1 Introduction

Antifouling paints are applied on the outer layer of ship hulls to prevent the settlement and growth of aquatic organisms. From the moment a ship enters seawater, organic molecules accumulate on its surface, and a biofilm (consisting of bacteria and diatoms) settles on the organic molecules within hours to 1 day (Almeida et al. 2007). During the first week, the growth of soft foulers (e.g., slime-growing algae) can be observed, and within weeks, macrofoulers such as barnacles and mussels are visible (Almeida et al. 2007). When these organisms grow on a ship’s hull, the drag they produce slows the ship’s movement, increasing the consumption of fuel by up to 40% and leading to higher emissions of greenhouse gases (Champ 2000). Furthermore, the growth of these organisms can lead to transfer of “invasive species” from one area to another, which could lead to the destruction of local ecosystems (Nehring 2001).

Problems related to biofouling can be addressed by using antifouling paints on ship hulls, but these paints pose a potential threat to the marine environment if biocides leach out of the paint, particularly in areas with a high density of boats (Maraldo and Dahllöf 2004). Biocides that leach into the marine environment can be toxic to marine organisms. For example, tributyltin (TBT) causes imposex in marine mollusks, such as oysters and sea snails (Champ 2000; Horiguchi et al. 1994, 1995, 1997; Shim et al. 2000). Other alternative biocides, such as Diuron, Irgarol 1051, and Sea-Nine 211, have been tested and appear to have toxic effects in many organisms (Garaventa et al. 2010; Sanchez-Bayo and Goka 2006; Turley et al. 2005). Good antifouling paints should be able to prevent fouling without persisting in the marine environment after they are released (Thomas and Brooks 2010).

Zinc pyrithione (ZnPT) is one of the most widely used alternative biocides (Okamura and Mieno 2006; Thomas 1999). However, many studies have shown that ZnPT is highly toxic to some species of fish, plankton, shrimp, and sea urchins (Bellas 2005; Goka 1999; Jung et al. 2017; Okamura et al. 2002). Because of the active usage of ZnPT in antifouling paints, as well as its toxic effects on non-target organisms, its environmental release, distribution, and fate need to be carefully understood. In this review, the overall chemical, toxicological, and environmental aspects of ZnPT are discussed, along with the challenges of analyzing ZnPT in the marine environment.

2 History of Antifouling Paints

The use of antifouling paints dates back to 412 B.C. when a mixture of oil, arsenic, and sulfur was used to prevent the attachment of marine organisms to the hulls of ships (Culver 1992). From the sixteenth to the twentieth century, lead sheathing, copper sheathing, mixtures of tar and resin, and shellac were used to protect ships. However, none of these successfully prevented biofouling until TBT was introduced globally in the early 1960s (Omae 2003). Although TBT was the most effective antifouling agent (Almeida et al. 2007), its use decreased dramatically when it was reported to cause toxic effects on marine mollusks (Champ 2000; Horiguchi et al. 1994, 1995, 1997). TBT also has high octanol-water and organic carbon partitioning coefficients and persists in the marine environment (Bangkedphol et al. 2009; Hoch 2001). In the late 1980s, several countries, including the UK, France, the USA, Canada, Australia, Switzerland, the Netherlands, Denmark, Japan, and Hong Kong, banned the usage of TBT on ships less than 25 m in length (Alzieu 1991; Kannan and Tanabe 2009). Through the International Convention on the Control of Harmful Antifouling Systems (AFS Convention), the International Maritime Organization (IMO) enforced a ban on applying or reapplying organotin compounds for all ships. Organotin compounds were completely banned in 2008, with 28 participating states rectified at that time: Antigua and Barbuda, Australia, the Bahamas, Bulgaria, Cook Islands, Croatia, Cyprus, Denmark, France, Greece, Iceland, Japan, Kiribati, Latvia, Lithuania, Luxembourg, Mexico, Nigeria, Norway, Panama, Poland, Romania, Saint Kitts and Nevis, Sierra Leone, Slovenia, Spain, Sweden, and Tuvalu (Senda 2009). Since then, other antifouling systems have been introduced, including tin-free self-polishing copolymer (SPC) coating, tin-free conventional coatings, alternative/booster biocides, foul-release coatings, and biomimetics (Table 1). Copper was used as a co-biocide to boost the performance of some TBT paints when they dominated the market (Nichols 1988; Young et al. 1979). Since the ban of TBT, copper has become the main biocide in antifouling paints (Srinivasan and Swain 2007).

Table 1 Major antifouling strategies and their advantages and disadvantages (source: Dafforn et al. 2011)
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3 Alternative Biocides

Alternative biocides have been developed in response to both the ban on TBT and the inability of copper to target some marine organisms (especially algal groups) that are tolerant to its effects (Dafforn et al. 2011; Foster 1977; Omae 2003; Reed and Moffat 1983). To augment the action of copper in antifouling paint, various biocides such as Irgarol 1051, Diuron, Sea-Nine 211, dichlofluanid, zinc pyrithione, chlorothalonil, ziram, and zineb were developed, and they have been widely used (Table 2). Because of previous experiences with organotin compounds, however, legislation and controls have been developed for these alternative biocides. Irgarol 1051 and Diuron were first accepted for use in some European countries (the UK, France, Greece, etc.) but were later prohibited in the UK and Denmark for ships less than 25 m in length (Price et al. 2013). Irgarol 1051 is currently under review by the Marine Environment Protection Committee (MEPC) of IMO to be included in Annex 1 to the AFS Convention because of the reported adverse effects of Irgarol 1051 and its metabolites on non-target organisms. Ships coated with Sea-Nine 211 and dichlofluanid were banned from entering some European waters (Price et al. 2013). Zinc pyrithione was allowed in most European waters, except Sweden’s (Readman 2006), and it is currently under review by the Biocidal Products Regulation (BPR) of the EU. Under the classification, labelling, and packaging (CLP) regulation, it is slated to be classified as reprotoxic 1B (ECHA 2018). Although still controversial, this classification may lead to the ban of its use in the future. Chlorothalonil was only allowed in France and Greece, whereas both ziram and zineb were allowed in the Netherlands. The use of alternative biocides in European waters is now regulated by the Biocidal Products Regulation put forth by the European Chemical Agency (Table 3). In Japan, alternative biocides registered under the Japan Paint Manufacturers Association (JPMA) include copper pyrithione (CuPT), ZnPT, Diuron, Sea-Nine 211, Irgarol 1051, chlorothalonil, and others.

Table 2 Chemical properties of alternative biocides
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Table 3 List of alternative biocides approved or registered in Europe, Japan, and New Zealand
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Of the alternative biocides, ZnPT is the most frequently used to boost the action of copper in antifouling paints (Lamore and Wondrak 2011). ZnPT (CAS No. 13463-41-7, EINECS No. 236-671-3) is a complex of zinc salt (Fig. 1) that consists of two pyrithione rings bound to a central metal ion through zinc-oxygen bridges (Dinning et al. 1998). Other names for ZnPT are zinc pyridinethione, bis(2-pyridylthio)zinc, 1,1′-dioxide bis(1-hydroxy-2(1H)-pyridinethionato-O,S) zinc, and 2-mercaptopyridine N-oxide zinc salt. The product names for ZnPT include Danex, Desquaman, Zinc Omadine, and Vancide ZP. ZnPT was introduced in 1991, and its popularity increased when 95% of the global antifouling market turned to biocidal coatings in 2009 (Ciriminna et al. 2015; Maraldo and Dahllöf 2004). ZnPT has been introduced as a stand-alone biocide or as a mixture with zinc dithiocarbamate, Sea-Nine 211, or copper thiocyanate, with concentrations up to 7% in antifouling paints for different types of ships (Cima and Ballarin 2015).

Fig. 1
figure 1

Schematic diagram of the fates of zinc pyrithione (ZnPT) and its metabolites in the marine environment

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4 Toxicity of ZnPT

ZnPT has been widely used as an algaecide, a bactericide, and a fungicide (Bao et al. 2011; Dafforn et al. 2011; Yebra et al. 2004). ZnPT works by restricting ATP synthesis in prokaryotic cells and hindering membrane transport in bacteria. It does this through transchelation with several metals and proteins, as well as the disruption of cell membranes and pH gradients (Chandler and Segel 1978; Dinning et al. 1998). Its acute toxicity in several marine organisms exposed to ZnPT is summarized in Table 4. Toxicity was evaluated and compared using several values such as median lethal concentration (LC50), median effective concentration (EC50), lowest observed effect concentration (LOEC), and no observed effect concentration (NOEC).

Table 4 Toxic effects of zinc pyrithione on marine organisms
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Plankton

ZnPT was shown to have high toxicity in phytoplankton communities at a concentration between 0.7 and 19 μg/L (Hjorth et al. 2006; Maraldo and Dahllöf 2004). These studies highlighted that despite the short half-life of ZnPT in the environment, a short pulse of exposure at a low concentration was sufficient to have significant effects on a phytoplankton community. A recent study by Jung et al. (2017) reported a 96-h EC50 value on the growth rate (cell number, optical density, and chlorophyll a) of the phytoplankton species Nitzschia pungens of 5.51 ± 1.36 μg/L and obtained a 48-h LC50 value of 3.15 ± 0.19 μg/L using larvae of the zooplankton Artemia. The toxicity of ZnPT in the zooplankton larvae was higher than the toxicity of Diuron and Irgarol 1051 (48-h LC50 of 30.57 ± 1.73 and 9.73 ± 2.27 μg/L, respectively).

Microalgae

ZnPT was found to alter germination and rhizoid growth in the microalgae Hormosira banksia, with 48-h EC50 values of 210 μg/L and 310 μg/L, respectively (Myers et al. 2006). When the microalgae were exposed to Diuron, the 48-h EC50 values were 6290 and 6750 μg/L in the same tests, indicating that ZnPT is more toxic than Diuron. ZnPT also showed significant toxicity for growth response in Chaetoceros gracilis spores, with a 72-h IC50 (half-maximal inhibitory concentration) value of 3.2 μg/L (Koutsaftis and Aoyama 2006). In addition, the growth of other microalgae, including Selenastrum capricornutum, was inhibited when exposed to ZnPT for 5 days at 28 μg/L (Koutsaftis and Aoyama 2006). In contrast, Madsen et al. (2000) reported that algae were the least sensitive to ZnPT, with the highest concentration of no effects (NOEC) observed at 7.8 μg/L for Selenastrum capricornutum.

Bivalves

There is little information on the toxicity of ZnPT in bivalves. Bellas et al. (2005) found that ZnPT affected the embryonic development of the mussel Mytilus edulis, with a 48-h EC50 of 2.54 μg/L. Previous studies have reported that zinc chloride is toxic to the embryos of the oyster Crassostrea virginica when exposed to concentrations of 75 μg/L (Calabrese et al. 1973). Additionally, larval growth of the flat oyster Ostrea edulis was found to be significantly reduced in the presence of 500 μg/L zinc salts (Alzieu et al. 1980). Based on these values, ZnPT produces stronger developmental toxic effects in bivalves than zinc chloride or zinc salts.

Sea Urchins and Sea Squirts

ZnPT appears to be very toxic to the sea urchin Anthocidaris crassispina, with only a value of 0.01 fg/L for the maximum concentration of no observed effects on embryonic development, which is much lower than those of CuPT (1 pg/L), Diuron (1000 μg/L), Irgarol 1051 (10 μg/L), and Sea-Nine 211 (1 fg/L) (Kobayashi and Okamura 2002). Paracentrotus lividus also showed high sensitivity to ZnPT, with 48-h EC50 values ranging from 2.45 to 9.46 μg/L (Bellas et al. 2005; Bellas 2008). Bellas (2005) reported morphological abnormalities in larvae of the sea squirt Ciona intestinalis after 24 h of exposure to 72–187 μg/L ZnPT. Exposure of unfertilized eggs to ZnPT did not inhibit fertilization but appeared to promote transmissible damage to embryonic or larval stages. Although there is not much information available on the effects of ZnPT on sea urchins and sea squirts, sea squirts seem to be less sensitive to ZnPT than sea urchins.

Polychaetes

The effect of ZnPT on polychaetes has been examined in only two articles (Bao et al. 2011; Marcheselli et al. 2010). ZnPT was found to be very toxic to Hydroides elegans and Dinophilus gyrociliatus, with lethal concentrations of 7.6 and 2.47 μg/L (48-h LC50 and 96-h LC50), whereas Diuron and Irgarol 1051 were reported to have 48-h LC50 toxicity levels of 16,000 and 2600 μg/L, respectively.

Crustaceans

The lethal concentrations of ZnPT in crustaceans vary widely. ZnPT is not very toxic to Cypretta seuratti and the water flea Daphnia magna, with lethal concentrations reported at 524–2415 μg/L and 98–145 μg/L (48–24-h LC50), respectively (Sanchez-Bayo and Goka 2006). In contrast, ZnPT appears to be extremely toxic to mysid shrimp, with a 96-h LC50 of 6.3 μg/L (Yamada 2006), and moderately toxic to Elasmopus rapax when exposed to 21.5 μg/L (4-day LC50) and 29 μg/L (96-h LC50) (Bao et al. 2011, 2012). However, ZnPT is slightly less toxic than CuPT (96-h LC50 for Elasmopus rapax at 11 μg/L). ZnPT was found to have a similar toxicity level as Sea-Nine 211. The 96-h LC50 of Sea-Nine 211 for mysid shrimp was 4.7 μg/L (Yamada 2006).

Fish

After 96 h of exposure to ZnPT at a concentration of 98.2 μg/L, the survival rate of Pagrus major was reduced by 50% (Mochida et al. 2006). This reduction in survival occurred at the much lower concentration of 2.6 μg/L for Pimephales promelas (Yamada 2006). When juvenile rainbow trout were exposed to ZnPT, the LC50 values for 7 days, 14 days, 21 days, and 28 days were 8.4, 5.6, 4.9, and 4.6 μg/L, respectively (Okamura et al. 2002). In this study, ZnPT was more toxic to juvenile rainbow trout than Diuron, Irgarol 1051, and Sea-Nine 211, while the toxicity of ZnPT was quite similar to that of CuPT. Marine fish were less sensitive to ZnPT than freshwater fish in studies focused on lethal toxicity. In addition to its lethal effects, ZnPT has been shown to alter embryogenesis in fishes and cause spinal deformities in fish larva when eggs are exposed to ZnPT at 10 μg/L (Goka 1999). These studies imply that ZnPT is toxic to marine fishes at very low concentrations.

Marine organisms are very sensitive to ZnPT, usually at μg/L levels. However, ZnPT concentrations in the marine environment are below the toxicity levels based on the values reviewed in section 6, suggesting that ZnPT found in the marine environment is not capable of threatening marine organisms. Once ZnPT is released into the marine environment, it will be diluted with ambient seawater and degraded, resulting lower environmental concentrations. However, the concentration of ZnPT may exceed the toxicity levels at sites of high leaching of ZnPT, such as shipping ports, shipyards, and hull cleaning sites (Dahllöf et al. 2005; Hasan et al. 2014; Kim et al. 2019). Exposure to the elevated concentrations at these high leaching sites will cause adverse effects on several marine organisms. ZnPT is classified as “very toxic” according to OECD guidelines (Yamada 2006). These indicate that ZnPT is a potential threat to the marine environment. In addition, ZnPT can be degraded into various metabolites with more stable forms in the environment. The toxicity of these metabolites also needs to be further studied.

The risk posed by ZnPT emission into the marine environment can be assessed by comparing the environmental concentration with the predicted no-effect concentration (PNEC). PNEC is the concentration of a substance below which no adverse effect on an environment is anticipated, and it is determined through a hazard assessment using available toxicity data. The environmental concentration can be directly measured or predicted using a mathematical model, such as the Marine Antifoulant Model to Predict Environmental Concentration (MAMPEC), which can determine the environmental loading by taking into account all the parameters defined in a biocide emission scenario. If the environmental concentration is equal to or greater than PNEC, it is considered to pose a possible risk to the marine environment (ISO 2013). Environmental risk assessments for the usage of antifouling biocides were performed and reported in several studies (e.g., Chen and Lam 2017; Dahllöf et al. 2005; Wang et al. 2014).

5 Fates of ZnPT in the Marine Environment

When ZnPT is exposed to different environmental conditions, it can be degraded through various chemical and/or biological processes, forming metabolites that have their own toxic properties (Fig. 1). For example, ZnPT can interact with other free metal ions in the water column and is transchelated into other metal pyrithione complexes based on the metal ion uptake. On the other hand, ZnPT shows less persistence in water because of its rapid photodegradation. ZnPT can also be hydrolyzed or biodegraded under specific conditions. These processes breakdown and transform pyrithione complexes into other metabolites; therefore, there is a need to understand all of the degradation processes and potential products of ZnPT. The different degradation processes of ZnPT are discussed in this section and summarized in Table 5.

Table 5 Summary of degradation tests of zinc pyrithione and reported half-lives
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5.1 Photodegradation

ZnPT is easily photodegraded under the influence of sunlight, especially when exposed to UV wavelengths of 320–355 nm (Zepp and Cline 1977). In a research conducted by Maraldo and Dahllöf (2004) and Dahllöf et al. (2005), the half-life of ZnPT in sterile seawater was reported to be 7–8 min in sunlight but more than 48 h in dark conditions. Grunnet and Dahllöf (2005) tested ZnPT at different seawater depths and reported that ZnPT tends to degrade faster at shallower depth, and as the depth of seawater increases, the half-life of ZnPT increases as well. No significant photodegradation was observed at a seawater depth greater than 2 m in their study. These authors also compared ZnPT photodegradation in artificial seawater under xenon light and sunlight. ZnPT was degraded faster under sunlight (half-life, 7 min) than under xenon light (half-life, 17.5 min). Turley et al. (2000) tested ZnPT in artificial seawater under abiotic conditions with sunlight and xenon light. These authors found that ZnPT remains intact for less than 2 min before being photodegraded under sunlight and for approximately 17.5 min under xenon light, which is similar to the findings reported by Grunnet and Dahllöf (2005). Furthermore, Sakkas et al. (2007) tested the effect of dissolved organic matter (DOM) and nitrate (NO3−) on ZnPT photochemical degradation under xenon light. The presence of DOM and NO3− accelerated the photolysis of ZnPT. The half-life of ZnPT in natural seawater was 9.3 min, and that in distilled water was 15 min. The authors also reported that higher concentrations of DOM and NO3− can further enhance the degradation of ZnPT. Interestingly, if only NO3− is present in the seawater, the degradation rate of ZnPT will be slower than when both DOM and NO3− are present. Slightly different photodegradation rates of ZnPT have been reported in different studies, possibly because of the variation in the photon flux density and spectral distribution of light sources. For example, in the report by Thomas (1999), ZnPT was exposed to room light in filtered seawater, and the half-life was found to be 4 h. However, when Marcheselli et al. (2010) exposed ZnPT to room light, approximately 50% of ZnPT remained in the artificial seawater after 48 h of exposure. Under dark conditions, no significant degradation was observed for ZnPT, even after 48 h.

When ZnPT is photodegraded, it can form different metabolites through bond breakage and reactions with other substances in water. When exposed to sunlight, ZnPT can break down and form pyridine sulfinic acid, pyridine sulfonic acid, pyridine disulfide, and pyridine/pyrithione mixed disulfide (Maraldo and Dahllöf 2004). These four metabolites are reported to reach concentrations > 10% of the original ZnPT concentration after being exposed to sunlight (NRA 2001). In addition to these four metabolites, ZnPT can be degraded into sodium pyrithione, pyrithione disulfide, and a tertiary-butylamine pyrithione derivative when exposed to sunlight in seawater (Grunnet and Dahllöf 2005). Turley et al. (2000) reported that the main metabolite from photolysis is pyridine-2-sulfonic-acid. In a study by Sakkas et al. (2007), the photodegradation of ZnPT yielded the six metabolites of 2-pyridinesulfonic-acid (PSA), pyridine-N-oxide (PO), 2-mercaptopyridine (PS), 2,2′-dithiobis(pyridine-N-oxide) ((POS)2), 2,2-dipyridyl disulfide ((PS)2), and 2,2′-dithiobispyridine mono-N-oxide (PPMD). The metabolite (PS)2 was found to be the major metabolite from photodegradation in natural seawater, as well as in distilled water.

The general consensus among all reported findings is that sunlight penetration in seawater directly affects the fate and distribution of ZnPT in the environment. In clear water with low turbidity and calm conditions, where direct sunlight can penetrate a deeper depth, ZnPT can be easily photodegraded, forming several metabolites with their own toxicity and stability characteristics. However, in moderately turbid coastal water, a 1–2 m depth is sufficient to remove or weaken UV light, causing the accumulation of ZnPT in seawater or sediment. Marinas or harbors, for example, can be high-risk areas because high levels of boating activities lead to the continuous release and accumulation of antifouling biocides (Thomas et al. 2000). The stability and toxicity of ZnPT metabolites are mostly unknown, especially in the marine environment. The metabolites may undergo further degradation or persist in the environment. The fates of ZnPT metabolites in the environment also need to be further studied along with their potential toxicities.

5.2 Biodegradation

Ritter (1996) tested the biodegradation of ZnPT under aerobic conditions for 30 days using water and sediments collected from freshwater and marine harbors. They found that ZnPT fully degraded into other metabolites within that time frame. Biodegradation is expected to result in a continuous reduction in ZnPT concentrations throughout an experiment. However, in a study by Maraldo and Dahllöf (2004), the decrease in ZnPT concentrations occurred in a step-like process and began after 10 h of exposure. Although the authors confirmed that the biodegradation of ZnPT occurred, the exact degradation time was uncertain but possibly more than 48 h. During degradation tests under dark conditions performed by Turley et al. (2000), the degradation of ZnPT was facilitated under biotic conditions, with half-lives of 4 days and 7–8 h in natural seawater and river water, respectively, indicating the possible biodegradation by microorganisms. The authors further indicated that the degradation of ZnPT was faster when there was sediment in the water, possibly because of biotic degradation by microorganisms on the sediment surface or abiotic degradation catalyzed by the sediment itself. These shorter half-lives are comparable to the half-life of more than 90 days found in the abiotic test in sterilized artificial seawater. On the other hand, Sakkas et al. (2007) tested the biodegradation of ZnPT under dark conditions in the presence of DOM and NO3− and observed no decrease in ZnPT concentrations during the experiment (the duration of exposure was not reported).

Ritter (1996) found omadine disulfide forms during biodegradation in aerobic sediment. In another study by Ritter (1999c), the aerobic degradation of 0.05 μg/g ZnPT produced omadine sulfonic acid, pyridine sulfonic acid, and two other metabolites called NP1 and NP2, which are heterocyclic metabolites with one ring. The production of NP2 was at first in question because it could have been formed during extraction. After comparing data generated for copper pyrithione, Ritter (1999a, b, c, d, e) concluded that NP2 was present in the sediment before the extraction. In anoxic sediment, the degradation of 3 μg/g ZnPT produced omadine disulfide, 2-mercaptopyridine N-oxide, and another metabolite called NP3 (Ritter 1999a, b, c, d, e). Omadine disulfide rapidly degraded, leaving only NP3 and 2-mercaptopyridine N-oxide remaining throughout the entire test period of 91 days. The degradation of 0.05 μg/g ZnPT produced NP1 and NP3. NP1 was formed in a very small amount, degraded into other compounds, and could not be detected after 14 days. NP3 was also degraded into other metabolites, such as pyridine sulfonic acid, NP4, and NP5.

Although photo-oxidation is a major pathway for the degradation of ZnPT in marine environments, biodegradation can be a significant mechanism by which ZnPT is removed in certain circumstances (e.g., in sediment). However, there is little available information on the biodegradation of ZnPT. Most of the products formed during biodegradation processes have not been identified. The metabolites may also have harmful effects on marine organisms.

5.3 Hydrolysis

ZnPT hydrolyzes under dark conditions with a half-life of more than 90 days in sterilized artificial seawater (Turley et al. 2000). ZnPT hydrolyzes slightly in sterile buffers (pH 5, 7, and 9) under dark conditions, and in artificial seawater at pH 8.2, it hydrolyses within 96 to 123 days (NRA 2001). A study performed according to the US EPA guidelines reported that six metabolites (pyrithione disulfide, pyrithione sulfinic acid, pyrithione sulfonic acid, pyridine sulfonic acid, pyridine sulfinic acid, and pyridine disulfide) were produced when ZnPT was exposed to 25 °C under dark conditions (NRA 2001). In that study, there were only two metabolites that exceeded 10% of the initial dosage of ZnPT: pyrithione disulfide and pyrithione sulfinic acid. On the other hand, according to Reynolds (1995), ZnPT was stable against the hydrolysis process at all pH values tested, and the concentrations remained constant.

The hydrolysis of ZnPT is poorly understood, although it has been confirmed by several studies. The results obtained for the hydrolysis of ZnPT have been inconsistent. In the marine environment, it is even more difficult to differentiate hydrolysis from other degradation processes. More studies need to be performed to elucidate the hydrolysis process and the metabolites formed during the process.

5.4 Transchelation

ZnPT is a complex of a zinc ion (Zn2+) and pyrithione, where two pyrithione ligands bind to the center of the zinc ion (Fig. 1). The zinc ion is interchangeable with other metal ions, such as Cu2+, Na+, Fe2+, and Mn2+. ZnPT usually transchelates into a more stable form of copper pyrithione (CuPT), which has a lower dissociation constant. The complexes vary in strength in the order of Cu > Zn > Mn > Fe > Na. Further evidence for the variable strengths of pyrithiones was provided by Dahllöf et al. (2005). In this study, equimolar concentrations of copper to ZnPT were added to seawater, and 100% transchelation was observed, as only CuPT was found in the seawater after the experiment. However, if copper was added to seawater with CuPT, no changes in concentration were observed, indicating that the formation of a single Cu-PT+ complex is unlikely. This finding suggests that CuPT is a strong complex, and that the substitution of copper in the ligand is unlikely. In the marine environment, organic and inorganic ligands influence the distribution of pyrithione complexes. Once ZnPT is released into the marine environment, it can easily be transchelated into other metal pyrithiones by releasing the zinc ion in the complex and absorbing other free metal ions in seawater (Maraldo and Dahllöf 2004; Thomas 1999; Turley et al. 2000). Even during analytical processes, ZnPT can be transchelated into CuPT inside an HPLC column (Yamaguchi et al. 2006). Preventative measures such as changing buffer solutions or adjusting the column temperature have been employed to minimize and prevent transchelation during the analysis of ZnPT (Yamaguchi et al. 2006).

The transchelation of ZnPT released into the marine environment can make it challenging to detect ZnPT, and thus, the potential effects of ZnPT will be underestimated if only ZnPT is targeted. The transchelated compound, such as CuPT, can be more toxic to marine organisms (Bao et al. 2014; Dahllöf et al. 2005). There needs to focus on all possible transformed chemicals, as well as the original target compound, to comprehensively understand the risk caused by the input of ZnPT into the marine environment. In addition, the transchelation process in natural seawater has not been fully described. Transchelation rates under different environmental conditions are mostly unknown and need to be explored.

6 ZnPT Analysis Methods

ZnPT is the least-studied biocide used in antifouling paints, despite its worldwide usage (Lamore and Wondrak 2011; Thomas and Langford 2009). Because of its fast photodegradation and transchelation characteristics, it is difficult to determine its environmental concentrations. Though studies on ZnPT are relatively scarce, several analytical techniques to overcome these constraints have been evaluated (Table 6). Thomas (1999) used one technique that can solve the problems posed by measuring ZnPT in the environment. In this study, ZnPT was transchelated into the more stable copper (II) complex prior to analysis. Water samples were extracted by shaking in a mechanical shaker after the addition of Cu(NO3)2 solution and dichloromethane (DCM). The solvent phase was collected and centrifuged at 3000 rpm to separate out the residual water and DCM in the sample. The DCM phase was then concentrated and analyzed using high-performance liquid chromatography atmospheric-pressure chemical ionization mass spectrometry (HPLC-APCI-MS) for molecular and fragment ions (m/z 221 + 316) of CuPT. The mean recovery rate was 77%, and the limit of detection in this study was 20 ng/L. A total of 21 water samples were analyzed, and all of the samples showed concentrations below 20 ng/L. The results reported by Thomas (1999) suggested that although the samples were collected at enclosed marinas where flushing was limited, ZnPT could be removed from the water column relatively quickly via photodegradation.

Table 6 Summary of analytical methods for zinc pyrithione
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Another method that can be used to simultaneously measure ZnPT and CuPT involves adding a small excess of copper ions to the sample solution (Bones et al. 2006). This approach allows the transchelation of ZnPT to CuPT to be quantified by comparing transchelated ZnPT with ZnPT standards. In the study by Bones et al. (2006), samples were extracted using online solid phase extraction (SPE) and analyzed using HPLC-APCI-MS. The recovery rate in this study was 81%, and the detection limit was 18 ng/L, which is similar to that of the method employed by Thomas (1999). The environmental ZnPT concentration was found to be below the detection limit (< 18 ng/L) in this study. Additionally, Bones et al. (2006) suggested that samples be processed under minimum light exposure conditions to prevent the loss of ZnPT due to photolysis.

Yamaguchi et al. (2006) suggested a direct HPLC-MS analysis method for ZnPT without transchelating ZnPT to CuPT, which, as described above, was a critical step in Bones et al. (2006) and Thomas (1999). Yamaguchi et al. (2006) used ammonium acetate in the mobile phase to decrease the activity of metallic ions, especially copper ions. The ammonium formed a stable complex with copper ions. Ammonium acetate also increased the pH in the mobile phase, which further stabilized the copper ions. In addition to adding ammonium acetate, the researchers found that maintaining the column temperature at 293 K could also reduce the transchelation rate. Additionally, it was recommended that the temperatures of the nebulizer and drying gas not exceed 523 K to prevent the decomposition of ZnPT. Although their work prevented transchelation, the detection limit of the method was relatively high (LOD = 1 mg/L) compared with those of other studies (ranging from 18 to 40 ng/L) (Bones et al. 2006; Eguchi et al. 2010; Harino et al. 2005; Thomas 1999).

Harino et al. (2005, 2007, 2009) developed a method to detect several types of alternative biocides simultaneously, including Sea-Nine, Diuron, dichlofluanid, Irgarol 1051, and pyrithiones. Water samples were extracted by DCM using a mechanical shaker and dried with anhydrous sodium sulfate. The solvent was exchanged with methanol and spiked with a standard, and the extracts were concentrated into a 2 mL volume and analyzed with liquid chromatography tandem mass spectrometry (LC/MS-MS). Sediment samples were extracted with acetonitrile using a mechanical shaker. Then, the supernatants were concentrated into 5 mL, and 45 mL of distilled water was added. The organic layer was dried with anhydrous sodium sulfate, and the solvent was exchanged with methanol and concentrated into a 2 mL volume prior to analysis with LC/MS-MS. The recovery rates in that study were approximately 80–98%. The detection limit for water samples using this method was 40 ng/L, and the detection limit for sediment samples was 8 to 20 μg/kg d.w. ZnPT, however, was not detected in any of the water or sediment samples in this study because the environmental concentrations were all lower than the detection limits (Harino et al. 2007). Similar methods have also been used in research by Eguchi et al. (2010) and Kim et al. (2014, 2015). In their studies, ZnPT was not detected in the environmental samples.

Dahllöf et al. (2005) used a different method in which seawater samples were extracted using solid phase extraction (SPE) at a flow rate of 15 mL/min. Target compounds were extracted from the column using a mixture of acetonitrile, methanol, and water at a ratio of 7:2:1. The extracts were then concentrated prior to analysis by HPLC. The detection limit for this method was 2 nM (635.4 ng/L), and the recovery rate was 85%. However, no ZnPT concentration was reported in the paper, as the study focused on the fate and effects of ZnPT in the environment.

In addition to using HPLC-MS for analyzing ZnPT, Mackie et al. (2004) developed a method using cathodic stripping voltammetry (CSV) in the presence of Triton-X-100 to measure pyrithione in natural waters. Triton-X-100 is a commonly used detergent in laboratories and is used to separate interfering peaks against thiol compounds because of its nonionic surfactant properties. The detection limit of the method was 1.5 nM (190.7 ng pyrithione/L) for a submersion of 60 s in UV-digested seawater. In their study, pyrithione was measured in marina water of high level of boating activity with concentrations of 105 ± 5 nM (13.4 ± 0.6 μg/L).

Although several methods to analyze ZnPT in the marine environment have been proposed, the environmental distributions of ZnPT have not been well-described using these methods. The paucity of environmental data for ZnPT is reasonable considering that ZnPT is easily degraded in the water column and even during analytical processes despite the efforts taken to reduce the degradation processes. The environmental occurrence, fate, and effects of ZnPT are unlikely to be fully understood based on the limited data available. Reliable and sensitive analytical tools need to be developed to keep ZnPT more stable during the whole analytical procedure. Measuring the breakdown products generated during the analytical process may be an approach to more accurately quantify ZnPT in environmental samples.

7 Summary and Conclusion

ZnPT is one of the most widely used alternative antifouling biocides as a replacement for TBT, which has been banned since the late 1980s. The toxic effects of ZnPT in the environment are unquestionable. In the aquatic environment, ZnPT is rapidly degraded, mostly by photolysis. However, where UV light is not able to penetrate through the water column because of turbidity or a deeper water depth, ZnPT may accumulate and persist in the sediment, continuously exerting toxic effects on the marine environment. While the photolysis and transchelation processes have been relatively well-studied, little information about the biodegradation and hydrolysis of ZnPT is available. Products formed during the processes are mostly unidentified. The quantification of ZnPT has mainly been performed using LC/MS-MS analysis, but it is mostly undetectable in the marine environment. The detection of ZnPT is constrained by the sensitivity of the analytical methods and the rapid degradation and transchelation of ZnPT to other compounds in the presence of other metal ions.

Although some environmentally friendly techniques, such as foul-release or biomimetic coatings, have been introduced into the market, they may not result in the extensive replacement of biocidal products, such as ZnPT, in the near future because of the insufficient durability and effectiveness of the coatings, as well as the cost for application. Because of active usage of ZnPT in antifouling paints, as well as its toxic effects on a wide range of marine organisms, it is recommended that the usage of ZnPT be controlled under a strict authorization process and that the level of ZnPT in the environment be carefully verified. Unfortunately, the analytical determination of ZnPT in the environment is challenging, making it difficult to evaluate its environmental distribution. More reliable and sensitive analytical methods need to be developed to accurately describe the environmental release and distribution of ZnPT and to understand its environmental fates and effects. In addition, the degradation of ZnPT supplies new chemicals into the environment; however, the degradation byproducts have not been adequately identified or quantified, and their toxic effects in the marine environment remain mostly unknown. The byproducts may have more harmful effects than the original chemical. Their toxic effects on marine organisms need to be explored, as well as their fates in the environment. The potential effects of ZnPT in the marine environment will be underestimated if only the parent compound is targeted. It is necessary to understand the total degradation process and focus on all potential transformed products, as well as the original target compound, to comprehensively understand the risk caused by the input of ZnPT into the marine environment.