, biochemical and physiological responses tend to precede those that occur at the individual or higher levels (Lam and Gray, 2003). This function is one possible reason for limiting the term “biomarker” to sub- organismic changes.

According Hence, in 1987, the biomarkers may be classified into three classes that were proposed by the Committee on Biological Markers of the National Research Council (NRCm) such as (i) markers of exposure, (ii) of effect and (iii) of susceptibility; (National Research Council, 1987). The definitions of the considered classes are:

(i) Biomarkers of exposure – exogenous substances, or their metabolites or products of the interaction between xenobiotic agents and some cells or target molecules, that are measured in a compartment within an organism;

(ii) Biomarkers of effect – measurable physiological, biochemical, behavioural or other alterations within an organism that, depending upon the magnitude, can be recognized as associated with an established or possible health impairment or disease;

(iii) Biomarkers of susceptibility – indicate an inherent or acquired ability of an organism to respond to the challenge of exposure to specific xenobiotic substances.

Imposex has been recognized as the biomarker of exposure and effect (Picado et al., 2007; Hagger and Galloway, 2009). This biomarker has been successfully applied in

57 ecotoxicology studies as biomonitors of organotin pollution and the most studied gastropod species used in TBT biomonitoring studies is the dogwhelk, Nucella lapillus.

1.3.2.1 Vantages and disadvantages

There are advantages and disadvantages in using this biomarker. Is considered as one of the main advantages is imposex specificity to TBT since there is no other definitive description of the phenomenon development by another agent, except for parasitism (Schulte-Oehlmann et al., 1997) and for TPT that occurs at very low environmental concentrations (Barroso et al., 2002a; Sousa et al., 2007). But the TBT is the major causative agent of imposex in dogwhelks.

The robustness of imposex as a specific biomarker led the OSPAR Commission to recommend its use to monitor TBT pollution levels across Europe (OSPAR, 2003). The imposex expression intensity is related to the TBT exposure concentration, so by the phenomenon quantification it is possible to estimate the regime to which the dogwhelk was exposed (Gibbs, 1999). The quantification of imposex is of simple observation and determination using classification schemes. The quantification this phenomenon is also of low cost and less time-consuming method, without requiring any sophisticated equipment, which shows the pollutant effects in a given ecosystem from the individual to the community level (Barroso et al., 2000).

Other advantage is that the imposex intensity is a measurement of the extent of female masculinization and can be described in many different ways (Titley-O’Neal et al., 2011). By using the traditional imposex monitoring techniques, that comprise the determination/calculation of indices developed to quantify the intensity of the imposex expression in response to the TBT environmental concentration, it is possible to compare different populations and thus the TBT concentration an the respective location. The simplest approach is determination of the percentage of imposex incidence (%I) by calculating the percentage of imposex-affected females (Titley-O’Neal et al., 2011).

Also the use of high sensitivity of biomonitoring as a method to evaluate TBT pollution levels minimizes some of the problems related to the compound chemical determination such as the high variation of the TBT concentrations in water (Galante- Oliveira et al., 2009; Galante-Oliveira et al., 2010) and the fact that the imposex may be initiated at TBT concentrations below the chemical detection limit (Gibbs and Bryan, 1986; Barroso et al., 2000). Beyond the already mentioned advantages, this response also presents the remarkable ability of allowing the detection of TBT pollution point sources at a high spatial resolution (Axiak et al., 2002).

The use of imposex biomarker also has some disadvantages. A disadvantage is, imposex is an irreversible phenomenon, and thus its expression intensity may indicate levels of TBT higher than those that actually occur i.e. sampling moment. The use of this method also has limitations that are extended to observation: it has to be carried out on living animals, without being preserved / frozen (Minchin and Davies, 1999a).

Anyhow, the most relevant problem of using this biomarker is the fact that several studies have recently reported imposex development in some species by compounds other than OTs. Pseudo or “natural” imposex may occur from the action of steroids (Bettin et al. 1996; Oehlmann et al. 1996; Oberdörster and McClellan 2000; Oberdorster and McClellan 2002; Santos et al., 2005) and some relatively recent reports on the phenomenon development by exposure to contaminants such as Aroclor 1260 (Garaventa et al., 2008) and nonylphenol (Evans et al., 2000) do question the specificity criterion hitherto accepted for this biomarker.

Moreover, Schulte-Oehlmann et al. (1997) mentioned that the first description of masculinization in Hydrobia ulvae females was done before the problem of the TBT environmental pollution. Other authors such as Krull (1935) and Rothschild (1938) described the presence of non-functional small penises in Hydrobia specimens infested by parasites, while Schulte-Oehlmann and co-authors (1997) noted only a small increase of parasitism in imposex-affected females when compared with non-affected ones. Even so,

59 according to these authors, parasites are not the only cause for females’ masculinization since they were not observed in 100% of the imposex-affected females. Therefore, parasitized specimens have been discarded from imposex analysis in biomonitoring programs.

1.3.2.2 Indicator species

Imposex phenomenon occurrence was already described in more than 200 gastropods species, representing 28 families, around the world (Shi et al., 2005), these gastropods species were used as bioindicators of TBT environmental levels by the quantification of the exhibited masculinization degree. The relationship between imposex occurrence and OTs, mainly TBT, was first proved in England by Smith (1981), and, since then, many species were listed as indicators of imposex.

However the bulk of studies have used gastropods as bioindicators of TBT or TPT exposure focusing on three families Muricidae > Buccinidae > Nassariidae, and this accounts for < 10% of the total number of the families investigated (Titley-O’Neal et al., 2011).

Later this work was then confirmed by Bryan et al. (1987), through a series of laboratory and field experiments using the dogwhelk N. lapillus, was this working team that ended up proposing the use of this species as a bioindicator of TBT environmental pollution (Gibbs et al., 1987). The N. lapillus is bioindicator species of environmental pollution by TBT most used in the entire world (Titley-O’Neal et al., 2011) and belongs to the order Neogastropoda. Each region of the world has its bioindicator species of imposex. These species belong to the order Neogastropoda, such as Nucella lapillus, and develop imposex signals when exposed to TBT (Titley-O’Neal et al., 2011). As Nucella lapillus was the first species proposed as a bioindicator of imposex (Bryan et. al., 1986; Gibbs et al., 1987), it is used here as the bioindicator model species. Various parameters are recommended to monitor TBT pollution biological effects in this species. The main

60 parameters are: mean females penis length (FPL), relative penis size index (RPSI), vas deferens sequence index (VDSI), percentage of imposex-affected females (%I) and percentage of sterile females (%S), (Gibbs et al., 1987).

RPSI is calculated by the formula [FPL³ / mean male penis length (MPL)³] x100 and was proposed by Gibbs et al. (1987) to compare imposex expression in different N. lapillus populations. In general, the imposex development can be followed by the vas deferens sequence (VDS) and its intensity quantified by using a VDS classification scheme. The average of the VDS stages obtained for a certain population results in the vas deferens sequence index constituting an expression of the intensity of the phenomenon in that population. According (Stroben et al., 1996) linear correlation between VDSI and TBT levels in the environment and tissue samples is higher than those for other imposex indices and this parameter (or index) is the one with the highest biological meaning. In N. lapillus, VDSI recognises the development of male sexual caracters (vas deferens and penis) in a female in 7 stages (0-6), see figure 1.13 (Gibbs et al., 1987) and two distinct pathways of completing the phenomenon development (A- and B-path). In the stage 0 the female is normal (no male characteristics), in the stage 1 the proximal section of vas deferens is formed, in the stage 2 occurs initiation of penis development and further development of vas deferens, in the stage 3 occurs the formation of a small penis and the development of the distal section of the vas deferens, the stage 4 involves mainly the fusion of the vas deferens, in the stage 5 occurs the overgrowth of the vas deferens on the genital papilla and hence the female become sterile, and in the stage 6 egg capsules are aborted and can be seen inside the capsule gland (Mallon and Manga, 2007).

Figure 1.13: Nucella lapillus VDS classification scheme. Abbreviations: a, anus; b, blister; gp, genital papilla; n, nodule; p, penis; v, vulva; vd, vas deferens; A, either blisterslike protuberances around the papilla; B, nodules of hyperplasic tissue. Adapted from Gibbs et al. (1987).

Females exhibiting VDS stage 5 or 6 are prevented from releasing egg capsules from their genital papilla and are sterile. However, is the VDS stage 6 that corresponds to the complete development of imposex, i.e., formation of a continuous vas deferens from the genital papilla to the penis base, which overgrowth covers of the genital papilla and blocks of the vulva, becoming sterile females and resulting in an accumulation of aborted egg capsules inside the capsules gland of N. lapillus (Gibbs et al., 1987). The distinction between A- and B-paths was made from of the vas deferens tissue proliferation overgrowing the genital papilla in the stage 5 is in the form of blisters-like protuberances around the papilla in the A-path, while in the B-path it appears like nodules of hyperplasic tissue. However, generally the most common in Nucella lapillus is that the imposex development finalizes by the A-path.

Some authors believe that the VDSI is the most consistent index to measure imposex intensity, as it is much more reliable than for example the RSPI (Bennett, 2004). The VDSI does not rely only on the penis length but it depends on the progress of the development of the male characters (penis and vas deferens) onto the female and allows the evaluation of the female reproductive capability, so it is an index biologically meaninful.

Although Nucella lapillus has been the 1st bioincator species of imposex validated, this species has a limited geographical distribution and is common on rocky shores and can be found in both sheltered and heavily wave exposed areas. Nucella lapillus occurs within a salinity range from 18 to 40 psu and between 0 and 20°C isotherms, throughout the North Atlantic littoral zone: from the Arctic to the south of Portugal in the east, including the Faroe Islands and Iceland, and from the south west of Greenland to the north of Long Island in the west (Crothers, 1985; Tyler-Walters, 2008), (Figure 1.14).

Figure 1.14: Geographical distribution of Nucella lapillus in the world. Extracted from Tyler- Walters (2008).

Due to the limited geographical distribution of Nucella lapillus in the world other bioindicator species, of prosobranch gastropods, were being described in other environments and regions as Hexaplex trunculus (Chiavarini et al, 2003; Axiaka et al., 2003; Terlizzi et al., 2004), Nassarius reticulatus (Stroben et al., 1992; Bryan et al., 1993; Barroso and Moreira, 1998; Barroso et al., 2002b; Rato et al., 2006), Stramonita haemastoma (Spence et al., 1990; Rilov et al., 2002) and Thais clavigera (Hung et al., 2001), inter other.

However, different species do not exhibit the same sensitivity to TBT, i.e. it was reported different levels of imposex in females of different species exposed to a given concentration of the compound. It was the case of Stroben et al. (1995) work that found, in a given highly contaminated area, dissimilar imposex levels in Ocenebra spp. and Nassarius spp. female specimens. Thus, these authors have ranked such bioindicator

63 species in the following descending order of TBT sensitivity (i.e., from the most to the least sensitive): Ocenebra aciculata > Nucella lapillus > Ocenebra erinacea > Trivia arctica > T. monacha > Nassarius reticulatus > Nassarius incrassatus (Stroben et al., 1995). According to other studies of the same working team, when a comparison between the sensitivity of mesogastropods and neogastropods to TBT was made, mesogastropods exhibited signs of imposex when exposed to higher concentrations of TBT and so neogastropods were proved to be more sensitive to these compounds (Stroben et al., 1992).

1.3.2.3 Biomonitoring guidelines

The prosobranch gastropods are collect randomly by hand at each site. Sampling of Nucella lapillus is carried out whilst the tide is at low shore due to the Nucella binding to rocks further down the shoreline (Mallon and Mangam, 2007). Sites are located at varying distances from areas of boating traffic (example harbours and marinas). The abundance of N. lapillus is observed at each site using the 0–6 scale adopted by Evans et al. (1996), where 0 is absent; 3 is common; and 6 is superabundant. At each of the juvenile and sub adult survey sites, an attempt is made to obtain a determined number of individuals from each of the size classes (Birchenough et al., 2002), example about 45-60 adult Nucella lapillus were collected at the intertidal rocky shore from May to August 2003 at sites 1-17 along the Portuguese coast (Galante-Oliveira, 2010). In addition, samples of a given interval of individuals are made by collecting N. lapillus at various points at a given distance (example 10 m), horizontally along the shore (Birchenough et al., 2002). The prosobranch gastropods are placed in a container containing sea water and sealed for transit. It should be noted which juveniles specimens and adults and in which regions these were collected (Mallon and Mangam, 2007). For classified the prosobranch gastropods as adult or juvenile, the length of each individual is measured from the tip of the spire to the end of the operculum (see Crothers, 1985). Sub-adults are classified as juveniles.

Following collection, specimens are transported to the laboratory for immediate analysis when they were fresh (Bryan et al., 1986; Douglas et al., 1993; Evans et al., 1994, 1996; Mallon and Mangam, 2007) or stored in a refrigerator until analysis is place (Evans et al., 1998; Mallon and Mangam, 2007). In the laboratory the shell height of Nucella lapillus and other prosobranch gastropods (SH –length from the apex to the siphonal canal) is measured with vernier callipers to the nearest 0.1 mm (Galante-Oliveira, 2010). After narcotization of the gastropods (except Nucella lapillus) their shells are broken with the aid of a vise and removed. The following is measured the penis length of the prosobranch gastropods collected using a stereo microscope and is determined the degree of imposex these gastropods. The procedure for measuring imposex was that described by Gibbs et al. (1987). Gibbs and his team recommend two indexes of imposex to determine the degree of imposex in N. lapillus: the Relative Penis Size Index (RPSI) and the Vas Deferens Sequence Index (VDSI) and are measured using international standard techniques (Harding, Rodger and Davies, 2001). The comparison between these two indices is made between measures performed in the present study and those in previous studies (Birchenough et al., 2002). According Birchenough et al. (2002) the RPSI values in earlier work were often in the range of 50–100 in samples collected from areas of severe TBT contamination. These values were associated with high VDSI indices and often sterility of individual females (Birchenough et al., 2002).

Other indices also recommended by Gibbs et al. (1987).to determine the degree of imposex in gastropods of N. lapillus are the percentage of imposex affected females (%I) and the percentage of sterile females (%S). These indices are determined for each site, according to Gibbs et al. (1987). Is generally used the soteag rocky shore monitoring programme. TBT contamination in sullom voe, shetland 2001 dogwhelk survey. Fisheries Research Services Report No 04/01 (Harding, Rodger and Davies, 2001).

Generally the TBT and its metabolites are measured by atomic absorption spectroscopy (AAS) and by gas chromatograph (GC) in homogenized whole tissues of

65 females from each site. Are used analytical procedures based on the methods of Ward et al. (1981) and are fully described by Bryan et al. (1986).

The OSPAR Commission adopted the use of the biomarker imposex in N. lapillus to monitor TBT pollution levels in coastal areas of the Europe (OSPAR, 2003). Due to absence this species in some coastal areas, the OSPAR Commission recommended the use of others species of dogwels for monitoring the TBT pollution and their effects, namely: Nassarius reticulatus, Buccinum undatum, Neptunea antiqua and Littorina littorea (OSPAR, 2003). The Vas Deferens Sequence Index (VDSI) and Intersex Index (ISI) are used to determine the levels of imposex these five species (OSPAR, 2004). For other regions of the world, different species are used to monitor TBT pollution levels.

The development, within OSPAR, of an Ecological Quality Objective (EcoQO) for imposex in dog-whelk N. lapillus is a TBT-specific biological effects assessment criterion. Is the desired level of Ecological Quality (EcoQ) that defines the EcoQO and the reference level of EcoQ is the one to which the anthropogenic influence on the ecological system is minimal. The EcoQO was developed within OSPAR since the EcoQ reference level achievement is under the Commission duties with the objective of protect and preserve the overall expression of the function and structure of the marine ecosystem, taking into account the biological community, natural factors and conditions, as those resulting from human activities (OSPAR, 2009).

The EcoQO was developed considering that the average imposex level, in a sample ≥10 females of dogwelks of the recommended species by OSPAR Commission, should be consistent with the exposure to TBT concentrations below the Environmental Assessment Criteria (EAC) derived for TBT, that is VDSI <2.0. According OSPAR (2009) where N. lapillus does not occur naturally, or where it has become extinct, the species N. antiqua, N. reticulatus, B. undatum or L. littorea should be used, with an exposure criteria on the same, or a correspondent, VDSI <2.0 (in Neptunea) and VDSI <0.3 (in Nassarius, Littorina and Buccinum).

The monitoring taking into account the EcoQO on imposex in dog-whelks is a mandatory commitment of Contracting Parties under the Co-ordinated Environmental Monitoring Programme (CEMP) and should be carried out in accordance with Technical Annex 3 of the Joint Assessment and Monitoring Program (JAMP) Guidelines for contaminant-specific biological effects monitoring (OSPAR, 2009 and Galante-Oliveira, 2010).

The table 1.3 was presented by CEMP- a colour code that was assigned to assessment classes aiming to ease, in any graphical representation, the recognition of the locations where the EcoQO was achieved.

In the number 1 the species cannot be used to distinguish between class A and B. The assessment class is therefore by definition B; in the number 2 the species cannot be used to distinguish between classes A, B and C. The assessment class is therefore by definition C; in the number 3 the species cannot be used to distinguish between class C and higher classes. For a VSDI = 4.0, additional observations are required to determine the

67 assessment class e.g. by using another species. When is observed one VDSI of 4.0, the assessment class is by definition F; in the number 4 the species cannot be used to distinguish between classes E and F. Are required more observations to determine the assessment class e.g. by using another species. If the VDSI (Nassarius) or the PCI (Buccinum) is >3.5, the assessment class is therefore by definition F.

1.4 Restrictive measures on the use of Organotins (OTs)

First regulations on the use of OTs were adopted by countries that experienced negative consequences due to TBT environmental pollution (see section 1.2). The first country where OTs were restricted was France, in 1982, due to the collapse of the oyster farming industry of the Arcachon Bay, in an attempt to reduce environmental concentrations of TBT. The French Ministry of Environment announced a temporary two years ban on the use of TBT-based AF paints containing more than 3% wt. of the biocide in vessels less than 25 m in length, for both the Atlantic coasts and the English Channel (Champ, 2000). Therefore, these paints application was only allowed on boats and marine craft with an overall length greater than 25 m (Champ, 2000). Four years later, a similar measure was taken by the United Kingdom Parliament in order to reduce the environmental impact of the pollutant TBT, namely on mollusk populations (Alzieu, 1998). This regulatory action restricted the application of TBT as biocide on small boats AF systems but also on aquaculture structures and equipment. This action consisted in five steps: 1) develop regulations starting on the 1st January 1986 to control the sale of the most damaging OT-containing paints, to ban the use of ‘free association’ OT-based AF paints by small boat owners, and to set the maximum levels for the OT content of ‘copolymer’ paints; 2) create a notification scheme for all new antifouling agents; 3) develop guidelines for the cleaning and painting boats coated with OT-based AF paints; 4) establish a provisional environmental quality target for the concentration of TBT in water; 5) coordinate and further develop OTs monitoring and research programs (Champ, 2000).

The United Kingdom Parliament also enacted the Food and Environment Protection Act (FEPA) to ensure that, in the future, all AF agents (of any kind) would be screened in the same way as other pesticides (Champ, 2000). This enactment was coordinated with the Control of Pesticides Regulations (COPR), which provided for the statutory screening of AF paints from the 1st July 1987 (Champ, 2000). These regulations prohibited the sale, advertisement, supply, storage or use of any biocide, including TBT-based AF paints. In addition, similar restrictions were adopted in other countries, namely: USA and New Zealand in 1988; Australia, Norway and Canada in 1989 and Japan in 1990 (Sousa et al., 2013). In Japan, given the results of laboratory and field studies, seven TPT compounds and thirteen TBT compounds were designated in 1990 as Class II Specified Chemical Substances (Champ, 2000). Subsequently, the production, import and use of TBT and TPT came under the domestic law concerning the manufacture control and respective regulation (Champ, 2000). Japanese government ministries established domestic counter-measures to prohibit the application of TBT antifoulants on all vessels (including both small boats and larger ships) and maritime structures. Switzerland and Austria, which have no direct access to the ocean, have banned the use of TBT-based AF paints in freshwater environments. In the Federal Republic of Germany, the following measures entered into force in 1990: ban on the use of OT-based AF paints in boats less than 25 m long; ban on retail sale; ban the use on structures for mariculture; TBT limit of 3.8% (wt.) in copolymeric paints and regulation for the safe disposal of AF paints after removal (MEPC 30/20/2-IMO, 1990; Champ, 2000). Within the European Union, an amendment to the Council Directive 76/769/EEC was proposed in February 1988, to restrict the marketing and use of certain dangerous substances such as OTs. Since then, the application of OTs as substances and constituents of preparations intended for use to prevent the fouling by micro-organisms, plants or animals became prohibited in hulls of boats with an overall length <25 m (as defined by ISO 8666) and in floats, cages, nets and any other appliances or equipment used for fish or shellfish farming (Davies et al., 1986, 1987; Davies and McKie, 1987). In addition, in 1989 the European Union introduced the Directive 89/677/EEC banning the use of TBT and TPT on boats of an overall length less than 25 m (Sousa et al., 2013). The Directive

89/677/EEC was posteriorly transposed to each member state internal law and implemented. With this TBT partial ban, some recovery of severely affected gastropod and oyster populations was reported, as well as of other marine communities (Minchin et al., 1995; Evans et al., 1995, 1996; Harding et al., 1997; Alzieu, 1998; Blanck and Dahl 1998; Miller et al., 1999; Rees et al., 2001). However, these restrictions were not totally effective and there are several reports describing that imposex and TBT levels around ports were not effectively decreasing around those TBT pollution hotspots (Bailey and Davies 1989; Davies and Bailey 1991; Minchin et al., 1995, 1996; Smith, 1996; Morgan et al., 1998). Moreover, TBT pollution was not only restricted to harbor areas, and populations from offshore areas as well as the ones from the open ocean, affected by TBT pollution (Ten Hallers-Tjabbes et al., 1994; Mensink et al., 1996). Noteworthy that TBT was also detected in remote and supposedly undisturbed ecosystems such as Antarctica (Negri et al., 2004) and the Great Barrier Reef in Australia (Marshall et al., 2002; Haynes et al., 2002; Haynes and Loong 2002), proving that the pollution by these compounds was a problem on a global scale. In 1998, at Marine Environmental Protection Committee 42 (MEPC 42-IMO, 2002), several countries as Belgium, Denmark, France, Germany, Norway, the Netherlands, Sweden and United Kingdom, joined Japan in requesting a global ban and proposed that the MEPC recommend a ten years period for phasing out and totally ban of the use of TBT as biocide in AF boat bottom paints on ships worldwide (MEPC 42/22, annex 5). It was proposed that a legal instrument should be developed by the International Maritime Organization (IMO) in the form of a free standing convention; it should be effective, legally binding, global in scope and that should ensure an expeditious entry-into-force process. Furthermore, the instrument should include a mechanism for addressing AF systems other than the OTs-based (Champ, 2000). In the session from 28 June to 2 July, 1999, the MEPC agreed to use the framework and the basic text contained in its document Ref. 43/3/2 as a basis for developing that legal instrument. This committee also reviewed the documents by the Marshall Islands meeting (MEPC 43/3/6) and joint documents by BIMCO, INTERCARGO, ICS, INTERTANKO, OCIMF and SIGTTO (MEPC 43/3/9) related with dates of TBT phasing out process (2003 and 2008). The MEPC

70 requested an IMO Council meeting that was held in London in November 1999 to approve the diplomatic conference on AF systems that adopted the legal instrument that regulated the use of shipboard AF systems and that phased out those containing OTs such as TBT (Champ, 2000 and Champ, 2001).

Accordingly, following that diplomatic conference, the definite IMO legal instrument was adopted on the 5th October, 2001: the International Convention on the Control the use of Harmful Antifouling Systems on Ships (AFS Convention), (Champ, 2001a). The IMO AFS Convention had as objective to prohibit the use of harmful OTs in AF paints used on ships and to establish a mechanism to prevent the potential future use of other harmful substances in AF systems (Champ, 2003). The AFS Convention foresaw to prohibit the application or re-application of TBT-based AF paints from the 1st January 2003 and the circulation of ships with TBT from the 1st January 2006 (Champ, 2003; Sousa et al., 2013). Even so, the document would only enter into force 12 months after being ratified by 25 members representing at least 25% of the world’s merchant shipping tonnage. These conditions were only verified on 17 September, 2007. Consequently, the AFS Convention entered into force on 17 September, 2008, being lost the effectuation of the application and circulation prohibitions at the initially proposed dates (2003 and 2006, respectively). Meanwhile, in March 2002, Sweden, Netherlands, United Kingdom and Finland, had announced the cancellation of the production and registration of all TBT-based AF products. In the same year the United States had also adopted the legal procedures necessary to implement the AFS Convention, so that, once ratified, it could immediately be applied (IMO, 2002). Also due to the delay in the AFS Convention ratification process, the European Union adopted the Directive 2002/62/EC and Regulation No. 782/2003 to implement the convention on EU member states yet in 2003 (from the 1st July) and the complete prohibition of OTs AF paints from January 1, 2008 (Champ, 2003; Sousa et al., 2013). Therefore, the circulation of ships with OTs as biocides on ships hulls (or on any of their external parts or surfaces) was banned from January 1, 2008 on EU member states national fleets, to any vessel flying their flags and on any vessel entering a member state port (Champ, 2000; Sousa et al., 2013). This interdiction was applied to all ships,

71 regardless of size and registered tonnage, but also to fixed and floating platforms, floating storage units (FSUs) and floating production storage and offtake units (FPSOs). Were 17 countries that ratified the AFS Convention from the 1st January 2003 by the end of 2003 / early 2004 (see table 4). In Europe, the application and re-application of TBT-based AF products on ships was not allowed from the 1st January 2003 and only Denmark (1.24% of the world’s merchant fleet tonnage) had completely ratified the AFS Convention by that date.

Other countries also followed this convention and, at the MEPC 49th session, in July 2003, five were those for which the convention had been ratified: Nigeria (0.07% of the world’s merchant fleet tonnage), West Indies and Barbados (0.81%), Japan (2.53%) and Norway (3.93%). These countries together accounting for only 8.58% of the tonnage required for the convention worldwide application.

Seven more countries signed the document, and were pending the ratification of the AFS convention as soon as the process of national legislation amendment was finished. These seven countries were: Australia (0.33%), Belgium (0.03%), Brazil (0.64%), Finland (0.28%), Morocco (0.51 %), Sweden (0.51%) and USA (1.91%), whose total contribution, together with the above indicated, would totalize 12.79%, that continued to be insufficient.

By the end of 2003 / early 2004 other countries had formally announced the convention signing with such: Greece (4.99%), Spain (0.40%) and Italy (1.68%). United Kingdom (1.05%) and Panama (20.80%) had informally indicated their willingness to ratify the Convention in 2003 (IMO, 2009). In the table 4 also is possible observe that from the early 2004 by the end of 2013 were 51 new countries that ratified the AFS convention and that from the end of 2013 until 25 June 2014 were 3 new countries that ratified this convention. The total number of countries that have ratified the AFS Convention from the 1st January 2003 until 25 June 2014 was 71.

Table 1.4: Countries that have ratified the AFS Convention from the 1st January 2003 until 25 June 2014. Adapted from: IMO (2009); status of conventions as at 2 December 2013 and status of conventions as at 25 June 2014 – the 1991 and 1993 amendments to the IMO Conventions are in force for all IMO Members. Continent/ period of Africa America Asia Europe Oceania ratification Morocco; West Indies Japan; Denmark; Australia; Nigeria; and Norway; From the Barbados; Belgium; 1st Brazil; Finland; January USA; Sweden; 2003 by Greece; the end of Spain; 2003 / Italy; early 2004 United Kingdom; Panama Egypt; Antigua and China; Bulgaria; Cook Islands; Liberia; Barbuda; Iran (Islamic Croatia; Kiribati; Sierra Leone; Bahamas; Republic of); Cyprus; Marshall Tunisia; Canada; Ireland; Estonia; Islands; Mexico; Jordan; Ethiopia; Niue; Republic of Lebanon; France; Palau; Korea; Macao; Germany; Tuvalu; Saint Kitts Malaysia; Hungary; Vanuatu; From the and Nevis; Mongolia; Iceland; early 2004 Trinidad and Russian Latvia; by the end Tobago; Federation; Lithuania; of 2013 Uruguay; Singapore; Luxembourg; Syrian Arab Malta; Republic; Montenegro; Netherlands; Poland; Romania; Serbia; Slovenia; Switzerland; Faroe Islands From the South Africa; Tonga end of Congo 2013 until 25 June 2014

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Chapter 2: The dissertation

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Chapter 2

The dissertation

______Table of Contents Chapter 2: The dissertation ...... 102 2.1 Thesis aims and organization ...... 103 2.2 Methods for the data obtaining and processing ...... 104

2.1 Thesis aims and organization

The main objective of the present study is the analysis of the global trend of tributyltin (TBT) environmental pollution and of the effort of its biomonitoring by the biomarker imposex in the last decade, through the collection of bibliographic information published in this topic between 2003 and 2013 and the respective data processing. This period comprises the most definite legislative restrictions on the use of organotins compounds (OTs) worldwide, specially TBT, as explained in chapter 1 (section 1.5): 2003 is considered the European ban on the use of these compounds (by the EU Regulation No. 782/2003) and the worldwide ban was in 2008 by the entry into force of the International Maritime Organization (IMO) ‘‘International Convention on the control of harmful antifouling systems in the Ships’’. Therefore, this work ultimately pretends the analysis of these compounds phase out period and the evaluation of these measures effectiveness in reducing TBT pollution on a global scale.

After an initial “General Introduction” on the theme in chapter 1, the dissertation “aims and organization” is presented in chapter 2 (section 2.1) and the methods applied in (i) the bibliographic information collection and (ii) the respective results analysis, are described in section 2.2. The results on the “evolution of the worldwide pollution by TBT” in the period under analysis are presented in chapter 3, initially showing the “global trend

103 of TBT pollution monitoring in the last decade” (section 3.1) and then the “TBT pollution evolution from 2003 to 2013” on a global scale (section 3.2). The entire data set was analysed always having in mind the studies geographical distribution, enabling a global and comprehensive overview of the problem under study.

2.2 Methods for the data obtaining and processing

The data processed in the present work was obtained from an initial comprehensive literature search. Two different search engines were used namely: the Web of Science from Thompson Reuters (formerly the ISI Web of Knowledge) and the Google Scholar. The keywords used in these two search engines were the following, and by the following order: Imposex, Tributyltin, TBT and organotin compounds. All the results retrieved were explored and their abstracts read. Works describing imposex levels biomonitoring and chemical monitoring of OTs concentrations in different ecosystem compartments (biota, water, sediment) were carefully analysed. The information contained therein was gathered in an Excel matrix built for this purpose. The information was organized as follows: (i) general data of the publication (i.e., author; title; year of publication); (ii) imposex monitoring data (i.e. bioindicator(s) order/family/genus/species; year(s) of specimens collection; geographic region monitored, namely continent and country; imposex levels indices and respective evolution; (iii) chemical monitoring data (i.e. OTs concentrations in the bioindicator(s) tissues, water and sediments; and (iv) reference to the risk of the analysed populations decline and/or the bioindicator extinction or mention to the analysed populations recovery. The imposex indices considered were the vas deferens sequence index (VDSI), the percentage of imposex-affected females (%I), the percentage of sterile females (%S), the mean females penis length (FPL) and the relative penis size/length index (RPSI or RPLI, respectively), these latter depending on the bioindicator species in use.

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Regarding OTs concentrations in tissues, water and sediments, special attention was given to the concentration of TBT and its metabolites (DBT and MBT) and, in addition and when mentioned, the butyltins degradation index (BDI) was also taken into account (as it allows the evaluation of fresh TBT inputs: BDI<1 means TBT recent input; BDI> 1 means TBT old input).

The exel matrix was then reorganized for the data processing. The analysis envolved the construction of the following graphs: number of publications vs. year; percentage of papers vs. continent (considering the following: America, Europa, Asia, Africa and Oceania); percentage of papers vs. country (by continent); number of species used as bioindicator of imposex levels in studies published from 2003 to 2013 (total and by continent and country); number of species vs. order (total and by continent and country); number of species vs. family (total and by continent and country); number of populations in decline (by continent); number of extinct populations (by continent); evolution of imposex levels (imposex levels vs. year by continent) and evolution of OTs concentrations in tissues, water and sediments (OTs concentrations vs. year by continent).

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Chapter 3: Evolution of the worldwide pollution by TBT in the last decade

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Chapter 3

Evolution of the worldwide pollution by TBT in the last decade

Table of Contents Chapter 3: Evolution of the worldwide pollution by TBT in the last decade...... 106 3.1 Global trend of TBT pollution monitoring in the last decade ...... 108 3.1.1 General results on the bibliographic search ...... 108 3.1.1.1 Number of studies published between 2003 and 2013 ...... 108 3.1.1.2 Worldwide distribution of the studies published ...... 109 3.1.2 Species used as bioindicators of imposex from 2003 to 2013 ...... 113 3.1.2.1 By taxonomic group...... 119 3.1.2.2 By geographic region ...... 122 3.2 TBT pollution evolution from 2003 to 2013 ...... 126 3.2.1 Imposex levels in gastropod females between 2003 and 2013 ...... 126 3.2.2 OTs concentrations in biota, sediment and water between 2003 and 2013 ...... 132 References ...... 142

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Chapter 3

Evolution of the worldwide pollution by TBT in the last decade

3.1 Global trend of TBT pollution monitoring in the last decade

3.1.1 General results on the bibliographic search

The bibliographic search performed in the current work retrieved 216 scientific articles published on the topic “TBT pollution environmental monitoring” between 2003 and 2013, of which 161 applied imposex as a biomarker of TBT pollution, and 55 performed TBT pollution chemical monitoring (chemical analysis of the matrices water and/or sediment).

3.1.1.1 Number of studies published between 2003 and 2013

The temporal evolution of the number of studies published on the topic “TBT pollution environmental monitoring” from 2003 to 2013 is presented in Figure 3.1.

Figure 3.1: Number of publications on TBT pollution environmental monitoring using imposex as a biomarker from 2003 to 2013.

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An evident peak in the number of publications in 2008 is perceived (see Figure 3.1). This high number of publications in 2008 is coincident with the AFS Convention entry into force (i.e. the worldwide ban on the use of TBT-based AF paints). This fact reveals the greater interest of the scientific community in creating baselines on TBT pollution levels at the precise moment when the AFS Convention entered into force. From 2008 onwards, the number of papers published has gradually decreased.

3.1.1.2 Worldwide distribution of the studies published

Most of the studies published in this topic between 2003 and 2013 were performed in Europe (37%), followed by America (27%), Asia (24%), Africa (9%) and Oceania (3%), (Figure 3.2).

Figure 3.2: Percentage of articles on the topic “TBT environmental pollution monitoring” by continent, during the period from 2003 to 2013.

Several hypotheses for this distribution can be raised as, for instance, the shoreline extension. America is the continent with the largest coastline, followed by Asia, Africa, Europe, and at last, with a substantial difference to the extension of the European coastline, Oceania (Atlas Geográfico Mundial, 1995). While this could be considered the reason for the very low percentage of publications registered in Oceania between 2003 and 2013, it does not justify the highest percentage of studies in Europe, since the

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America, Asia and Africa have longer shorelines. Thus, the number of publications is not proportional to the shoreline extension and it seems more plausible to consider that a combination of factors is influencing the worldwide distribution of publications on TBT pollution monitoring using imposex as a biomarker from 2003 to 2013. Africa is the continent with the highest percentage of least developed countries (UNDP, 2014), with a manifest very poor scientific infrastructure and where the protection of coastal areas and water bodies’ policies are virtually nonexistent. So it seems obvious that the low number of scientific publications reporting imposex levels in this continent might be attributed to these factors and is not related with the continent shoreline extension. Regarding the insignificant number of studies using imposex in gastropods as a biomarker of TBT pollution in Africa, some authors have recently suggested that more species should be investigated along the African coastline since is a geographical area where there is a significant shipping activity and thus prone to TBT pollution impacts (Titley-O’Neal et al., 2011). Europe, America and Asia have similar and more relevant scientific production record in this topic. Despite the location of some of the least developed countries in Asia, some of the most developed economies are also located in this continent (e.g. Japan), which is patent by the higher scientific production in Asia than in Africa. Harbors, dockyards and marinas, generally located in coastal (marine or/and estuarine) environments, have been described as the main sources of TBT pollution worldwide (Berge et al., 1997; Green et al., 2001). Thus, despite several reasons might be seen as hypotheses for the ≈10% difference on the number of publications in this topic between America and Europe, one cannot rule out that the number of countries in the shore (29 in Europe and 20 in America, 5 in North America and 15 in South America), (Atlas Geográfico Mundial, 1995) can certainly imply a higher number of TBT sources requiring pollution monitoring, and thus a higher number of studies in Europe. The distribution of the published studies by country for each continent is shown in Figure 3.3. In Europe, the majority of the studies were performed in Portugal (29%), followed by Spain (18%), Italy (12%), France (11%), Croatia (7%), Iceland (5%), Ireland (4%), Germany (4%), England (4%), Netherlands (1%), Malta (1%), Denmark (1%), Sweden

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(1%), Romania (1%) and Poland (1%). In America, most of the studies were performed in Brazil (45%), followed by the USA (12%, of which 4% in Virginia State, 4% in Florida, 2% in South Carolina and 2% in Texas), Argentine (9%), Canada (7%), Greenland (4%), Peru (4%), Alaska (4%), Venezuela (4%), Chile (4%), Ecuador (2%), Mexico (2%), Panama (2%), the Virgin Islands (2%) and Costa Rica (2%). In Asia, most of the studies published in this topic between 2003 and 2013 were performed in Japan (23%), followed by China (21%), Korea (13%), India (12%), Taiwan (11%), Malaysia (10%), Thailand (4%), Pakistan (2%), Philippines (2%) and Iran (2%). In Africa, most of the studies were performed in Tunisia (84%), and then in Morocco (11%) and South Africa (5%), while in Oceania, 100% of the studies were performed in Australia (86% in the country shore and 14% in the Great Barrier Reef).

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3.1.2 Species used as bioindicators of imposex from 2003 to 2013

During the period under study, 96 species were used worldwide as bioindicators of imposex for TBT pollution monitoring. The 5 most used bioindicators were Nucella lapillus (11% of the studies analysed); Hexaplex trunculus (10%); Nassarius reticulatus (9%); Stramonita haemastoma (8%) and Thais clavigera (6%) (see Figure 3.4).

Thais clavigera Nucella lapillus 6% 11% Stramonita Haemastoma 8% Nassarius reticulatus 9% Hexaplex trunculus 10%

Figure 3.4: Bioindicators of imposex used worldwide for TBT pollution monitoring between 2003 and 2013.

Generally, Nucella lapillus and Nassarius reticulatus were the bioindicators most used in Europe, Hexaplex trunculus was the most frequently used in Africa, but was also often used in Europe, while Stramonita haemastoma was the most widely used in America, and Thais clavigera in Asia.

The choice of the bioindicator is related with several factors and the species distribution (i.e. availability) and abundance at the site being monitored are the most relevant ones.

Nucella lapillus is the most frequently used bioindicator of imposex levels worldwide. This species is widely distributed along the North Atlantic rocky shores, from Iceland in

113 the north, to Portugal in the south, and includes the Irish Sea and the North Sea coasts. This species is also the bioindicator recommended by the Oslo and Paris Commission (OSPAR) and by the Joint Assessment Monitoring Program (JAMP) guidelines to monitor imposex levels in the Northeast Atlantic (OSPAR Agreement 2008–09), being this the most probable reason why it was the bioindicator most frequently used in Europe and, because this was the region of the world where the highest number of studies in TBT pollution monitoring were published between 2003 and 2013 (i.e. as most of the studies published in this topic and period were in Europe) it was expected that Nucella lapillus would be the most frequently used bioindicator of imposex levels for TBT pollution monitoring globally.

A total of 12 different bioindicators of imposex were used in Europe from 2003 to 2013 (see Table 3.1). The three bioindicator species most used were Nucella lapillus, Nassarius reticulatus and Hexaplex trunculus. Nassarius reticulatus is also recommended by OSPAR to monitor the biological effects of this kind of pollution in estuaries inner areas and in southern locations where N. lapillus is absent due to its geographical distribution (OSPAR Agreement 2008–09), and so was also frequently used in Europe. Hexaplex trunculus was the third most used species. It is also widely distributed in Europe, namely around France, Italy, and in the Mediterranean Sea (Bouchet, Houart and Gofas, 2014).

Table 3.1: Bioindicator species of imposex used in studies published from 2003 to 2013 in Europe. Bioindicator species of imposex used in Europe % of studies Nucella lapillus 34.3 Nassarius reticulatus 30.1 Hexaplex trunculus 13.7 Bolinus brandaris 4.1 Nassarius nitidus 4.1 Buccinum undatum 2.7 Hydrobia ulvae 2.7 Thais haemastoma 2.7 Bembicium auratum 1.4 Cronia margariticola 1.4 Nassarius incrassatus 1.4 Rapana venosa 1.4

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In America, 44 different bioindicators were used (see Table 3.2). The three most used species were Stramonita haemastoma, Thais rustica and Nassarius vibex, mainly due to their wide distribution along the American coastline.

Table 3.2: Bioindicator species of imposex used in studies published from 2003 to 2013 in America. Bioindicator species of imposex used in America % of studies Stramonita haemastoma 21.5 Thais rustica 6.3 Nassarius vibex 5.1 Leucosonia nasa 3.8 Nucella lima 3.8 Acanthina monodon 2.5 Adelomelon brasiliana 2.5 Chicoreus brevifrons 2.5 Odontocymbiola magellanica 2.5 Rapana venosa 2.5 Stramonita rustica 2.5 Thais brevidentata 2.5 Thais deltoidea 2.5 Adelomelon ancilla 1.3 Adelomelon beckii 1.3 Adelomelon ferussacii 1.3 Bostrycapulus calyptreaformis 1.3 Buccinanops cochlidium 1.3 Buccinanops globulosum 1.3 Buccinanops globulosus 1.3 Buccinanops squalidum 1.3 Buccinum finmarkianum 1.3 Buccinum undatum 1.3 Cerithium lutosum 1.3 Crepidula aculeata 1.3 Echinolittorina ziczac 1.3 Nassarius coppingeri 1.3 Nucella emarginata 1.3 Nucella lamellosa 1.3 Nucella lapillus 1.3 Pareuthria plumbea 1.3 Phylonotus margaritensis 1.3 Pisania auritula 1.3 Prunum martini 1.3 Pugilina morio 1.3 Purpura patula 1.3 Searlesia dira 1.3 Thais biserialis 1.3 Thais chocolata 1.3 Thais kiosquiformis 1.3 Trophon geversianus 1.3 Voluta ebraea 1.3 Xanthochorus buxea 1.3 Ximenopsis muriciformis 1.3

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Stramonita haemastoma is widely distributed along the coast of several countries in South America and the Caribbean Sea (Houart and Gofas, 2014), the geographical areas where most of the studies published in the American continent were performed. It is also important to mention that imposex monitoring has increased substantially in recent years in Latin America (Central America, South America and the islands of the Caribbean), (Titley-O’Neal et al., 2011), and that the majority of these studies have mostly used Stramonita haemastoma as bioindicator. The other two species frequently used in America are also widely distributed in this geographical area: Thais rustica in coastal areas of several countries in South America (Brazil, Mexico, Peru, Venezuela, Ecuador, Panama and Argentine) and in the Caribbean Sea (Houart, 2014); Nassarius vibex in Colombia, Costa Rica, Cuba, Jamaica, Panama, Venezuela, around the Caribbean Sea and the Gulf of Mexico (Rosenberg, 2014).

Regarding Asia, the number of different bioindicator species of imposex used between 2003 and 2013 was 33 (see Table 3.3). The most used species was Thais clavigera. This is a well-established bioindicator of imposex for TBT pollution monitoring in Asia, and had inclusively been validated by imposex induction under laboratory conditions by exposure to different organotin compounds (Horiguchi et al., 1995; Horiguchi et al., 1997). Beyond this reason, this species is also widely distributed in Asia: it can be found on rocky shores around Malaysia, Singapore, Japan and in Indo-China region (Horiguchi et al., 1995; Horiguchi et al., 1997).

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Table 3.3: Bioindicator species of imposex used in studies published from 2003 to 2013 in Asia. Bioindicator species of imposex used in Asia % of studies Thais clavigera 26.9 Babylonia areolata 3.9 Cantharus cecillei 3.9 Strombus canarium 3.9 Thais bitubercularis 3.9 Thais gradata 3.9 Thais tuberosa 3.9 Babylonia spira 1.9 Babylonia japonica 1.9 Gyrineum natator 1.9 Hemifusus ternatanus 1.9 Lataxiena blosvillei 1.9 Littorina brevicula 1.9 Littorina mandshurica 1.9 Mauritia arabica 1.9 Morula musiva 1.9 Murex altispira 1.9 Murex occa 1.9 Murex trapa 1.9 Nassarius jacksonianus 1.9 Nassarius siquijorensis 1.9 Nassarius stolatus 1.9 Neptunea arthritica 1.9 Nucella heyseana 1.9 Pomacea canaliculata 1.9 Pugilina cochlidium 1.9 Rapana venosa 1.9 Semiricinula muricoides 1.9 Thais distinguenda 1.9 Thais hippocastanum 1.9 Thais lacera 1.9 Thais luteostoma 1.9 Turricula javana 1.9

In Africa, the 11 different species used are listed in Table 3.4. The three bioindicator species most used in TBT pollution biomonitoring studies in Africa were Hexaplex trunculus, Bolinus brandaris and Stramonita haemastoma. Hexaplex trunculus is found in the Mediterranean Sea and along the Atlantic coasts of Europe and Africa (Bouchet, Houart and Gofas, 2014). Imposex in this species was described and validated in 1988 (Martoja et al., 1988) and it has been the most used bioindicator where most of the studies were performed in Africa, i.e. in the northern coast (Rilov et al., 2000). Bolinus brandaris is found in the Mediterranean Sea - Western

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Basin (Houart and Gofas, 2014) and Stramonita haemastoma is found along the Atlantic coast of Africa (Houart, Gofas 2014).

Table 3.4: Bioindicator species of imposex used in studies on the topic under analysis in Africa from 2003 to 2013. Bioindicator species of imposex used in Africa % of studies Hexaplex trunculus 48.2 Bolinus brandaris 14.8 Stramonita haemastoma 7.4 Conus mediterraneus 3.7 Cyclope neritea 3.7 Murex brandaris 3.7 Murex trunculus 3.7 Nassarius kraussianus 3.7 Nassarius mutabilis 3.7 Nassarius nitidus 3.7 Thais haemastoma 3.7

Finally, in Oceania were used 5 different species (Table 3.5). Bembicium auratum is endemic to Australia, being distributed in both central and southern regions of the Great Barrier Reef, through the southeastern and south coasts of the country, in the Bass Strait (including the entire Tasmanian shoreline) up to Perth in the western shore. Cymbiola nobilis and Melo melo are common in the Southeast Asia, from Thailand and Malaysia, to the South China Sea and the Philippines (National Parks Board, 2014; Poutiers, 1998), being also present in the North coast of Australia (Swennen and Horpet, 2008). Morula marginalba is distributed along the north and east coasts of Australia and around the islands in the central Indo-Pacific Ocean. In Australia its range extends from the northwest tip to Twofold Bay in New South Wales. It is common on rocks in the intertidal zone (Davey, 2000). Thais orbita is abundant in the intertidal and sub littoral zone, occurring on rocks and among seaweed around the coasts of Australia, New Zealand and Lord Howe Island (Hardy and Eddie 2011; Grove and Simon, 2012).

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Table 3.5: Bioindicator species of imposex used in studies published from 2003 to 2013 in Oceania. Bioindicator species of imposex used in Oceania % of studies Bembicium auratum 20.0 Cymbiola nobilis 20.0 Melo melo 20.0 Morula marginalba 20.0 Thais orbita 20.0

Despite of, in Europe, more studies have been published than in the other continents, the bioindicators used are not very diverse (12, see Table 3.1) as those used in America (44, Table 3.2) and Asia (33, Table 3.3), where less studies were published. There is greater diversity of bioindicators in America and Asia compared to Europe, probably because America and Asia are continents with greater coastal zone and with different climatic regions (Titley-O’Neal et al., 2011). Each climatic region in America and in Asia has its dominant gastropod species, i.e. different species in different areas with different climates, to which each species is adapted (Titley-O’Neal et al., 2011) resulting in a greater diversity of bioindicators of imposex for TBT pollution monitoring.

3.1.2.1 By taxonomic group

The species used as bioindicators in Europe are caenogastropods (Subclass Caenogastropoda) of two different orders, Neogastropoda (n=10) and Littorinimorpha (n=2), and 5 different families: Muricidae (n=8), Nassariidae (n=3), Buccinidae (n=1), Hydrobiidae (n=1) and Littorinidae (n=1), (see Figure 3.5).

Figure 3.5: Number of species used as bioindicators of imposex, in studies published in Europe from 2003 to 2013, by Family.

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The species used as bioindicators in America belong to 3 different caenogastropod orders – Neogastropoda (n=38), Littorinimorpha (n=4), and Caenogastropoda (n=2) – and 10 different families: Muricidae (n=20), Nassariidae (n=6), Buccinidae (n=5), Volutidae (n=5), Calyptraeidae (n=3), Fasciolariidae (n=1), Marginellidae (n=1), Littorinidae (n=1), Cerithiidae (n=1) and Melongenidae (n=1), (Figure 3.6).

Figure 3.6: Number of species used as bioindicators of imposex, in studies published in America from 2003 to 2013, by Family.

The gastropods used as bioindicators of imposex in Asia are caenogastropods of 3 different orders – Neogastropoda (n=27), Littorinimorpha (n=5), and Caenogastropoda (n=1) – and 11 different families: Muricidae (n=16), Babyloniidae (n=3), Buccinidae (n=3), Nassariidae (n=3), Littorinidae (n=2), Melongenidae, (n=2), Ampullariidae (n=1), Cypraeidae (n=1), Ranellidae (n=1), Strombidae (n=1), and Turridae (n=1), (see Figure 3.7).

Figure 3.7: Number of species used as bioindicators of imposex, in studies published in Asia from 2003 to 2013, by Family.

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In Africa, the bioindicators used are all neogastropods but of 3 different families: Muricidae (n=6), Nassariidae (n=4), and Conidae (n=1), (Figure 3.8).

Figure 3.8: Number of species used as bioindicators of imposex, in studies published in Africa from 2003 to 2013, by Family.

The gastropods used as bioindicators of imposex in Oceania are caenogastropods of 2 different orders – Neogastropoda (n=4) and Littorinimorpha (n=1) – and of 3 different families: Muricidae (n=2), Volutidae (n=2), and Littorinidae (n=1), (Figure 3.9).

Figure 3.9: Number of species used as bioindicators of imposex, in studies published in Oceania from 2003 to 2013, by Family.

Regardless of the geographic region where the studies were performed, and despite the wide diversity of species used, the bioindicators were all caenogastropods (Subclass Caenogastropoda) and most frequently of Order Neogastropoda and Littorinimorpha, being the majority from family Muricidae.

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3.1.2.2 By geographic region

The distribution of the different bioindicators of imposex, applied in TBT pollution biomonitoring worldwide from 2003 to 2013, was then analysed by country. This analysis is here presented separately for each continent.

EUROPE The number of different bioindicator species, belonging to a certain number of different orders and families, used in European monitoring campaigns of imposex levels and which results were published between 2003 and 2013, is given in Table 3.6.

Table 3.6: Number different species (Ns) used as bioindicators of imposex levels for TBT pollution monitoring between 2003 and 2013 in Europe, indicated per country, Order and Family.

Country Ns Order Ns per Order Family Ns per Family Portugal 6 Neogastropoda 4 Nassariidae 2 Muricidae 2 Littorinimorpha 2 Littorinidae 1 Hydrobiidae 1 Italy 4 Neogastropoda 4 Muricidae 3 Nassaridae 1 Spain 4 Neogastropoda 4 Nassariidae 2 Muricidae 2 France 3 Neogastropoda 3 Muricidae 2 Nassaridae 1 Germany 2 Neogastropoda 2 Nassaridae 1 Muricidae 1 Iceland 2 Neogastropoda 2 Muricidae 2 Croatia 1 Neogastropoda 1 Muricidae 1 Denmark 1 Neogastropoda 1 Buccinidae 1 England 1 Neogastropoda 1 Muricidae 1 Irland 1 Neogastropoda 1 Muricidae 1 Malta 1 Neogastropoda 1 Muricidae 1 Romania 1 Neogastropoda 1 Muricidae 1 Sweden 1 Neogastropoda 1 Buccinidae 1

Portugal was the country where more different species (6) were used as bioindicators of imposex in Europe, and the only country where others than neogastropods were applied.

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AMERICA The number of different bioindicator species used in the American countries where imposex levels biomonitoring was performed and published from 2003 to 2013 is given in Table 3.7.

Table 3.7: Number different species (Ns) used as bioindicators of imposex levels for TBT pollution monitoring between 2003 and 2013 in America, indicated per country, respective Orders and Families.

Country Ns Order Ns per Order Family Ns per Family Argentine 13 Neogastropoda 12 Volutidae 4 Buccinidae 3 Nassariidae 2 Muricidae 2 Marginellidae 1 Littorinimorpha 1 Calyptraeidae 1 Brazil 9 Neogastropoda 9 Muricidae 4 Volutidae 1 Fasciolariidae 1 Buccinidae 1 Nassariidae 1 Melongenidae 1 Canada 6 Neogastropoda 6 Muricidae 4 Buccinidae 2 Venezuela 3 Neogastropoda 3 Muricidae 2 Fasciolariidae 1 Virgin Islands 3 Neogastropoda 3 Muricidae 3 Mexico 3 Neogastropoda 1 Muricidae 1 Littorinimorpha 1 Littorinidae 1 Caenogastropoda 1 Cerithiidae 1 Ecuador 3 Neogastropoda 3 Muricidae 3 Chile 2 Neogastropoda 2 Muricidae 1 Nassariidae 1 Peru 2 Neogastropoda 2 Muricidae 2 Texas 1 Neogastropoda 1 Muricidae 1 Pensacola 1 Neogastropoda 1 Muricidae 1 Florida 1 Neogastropoda 1 Muricidae 1 Greenland 1 Neogastropoda 1 Buccinidae 1 Costa Rica 1 Neogastropoda 1 Muricidae 1 Virginia 1 Neogastropoda 1 Muricidae 1 Panama 2 Neogastropoda 1 Muricidae 1 Littorinimorpha 1 Calyptraeidae 1 Alaska 1 Neogastropoda 1 Muricidae 1

Argentine and Brazil were the countries where a high diversity of bioindicators were used for imposex levels assessment in America.

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ASIA The number of different bioindicator species, belonging to a certain number of different orders and families, used in the Asian countries where imposex levels biomonitoring was performed and published from 2003 to 2013 is given in Table 3.8. Thailand was the country where more different species were used in Asia.

Country Ns Order Ns per Order Family Ns per Family Thailand 15 Neogastropoda 15 Muricidae 8 Nassariidae 3 Melongenidae 2 Buccinidae 1 Turridae 1 Japan 6 Neogastropoda 4 Muricidae 2 Babyloniidae 1 Buccinidae 1 Littorinimorpha 2 Littorinidae 2 China 5 Neogastropoda 4 Muricidae 2 Buccinidae 1 Babyloniidae 1 Littorinimorpha 1 Cypraeidae 1 Malaysia 5 Neogastropoda 4 Muricidae 4 Littorinimorpha 1 Strombidae 1 Taiwan 3 Neogastropoda 2 Muricidae 2 Caenogastropoda 1 Ampullariidae 1 Hong Kong 2 Neogastropoda 2 Muricidae 2 India 1 Littorinimorpha 1 Ranellidae 1 Korea 1 Neogastropoda 1 Muricidae 1 Pakistan 1 Neogastropoda 1 Babyloniidae 1

AFRICA The number of different bioindicators of imposex between 2003 and 2013 used in TBT pollution monitoring studies in Africais given in Table 3.9. Tunisia was the country where more different species were.

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Table 3.9: Number different species (Ns) of bioindicators of imposex levels for TBT pollution biomonitoring between 2003 and 2013 in Africa, indicated per country, respective Orders and Families.

Country Ns Order Ns per Order Family Ns per Family Tunisia 8 Neogastropoda 7 Muricidae 3 Nassariidae 3

Conidae 1

Morocco 4 Neogastropoda 4 Muricidae 4 South Africa 1 Neogastropoda 1 Nassariidae 1

OCEANIA The number of different bioindicators applied in Oceania during the period under study is given in Table 3.10.

Country Ns Order Ns per Order Family Ns per Family Australia 5 Neogastropoda 4 Muricidae 2 Volutidae 2 Littorinimorpha 1 Littorinidae 1

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3.2 TBT pollution evolution from 2003 to 2013

3.2.1 Imposex levels in gastropod females between 2003 and 2013

Several indices have been used to assess imposex expression in gastropods (see review by Titley O’Neil et al., 2011). Among them, the most commonly used were the percentage of imposex-affected females (%I), the vas deference sequence index (VDSI), the female penis length (FPL), the relative penis length (or size) index (RPLI or RPSI, respectively) depending on the bioindicator species (see explanation on the difference between these later indices, and when to use each of them, in Chapter 1), and the percentage of sterile females (%S). In the current work, these were the indices considered to study the evolution of the imposex levels between 2003 and 2013 worldwide. The studies published during this period were analysed and the evolution of the imposex levels reported were followed by geographic region (continent and respective countries).

EUROPE Imposex levels registered in Europe between 2003 and 2013 are summarized in Table 3.11. The table is sorted by the sampling year ascending order and, from then, by the bioindicator species used in each study and the respective country where the study took place. Imposex indices evolution reported in each publication analysed is given by the colored arrows ↓ (decreasing trend), ↔ (levels maintenance) and ↑ (increasing trend) for each imposex index calculated in the respective study.

Generally, a decreasing trend is perceived from 2003 to 2013 in Europe, although increases of some of the imposex indices were still registered until 2008, year from which decreases in all the indices were consistently reported. This trend could be justified by the entry into force of the AFS Convention, which only happened in September 2008. However, the European Union adopted the Regulation No. 782/2003 that anticipated the AFS Convention entry into force to the 1st July 2003 in EU member states, i.e., prohibited

126 the application or reapplication of TBT coatings on EU member states’ national mercantile fleets, and on ships operating under their authority, from the 1st July 2003. Therefore, it would be expected that consistent decreases would be registered from 2003-2004 onwards but, while some decreases were registered since then (VDSI, FPL and RPSI in Nucella lapillus in 2003 in Portugal, France, England and Iceland; VDSI, FPL and RPLI in Nassarius reticulatus in 2005 and 2006 in Portugal), the consistent decreasing trend in imposex intensity, and thus in TBT pollution, was only observed from 2008 to 2013.

Sampling period Sampling year Specie Country %I VDSI FPL RPLI RPSI %S 2003 Hydrobia ulvae Portugal ↑ ↑ ↔ ↓ 2003 Nassarius reticulatus Portugal ↑ ↑ ↑ 2003 Nucella lapillus France ↓ ↓ 2003 Nucella lapillus England ↑ ↓ 2003 Nucella lapillus Portugal ↑ ↓ ↓ ↓ June 2003 Nucella lapillus Iceland ↓ ↓ 2003 Nucella lapillus Iceland ↓ ↓ September 2004 Hexaplex trunculus Italy ↑ October-December 2004 Nassarius nitidus Italy ↓ ↑ ↑ 2004 Nassarius reticulatus France ↔ 2005 Nassarius reticulatus Portugal ↓ ↑↓ ↓ 2005 Nassarius reticulatus Portugal ↓ May-July 2005 Hexaplex trunculus Croatia May -June 2005 Nassarius reticulatus Portugal ↓ ↓ ↓ ↓ August 2005 Nucella lapillus Portugal ↑ ↑ ↑ June -October 2006 Nassarius reticulatus Portugal ↓ ↓ ↓ ↓ February - March 2006 Nucella lapillus Irland ↓ December - June 2007 Nucella lapillus Portugal ↑ ↓ ↓ ↓ February–June 2007 Nassarius reticulatus Spain ↑ ↑ ↑ ↑ April 2007 Nucella lapillus Irland ↑ ↑ 2008 Nucella lapillus Portugal ↓ March 2008 Nassarius reticulatus Portugal ↓ 2008 Nucella lapillus France ↓ ↓ March 2008 Nucella lapillus Portugal ↓ August 2008 Rapana venosa Romania ↓ 2008 Nucella lapillus Iceland ↓ ↓ October - September 2009 Bolinus brandaris Portugal ↓ ↓ ↓ August 2009 Nassarius reticulatus Portugal ↓ ↓ ↓ ↓ ↔ August 2009 Nucella lapillus Portugal ↓ ↓ ↓ ↓ 2009 Nucella lapillus France ↓ ↓ January 2009 Nucella lapillus Portugal ↓ August 2010 Nucella lapillus Portugal ↓ ↓ ↓ ↓ February 2011 Nucella lapillus France ↓ 2012 Hydrobia ulvae Portugal ↓ ↓

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AMERICA Given the American continent dimension and its heterogeneity, not only regarding the socio-economic environment but also its scientific output in this topic in the period under study, the results on the evolution of the imposex levels from 2003 to 2013 are given for North and South America separately (Tables 3.12 and 3.13, respectively).

In North America, all studies published between 2003 and 2013 revealed a consistent decrease of imposex levels, and thus in TBT pollution.

Sampling period Sampling year Specie Country %I VDSI FPL RPSI June-July 2005 Nucella lima Alaska ↓ June-July (2007) -June (2008) 2008 Nucella lapillus Canada ↓ ↓ ↓ ↓ 2009 Nucella lima Alaska ↓ ↓

However, the trend was totally different in South America. Only sporadic decreases were registered from 2004 to 2008, namely in Argentine, Chile and Mexico. An irregular pattern of decreases and increases were reported in Brazil, depending on the region and on the bioindicator species applied. In the Virgin Islands and Peru, countries that have not ratified the AFS Convention still today, consistent increases were registered during the period under study. In fact, most of the South American countries had not ratified the AFS Convention by 2008 not being subjected to any restriction to the use of TBT compounds. From 2008 onwards, after the ratification process was complete, some decreases were finally registered in Brazil and Ecuador

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Sampling period Sampling year Specie Country %I VDSI FPL RPLI RPSI 2003 Stramonita haemastoma Brazil ↑ 2004 Odontocymbiola magellanica Argentine ↑ ↑ ↑ 2004 Stramonita haemastoma Brazil ↑ ↑ September–November 2004 Adelomelon beckii Argentine ↓ ↓ July 2004 Chicoreus brevifrons Venezuela ↑ July 2004 Stramonita haemastoma Brazil ↑ ↑ ↑ February- November 2004 Stramonita haemastoma Brazil ↑ ↑ ↑ ↑ November 2004 Stramonita rustica Brazil ↓ ↓ ↓ ↓ June 2005 Acanthina monodon Chile ↓ ↓ February - April 2005 Bostrycapulus calyptreaformis Panama ↑ 2005 Nassarius vibex Mexico ↓ ↓ 2005 Stramonita haemastom Brazil ↑ ↑ 2006 Stramonita haemastoma Brazil ↔ ↓ ↓ ↓ 2007 Nassarius vibex Brazil ↑ ↑ ↑ ↑ March 2007 Purpura patula Virgin Islands ↑ ↑ March 2007 Thais deltoidea Virgin Islands ↑ ↑ March 2007 Thais rustica Virgin Islands ↑ ↑ 2008 Adelomelon brasiliana Argentine ↑ ↑ 2008 Stramonita haemastoma Brazil ↑ ↑ ↑ 2008 Nassarius vibex Brazil ↓ ↓ February -May 2008 Pugilina morio Brazil ↑ 2008 Stramonita haemastoma Brazil ↓ September 2009 Thais biserialis Ecuador ↓ ↓ ↓ ↓ October 2009 Thais chocolata Peru ↑ ↑ ↑ October 2012 Xanthochorus buxea Peru ↓ ↓

ASIA By the visual analysis of Table 3.14 it is possible to observe the consistent decrease in all imposex indices from 2008 onwards, reinforcing the success of the AFS Convention implementation. In 2003 and 2004, increases were reported in Japan, China and Malaysia, while the decreasing trend was reported in Korea as early as 2004 probably as a result of the legislative action MOE Regulation 154/2000, restricting the use of TBT compounds in this country (Choi et al. 2010).

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Sampling period Sampling year Specie Country %I VDSI FPL RPLI RPSI %S January 2003 Thais clavigera Japan ↑ June-July 2004 Thais clavigera China ↑ ↑ ↑ March-May 2004 Thais clavigera Korea ↓ ↓ November - April 2004 Thais gradata Malaysia ↑ ↑ December 2005 Thais clavigera China ↓ ↓ ↓ ↓ June -November 2006 Murex trapa Thailand ↑ June 2006 Thais clavigera China ↑ ↑ ↑ ↑ 2007 Strombus canarium Malaysia ↓ September 2007 Thais gradata Malaysia ↑ ↑ 2008 Babylonia spira Pakistan ↓ ↓ ↓ ↓ 2010 Thais clavigera Japan ↓ ↓ June and July from 2010 Thais clavigera Japan ↓ ↓ 2010 Thais tuberosa Malaysia ↓ ↓ April 2011 Rapana venosa China ↓ ↓

AFRICA The evolution of imposex levels from 2003 to 2013 in Africa is presented in Table 3.15.

Sampling period Sampling year Specie Country %I VDSI FPL RPLI %S May–June 2004 Hexaplex trunculus Tunisia ↑ ↑ ↑ ↑ May–June 2004 Hexaplex trunculus Tunisia ↓ ↑ ↓ September 2004 Hexaplex trunculus Tunisia ↑ ↑ 2004 Hexaplex trunculus Tunisia ↔ ↓ July 2004 Hexaplex trunculus Tunisia ↑ ↑ ↑ 2007 Bolinus brandaris Tunisia ↓ ↓ ↓ February- December 2007 Bolinus brandaris Tunisia ↓ ↓ ↓ ↓ September 2007 Nassarius mutabilis Tunisia ↓ ↓ ↓ ↓ January 2008 Hexaplex trunculus Tunisia ↑ January 2009 Hexaplex trunculus Tunisia ↓ ↓ May 2010 Bolinus brandaris Tunisia ↓ ↓

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Other three studies were published in Africa between 2003 and 2013: (i) imposex levels in Murex trunculus, Murex brandaris and Thais haemastoma were reported in Morocco in 2000-2001 by Lemghich and Benajiba (2007), and in Stramonita haemastoma in 2005-2006 by Mortaji et al. (2011); (ii) another study reporting imposex levels in Nassarius kraussianus in 2002 in South African harbors was published in 2003 by Marshall and Rajkumar. Even so, these publications only reported levels of imposex in isolated campaigns and any of them mention the temporal evolution of any index, reason why they are not included in Table 3.15. A consistent decrease in imposex levels from 2007 onwards is perceived in Tunisia by the analysis of Table 3.15. Imposex in Hexaplex trunculus revealed successive increases until 2006, and even in FPL in 2008, decreasing in 2009. However, the use of different bioindicators revealed decreases in imposex levels starting in 2007.

OCEANIA In Oceania (see Table 3.16), Andersen (2004) has described a decrease in imposex levels (namely in the percentage of imposex-affected females, %I) since 2003 using Morula marginalba as a bioindicator. However, in 2005 the phenomenon was still affecting around 90% of Bembicium auratum females near TBT main sources in Australia (Wilson and Roach, 2009). Nevertheless, imposex levels in this species have declined rapidly and no signs of severe impacts were registered in populations since sterile females were never registered (Wilson and Roach, 2009).

Table 3.16: Evolution of the imposex levels in gastropod females reported in Oceania from 2003 to 2013. The sampling year, the bioindicator species used, the country where the study took place, and the trend of several imposex indices (%I, VDSI, FPL, RPLI, RPSI and %S) indicated by colored arrows (↓, decreasing trend; ↔, levels maintenance; and ↑, increasing trend) are presented by the sampling year ascending order.

Sampling period Sampling year Specie Country VDSI %I October 2003 Morula marginalba Australia ↓ 2005 Bembicium auratum Australia ↓

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3.2.2 OTs concentrations in biota, sediment and water between 2003 and 2013

The evolution of OTs concentrations in females’ tissues, sediments and water reported worldwide between 2003 and 2013 are here presented by country and for each continent separately. OTs concentrations in the tissues of females accompany some of the data of imposex, that are analyzed the tissues of females to correlate these values with imposex levels displayed by females and validate like this Imposex using the biomarker, and that therefore these results of OTs concentrations in tissues result from the analysis of 161 articles of biomonitoring. The data of OTs concentrations in sediments and water, resulting from 161 articles of biomonitoring of the imposex and of the 55 articles of chemical monitoring.

EUROPE The evolution of OTs females’ body burdens from 2003 to 2013 in Europe, indicated per country and per bioindicator species, is given in Table 3.17 while OTs concentrations trends in sediments and water samples are presented in Tables 3.18 and 3.19, respectively.

Generally, OTs females’ body burdens decreased in all bioindicator species from 2008 onwards. However, some species were already revealing some reduced OTs content earlier, since 2003 (e.g. Nucella lapillus in England, France and Spain in 2003; Nassarius reticulatus in Portugal in 2005 and 2006). As a matter of fact, the AFS Convention was adopted in 2001 and opened for signature in 2002, with the purpose of restricting TBT- based antifouling systems on ships by 2003. However, the ratification process was long and delayed the document entry into force until September 2008 generating a phasing out period of these compounds during the years prior to the entry into force as some companies started to phase out their commercialization due to the upcoming ban (Sousa 2009). Thus, a declining trend is apparent from 2003 to 2007 but the definite reduction of OTs in gastropods tissues was observed from 2008 onwards.

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Sampling period Sampling year Specie Country TBT DBT MBT BDI 2003 Nucella lapillus England ↓ 2003 Nucella lapillus France ↓ late spring–summer 2003 2003 Nassarius nitidus Italy ↑ ↑ ↑ September 2003 Nassarius reticulatus Portugal ↑ 2003 Nassarius reticulatus Portugal ↑ July 2003 Nucella lapillus Spain ↓ ↑ Winter 2004 Hexaplex trunculus Croatia ↑ November -December 2004 Hexaplex trunculus Croatia ↑ 2004 Nassarius reticulatus France ↑ October-December 2004 Nassarius nitidus Italy ↓ ↓ ↓ ↑ 2005 Nassarius reticulatus Portugal ↓ ↓ ↓ May -June 2005 Nassarius reticulatus Portugal ↓ ↓ ↓ 2005 Nassarius reticulatus Portugal ↓ August 2005 Nucella lapillus Portugal ↑ February - March 2006 Nucella lapillus Irland ↓ June -October 2006 Nassarius reticulatus Portugal ↓ ↓ ↓ August 2006 Nassarius reticulatus Spain ↓ July–September 2006 Nucella lapillus Spain ↑ ↑ ↑ ↑ July 2006 - May 2007 2006 Nucella lapillus Iceland ↓ April 2007 Nucella lapillus Irland ↑ January–May 2007 Nassarius nitidus Portugal ↑ December - June 2007 Nucella lapillus Portugal ↓ February–June 2007 Nassarius reticulatus Spain ↑ ↑ January–May 2007 Nassarius reticulatus Spain ↑ 2008 Nucella lapillus France ↓ March 2008 Nassarius reticulatus Portugal ↓ March 2008 Nucella lapillus Portugal ↓ August 2008 Rapana venosa Romania ↓ February - March 2008 Nassarius reticulatus Spain ↓ ↓ ↓ ↑ May 2008 Nassarius reticulatus Spain ↓ 2009 Nucella lapillus France ↓ October - September 2009 Bolinus brandaris Portugal ↓ August 2009 Nassarius reticulatus Portugal ↓ ↓ ↓ February 2011 Nucella lapillus France ↓ 2012 Hydrobia ulvae Portugal ↓

The same trend is also observed in sediments and water OTs concentrations (see Tables 3.18 and 3.19, respectively). Increases in OTs concentrations in sediments were still reported in 2003 in Italy and in 2007 in Ireland, while decreases were registered in Portugal from 2005 and generally from 2007 onwards.

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Sampling period Sampling year Country TBT DBT MBT BDI late spring–summer 2003 Italy ↑ ↑ ↑ May -June 2005 Portugal ↓ ↓ ↓ April 2007 Irland ↑ 2007 Spain ↓ ↓ ↓ May 2008 Spain ↓ 2008 Spain ↓ ↑ 2008 Spain ↓ ↓ ↓ February 2009 Poland ↓ ↓ ↓ ↓ August 2009 Portugal ↓ ↓ ↓ May 2011 Germany ↓

Sampling period Sampling year Country TBT DBT MBT BDI May and August 2003 Spain ↑ late spring–summer 2004 Italy ↑ ↑ ↑ 2006 France ↑ July 2007 England ↑ 2007 Spain ↓ ↑ 2008 Spain ↓ ↑ February 2009 Poland ↓ ↓ ↓ ↓ 2010 Croatia ↓ May 2011 Germany ↓

From 2003 to 2007, increases in OTs concentrations in water were reported in Spain, Italy, France and England, and decreases, from 2007 onwards, in Spain, Poland, Croatia,

134 and Germany. Regarding the Butyltins Degradation Index (BDI), calculated by the formula [TBT]/([DBT]+[MBT]), recent inputs of TBT were still detected in 2007 and 2008 in Spain.

AMERICA The evolution of OTs females’ body burdens from 2003 to 2013, indicated per country and per bioindicator species, is given in Table 3.20 for North America and in Table 3.20 for South American countries.

Sampling period Sampling year Specie Country TBT DBT MBT 2005 Nucella lima Alaska ↓ June 2005 Nucella lima Alaska ↓ February -May 2008 Nucella lapillus Canada ↓ ↓

In North America, and corroborating the imposex levels in the same period and geographical region (see section 3.2.1), a decreasing trend was reported from 2005 to 2008 in Alaska and Canada (see Table 3.20). However, a completely different scenario was observed in South America. Generally, until 2008/2009, the only countries where concentrations of TBT decreased in females’ tissues were in Ecuador, some sites in Brazil and in Argentine, in 2003/2004. Beyond these exceptions, increases were extensively registered in numerous South American countries until 2008-2009, date from when OTs concentrations in gastropods begin to decrease in a consistent way. Increases in TBT concentrations from 2009 onwards were only registered in Thais chocolata tissues in Peru.

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Sampling period Sampling year Specie Country TBT DBT MBT BDI 2003 Nassarius vibex Brazil ↑ 2003 Stramonita haemastoma Brazil ↑ 2003 Thais biserialis Ecuador ↓ ↓ ↓ ↓ 2004 Stramonita haemastoma Brazil ↓ September–November 2004 Adelomelon beckii Argentine ↓ July 2004 Odontocymbiola magellanica Argentine ↑ July 2004 Pareuthria plumbea Argentine ↑ July 2004 Stramonita haemastoma Brazil ↑ July 2004 Acanthina monodon Chile ↓ November 2004 Bostrycapulus calyptreaformis Panama ↑ 2004 Xanthochorus buxea Peru ↑ February - April 2005 Buccinanops cochlidium Argentine ↑ 2005 Ximenopsis muriciformis Argentine ↑ June-July 2005 Stramonita haemastoma Brazil ↑ February 2006 Leucozonia nasa Venezuela ↑ 2007 Buccinanops globulosus Argentine ↑ September 2007 Prunum martini Argentine ↑ 2007 Trophon geversianus Argentine ↑ March 2007 Pugilina morio Brazil ↑ March 2007 Phylonotus margaritensis Venezuela ↑ March 2007 Purpura patula Virgin Islands ↑ 2008 Adelomelon ancilla Argentine ↑ 2008 Adelomelon brasiliana Argentine ↑ 2008 Adelomelon ferussacii Argentine ↑ 2008 Buccinanops globulosum Argentine ↓ 2008 Buccinanops squalidum Argentine ↑ 2008 Crepidula aculeata Argentine ↑ 2008 Odontocymbiola magellanica Argentine ↑ 2008 Stramonita haemastoma Brazil ↑ 2008 Stramonita haemastoma Brazil ↓ ↓ ↓ 2008 Stramonita rustica Brazil ↓ 2008 Thais deltoidea Virgin Islands ↑ 2008 Thais rustica Virgin Islands ↑ June-July (2007) -June (2008) 2008 Stramonita haemastoma Brazil ↑ 2008 Nassarius vibex Mexico ↓ 2008 Nassarius vibex Brazil ↓ 2009 Stramonita haemastoma Brazil ↓ September 2009 Thais chocolata Peru ↑ ↑ ↑ October 2009 Chicoreus brevifrons Venezuela ↓ October 2012 Thais deltoidea Virgin Islands ↓

In fact, Peru was an exception on the effectiveness of the global ban on TBT-based antifouling paints by the AFS Convention entry into force in 2008 at a moment when TBT pollution amelioration has being reported in Europe (Sousa et al., 2009; Gipperth, 2009), Asia (Choi et al., 2010), North Africa (Lahbib et al., 2009), North America (Wade et al., 2008), and other parts of South America (Oliveira et al., 2010; Castro et al., 2012).

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The evolution of the OTs concentrations in sediments, reported from 2003 to 2013 in South America, is given in Table 3.22. The authors reported increases in OTs concentrations sediments, in 2007 and 2008 in Brazil, and from 2008 onwards, decreases in Brazil and Ecuador.

Sampling period Sampling year Country TBT DBT MBT BDI October 2007 Brazil ↑ ↑ ↑ ↓ 2008 Brazil ↑ April 2008 Brazil ↓ ↓ ↓ ↓ July 2008 Brazil ↓ ↓ ↓ ↓ September 2009 Ecuador ↓ ↓ ↓ ↓

Despite the decreasing trend in OTs concentrations in sediments from 2008 onwards, fresh inputs of TBT were still registered, as indicated by the BDI>1 reported.

ASIA OTs concentrations quantified in gastropod females’ tissues in studies of TBT pollution biomonitoring using imposex as a biomarker from 2003 to 2013 in Asia are given in Table 3.23. Increases in TBT concentrations were reported for the majority of the bioindicators used in several countries from 2003 to 2007, with the exceptions of Rapana venosa and Thais clavigera in China in 2006, T. clavigera in Japan and Korea in 2006 and Strombus canarium and Thais tuberosa in Malaysia in 2006. From 2008 onwards consistent decreases in several bioindicators were reported in Pakistan, Thailand and Malaysia.

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Sampling period Sampling year Specie Country TBT DBT MBT BDI January 2003 Murex occa Thailand ↑ July 2004 Thais clavigera China ↑ ↑ ↑ June-July 2004 Babylonia areolata Thailand ↑ March-May 2004 Nassarius jacksonianus Thailand ↑ June-July 2004 Thais bitubercularis Thailand ↑ November 2001- April 2004 2004 Thais tuberosa Malaysia ↑ November 2001- April 2004 2004 Nassarius siquijorensis Thailand ↑ November 2001- April 2004 2004 Pugilina cochlidium Thailand ↑ November 2001- April 2004 2004 Thais lacera Thailand ↑ December 2005 Murex altispira Thailand ↑ 2005 Morula musiva Thailand ↑ June -November 2006 Rapana venosa China ↓ June -November 2006 Thais clavigera China ↑ ↑ ↑ June -November 2006 Thais clavigera Hong Kong ↑ June -November 2006 Thais clavigera Hong Kong ↑ June -November 2006 Thais clavigera Hong Kong ↓ ↓ ↓ ↑ June -November 2006 Babylonia japonica Japan ↑ June -November 2006 Thais clavigera Japan ↑ June -November 2006 Thais clavigera Japan ↓ June -November 2006 Thais clavigera Korea ↓ ↓ June -November 2006 Strombus canarium Malaysia ↓ June -November 2006 Thais bitubercularis Malaysia ↑ June -November 2006 Thais gradata Malaysia ↑ ↑ ↑ ↓ June -November 2006 Thais tuberosa Malaysia ↓ June 2006 Hemifusus ternatanus Thailand ↑ March 2006 Lataxiena blosvillei Thailand ↑ June -November 2006 Semiricinula muricoides Thailand ↑ June -November 2006 Turricula javana Thailand ↑ October 2005-January 2006 2006 Thais hippocastanum Malaysia ↑ September 2007 Nassarius stolatus Thailand ↑ 2008 Babylonia spira Pakistan ↓ 2010 Thais luteostoma Thailand ↓ June and July from 2003 to 2010 2010 Murex trapa Thailand ↓ April 2011 Thais gradata Malaysia ↓

Regarding sediments, concentrations of OTs for the same period and geographic region are given in Table Table 3.24. Increases of OTs concentrations were reported until 2007 in several countries, exception made to Thailand in 2004, and Japan, Korea and Manila in 2005 and Taiwan in 2006. OTs concentrations reductions in sediments were consistently reported from 2005 onwards in Korea, from 2007 in India, and from 2008 onwards in Taiwan and China. Even though, fresh inputs of TBT were still registered in India in 2008-2009 as revealed by the BDI.

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Sampling period Sampling year Country TBT DBT MBT BDI November 2003 Korea ↑ ↑ ↑ ↓ 2003 Taiwan ↑ March-May 2004 Korea ↑ ↑ ↑ ↓ 2004 Thailand ↓ 2005 Japan ↓ 2005 Korea ↓ January 2005 Manila ↓ 2005 Taiwan ↑ August 2006 India ↑ ↑ ↑ ↑ 2006 Malaysia ↑ ↑ ↑ ↓ May 2006 Taiwan ↓ ↓ ↑ ↓ March 2007 China ↑ ↑ ↑ 2007 China ↑ ↑ November 2007 India ↑ 2007 India ↑ ↑ ↑ ↓ 2007 Korea ↓ 2008 India ↓ ↓ ↓ ↑ January–February 2009 India ↓ ↓ ↓ ↑ February - May 2009 Korea ↓ ↓ January–February 2009 Taiwan ↓ April 2011 China ↓

Regarding the chemical analysis of water samples between 2003 and 2013, Asia is the most documented continent. Values of OTs concentrations in water are presented in Table 3.25 by sampling year and country where the study was performed. Decreases of OTs concentrations in water were registered from 2003 to 2010. Nevertheless, TBT fresh inputs were detected in India in 2008 and in China in 2010, and exceptional increases in TBT concentrations were reported in Japan in 2003, in China in 2006, in India in 2007 and in Korea in 2008.

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Sampling period Sampling year Country TBT DBT MBT BDI May -September 2003 China ↓ ↓ ↓ ↓ 2003 Japan ↓ 2003 Japan ↑ October 2004-February 2005 2005 Korea ↓ 2005 Korea ↓ 2005 Thailand ↓ 2005 Thailand ↓ March 2006 China ↑ ↑ ↑ 2006 Japan ↓ 2006 Thailand ↓ November 2007 India ↑ 2007 India ↑ ↑ ↑ ↓ 2008 India ↓ ↓ ↓ ↑ February - May 2008 Korea ↑ ↑ January–February 2009 Taiwan ↓ March 2010 China ↓ ↑

AFRICA OTs concentrations quantified in gastropod females’ tissues from 2003 to 2013 in African populations are given in Table 3.26. Generally, decreases in TBT concentrations in gastropods tissues were only observed in Tunisia from 2009 onwards. Increases were reported in 2004, 2006 and 2008 in Hexaplex trunculus in Tunisia, and in Stramonita haemastoma in Morocco in 2006 and 2007.

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Sampling period Sampling year Specie Country TBT May–June 2004 Hexaplex trunculus Tunisia ↓ March - July 2004 Hexaplex trunculus Tunisia ↑ September 2004 Hexaplex trunculus Tunisia ↑ 2004 Hexaplex trunculus Tunisia ↓ July 2004 Hexaplex trunculus Tunisia ↑ 2006 Stramonita haemastoma Morocco ↑ February 2006 Hexaplex trunculus Tunisia ↑ 2007 Bolinus brandaris Tunisia ↓ February- December 2007 Bolinus brandaris Tunisia ↓ September 2007 Conus mediterraneus Tunisia ↑ September 2007 Cyclope neritea Tunisia ↑ September 2007 Hexaplex trunculus Tunisia ↑ September 2007 Hexaplex trunculus Tunisia ↑ September 2007 Nassarius mutabilis Tunisia ↑ September 2007 Nassarius mutabilis Tunisia ↓ September 2007 Stramonita haemastoma Morocco ↑ January 2008 Hexaplex trunculus Tunisia ↑ January 2009 Hexaplex trunculus Tunisia ↓ May 2010 Bolinus brandaris Tunisia ↓ June 2010 Nassarius nitidus Tunisia ↓

Regarding sediments analyses, Lahbib and his team (2011) reported TBT concentrations increases in Tunisia from 2008 to 2010.

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Chapter 4: Final considerations

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Chapter 4

General conclusions ______Table of Contents 4.1 General conclusions ...... 167 References ...... 171

4.1 General conclusions

The current work pretends to be a contribution to describe the global trend of evaluation of TBT pollution in the last decade. There were a number maximum of publications about TBT pollution in 2008 that was coincident with the AFS Convention entry into force (i.e. the worldwide ban on the use of TBT-based AF paints). This fact reveals the greater interest of the scientific community in creating baselines on TBT pollution levels at the precise moment when the AFS Convention entered into force. From 2008 onwards, the number of papers published has gradually decreased.

Most of the studies about pollution by TBT published in this topic between 2003 and 2013 were performed in Europe (37%), followed by America (27%), Asia (24%), Africa (9%) and Oceania (3%).

During the period 2003-2013, 96 species were used worldwide as bioindicators of imposex for TBT pollution monitoring. Being the 5 bioindicators most used in this period: Nucella lapillus (11% of the studies analysed); Hexaplex trunculus (10%); Nassarius reticulatus (9%); Stramonita haemastoma (8%) and Thais clavigera (6%).

167 the species distribution (i.e. availability) and abundance at the site being monitored are the most relevant ones.

In Europe the three bioindicator species most used between 2003 and 2013 were Nucella lapillus, Nassarius reticulatus and Hexaplex trunculus; while in America the three most used species were Stramonita haemastoma, Thais rustica and Nassarius vibex; in Asia the most used species was Thais clavigera, this is a well-established bioindicator of imposex for TBT pollution monitoring in Asia, and had inclusively been validated by imposex induction under laboratory conditions by exposure to different organotin compounds (Horiguchi et al., 1995; Horiguchi et al., 1997); in Africa the three bioindicator species most used in TBT pollution biomonitoring studies were Hexaplex trunculus, Bolinus brandaris and Stramonita haemastoma. In Oceania were used 5 different species: Bembicium auratum, Cymbiola nobilis, Melo melo, Morula marginalba and Thais orbita.

In this current work is possible to verify greater diversity of bioindicators in America and Asia compared to Europe, probably because America and Asia are continents with greater coastal zone and with different climatic regions (Titley-O’Neal et al., 2011).

Regarding to the taxonomic groups, of bioindicator species, most used in each continent: the species used as bioindicators in Europe are caenogastropods (Subclass Caenogastropoda) of two different orders, Neogastropoda (n=10) and Littorinimorpha (n=2), and the 2 different families more used are Muricidae (n=8) and Nassariidae (n=3); the species used as bioindicators in America belong to 3 different caenogastropod orders – Neogastropoda (n=38), Littorinimorpha (n=4) and Caenogastropoda (n=2), and the 3 different families more used are Muricidae (n=20), Nassariidae (n=6) and Buccinidae (n=5); the gastropods used as bioindicators of imposex in Asia are caenogastropods of 3 different orders – Neogastropoda (n=27), Littorinimorpha (n=5) and Caenogastropoda (n=1), and the 3 different families more used are Muricidae (n=16), Babyloniidae (n=3) and Buccinidae (n=3); in Africa, the bioindicators used are all neogastropods but of 3 different families are Muricidae (n=6), Nassariidae (n=4) and Conidae (n=1); the gastropods used as bioindicators of imposex in Oceania are caenogastropods of 2

168 different orders – Neogastropoda (n=4) and Littorinimorpha (n=1) – and of 3 different families: Muricidae (n=2), Volutidae (n=2) and Littorinidae (n=1).

Generally, a decreasing trend of imposex levels is perceived from 2003 to 2013 in Europe, although increases of some of the imposex indices were still registered until 2008, year from which decreases in all the indices were consistently reported. This trend could be justified by the entry into force of the AFS Convention, which only happened in September 2008.

In North America, all studies published between 2003 and 2013 revealed a consistent decrease of imposex levels, and thus in TBT pollution. In South America the Virgin Islands and Peru, countries that have not ratified the AFS Convention still today, consistent increases were registered during the period under study. In fact, most of the South American countries had not ratified the AFS Convention by 2008 not being subjected to any restriction to the use of TBT compounds.

In Asia there was a consistent decrease in all imposex indices from 2008 onwards, reinforcing the success of the AFS Convention implementation. In 2003 and 2004, increases were reported in Japan, China and Malaysia, while the decreasing trend was reported in Korea as early as 2004 probably as a result of the legislative action MOE Regulation 154/2000, restricting the use of TBT compounds in this country (Choi et al. 2010).

A consistent decrease in imposex levels from 2007 onwards is perceived in Tunisia. Imposex in Hexaplex trunculus, collected in Tunisia, revealed successive increases until 2006, and even in FPL in 2008, decreasing in 2009. However, the use of different bioindicators revealed decreases in imposex levels starting in 2007.

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In Oceania, Andersen (2004) has described a decrease in imposex levels in Morula marginalba since 2003. However, in 2005 the phenomenon was still affecting around 90% of Bembicium auratum females near TBT main sources in Australia (Wilson and Roach, 2009). Nevertheless, imposex levels in this species have declined rapidly and no signs of severe impacts were registered in populations since sterile females were never registered (Wilson and Roach, 2009).

In Europe some species revealed some reduced OTs, since 2003 (e.g. Nucella lapillus in England, France and Spain in 2003; Nassarius reticulatus in Portugal in 2005 and 2006). As a matter of fact, the AFS Convention was adopted in 2001 and opened for signature in 2002, with the purpose of restricting TBT-based antifouling systems on ships by 2003.

In South America there was an increase in TBT concentrations from 2009 onwards were only registered in Thais chocolata tissues in Peru. In fact, Peru was an exception on the effectiveness of the global ban on TBT-based antifouling paints by the AFS Convention entry into force in 2008 at a moment when TBT pollution amelioration has being reported in Europe. Despite the decreasing trend in OTs concentrations in sediments in Brazil and in Ecuador from 2008 onwards, fresh inputs of TBT were still registered, as indicated by the BDI>1 reported.

In Asia there were increases in TBT concentrations and these increases were reported for the majority of the bioindicators used in several countries from 2003 to 2007, with the exceptions of Rapana venosa and Thais clavigera in China in 2006, T. clavigera in Japan and Korea in 2006 and Strombus canarium and Thais tuberosa in Malaysia in 2006.

In sediments from India were still registered in 2008-2009 the fresh inputs of TBT as revealed by the BDI. There were also TBT fresh inputs in water samples from India in 2008

170 and in China in 2010, and exceptional increases in TBT concentrations were reported in Japan in 2003, in China in 2006, in India in 2007 and in Korea in 2008.

In Africa, decreases in TBT concentrations in gastropods tissues were only observed in Tunisia from 2009 onwards. Regarding sediments analyses, Lahbib and his team (2011) reported TBT concentrations increases in Tunisia from 2008 to 2010.

The imposex response in prosobranch gastropods is an easily determined biomarker, which provides useful insights into the biological impact of TBT (Ruiz, 2005). Also monitoring of TBT and the use of gastropods as sentinel organisms (Oehlmann et al., 1996b and Minchin et al., 1997) could be a useful tool for establishing the extent of TBT pollution on the coast of several countries, along with an analysis of the chemical composition of both water and sediments.

Even with the measures of restriction on the use of TBT in antifouling paints are several reports that described imposex and tributyltin levels around ports were not decreasing and that those locations were hotspots of tributyltin pollution (Bailey and Davies, 1989; Davies and Bailey, 1991; Minchin et al., 1995, 1996; Smith, 1996; Morgan et al., 1998; Santos et al., 2002; Barroso and Moreira, 2002; Gibson and Wilson 2003). Simultaneously, other studies demonstrated that TBT pollution was not restricted to harbor areas, and populations from the open ocean as well as the ones from offshore areas were also affected (Ten Hallers-Tjabbes et al., 1994; Mensink et al., 1996; Michel et al., 2001; Barreiro et al., 2001). Tributyltin was even detected in remote and supposedly undisturbed ecosystems (Negri et al., 2004) such as Great Barrier Reef in Australia (Marshall et al., 2002; Haynes et al., 2002; Haynes and Loong, 2002).

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