Ecotoxicology of Antifouling Biocides

Organotin compounds were first developed as moth-proofing agents in the 1920s, and were only later used more widely as bactericides and fungicides (WHO 1980). Organotin compounds produced for commercial applications include methyltins, butyltins, phenyltins, octyltins, and cyclohexyltins. The major uses of organotins are as polyvinyl chloride (PVC) heat stabilizers, catalysts (for silicone and poly-urethane production), biocides, agrochemicals, and glass coatings. The use of tributyltin (TBT) in marine antifouling paints dates from the 1960s, initially as a booster biocide in copper-based formulations. As a result of TBT's efficacy over copper, the use of TBT-based paints accelerated greatly in the 1970s. Annual production of organotin compounds increased from <5,000 t in 1955 to >50,000 t in 1995 (Fent 1996; OECD 2001), with 15–20% of the production accounted for by triorganotins (Bennett 1996). The global annual production of TBT alone was estimated to be 4,000 t in the late 1990s (OECD 2001). In addition to antifouling uses, TBT was used in wood and material preservatives, and slimicides. The use of TBT in antifouling paints applied to hulls of ships and boats, fish-nets, crab pots, docks, and water cooling towers contributed to the direct release of organotins into the aquatic environment. These antifouling usages have caused the greatest environmental concern, because of TBT's high aquatic toxicity. Since the widespread use of TBT-based paints began in the early 1970s, several researchers have reported the harmful effects of TBT on economically important marine food species such as oysters and mussels (for reviews, see Fent 1996; Champ and Seligman 1996 and other chapters in this volume).

Tributyltin (TBT) compounds have been used extensively as a biocide in marine anti-fouling paints. These compounds are persistent in the marine environment, especially in sediments, due to slow degradation rates and consistent flux (Stewart and de Mora 1990; Michel and Averty 1999). Further, they can accumulate in a variety of marine organisms, from plankton and fish, to various marine mammals (Harino et al. 1999, 2003, 2007a, b, c). Numerous deleterious biological effects of TBT on non-target organisms have been observed (Fent 1996). The most obvious manifestations of TBT contamination have been shell deformation in Pacific oysters (Alzieu 1996) and the development of imposex/intersex in gastropods (Gibbs and Bryan 1996). BT compounds may potentially affect human health through consumption of contaminated seafood.

Owing to the widespread deleterious effects on non-target organisms, the use of TBT as an antifouling agent has been regulated in developed countries for over 20 years (Bosselmann 1996). France was the first country to implement a ban on the use of TBT antifouling paints on ships of less than 25 m at the beginning of 1982. Most European countries, the USA, Canada, Australia, and New Zealand implemented similar limited legislation. However, only a few countries or regions in Asia have such regulations, although Japan banned the use of TBT on all vessels in 1991. Considering these facts, it is necessary to understand the present status of BT contamination in Asian coastal waters before implementation of the global ban to evaluate the effectiveness in the future.

The use of TBT and TPT as constituents of antifouling compounds, over more than three decades, has left many coastal environments with a longstanding legacy of contamination. Inputs to water should no longer be an acute threat. However, despite recent actions by IMO to ban the use of these compounds, recovery will not be instantaneous and contamination could even increase at some locations where the legislation is ineffective, or where coatings are replaced, or sediments re-mobilised. Partitioning to solids is reversible and hence sediments may act as a persistent sink and secondary source of adsorbed organotins, as well as those residues entrained as paint flakes from boatyards and docks. Estimates of TBT half-times in sediments range from a few months to decades (in anoxic sediments), indicating that this ‘reservoir’ of organotins is likely to remain biologically relevant, and will require management, for a considerable period. For some ports and harbours, appropriate dredging and disposal of TBT-enriched sediments represents an extremely costly option to maintain viability, and it will be important to ensure that further harm to the environment does not ensue from remobilization of these residues. Based on example in the UK, we review here some of the factors which influence long-term partitioning behaviour and persistence of organotions (predominantly TBT)–and the likely timescales for recovery.

Deep sea environments are divided into the bathyal zone (200–2,000 m), the abyssal (2,000–6,000 m) and the hadal zone (over 6,000 m). These zones cover the largest part of the ocean biome (more than 80%). Until now, it was considered that these zones were deserts because sunlight could not reach to such depths and pressures were too high for biota. Advances in deep sea submersibles and image capturing technologies are now increasing the opportunities for marine biologists to observe and uncover the mysteries of the deep ocean realm. Remotely operated vehicles (ROVs) have been used underwater since the 1950s. ROVs are basically unmanned submarine robots with umbilical cables used to transmit data between the vehicle and researcher for remote operation in areas where diving is constrained by physical hazards. ROVs are often fitted with video, cameras, mechanical tools for specimen retrieval and measurements. Subsequently, manned deep sea submersi-bles have been developed and research has progressed. Although the deep sea is in total darkness, is extremely cold, and subjection to great pressure, the marked development of bathyscaphes has revealed the presence of many deep-sea organisms such as bivalves and gastropods in water depths of 3,000 m and more, and has permitted the collection of sediment and marine organisms (e.g. Endo et al. 1999; Okutani et al. 2002; Okutani and Iwasaki 2003).

The contamination of deep-sea ecosystems by man-made chemicals has also been clarified by progress in diving technology. Organochlorine insecticide residues were measured in the livers of Antimore rostrata, a deep sea fish collected from 2,500 m in 1972, 1973, and 1974 off the east coast of the United States (Berber and Warlen 1979). Subsequently, metals, polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and chlorinated pesticides in were determined in tilefish (Lopholatilus chamaeleonticeps) collected from Lydonia Canyon (on the Georges Bank) in 1981–1982 (Steimle et al. 1990). Persistent orga-nochlorines such as PCBs, DDT and its metabolites (DDTs), chlordane compounds (CHLs), and hexachlorobenzene (HCH) were detected in deep-sea organisms from a water depth of 180–980 m in Suruga Bay (Lee et al. 1997), whilst PCBs have also been detected (22 mg kg⁻¹) along with DDTs (13 mg kg⁻¹) in amphipods collected from a water depth of 2,075 m in the Arctic Ocean (Hargrave et al. 1992). More recently, detection of persistent organic pollutants has been reported in many kind of samples from various water depths (e.g., Takahashi et al. 1997a). It has thus been concluded that persistent organic pollutants (POPs) and various other contaminants can be transferred to deep-sea areas where they may be accumulated by deep-sea organisms.

Certain environmental chemicals cause feminization of males and/or masculinization of females, and such phenomena are generally called endocrine disruption (Colborn et al. 1996). The current status of studies of endocrine disruption both in wildlife and humans is reviewed by the International Programme on Chemical Safety (IPCS) under the joint work of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO) and the World Health Organization (WHO) (International Programme on Chemical Safety 2002). Here, the author will review the masculinization of female gastropod mollusks, called imposex, in terms of the basic biology and induction mechanism of imposex.

The first report of masculinized female gastropods was made by Blaber (1970), describing a penis-like outgrowth behind the right tentacle in spent females of the dog-whelk, Nucella lapillus around Plymouth, UK. The term imposex, however, was coined by Smith (1971) to describe the syndrome of a superimposition of male type genital organs, such as the penis and vas deferens, on female gastropods. Imposex is thought to be irreversible (Bryan et al. 1986). Reproductive failure may occur in females with severe imposex, resulting in population decline or even mass extinction (Gibbs and Bryan 1986, 1996). In some species, imposex is typically induced by tributyltin (TBT) and triphenyltin (TPT), chemicals released from antifouling paints used on ships and fishing nets (Bryan et al. 1987, 1988; Gibbs et al. 1987; Horiguchi et al. 1995, 1997a).

The potential effects of so-called endocrine disrupting chemicals (EDCs) on fish reproduction have been a growing concern since the early 1990s (Colborn et al. 1993; Jobling et al. 1996). Adverse reproductive health effects, possibly a result of EDCs, have been observed in wild fish populations (Jobling et al. 1996; Hashimoto et al. 2000; Larsson et al. 2000). Although much of the research on EDCs has focused on estrogenic compounds, tributyltin (TBT) is also thought to act as an endocrine disruptor (Matthiessen and Gibbs 1998; McAllister and Kime 2003; Nakanishi et al. 2006). Indeed, masculinization has been induced in the Japanese flounder (Paralichthys olivaceus) by TBT administration (Shimasaki et al. 2003). TBT exposure has also induced masculinization, as well as irreversible sperm damage, in the zebrafish (Danio rerio), although direct evidence of aromatase inhibition was not shown (McAllister and Kime 2003). Additionally, Shimizu and Kimura (1987) observed that long-term exposure to tributyltin oxide (TBTO) resulted in significant depression of the gonad somatic index in male goby (Chasmicthys dolichognathus). Thus, there is evidence for TBT-induced reproductive toxicity in fish. In view of the need to maintain and protect wild populations of fish, it is essential that scientists clarify the adverse effects of TBT, especially with regard to reproduction.

In this chapter, I would like to focus on the toxicity of TBT on fish reproduction and review progress in this research field. Sublethal effects (e.g., growth inhibition, histological changes) caused by acute and chronic toxicity of TBT and other organotin compounds to several marine fish species have been described elsewhere (see the review by Fent (1996) ) and are not addressed in this chapter.

During the past several decades, butyltin compounds (BTs), one of the representative groups of organotin compounds (OTs), have been widely used as an antifouling agent in paints for boats, ships, and aquaculture nets (Fent 1996, Champ and Seligman 1996), thus these compounds have been found in a variety of marine organisms, often at concentrations exceeding acute or chronic toxicity levels (Bryan and Gibbs 1991; Alzieu 1996). The hazardous effects of antifouling paints containing BTs in marine ecosystem have become a significant environmental issue all over the world (Champ and Wade 1996; Bosselmann 1996). To prevent the destruction of marine ecosystems, BT application to small boats and fish farming equipment has been banned or regulated in developed countries since the late 1980s (Champ and Wade 1996; Bosselmann 1996). Nevertheless, significant accumulation of BTs has been noted at various trophic levels in the marine food chain including plankton, algae, crustaceans, fishes and cetaceans, indicating that BTs impact continues to be felt in marine ecosystems.

Tri-organotins, tributyltin (TBT) are reported to be the most toxic compounds, and at nanogram-per-liter levels, TBT has adverse effects on many aquatic organisms, for example, producing retardation of regenerative growth, delayed molting, reduction in burrowing activity and deformities in limbs in the fiddler crab (Weis and Perlmutter 1987; Weis et al. 1987; Weis and Kim 1988), impairment of egg production in the calanoid copepod (Johansen and Møhlenberg 1987), reduction in larval growth in the silverside (Hall et al. 1988) and avoidance reactions in the Baltic amphipod (Laughlin et al. 1984). Recently, a relationship between metabolic capacity, accumulation and toxicity of BTs in marine organisms has been reported in terms of comparisons of BT residue levels in organisms at various trophic levels in the food chain (Fent 1996; Takahashi et al. 1999; Ohji et al. 2002a). The results indicate that though BTs accumulated in most organisms at levels up to 70,000 times higher than those in seawater, no significant biomagnification was observed in the higher levels of the food chain (Takahashi et al. 1999). High concentrations have, however, been found in lower trophic animals such as caprellids. It seems that TBT accumulates specifically for the caprellids in the marine ecosystem regardless of the trophic level in the food chain, and it can be a break point for the disturbance in the natural food chain structure. It is considered causing them to accumulate BTs at elevated concentrations because of their lower metabolic capacity to degrade TBT (Ohji et al. 2002a). The BTs seem to be accumulated in a species specific manner. Thus, studying the implications of species-specific accumulation and the biological effects of BTs on caprellids may provide some clues to understanding the accumulation mechanisms in the coastal ecosystem as well as the mode of action of BTs in organisms.

In recent years, discharges of anthropogenic chemicals to the environment have been increasing in association with industrial development. These chemicals and their degradation products are released to the environment, discharged into water, and may ultimately contaminate aquatic organisms. Polychlorinated biphenyls (PCBs) and tributyltin (TBT) are particularly ubiquitous pollutants.

A number of field studies have provided evidence that environmental toxicants can modulate immune parameters of exposed fish (for overviews see Snieszko 1974; Dunier and Siwicki 1993; Zeeman 1994; Zelikoff 1994; Rice 2001). Also, laboratory studies demonstrated that toxicants impact the immune system of fish (e.g., Weeks et al. 1988; Thuvander and Carlstein 1991; Kaattari et al. 1994; Rice and Schlenk 1995; Sanchez-Dardon et al. 1999; Carlson et al. 2004; Quabius et al. 2005). Consequently, it has been suggested to utilize immune parameters of fish as indicators of environmental pollution (Anderson 1990; Weeks et al. 1992; Wester et al. 1994).

In fish immunotoxicological studies, emphasis has been given to the measurement of single endpoints or functions such as depressed phagocytic activity of macrophages or the alteration of oxidative burst activity of immune cells. However, the ultimate concern is that toxic exposure might increase the susceptibility of fish to pathogen infections. Thus, it is important to elucidate the consequences of changes in molecular or cellular immune parameters for the overall immune system function, since, as pointed out by Rice (2001), alterations at the molecular and cellular level do not necessarily translate into immune modulation at the system level. Rather few studies directly correlated alterations of specific molecular and cellular immune parameters with altered immune system function and/or altered susceptibility to pathogens (Palm et al. 2003; Carlson et al. 2002; Burki et al. 2008). However, some more indirect evidence is available that toxic impact on immune parameters increases the susceptibility of fish to disease. One example comes from the decline of wild Pacific salmon populations in the USA and Canada. When juvenile chinook salmon (Oncorhynchus tshawytscha) collected from estuaries in the Puget Sound area and showing suppression of various immune parameters were infected with Vibrio anguillarum in the laboratory, they were more susceptible than fish from non- or less-polluted areas (Arkoosh et al. 1998). Significant differences were seen even 2 months after removal from the contaminated areas, suggesting that the chemical exposure had a lasting effect on disease susceptibility. This assumption is supported by the study of Milston et al. (2003) who showed that short-term contaminant exposure of chinook salmon during early life-history stages resulted in long-term impairment of humoral immune competence. Under complex field conditions, as in Puget Sound, it is difficult if not impossible to provide conclusive evidence that the contaminant-induced immunosuppression is causative to the observed decline of chinook salmon populations. However, in a demographic modeling study, Spromberg and Meador (2005) could show that immune suppression acting through reduction of age-specific survival would produce pronounced changes in the population growth rate. This result highlights the potential of immunotoxicants to adversely affect organism health and population growth of aquatic wildlife.

The concept of endocrine disruption was introduced at the Work Session on “Chemically Induced Alterations in Sexual Development: The Wildlife/Human Connection” in 1991. At this session it was pointed out that a number of environmental chemicals affect hormonal systems and have adverse health effects on wildlife and probably on humans. Such chemicals are referred to as endocrine disrupting chemicals (EDCs), and their effects have emerged as a major environmental issue. The nuclear receptors (NRs) of intrinsic hormone systems are likely to be targets of EDCs, because their intrinsic ligands are fat-soluble and low-molecular-weight agents, as are the environmental pollutants. Examples can be found among persistent organochlorine pollutants (DDT, PCBs), plasticizers (pthalates), detergents (alkylphenols) and birth control pills (ethy-nylestradiol) (McLachlan 2001). The effects of synthetic chemicals on sex hormone receptors such as the estrogen receptor (ER) and androgen receptor (AR) have attracted much attention, focusing on the reproductive failures observed in wildlife.

Among them, the imposex phenomenon in marine gastropods provides one of the clearest examples of endocrine disruption in wildlife. While many field studies have demonstrated the adverse effects of organotins upon female gastropods, the mechanism underlying the imposex phenomenon has not been fully elucidated. Organotin compounds have been widely used as antifouling paints for ships and fishing nets since the 1960s and have thus been released into marine environments. Aquatic invertebrates, particularly marine gastropods, are extremely sensitive to organotin compounds and undergo changes in sexual identity in response to exposure. Most marine gastropods in organotin-polluted areas have shown reproductive failure due to oviduct blockage by vas deferens formation, resulting in population decline or mass extinction (Bryan et al. 1988; ten Hallers-Tjabbes et al. 1994). This phenomenon is called “imposex” as an abbreviation of “imposed sexual organs”, because male genital organs, such as the penis and vas deferens, are imposed upon female organs (Smith 1971). Approximately 150 species of imposex-affected gastropods have been found in the world (Fent 1996; Matthiessen et al. 1999). Despite several hypotheses on the cause of imposex induction, such as aromatase inhibition, testosterone excretion-inhibition, functional disorder of the female cerebropleural ganglia, and involvement of neuropeptide APGWamide (Bettin et al. 1996; Ronis and Mason 1996; Oberdörster and McClellan-Green 2000, 2002), the mechanism through which they induce and promote the development of a penis-like structure and a vas deferens in female gastropods remains obscure.

Drug metabolizing enzymes play important roles in terrestrial and aquatic organisms for the synthesis, metabolism and excretion of various kinds of chemicals, especially lipophilic compounds including natural and synthetic chemicals. Drug metabolizing systems consist of two phases. In the first phase reaction, oxidative, reductive and hydrolytic reactions are dominant, and water solubility of chemicals increase through these processes. The toxicity of chemicals, however, are not necessarily decreased in the first phase and may be toxified in some cases through metabolic activation. In the second phase, conjugate reactions are dominant, and chemicals modified in the first phase are biotransformed to more water soluble compounds, and are readily excreted. These two-phase biotransformation processes are therefore important in the metabolism of lipophilic compounds such as endogenous steroids in the bodies of all animals, and also in the fates and effects of medicines in the medical and pharmaceutical fields. Since the 1970s, drug metabolizing enzymes have been found to be important in evaluating the toxicities of various lipophilic environmental pollutants, including pesticides and industrial chemicals. Although much knowledge has been accumulated on xenobiotic metabolism using experimental and wild animals, that obtained from aquatic organisms is insufficient for the complete understanding of fates of environmental chemicals incorporated in the body compared with mammals such as rats and mice. Many polycyclic aromatic hydrocarbons (PAHs) are known to induce cytochrome P450(CYP) and the metabolites act as potent carcinogens and/or mutagens and are, therefore, considered as important risk factors in epidemiological and epizootiological cancer. The P450 system is universally distributed in all organisms and plays a key role in the metabolism of xenobiotic compounds such as PAHs, dioxins, pesticides etc., leading to their detoxification or bioactivation. However, the system is intrinsically important in the metabolism of endogenous substrates including steroids, arachidonic acid, prostaglandins and others.

Thus, drug metabolizing enzyme systems are important for the understanding of both detoxification and bioactivation.

Plankton constitutes the largest component of the world's biomass, exerting a vital influence on aquatic life as well as forming the basis of aquatic food webs. Plankton are potentially suitable indicators of any type of contamination in seawater, because of their quick responses to toxicants and other chemicals. Assessing their level of contamination can provide a strong explanation for level of contamination in higher tropic levels of the food chain. Even though many experts question the significance of tributyltin (TBT) biomagnification from plankton, as the primary assemblage in the food chain, plankton would be the first target for organotin compounds released into the water column and a source for assimilation in higher organisms.

Research over the last 25 years has highlighted that organotin accumulates in a variety of marine organisms, from plankton (Harino et al. 1998, 1999; Takahashi et al. 1999) to high-level predators (Tanabe et al. 1998). Organotin compounds have been responsible for many deleterious effects on non-target aquatic life (Fent and Meier 1994; Ohji et al. 2002a, b, 2003a, b, 2004, 2005; Grzyb et al. 2003). Many studies have been conducted regarding organotin impacts on plankton including phytoplankton (Laughlin et al. 1986a; Maguire et al. 1984; Maeda et al. 1990; Reader and Pelletier 1992; Beaumont et al. 1987; Avery et al. 1993; Mooney and Patching 1995; St-Louis et al. 1997; Tsang et al. 1999; Rumampuk et al. 2004); holozooplankton species, such as copepods (Linden et al. 1979; U'ren 1983; Bushong et al. 1987; Hall et al. 1987), rotifers (Cochrane et al. 1991; Snell et al. 1991 a, b; DelValls et al. 1997; Sun et al. 2001; Jeon et al. 2003; van den Brink and Kater 2006), daphnia (Steinhauser et al. 1985; Kline et al. 1989), mysids (Davidson et al. 1986a, b), and meroplankton such as larvae of amphipods and mussels (Laughlin et al. 1986b; Beaumont and Budd 1984), and early life stages of fish (Seinen et al. 1981; Pinkney et al. 1990). Compared to laboratory data, field surveys on accumulation of organotin compounds in plankton are very limited (Harino et al. 1999; Takahashi et al. 1999). Mostly, assessment of organotin impacts on plankton are based on laboratory toxicity tests. There has also been an attempt to use zooplankton as a tool to assess biological impacts of contaminated sediments (van den Brink and Kater 2006). Nevertheless, whilst biomagnification of organotin compounds in aquatic ecosystem has been considered (Cooney 1988; Guruge et al. 1996; Hu et al. 2006), there are very limited data on the transfer of organotin compounds from plankton to higher taxa, either in field surveys (Takahashi et al. 1999), or in experimental studies (Sun et al. 2001).

Concern over organotin (OT) bioaccumulation has focused on molluscs due to the effects of TBT and TPT on reproduction and recruitment in this phylum at extremely low concentrations. There have also been concerns because molluscs form an important component of food chains involving humans. Molluscs represent a significant, if variable, concentration step in the transfer of OTs from water (Bioconcentration Factor-BCF ̃10²–10⁵) and sediment (BCFsed up to 10²) – an attribute which has been harnessed in biomonitoring programmes. Few phyla display comparable abilities for bioconcentration of OTs, which accounts for their sensitivity. However, bioaccumulation is not always simply a function of adsorption of dissolved forms (except at the lowest trophic levels), but may also involve uptake from dietary sources including sediments, and modification by metabolism and excretion, giving rise to much variability. In this chapter we review the pathways and potential for bioaccumulation of OTs in three major classes of the Mollusca, namely gastropods, bivalves and cephalopods. Much of the knowledge gained from studies on OTs will have broader implications – in terms of understanding the processes and timescales of impacts of future persistent contaminants, and, hopefully, in the design of future risk assessment protocols.

Organotins (OTs) are used in a variety of consumer and industrial products such as marine antifouling paints, agricultural pesticides, preservatives, and plastic stabilizers. In particular, butyltins (BTs) and phenyltins (PTs) have been extensively used in boat paints because of their excellent and long-lasting antifouling properties. However, it is well known that BTs and PTs leaching from boats can accumulate in tissues of aquatic organisms causing various deleterious effects.

To understand the contamination status of OTs from fresh water to deep sea ecosystems, various fish species are used as bioindicators. Furthermore, many fish species are economically important as food, thus, to examine the pollution level is mandatory to evaluate risk assessment for human consumption as well as understanding aquatic contamination levels and bioaccumulation.

Alternative antifouling biocides to TBT were first detected in environmental surface waters in the early 1990s (Readman et al. 1993). Irgarol 1051 was first detected in the surface waters of marinas on the Côte d'Azur, France at concentrations of up to 1,700 ng l-1 (Readman et al. 1993) and in subsequent years the occurrence of Irgarol 1051 was reported in both fresh and marine waters (Scarlett et al. 1999; Thomas et al. 2000; Martinez et al. 2001; Lamoree et al. 2002) These reports established that the alternative antifouling biocides being used to replace the restricted TBT could also be accumulating in the environment and possibly posing a risk to aquatic habitats. Following Irgarol 1051, a number of other compounds were also used as biocidal additives to antifouling paints and methods have been developed to determine their occurrence in environmental waters (Thomas 1998; Piedra et al. 2000; Thomas et al. 2001). The early studies used GC-MS analysis of water extracts to analyse Irgarol 1051 alone; however, as the field developed, multi-residue LC-MS or LC-tandem MS techniques followed that allowed for the simultaneous analysis of the most commonly used biocides and their metabolites (Thomas 1998). However, for certain biocides (e.g. zinc pyrithione) specific methods are predominantly used due to the intrinsic physico-chemical properties that make it a difficult compound to quantitatively analyse (Thomas 1999).

Antifouling paints containing active biocides are typically used on the hulls of ships and boats to prevent the growth of fouling organisms. Antifouling paint biocides are therefore released directly into surface waters following their release from painted surfaces or from the inappropriate disposal of paint related waste. The levels of biocides found in surface waters are therefore directly related to the amount released from such surfaces. Once in the water column, antifouling biocides, as with all other contaminants, are subjected to a number of environmental processes that control their environmental fate. Depending on their physico-chemical properties, biocides can partition onto sediments and accumulate in biological material. In order to measure the occurrence of antifouling biocides in water, sediments and biota analytical methods have been developed and applied. This chapter will review the data available on the occurrence of the biocides listed below in surface waters, sediments and biota for Europe and the Americas, including Canada, USA and the Caribbean (Fig. 19.1).

The International Maritime Organization (IMO) has adopted the International Convention on the Control of Harmful Antifouling Systems (AFS Convention), which prohibits the use of organotins (OTs) as active ingredients in antifouling systems for ships. Following the international restrictions on the use of OT-based antifoulants, paint manufacturers have developed many products as alternatives to the use of OTs. More than 20 chemical substances have been used or proposed as alternative compounds. After release of these antifouling biocides from the hulls of ships, fishing nets etc. into the aquatic environment, these chemicals are distributed in water, sediment, and aquatic organisms. The coastal waters of European countries and the USA are already polluted by alternative biocide (Gough et al. 1994; Toth et al. 1996; Gardinali et al. 2002; Bowman et al. 2003).

Asia is geographically located in the eastern and northern hemispheres and is the world's largest and most populous continent. It covers 8.6% of the Earth's total surface area (or 29.4% of its land area) and, with almost four billion people, it contains more than 60% of the world's current human population. Economic growth in Asia since World War II to the present time had been concentrated in countries of the Pacific Rim such as Thailand, Malaysia, Singapore, Hong Kong, Taiwan, Japan, and Korea. Development of economies may result in marine pollution by antifouling biocides, as trade increases, as illustrated by the presence of OT pollution in Japan, Korea, Southeast Asia and India (Harino et al. 1998a,b, 1999, 2003; Kan-atireklap et al. 1997a, b, 1998; Midorikawa et al. 2004a, b; Prudente et al. 1999; Shim et al. 2002, 2005; Sudaryanto et al. 2002).

Organotin compounds have been strictly regulated in many countries for the last two decades because of their severe toxic effects on marine organisms (Fent 1996). Consequently, new antifouling biocides that substitute for organotin compounds have been developed. Many studies have surveyed the occurrence of these biocides in the environment. Diuron and Irgarol 1051 (herbicides used in popular slime-resistant antifouling paints) have been detected in environmental water samples from several European countries, the United States, various Caribbean countries, and Japan (Thomas et al. 2000, 2002; Konstantinou and Albanis 2004; Harino et al. 2005b; Carbery et al. 2006). The occurrence of pyrithiones (PTs), such as copper (CuPT) and zinc pyrithione (ZnPT), in the aquatic environment has only recently been reported; Harino et al. (2006) first detected CuPT in sediment from coastal northern Vietnam and, subsequently, in Otsuchi Bay, Japan (Harino et al. 2006, 2007).

Thus, it is clear that these alternative biocides are already in widespread use. The toxicity of these alternative biocides to various marine organisms has already been examined, as described below, but in view of the need to maintain wild populations of marine organisms, it is necessary to clarify the adverse effects of these alterna-tive biocides. Indeed, many studies have been conducted to elucidate the toxic effects of alternative biocides on aquatic organisms.

In Chapter 21, we used publicly available data to review progress in research on the toxicity of several of the alternative biocides on the list (ref. Table 21.1) of those registered and approved by the Japan Paint Manufacturers' Association, especially in plankton species and teleost fish. In the first section of this chapter, we review the toxicity of several of these listed alternative biocides, particularly to marine organisms other than plankton and teleost fish. In the next section, the discussion of sublethal effects of the alternative biocides is divided into two subsections. In the first subsection, teratogenicity to fish, which is exhibited by dithiocarbamates (DTCs) as well as by zinc pyrithione (ZnPT) and copper pyrithione (CuPT) and related products is the main feature reviewed. In the second subsection, the toxic effects of herbicides such as Irgarol 1051 and diuron on corals are reviewed. Additionally, in this subsection, the physiological changes reported in some fish species exposed to these biocides are discussed.

The ideal antifouling biocide, from a marine conservation perspective, should be degraded to compounds of lower toxicity in the environment to avoid impact on non-targed organisms. On the other hand, an effective biocide needs to have high toxicity to prevent fouling. These two suppositions are, to an extent, mutually exclusive. Nevertheless, some antifouling biocides released to the marine environment undergo hydrolysis, whilst others are degraded by sunlight in the photic zone.

Furthermore, a number of antifouling biocides are degraded by the many species of bacteria which inhabit water and sediment. Stable antifouling biocides are transported widely and can accumulate in sediment or are concentrated in aquatic organisms. Thus, in order to reduce the threat of magnification of residues, an ideal biocide should be degraded easily and rapidly to substances of lower toxicity, following their released into the aquatic environment. Information about their possible degradation mechanisms in the environment, whether by hydrolysis, sunlight or bacteria, is an important requirement in order to estimate the persistence of these compounds and to identify the factors that influence their behaviour. In this chapter, the degradation pathways and rated of representative alternative biocides in aquatic environment are reviewed.

Due to the restrictions on TBT usage in antifouling paints since 2003 and its complete ban on all vessels in 2008 (IMO 2001), copper has been increasingly used as the main biocide ingredient in antifouling paint coatings. Copper is toxic to a wide range of aquatic organisms, which makes it an ideal biocide, preventing the colonisation of biofouling organisms on the vessel surface. There has been much concern from regulators and scientists that copper concentrations may become elevated in areas of high boating density such as marinas and estuaries with potential damaging effects on the animal and plant communities. In certain European countries, copper has been banned from use on recreational vessels, although so far this is restricted to inland freshwaters, many countries are beginning to re-evaluate current copper risk assessments in marine coastal waters.

This chapter provides an outline of the concentrations of copper in the marine coastal environment as a result of its use as an antifouling biocide. The potential risk of copper to marine life has been evaluated with respect to copper bioavail ability, speciation and toxicity. The chapter outlines some of the shortfalls of current copper risk assessment and provides some suggestions for improvement.