Toxicity of the booster biocide Sea-Nine to the early developmental stages of the sea urchin Paracentrotus lividus

Aquatic Toxicology 83 (2007) 52–61

Toxicity of the booster biocide Sea-Nine to the early developmental stages of the sea urchin Paracentrotus lividus Juan Bellas ∗ Departamento de Ecolox´ıa e Biolox´ıa Animal, Universidade de Vigo, Estrada Colexio Universitario s/n, 36310 Vigo, Galicia, Spain Received 8 February 2007; received in revised form 12 March 2007; accepted 16 March 2007

Abstract The toxicity of the alternative antifouling compound Sea-Nine to the early developmental stages of the sea urchin Paracentrotus lividus was investigated. The inhibition of the fertilization rate and the induction of transmissible damages to the offspring, measured as inhibition of embryonic development and larval growth, were studied by preexposure of gametes to a range of Sea-Nine concentrations. Sperm and egg exposures resulted in a significant decrease of the fertilization rate and induced a transmissible damage to the offspring. The effects of Sea-Nine throughout the embryonic development were also studied by a 48 h exposure of fertilized eggs. The larval growth was the most sensitive response tested, with toxic effects detected at 8.6 nM = 2.4 ␮g/L (EC10 ). The inhibition of P. lividus embryonic development and larval growth was also used to study the loss of toxicity in Sea-Nine solutions exposed for 8 h to direct sunlight and maintained for 28 h in dark conditions. The results showed that the toxicity of Sea-Nine solutions did not decrease but a slight increase in toxicity was observed in comparison with control solutions. The risk of Sea-Nine maximum concentrations measured in marinas around Europe to P. lividus early developmental stages was calculated and the obtained risk quotient was 5.5, indicating that adverse ecological effects of this compound are likely to occur. © 2007 Elsevier B.V. All rights reserved. Keywords: Antifouling; Sea-Nine; Toxicity; Sea urchin; Fertilization; Embryo–larval bioassay

1. Introduction The term marine biofouling refers to the growth of unwanted organisms on the surfaces of artificial structures immersed in the sea (Kiil et al., 2001; Townsin, 2003). Biofouling on ships and boats causes important problems to governments and industry with severe economic consequences, related to the increased fuel consumption, increased workload on the machinery and reduced inter docking periods. In addition, ecological implications of biofouling include increased emission of carbon dioxide (and other “greenhouse” and “acid rain” gases) and increased risk of invasion by alien species (Rouhi, 1998; Callow and Callow, 2002; Godwin, 2003; Yebra et al., 2004). Antifouling paints have been widely used to prevent and control biofouling and, therefore, their economic and ecological relevance is unquestionable. Among the several antifouling biocides used until now, tributyltin (TBT)-based coatings, introduced in the early 70s, are the most effective and successful

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at reducing biofouling due to their high toxicity to fouling organisms. However, deleterious ecotoxicological effects derived from the environmental stability and extreme toxicity of TBT to non-target organisms such as imposex and decreased reproductive viability in gastropods, or increased shell thickness in oysters, were soon detected (Bryan and Gibbs, 1991; Alzieu, 2000; Evans et al., 2000). Those environmental problems resulted in the development of legislation by the International Maritime Organization (IMO) to ensure the international prohibition of the application of organotin-based coatings on ships by 1 January 2003, and a complete prohibition of the presence of organotin compounds in antifouling systems by 1 January 2008 (IMO Resolution A. 895 21, 25/11/1999). The Regulation 782/2003 of the European Parliament also prohibits the use of organotin compounds on ships registered in the EU with the same enforcement dates. As a result of the existing legislation an intense effort has been devoted to the development of alternative biocides. The global phase out of organotin-based antifouling coatings requires the development of environmentally acceptable antifouling compounds which maintain the same efficiency against fouling as TBT. Several formulations based on cop-

J. Bellas / Aquatic Toxicology 83 (2007) 52–61

per combined with organic booster biocides (e.g. Irgarol 1051, Diuron, chlorothalonil or zinc pyrithione) have been approved for use during last years and replaced TBT in antifouling products. However, some of those booster biocides have been banned or regulated in Europe because of their environmental stability and toxicity to non-target organisms (Voulvoulis et al., 1999; Ranke and Jastorff, 2000; Thomas et al., 2002; Konstantinou and Albanis, 2004) stressing the need to perform adequate risk assessment procedures for antifouling biocides. Thus, the European Parliament implemented in May 2000 a Directive (Biocidal Products Directive 98/8/EC) which establishes the requirements that a new biocide must meet before it is introduced into the market, and reviews the environmental performance of existing biocides through a rigorous evaluation to ensure that they pose no unacceptable risks to people, animals or the environment. Sea-Nine is a highly effective biocide against a wide range of fouling organisms which gained the “Green Chemistry Challenge Award” from the U.S. EPA in the category “Designing Safer Chemical Products” for its environmental safety as marine antifouling agent (http://www.rohmhaas.com/seanine/index. html). The environmental risk of Sea-Nine is considered to be low due to its very rapid environmental degradation (Jacobson and Willingham, 2000; Madsen et al., 2000). This compound has been reported to be rapidly degraded in water and sediment both biologically (mainly) and chemically, and its concentration is considered to be rapidly reduced below toxic levels (Jacobson et al., 1993; Callow and Willingham, 1996; Konstantinou and Albanis, 2004). Nevertheless, little information has been published about the toxicity of Sea-Nine to non-target marine invertebrates. Such information is urgently required in order to evaluate the environmental risk of this compound in view of reported high concentrations in coastal waters. For instance, significant concentrations (49–3700 ng/L = 0.2–13 nM) of SeaNine have been found in Spanish, Danish and Greek marinas due to its use as a biocide in antifouling applications (Mart´ınez et al., 2000, 2001; Thomas, 2001; Sakkas et al., 2002a). The embryonic and larval stages of marine invertebrates are less tolerant to toxicants than adults (e.g. Connor, 1972; Marin et al., 1991; Ringwood, 1991; His et al., 1999) and the embryo–larval bioassays, in particular with bivalves and sea urchins, have been used for decades as sensitive, simple, and reliable tools for assessing and monitoring marine pollution (Woelke, 1972; Kobayashi, 1981; His et al., 1997). In the present study, we have conducted bioassays with early developmental stages of the edible sea urchin Paracentrotus lividus (Lamarck, 1816). P. lividus (Echinodermata, Echinoidea) is a rather large sea urchin (up to 7 cm diameter) that occurs as the dominant echinoid species throughout the Mediterranean Sea and is widely distributed along the North-Eastern Atlantic coasts (Hayward and Ryland, 1990; Boudouresque and Verlaque, 2001). P. lividus is usually found on the lower rocky shore down to depths of 3 m playing key ecological roles in the general functioning of ecosystems by removing algal communities or by preventing their establishment, leading to dramatic changes in the structure of benthic assemblages (Hayward and Ryland, 1990). In some European countries (France, Spain, Italy, Ireland, Portugal and Croatia) P. lividus is exploited for its high valued gonads

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which are prized as a delicacy (Boudouresque and Verlaque, 2001). In view of the lack of information on the toxicity of Sea-Nine to marine invertebrates and particularly to the early developmental stages, the present work was undertaken with the aim of assessing the toxicity of Sea-Nine to the fertilization rate, embryonic development and early larval growth of P. lividus. Furthermore, the effects of the potential degradation on the toxicity of Sea-Nine were studied. This information is crucial for use in the ecological risk assessment of this biocide. 2. Materials and methods 2.1. Experimental solutions Stock solutions were made up by dissolving Sea-Nine® 211 (4,5-dichloro-2-n-octyl-4-isothiazolin-3-one) (Rohm and Haas) in an organic non-toxic dissolvent dimetylsulfoxide (DMSO) (Bellas et al., 2005), approximately 1 h before the beginning of the experiments. The experimental concentrations were obtained by diluting the stock solution in artificial seawater (ASW) prepared according to Zaroogian et al. (1969) but salinity was adjusted to 34 ppt adding distilled water. The range of experimental concentrations was chosen on the basis of literature toxicity data and the solubility of the compound (Shade et al., 1993; Jacobson and Willingham, 2000; Fern´andez-Alba et al., 2002; Yamada, 2006). All glassware was acid-washed (HNO3 10% vol.) and rinsed with acetone and distilled water before the experiments. The same amount of DMSO (less than 1 mL/L) was added to each treatment. Four replicates of each experimental concentration, four controls with ASW, and four DMSO controls were tested. Physicochemical conditions of the experiments were 34.44 ± 0.19 ppt salinity, 6.98 ± 0.25 mg/L O2 and 8.38 ± 0.05 pH (mean ± S.D., n = 15). 2.2. Biological material Experiments were performed during July–October 2006. Mature Paracentrotus lividus were collected in a pristine site from R´ıa de Vigo (Galicia, NW Spain). Animals were transported to the laboratory in a portable icebox and maintained in aquaria with running natural seawater filtered through 0.22 ␮m for at least 1 week until the experiments. Gametes were obtained by dissection according to Fern´andez and Beiras (2001). Mature oocytes from a single female were transferred to a 100 mL measuring cylinder and their quality was checked under microscope. Only batches of mature eggs that were spherical and undamaged were used for the experiments. Sperm mobility was checked under microscope and the sperm solution was stored at 4 ◦ C until use. 2.3. Effects on the fertilization rate The effects of Sea-Nine on the fertilization rate of P. lividus were evaluated by exposing eggs and sperm to a range of concentrations of this compound (4–8–16–25–32–50–64–100–128– 200–256–400–800–1600 nM). Aliquots (100 ␮L) of concentrated sperm suspension (ca. 108 sperm/mL) were added to 4 mL

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J. Bellas / Aquatic Toxicology 83 (2007) 52–61

vials containing 2 mL of the experimental solutions and were incubated for 45 min in a culture chamber at 20 ◦ C. Approximately 350–400 unfertilized eggs were delivered into glass vials with airtight Teflon-lined caps containing 20 mL of the experimental solutions and incubated for 4 h at 20 ◦ C. After incubation the eggs were rinsed carefully with ASW to remove the excess of Sea-Nine, by using a 50 ␮m mesh, and transferred to experimental vials with 20 mL of ASW. Untreated eggs and sperm followed the same exposure conditions. Thereafter, 50 ␮L of sperm suspension was added to glass vials containing 20 mL of ASW and 350–400 eggs and were carefully stirred to allow fertilization. The following crosses were made: treated sperm × untreated eggs, untreated sperm × treated eggs and treated sperm × treated eggs. The fertilization rate was determined 1 h after fertilization when the 2-cell stage was attained in the control vials. 2.4. Effects on the embryonic development and larval growth The effects of Sea-Nine on the subsequent embryonic development of P. lividus were evaluated by exposing eggs and sperm to a range of concentrations of this compound as explained above. After exposure of the gametes to the experimental concentrations, the following crosses were made: treated sperm × untreated eggs, untreated sperm × treated eggs and treated sperm × treated eggs. The vials were incubated at 20 ◦ C for 48 h in darkness. To test the effects of Sea-Nine throughout the embryonic development of P. lividus 350–400 newly fertilized eggs obtained by in vitro fertilization were delivered into glass vials with airtight teflon-lined caps containing the experimental solutions (4–8–16–20–30–32–40–50–60–64–70–80–90–100– 128–256–512 nM) and were incubated at 20 ◦ C for 48 h in darkness. After the incubation period larvae were preserved by adding a few drops of 40% buffered formalin and the percentage of fully developed 4-arm pluteus (n = 100) and the larval growth of the pluteus larvae (n = 35) were recorded. Alternatively, a batch of fertilized eggs was exposed to the Sea-Nine experimental concentrations (12–25–50–100–200– 400–800 nM) in order to follow the effects of Sea-Nine during different stages of the embryonic development of P. lividus. Aliquots of the egg suspensions containing approximately 200–250 embryos were fixed at various times after fertilization (1, 1.5, 2, 2.5, 3, 4, 5, 6, 20, 32 and 48 h), and the percentage of embryos at 1, 2, 4, 8, 16, 32, 64 cs, blastulae, gastrulae, pre-pluteus and pluteus stages was recorded. 2.5. Degradation of Sea-Nine Two different experiments were conducted to test the degradation of Sea-Nine. In the first experiment open borosilicate volumetric flasks with the experimental solutions (4, 8, 16, 32, 64, 128 and 256 nM) were exposed for 8 h to natural sunlight conditions during the first week of August at 42◦ 10 N, 8◦ 41 W. Light conditions outdoors were sunny during the experimental

period and maximum irradiance intensities were 1207.3 W/m2 . Sea-Nine solutions were simultaneously incubated for 8 h in dark conditions at 20 ◦ C. A second experiment was carried out to test the degradation of Sea-Nine in absence of light by incubating Sea-Nine solutions in a culture chamber at 20 ◦ C for 28 h in dark conditions. Artificial seawater used in the degradation experiments was allowed to grow a bacteria population for 3 weeks. The seawater bacteria were stained in solution with DAPI (4 ,6 -diamidino-2-phenylindole) and enumerated using a standard epifluorescence microscopy technique (Porter and Feig, 1980). Cell counts indicated a concentration of 1.05 × 106 bacteria/mL. After the exposure to natural light or incubation in darkness, the Sea-Nine solutions were used to incubate P. lividus fertilized eggs for 48 h using the experimental procedure described above. A parallel test with the same batch of embryos was conducted with control fresh Sea-Nine solutions for each experiment. 2.6. Statistical analyses The Lowest Observed Effect Concentrations (LOEC) were determined by one-way ANOVA and Dunnett’s test. Significant differences between parameters for pairs of curves from the degradation experiments were tested using an extra sum-of-squares F-test to determine if curves were statistically indistinguishable (Motulsky and Christopoulos, 2004). The raw data were arcsine-transformed to achieve normality (Hayes, 1991). The EC10 and EC50 (median effective concentration) reducing the recorded biological responses by 10 and 50%, respectively, were calculated according to the probit method. Previously to performing ECX calculations data were normalized to the control mean percentage of normal larvae using Abbot’s formula (Emmens, 1948). Control embryogenesis success was always above 90% normal hatched larvae. Statistical tests were performed according to Newman (1995) and Sokal and Rohlf (1995). 3. Results 3.1. Effects on the fertilization rate The effects of Sea-Nine on the fertilization rate of Paracentrotus lividus are shown in Fig. 1. When treated sperm was used to fertilize untreated eggs a significant reduction on the mean fertilization rate (12%) was observed at a concentration of 400 nM (LOEC) (Fig. 1A). The calculated EC10 and EC50 were 419 and 701 nM (118 and 198 ␮g/L) (Table 1). The fertilization rate decreased by approximately 56% at 800 nM and no fertilization was observed at 1600 nM. Treated eggs fertilized with untreated sperm resulted in a significant reduction of the mean fertilization rate (7%) at 100 nM (LOEC). The fertilization rate was reduced by 40% at 400 nM and by 65% at 800 nM (Fig. 1B). The EC10 and EC50 values were 120 and 410 nM (34 and 116 ␮g/L). Treated eggs fertilized with treated sperm yielded EC10 and EC50 values for the fertilization rate of 282 and 351 nM (80 and 102 ␮g/L). The fertilization rate was significantly reduced

J. Bellas / Aquatic Toxicology 83 (2007) 52–61

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by 60% at 800 nM. The EC10 and EC50 were 174 and 545 nM (49 and 154 ␮g/L). Fertilization of treated eggs with untreated sperm also resulted in a decrease of the percentage of 4-arm pluteus and larval growth (Fig. 2C and D). The percentage of 4-arm pluteus was significantly reduced (15%) at 128 nM (LOEC) and was 0 at 400 nM. EC10 and EC50 values were 212 and 281 nM (60 and 79 ␮g/L). The larval growth was significantly reduced by 75% at 100 nM (LOEC) and calculated EC10 and EC50 values were 142 and 256 nM (40 and 72 ␮g/L). When treated eggs were fertilized with treated sperm the percentage of 4-arm pluteus decreased significantly by 30% at 128 nM (LOEC) and was 0 at 200 nM (Fig. 2E). The obtained EC10 and EC50 values were 85 and 123 nM (24 and 35 ␮g/L). The mean larval growth was significantly reduced by 30% at 50 nM (LOEC) and by 75% at 128 nM (Fig. 2F). The EC10 and EC50 values were 39 and 89 nM (11 and 25 ␮g/L). The exposure of P. lividus fertilized eggs to Sea-Nine concentrations throughout the embryonic development resulted in a significant reduction of the percentage of 4-arm pluteus and larval growth (Fig. 3A and B). The mean percentage of 4-arm pluteus was reduced by 10% at 60 nM (LOEC) and by more than 90% at 80 nM. The calculated EC10 and EC50 were 63 and 69 nM (18 and 19 ␮g/L). The mean larval growth was reduced significantly by 15% at 30 nM (LOEC) and yielded EC10 and EC50 values of 25 and 73 nM (7 and 21 ␮g/L). The development of P. lividus embryos incubated in Sea-Nine solutions is shown in Fig. 4. Control embryos, and those exposed to Sea-Nine concentrations between 12 and 50 nM underwent similar development at the sampling times. Embryos exposed to 100 nM showed a delay in development compared to controls, whilst higher concentrations caused an irreversible impairment of the embryonic development. The embryonic development was arrested at the gastrula stage after 20 h exposure to 200 nM, and at the 8 cs stage after 3 h exposure to 400 nM, whereas embryos exposed to 800 nM did not develop over the 48 h period. An increase in abnormal embryos was also registered at increasing Sea-Nine concentrations from 10% abnormal embryos in controls, to 20% at 100 nM, and 100% at 400 and 800 nM. Fig. 1. Fertilization rate in P. lividus after 45 min exposure of sperm (A), 4 h exposure of eggs (B), and 45 min exposure of sperm and 4 h exposure of eggs (C) to different concentrations (nM) of Sea-Nine (n = 4).

at 200 nM (LOEC) and decreased by approximately 70% at 400 nM. No fertilization was observed above 800 nM. 3.2. Effects on the embryonic development and larval growth Significant effects of Sea-Nine on the embryonic development and larval growth were also observed after fertilization of untreated eggs with treated sperm (Fig. 2A and B). The mean percentage of 4-arm pluteus larvae decreased by 20% at 400 nM (LOEC) and by 60% at 800 nM. Calculated EC10 and EC50 values were 327 and 630 nM (92 and 178 ␮g/L). The mean larval growth was significantly reduced by 25% at 400 nM (LOEC) and

3.3. Degradation of Sea-Nine No significant differences in the percentage of normal larvae or in larval growth were detected after 8 h incubation in darkness of Sea-Nine solutions compared to control fresh solutions (Fig. 5A and B). Calculated EC10 , EC50 and LOEC are shown in Table 2. On the other hand, although no significant differences in the percentage of normal larvae were observed when Sea-Nine solutions were exposed to sunlight for 8 h, significant differences were found in the larval growth in comparison with control solutions (p < 0.0001; F = 11.7) (Fig. 5A and B). However, the results are different at low and high concentrations. Below 64 nM sunlight-exposed Sea-Nine solutions were more toxic than control solutions, and above 64 nM lower toxicity of sunlight-exposed solutions was observed. Based on EC10 and EC50 values, the sunlight-exposed solutions were more toxic than control solutions (Table 2). Similar results were observed

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Table 1 EC10 , EC50 and LOEC (nM) for the biological responses tested EC10

EC50

LOEC

Fertilization rate

Treated sperm (45 min) Treated eggs (4 h) Treated eggs and sperm

419 (363–468) 120 (90–148) 282 (260–298)

701 (641–770) 410 (345–497) 351 (337–364)

400 100 200

Normal 4-arm pluteus

Treated sperm (45 min) Treated eggs (4 h) Treated eggs and sperm Exposed embryos (48 h)

327 (259–385) 212 (191–226) 85 (77–92) 63 (60–64)

630 (555–717) 281 (268–293) 123 (116–130) 69 (67–72)

400 128 128 60

Larval growth

treated sperm (45 min) treated eggs (4 h) treated eggs and sperm exposed embryos (48 h)

174 (89–251) 142 (103–171) 39 (29–46) 25 (22–28)

545 (405–767) 256 (224–291) 89 (79–101) 73 (68–78)

400 100 50 30

The 95% confidence intervals (95CI) are given in brackets.

when Sea-Nine solutions were incubated for 28 h in darkness. No significant differences in toxicity were observed based on the percentage of normal larvae, but significant differences in larval growth were detected (p < 0.005; F = 5.8) (Fig. 5C and D). Based on EC10 and EC50 values, Sea-Nine solutions incubated for 28 h in darkness were more toxic than control solutions to the larval growth (Table 2). 4. Discussion Sea-Nine is a new antifouling substance that has recently been marketed as an environmentally acceptable alternative to organotin compounds. This broad-spectrum biocide is highly effective against a wide range of fouling organisms from bacteria to barnacles (Jacobson and Willingham, 2000). Also, interesting physicochemical features for the use of Sea-Nine in antifouling coatings such as fast degradation rates in water and sediment have been reported, but little is known about the toxicity of SeaNine to non-target marine invertebrates and, in particular, to their sensitive early stages of development. This information is urgently needed due to the expected increase on the use and the environmental levels of this biocide in particular after the TBT phase out (Voulvoulis et al., 1999; Konstantinou and Albanis, 2004; Yebra et al., 2004). In this study, we have investigated the

toxic effects of Sea-Nine during the early stages of development of the sea urchin Paracentrotus lividus in order to contribute to the assessment of the risk associated with the use of this alternative biocide. The preexposure of P. lividus gametes for 45 min (sperm) and 4 h (eggs) to Sea-Nine resulted in a significant decrease of the fertilization rate. The highest toxicity was observed when fertilization was conducted with treated sperm and eggs. Calculated EC10 and EC50 values ranged from 120 to 419 nM and from 351 to 700 nM, respectively. To our knowledge, we report the first data on the toxicity of Sea-Nine to the fertilization rate of marine organisms. The study of the effects of sperm and eggs preexposure on the subsequent embryonic development and larval growth also showed high toxicity of Sea-Nine, indicating that a very short exposure to relatively low concentrations (40–50 nM) of this biocide (45 min–4 h) may cause deleterious effects on the offspring of sea urchins. Furthermore, when fertilization was conducted with treated eggs or treated sperm and eggs lower EC10 and EC50 values were obtained for the percentage of 4-arm pluteus and larval growth than for the fertilization rate. Therefore, preexposure of eggs to Sea-Nine induced a transmissible damage to the embryo, impairing the embryonic development, giving rise to embryonic and larval malformations, and decreasing the larval growth. This inducible damage of Sea-Nine on

Table 2 EC10 , EC50 and LOEC (nM) for the degradation of Sea-Nine

Experiment 1 Normal 4-arm pluteus

Larval growth

Experiment 2 Normal 4-arm pluteus Larval growth

EC10

EC50

Control solution Sunlight-exposed solution (8 h) Dark incubated solution (8 h)

64 (63–66) 64 (61–67) 65 (63–66)

77 (72–92) 89 (87–92) 80 (77–84)

64 64 64

Control solution Sunlight-exposed solution (8 h) Dark incubated solution (8 h)

24 (16–30) 8.6 (5.3–12) 19 (14–24)

71 (59–85) 64 (52–81) 65 (57–76)

32 16 32

Control 4-arm pluteus Dark (28 h) 4-arm pluteus

68 (60–75) 63 (59–66)

91 (84–99) 77 (72–93)

128 64

Control larval growth Dark (28 h) larval growth

35 (28–42) 22 (15–29)

95 (84–107) 68 (55–85)

64 64

The 95% confidence intervals (95CI) are given in brackets.

LOEC

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Fig. 2. Percentage of 4-arm pluteus and larval growth in P. lividus after 45 min exposure of sperm (A and B), 4 h exposure of eggs (C and D), and 45 min exposure of sperm and 4 h exposure of eggs (E and F) to different concentrations (nM) of Sea-Nine (n = 4).

P. lividus eggs has been previously reported for metals and for the booster biocide zinc pyrithione on marine invertebrates (Kobayashi, 1981,1995; Pagano et al., 1996; Bellas et al., 2004; Bellas, 2005). This result was not observed when fertilization was conducted with treated sperm and untreated eggs, which indicates that no transmissible damage is induced to the embryo from treated sperm. The development of P. lividus embryos exposed to Sea-Nine was followed through different stages in order to understand the toxic effects of Sea-Nine on the embryogenesis. The embryos developed with good synchrony in controls whilst at increasing Sea-Nine concentrations the synchrony of cellular divisions was lost. A slight delay in development was observed at 25 and

50 nM after 3 and 4 h exposure, although embryos developed normally to the pluteus stage. However, embryos exposed to 100 and 200 nM developed normally only until the gastrula stage when the embryogenesis was impaired, at 400 nM the embryogenesis was inhibited at the 8-cell stage, and at the highest tested concentration (800 nM) only undeveloped eggs were observed. Therefore, low Sea-Nine concentrations may cause a delay in the embryonic development whereas higher concentrations completely impair the embryogenesis at critical stages such as the third cleavage and the formation of the gastrula. P. lividus fertilized eggs exposed to Sea-Nine throughout the embryonic development (48 h) were affected at 63 nM (EC10 ) and the EC50 was 69 nM. The larval growth was

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Fig. 3. Percentage of 4-arm pluteus larvae (A) and larval growth (B) after 48 h exposure of fertilized eggs of P. lividus to different concentrations (nM) of Sea-Nine (n = 4).

reduced at 25 nM (EC10 ) with an EC50 value of 73 nM, being the most sensitive response tested here. The effective concentrations reported here for the embryonic development of P. lividus are within the range of values previously registered for other marine invertebrates. EC50 values for oyster embryogenesis ranged from 25 to 85 nM (Shade et al., 1993; Willingham and Jacobson, 1996), and from 7 to 38 nM for mussel embryogenesis (Shade et al., 1993; Bellas, 2006). Ascidian embryos and larvae (EC50 = 153–372 nM) and barnacle larvae (EC50 = 1205 nM) were less sensitive to Sea-Nine (Jacobson and Willingham, 2000; Bellas, 2006). Regarding sea urchins, Kobayashi and Okamura (2002) reported extreme toxicity of Sea-Nine to the embryogenesis of Anthocidaris crassispina at 109 -fold lower concentrations than those reported here (LOEC = 10 fg/L = 0.03 fM). In contrast, a recent study (Bellas, 2006) reported similar values to those reported here for P. lividus embryogenesis (EC50 = 43–88 nM). This difference in toxicity of Sea-Nine to sea urchin species is difficult to explain. The bioassay methodology of the present and the above-mentioned studies is similar, and although Kobayashi and Okamura (2002) recorded minor morphological abnormalities that were not considered here, we report a sublethal response in early larval growth which is more sensitive to toxicants than the embryogenesis success in sea urchins. Marine invertebrate and in particular bivalve and sea urchin embryos and larvae are among the most sensitive organisms to Sea-Nine. Therefore, since the sea urchin embryo–larval bioassay is ecologically relevant, easily standardizable, quick and cost-effective, this bioassay may be a suitable tool to be used in standard toxicity testing for the evaluation of antifouling biocides. The reported fast degradation of Sea-Nine in the aquatic environment is the most interesting physicochemical feature for its use in antifouling technologies as an alternative to organotinbased coatings (Jacobson and Willingham, 2000; Thomas et al., 2002; Konstantinou and Albanis, 2004). Nevertheless, high concentrations of this biocide have been measured in marinas around Europe. The highest levels of Sea-Nine were reported in the Spanish Mediterranean (2600–3700 ng/L), and significant concentrations were also detected in Greece (49 ng/L) and Denmark (283 ng/L) (Mart´ınez et al., 2000, 2001; Thomas, 2001; Sakkas et al., 2002a). Several studies have investigated the

degradation of Sea-Nine in the aquatic environment and this compound is considered to be degraded both biologically and chemically, with reported half-lives between less than 1 h to several days (Jacobson et al., 1993; Callow and Willingham, 1996; Thomas et al., 2002). Although the main degradation pathway seems to be the biodegradation, Sea-Nine was also reported to present rapid photodegradation leading to the formation of several photoproducts (Willingham and Jacobson, 1996; Sakkas et al., 2002b). For instance, Shade et al. (1993) reported a biotic half-life of Sea-Nine (10 ␮g/L = 35 nM) of 11 h in seawater with 7 × 104 bacteria/mL, whereas tests performed with seawater with lower number of bacteria (80–600 bacteria/mL) resulted in longer half-lives (76–187 h). In the present study (1.05 × 106 bacteria/mL), Sea-Nine solutions were still highly toxic after direct sunlight exposure (8 h) or after 28 h incubation in darkness. Furthermore, we registered a slight increase in the toxicity of Sea-Nine on the larval growth in comparison with control solutions. The EC10 and EC50 for the larval growth were 8.6 and 64.1 nM after sunlight exposure whereas control solutions yielded EC10 and EC50 values of 23.6 and 70.7 nM. Thus, although the degradation of Sea-Nine in seawater may be fast, the eventual degradation products may exert some toxic effects to marine invertebrate early developmental stages. Also, degradation may be lowered in seawater with low bacteria concentration. Moreover, high levels of Sea-Nine detected in European marinas indicate that the degradation pathway is not rapid enough to remove Sea-Nine used in antifouling paints from the water during boating season. The environmental risk of a biocide may be calculated using a Risk Quotient (RQ) quantified as PEC/PNEC ratios, where PEC is the Predicted Environmental Concentration of the biocide and the PNEC (Predicted No Effect Concentration) is the concentration of the biocide that causes no deleterious effect to the environment. To evaluate the risk of Sea-Nine to the early developmental stages of P. lividus we can estimate the PNEC values from the lowest effect concentrations (EC10 ) for the biological responses tested here, applying an assessment factor of 10 (OECD, 1992), and using the worst case concentrations mentioned above as PEC. RQ values lower than 1 indicate that no adverse effects are anticipated, whilst RQ values greater than 1 indicate that adverse effects are likely to occur. The calculated

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Fig. 4. Embryonic development in P. lividus after 1 (A), 1.5 (B), 2 (C), 3 (D), 4 (E), 5 (F), 6 (G), 20 (H), 32 (I), and 48 h (J) exposure to different concentrations (nM) of Sea-Nine (n = 4).

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Fig. 5. (A and B) Percentage of 4-arm pluteus larvae and larval growth after 48 h exposure of fertilized eggs of P. lividus to Sea-Nine solutions exposed to direct sunlight for 8 h (squares) and to Sea-Nine solutions incubated in dark conditions for 8 h (triangles). Filled-in circles represent control Sea-Nine solutions (n = 4). (C and D) Percentage of 4-arm pluteus larvae and larval growth after 48 h exposure of fertilized eggs of P. lividus to Sea-Nine solutions incubated in dark conditions for 28 h (squares). Filled-in circles represent control Sea-Nine solutions (n = 4). Error bars represent standard deviations.

RQ obtained here is 5.5 (and 15.2 if we consider the degradation studies), therefore, on the basis of the present study, Sea-Nine concentrations in those marinas are likely to cause deleterious effects on P. lividus populations.

Nine is expected after the complete phase out of organotin-based antifouling paints.

5. Conclusions

I am grateful to Rocio Rendo for her technical assistance during experiments. I was supported by a Juan de la Cierva contract from the Spanish Ministry of Education and Science.

This study showed that short exposures to relatively low concentrations of Sea-Nine cause deleterious effects on sea urchin gametes, embryos and larvae. A recent paper reported TBT toxicity on the embryonic development of P. lividus at 0.9 nM (EC50 ) (Bellas et al., 2005) whereas the lowest EC10 reported here for Sea-Nine is 24 nM (and 8.6 nM after 8 h sunlight exposure). Therefore, TBT seems to be one or two orders of magnitude more toxic than Sea-Nine to sea urchin embryos and larvae. Furthermore, TBT presents less favourable physicochemical features for its use as an antifoulant such as a high environmental stability (Bryan and Gibbs, 1991). Although Sea-Nine may present a better environmental profile than TBT, maximum concentrations reported in literature are sufficient to endanger early developmental stages of sea urchin, and its suitability as an alternative antifoulant is questioned. As has been stated in previous studies, the continued use of Sea-Nine in the future should be considered carefully, and the search for environmentally acceptable antifouling methods should continue (Larsen et al., 2003), even more if we bear in mind that an increase on the use of Sea-

Acknowledgements

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