Chlorinated and brominated organic contaminants and metabolites in the plasma and diet of a captive killer whale (Orcinus orca)

Marine Pollution Bulletin 58 (2009) 1078–1095

Contents lists available at ScienceDirect

Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul

Baseline

Edited by Bruce J. Richardson The objective of BASELINE is to publish short communications on different aspects of pollution of the marine environment. Only those papers which clearly identify the quality of the data will be considered for publication. Contributors to Baseline should refer to ‘Baseline—The New Format and Content’ (Mar. Pollut. Bull. 42, 703–704).

Chlorinated and brominated organic contaminants and metabolites in the plasma and diet of a captive killer whale (Orcinus orca) Erin R. Bennett a, Peter S. Ross b, David Huff c, Mehran Alaee d, Robert J. Letcher a,* a

Great Lakes Institute for Environmental Research (GLIER), University of Windsor, Windsor, ON, Canada N9B 3P4 Institute of Ocean Sciences, Fisheries and Oceans Canada, P.O. Box 6000, Sidney, BC, Canada V8L 4B2 c Vancouver Aquarium Marine Science Centre, P.O. Box 3232, Vancouver, BC, Canada V6B 3X8 d Aquatic Ecosystem Protection Research Division, Water Science and Technology Directorate, Environment Canada, Burlington, ON, Canada L7R 4A6 b

As a consequence of their often long life spans and high trophic levels in aquatic food webs, many marine mammal species are vulnerable to accumulating high concentrations of lipophilic and persistent organic pollutants (POPs) including PCBs and organochlorine pesticides (e.g. DDTs, chlordanes). In the northeastern Pacific Ocean, two populations of fish-eating killer whales (Orcinus orca), the northern and southern residents, are at risk due to diminished food supply, noise and disturbance, and surprisingly high concentrations of toxic chemicals (Ross, 2006; Ford and Ellis, 2006). While the marine mammal-eating counterparts, the ‘‘transient” killer whales, are considered as among the most PCB-contaminated marine mammals in the world, the residents are also very contaminated, with average PCB concentrations in males of 37 and 146 mg/kg (lipid weight), respectively (Ross et al., 2000). This compares with the heavily contaminated beluga whales (Delphinapterus leucas) of the St. Lawrence estuary, with males averaging approximately 80 mg/kg (Hobbs et al., 2003; Letcher et al., 2000a; McKinney et al., 2006). While feeding ecology and trophic position partly explains this contamination in killer whales, long lifespan represents an important additional factor (Hickie et al., 2007). Polybrominated diphenyl ether (PBDE) flame retardants represent a more recent concern in marine mammals, with reports of these chemicals in blubber biopsies of killer whales from the northeastern Pacific Ocean (Rayne et al., 2004; Ross, 2006; Krahn

* Corresponding author. Present address: Wildlife and Landscape Science Directorate, Science and Technology Branch, Environment Canada, National Wildlife Research Centre, Carleton University, Ottawa, ON, Canada K1A 0H3. Tel.: +1 613 998 6696; fax: +1 613 998 0458. E-mail address: [email protected] (R.J. Letcher).

et al., 2007). PBDEs have emerged as a significant global concern for fish and marine mammals (Ross et al., 2009). The bioaccumulation and fate of POPs, including PCBs and PBDEs, in biota including (marine) mammals are influenced by a number of factors, including diet, age, gender, genetic variation, and metabolism. Metabolic capacity towards PCBs, PBDEs, and other organohalogens and subsequent metabolite formation is a function of xenobiotic-metabolizing enzymes, with the activation, substrate selectivity and profile of important Phase I cytochrome P450 monooxygenases (CYPs) and Phase II conjugation-mediating enzyme systems being key (Letcher et al., 2000a,b; Hakk and Letcher, 2003; Lewis et al., 1998). Cetacean species, including odontocetes, appear to have a limited ability to metabolize, e.g., non-planar PCB and PBDE congeners, as CYP2B-like enzyme activity is low compared to other vertebrates, including birds, pinnipeds, polar bears, and terrestrial mammals (Goksøyr, 1995; McKinney et al., 2004). Persistent hydroxylated (OH) PCB metabolites reported in e.g. mammals and birds are formed via CYP-mediated PCB metabolism (Letcher et al., 2000a,b). Methoxylated (MeO) PBDEs and OH-PBDEs, the latter of which may be formed metabolically from parent PBDEs in an organism, are also known to be accumulated in e.g., fish and marine mammals from natural marine sources (such as algae and sponges) (Hakk and Letcher, 2003; Teuten et al., 2005). OH-PCBs and/or OH-PBDEs have been reported in the liver of St. Lawrence River beluga whales (McKinney et al., 2006), and plasma of bottlenose dolphins (Tursiops truncatus) from the Atlantic coast of the United States (Houde et al., 2006), and the cerebrospinal fluid of short-beaked common dolphins (Delphinus delphis) and/or Atlantic white-sided dolphins (Lagenorhynchus acutus) from the western North Atlantic of the United States (Montie et al., 2009).

1079

Baseline / Marine Pollution Bulletin 58 (2009) 1078–1095

Various tribromo- to hexabromo-OH-PBDEs have been reported in rain, snow, surface water, and/or rain, as well as sewage treatment plant outfalls in southern Ontario in the basin of the Laurentian Great Lakes of North America (Hua et al., 2005; Ueno et al., 2008). Triclosan (2,4,40 -trichloro-2 20 -hydroxydiphenyl ether) is another halogenated phenolic contaminant (HPC), and is a synthetic, broad-spectrum antibacterial agent that is widely used in deodorants, toothpastes, cosmetics, polymers, and textiles. Over 95% of triclosan use has been in consumer products that are disposed of through residential drains (Reiss et al., 2002). To our knowledge, studies on triclosan in aquatic organisms has been limited to lower level organisms such as algae, crustaceans, and fish close to wastewater treatment plant discharges (Adolfsson-Erici et al., 2002; Hua et al., 2005; Valters et al., 2005). While obtaining blood samples from free-ranging killer whales is virtually impossible as a consequence of legal, ethical, and logistical constraints in much of the world, aquarium specimens may provide a glimpse into the relationship between POP and POP metabolite patterns in killer whales and their food. In the present study we measured organochlorines pesticides, PCBs, PBDEs, and various HPCs such as OH-PCBs, OH-PBDEs, pentachlorophenol and triclosan in whole blood from a captive adult female killer whale that had been fed a diet of wild-caught Pacific herring (Clupea pallasii), squid (Mastigoteuthis flammea) and capelin (Mallotus villosus) during a 20 year period in captivity. In addition, a pooled sample of her Pacific herring diet was analyzed for these same contaminants. An adult female killer whale (Bjossa) was captured in November, 1980 (404 cm in length) and the estimated year of birth was 1974. Bjossa died in 2001 after an infection, approximately one year after the blood sample was collected for the present study. She had been kept in captivity for 20 years at the Vancouver Aquarium Marine Science Centre (British Columbia, Canada), and had been fed a diet consisting primarily of wild Pacific herring. A single blood sample (10 mL) was collected in 2000, and plasma was collected following centrifugation for 20 min. at 400 g of heparinized whole blood. Plasma was stored at 4 °C until chemical analysis. For the present study, 1.4 mL of plasma was available and used. Samples of wild Pacific herring (n = 3), the principal food item for Bjossa, were also used in this study. The herring was purchased from a commercial fishery operating in the Strait of Georgia between mainland British Columbia and Vancouver Island, and stored at 20 °C until use. As detailed in the Supplemental information section, PCBs (47 congeners), OC pesticides (16 OCs; p,p’-DDT, p,p’-DDD, p,p’-DDE, a-, b-, and c-hexachlorocyclohexane (HCH), chlordanes (CHLs; oxychlordane, trans-chlordane, cis-chlordane, trans-nonachlor, cis-nonachlor and heptachlor epoxide), pentachlorobenzene and hexachlorobenzene, and Mirex and photo-Mirex), PBDEs (39 congeners), OH-PBDEs (15 congeners), MeO-PBDEs (15 congeners), OH-PCBs (14 congeners), 4-OH-heptachlorostyrene (4-OH-HpCS), triclosan and PCP were monitored for and the determination of these compounds has been described in extensive detail elsewhere (Gebbink et al., 2008; McKinney et al., 2006; Montie et al., 2009; Sandala et al., 2004; Valters et al., 2005). Congeners that were detected are listed in the footnotes in Table 1. Briefly, about 1.3 grams of plasma was spiked with standard solutions of CB83 and CB122, 4-OH-CB72, 40 -OH-BDE17 and BDE30. HPCs in the isolated phenolic phase were derivatized using diazomethane to MeO-containing analogues and purified on a silica/sulfuric acid (22%) column (5 g) and eluted with 50 mL of DCM:hexane. The lipid content in plasma was determined using a colorimetric method and pure olive oil as the calibration standard (Gebbink et al., 2008; Sandau et al., 2000). See the Supplementary information for further details and materials and methods.

Table 1 Concentrations of chlorinated and brominated organohalogensa in killer whale blood and whole wild Pacific herring homogenates.

Sample wet weight Lipid content (%)

Killer whale plasma (n = 1)

Herring homogenate (n = 3)

1.21 g 0.0781

4.34 g 12.1

Compounds

ng/g (wet weight)

ng/g (wet weight)

R-PCBb R-OH-PCBc R-CBz R-HCH R-CHL R-DDT R-Mirex

31.9 6.5 1.2 1.0 3.2 44.8 0.1 0.8 0.3 0.3 9.0 pg/g (wet weight)

10.0 n.d. 1.6 3.1 3.2 19.3 n.d. 1.4 0.1 n.d. n.d. pg/g (wet weight)

1070 7.6 7.92

n.a. n.a. n.a.

Dieldrin OCSd PCP Triclosan

R-PBDEe 6-OH-BDE47f 6-MeO-BDE47f

n.d. = not detected; n.a. = not analyzed. a The compounds and congeners are described in detail in methods section and in the Supplementary data section. b With the exception of CB28, 31 and 195, all 47 PCB monitored for were detected. However, CB52, 101, 118, 153, 138 and 180 mainly comprised the R-PCB. c Of the 14 congeners monitored for, the ROH-PCB concentration is comprised of 4’-OH-CB104, 4’-CB-CB120, 4’-OH-CB130, 4-OH-CB187 and 4-OH-CB193. d OCS is an abbreviation for octachlorostyrene. e Of the 39 PBDE congeners monitored for, the SPBDE concentrations is comprised of BDE47, 49, 88, 100, 138, 153, 154, 181 and 183. f Of the OH-PBDEs (15 congeners) and MeO-PBDEs (15 congeners) monitored for, only 6-OH-BDE47 and 6-MeO-BDE47 were detected.

Obtaining samples from wild killer whales is constrained by substantial technical, legal and ethical challenges. We had the exceptional opportunity to obtain blood from a single Pacific killer whale that had been maintained in captivity over the course of 20 years, and had been fed a diet of fish and squid obtained from the coastal waters of British Columbia. For the organohalogens and metabolites determined, results can provide insight into contaminants that are found in coastal food webs of British Columbia, such that they are likely to be found in free-ranging marine mammals in the same region. Bjossa gave birth three times in her 25 year lifetime, which appears to be consistent with the birth frequency of her wild conspecifics, where it has been reported that young are born at intervals of three to eight years (Clark and Odell, 1999). Female mammals can reduce their accumulated burden (relative to males) of lipophilic and bioaccumulative POPs, which has been shown for marine mammals including killer whales (Ylitalo et al., 2001). p,p’-DDE (which comprised most of the R-DDT concentration) and R-PCBs were the major contaminants in both killer whale blood and herring homogenate (Table 1). The PCB congener pattern in plasma (Fig. 1) was similar to the patterns observed in blubber biopsies from free-ranging killer whales, despite herring being a relatively minor component of the diet of wild killer whales (Ford et al., 1998; Hickie et al., 2007; Ross et al., 2000). Ross et al. (2000) reported that the six dominant PCB congeners (CB52, 101, 118, 153, 138, and 180) made up approximately 50% of the R-PCB concentration in both the resident and transient killer whale blubber biopsies. In our study, the sum of these six congeners made up 46% of the R-PCBs in the blood plasma sample (Fig. 1). Furthermore, the recalcitrant PCB congeners, CB153 and 138 dominated the congener profile observed in the blood sample, which is consistent with other reports of odonocete cetaceans (McKinney et al., 2006). The common basic PCB profile observed in our captive killer

1080

Baseline / Marine Pollution Bulletin 58 (2009) 1078–1095

14

A 12

PCB Congener Conc. as % of the Sum PCB Conc.

10 8 6 4 2 0

12

B 10 8 6 4 2 0 206 194 195 203 201 170/190 180 172 200 171/202/156 177 174 128 183 182/187 129/178 158 138 179 141 105 153 146 118 149 151 110 97 99 101 60 66/95 70 74 64 42 44 49 52 28

PCB Congener

Fig. 1. PCB congener pattern in killer whale plasma (A) (n = 1) and whole wild Pacific herring homogenate (B) (n = 3). The pattern is expressed as a percentage of the R-PCB concentration. See Table 1 for the sum-PCB concentration.

whale and its wild counterparts provides a strong indication of the definitive role that metabolism plays in shaping the pattern of PCBs and other organohalogens regardless of the diet. PCP and five OH-PCB congeners (i.e., 40 -OH-CB104, 40 -OHCB120, 30 -OH-CB130, 4-OH-CB187, and 4-OH-CB193) were detected and identified in the killer whale plasma (Fig. 2). In addition, three unidentified tetra-, penta, and hexa-chloro-OH-PCB congeners were detected (Fig. 2). In the case of the unidentified hexachloro-OH-PCB the electron capture negative ionization (ECNI)-mass spectrometry response was large (see also Supplemental information), which indicates that the R-OH-PCB concentration in Table 1 is underestimated. That is, an estimate of the R-OH-PCB concentration including these three unknowns would be 7.94 ng/g (ww) as opposed to 6.52 ng/g (ww) we quantified (Table 1). The major OH-PCB congeners quantified in whale plasma, 40 -OH-CB120 and 4-OH-CB187, are most likely metabolically-derived from CB118 and CB187, respectively (Kawano et al., 2005; Verreault et al., 2008). In the killer whale blood, the R-OH-PCB metabolites ranked fourth among contaminants measured (Table 1). The R-OH-PCB concentration was less than R-PCBs, R-DDTs, and triclosan, but was much higher than R-CHLs and R-HCHs. Furthermore, the R-OH-PCB to R-PCB concentration ratio was 0.20, and was comparable to ratios reported in other marine mammals. Ratios were reported as being as high as 8.00 in the plasma of polar bears from the Canadian and East Greenland Arctic and as low as 0.002 in ringed seal (Pusa hispida) plasma from the Resolute Bay area, beluga whale from western Hudson Bay and the St. Lawrence estuary,

and bottlenose dolphins from Sarasota Bay and Indian River Lagoon in Florida (Gebbink et al., 2008; Houde et al., 2006; McKinney et al., 2006; Sandala et al., 2004; Sandau et al., 2000). It is probable that the present killer whale had the capacity to oxidatively metabolize PCBs to OH-PCBs, as OH-PCBs were not likely accumulated as indicated by the lack of detectablity of these metabolites in its diet (Table 1). PBDEs have not been previously reported in plasma (or blood) of killer whales, although several studies have characterized their profiles in blubber biopsies (Rayne et al., 2004; Ross, 2006; Krahn et al., 2007). It is difficult to make a direct comparison of PBDE (or other POP) concentrations in plasma versus blubber without knowing the concentration relationship between plasma and blubber. Rayne et al. (2004) reported that PBDE congener patterns in free-ranging killer whales from the northeastern Pacific Ocean were dominated by BDE47. This is consistent with our results for plasma from Bjossa where the BDE47 concentration was 660 pg/g (ww) and accounted for 62% of the R-PBDE concentration (Table 1). Relative to other classes of organohalogens, the R-PBDE concentration in the plasma of Bjossa was low compared to R-PCBs, but comparable to R-CHLs, R-CBzs, and R-HCHs (Table 1). In the whales from the Rayne et al. (2004) report, the R-PBDE mean concentrations in blubber biopsies for transient females were 885 ± 706 ng/g (lw) and 415 ± 676 ng/g (lw) for resident females. Of the OH-PBDE and MeO-PBDE congeners monitored, only 6OH-BDE47 and 6-MeO-BDE47 were detected and quantifiable in the killer whale plasma. The very low 6-OH-BDE47 and 6-MeO-

1081

Baseline / Marine Pollution Bulletin 58 (2009) 1078–1095

12

2

Peak #7 503 499

5 1 1 10

3 1 8

6

14

16

18

9

7

4

20

22

13

24

26

28

GC/ECD: retention time (min.)

591 Peak #12 81

116 433

100

150

200

250

300

350

400

450

510 525 500

550

600

mass to charge ratio (ECNI response) Fig. 2. GC-lECD chromatogram of the halogenated phenolic compound (HPC) fraction from killer whale plasma (derivatized to their corresponding MeO-compounds). Peak assignment is as follows: 1 – PCP; 2 – unidentified tetra-OH-PCB; 3 – 4-OH-CB72 (ISTD); 4 – 4’-OH-CB104; 5 – unidentified penta-OH-PCB; 6 – 4’-OH-CB120; 7 – unidentified HPC (see inset); 8 – 3’-OH-CB130; 9 – unidentified hexa-OH-PCB; 10 – 4-OH-CB187; 11 – 4’-OH-CB159 (ISTD); 12 – unidentified HPC (see inset); 13 – 4-OH-CB193.

BDE47 concentrations (Table 1) equated to 6-OH-BDE47 to R-PBDE and 6-MeO-BDE47 to R-PBDE concentration ratios of 0.0071 and 0.0074, respectively. This suggests that killer whales have a low capacity to oxidatively metabolize PBDEs to OH-PBDEs were not analyzed in herring, but the killer whale had a limited ability to slowly metabolize PBDEs, or levels of OH-PBDEs in herring were present but too low for us to be able to detect (Table 1). However, low level accumulation of OH-PBDEs could have resulted from their presence in their aquarium water, as OH-PBDEs have been reported in surface waters and sewage treatment plant outfalls (Ueno et al., 2008). These low OH-PBDE and MeO-PBDE concentrations and their ratio to PBDEs in the present killer whale are consistent with reports in the plasma of other marine mammals, including Canadian beluga whales (McKinney et al., 2006) and even East Greenland polar bears (Gebbink et al., 2008). However, in a recent report by Kelly et al. (2008) blood samples collected from western Arctic beluga whales had R-MeO-PBDE to R-PBDE ratios ranging from 4.1 to 5.9, which likely reflects a naturally accumulated source of MeO-PBDEs in their diet (Teuten et al., 2005). PCP, a commonly used wood preservative and pesticide ingredient, was also detected in killer whale plasma (Fig. 2). The PCP concentration in killer whale plasma (0.28 ng/g, wet weight) was similar to that observed in Canadian and/or East Greenland polar bears (0.21 ng/g, wet weight) and ringed seals (0.24 ng/g, wet weight) (Sandala et al., 2004; Sandau et al., 2000). Two other major peaks were also detected in the GC-lECD chromatogram of the whale HPC fraction (Fig. 2). Fullscan GC–MS(ECNI) mass spectral data (50–600 m/z) revealed that one peak was a chlorinated HPC (Fig. 2, Peak #7), while the other late eluting peak was identified as an HPC containing 3 bromine and 2 chlorine atoms (Fig. 2, Peak #12). Numerous minor peaks presumed to be derivatized HPCs were also present in the GC-lECD chromatogram (Fig. 2), but were too low to obtain reliable MS data.

Triclosan was also found in the killer whale plasma at a concentration exceeded only by R-PCBs and R-DDTs, highlighting what appeared to be a significant accumulation of this antimicrobial compound (Table 1). Its lack of detection in the herring fed to this whale is likely due to its presence at concentrations less than our detection limit in herring and a significant biomagnification in killer whales, or possibly to its presence in another dietary item, since triclosan is considered a persistent, bioaccumulative substance (Valters et al., 2005). Despite the likely presence of triclosan in wastewater discharge (Adolfsson-Erici et al., 2002; Hua et al., 2005), it is unlikely that that this whale was significantly exposed to this compound via its aquarium water. OH-PCB metabolites, metabolic and/or naturally accumulated OH-PBDEs and MeO-PBDEs, PCP and triclosan were present in a captive killer whale and in some instances exceeding the concentrations of legacy organochlorine contaminants. Since this captive killer whale subsisted on a diet harvested from coastal British Columbia, we speculate that free-ranging resident and transient killer whales are exposed to some of the same HPCs observed in our study. The biological significance of the formation and presence of these compounds in killer whales is not clear. However, the thyroid hormone-disrupting potential of HPCs and other organohalogens suggest that killer whales may be at increased risk of thyroid-associated and possibly other endocrine-related effects (Brouwer et al., 1998). Meerts et al. (2004) showed that prenatal exposure of rat pups to 4-OH-CB107 can cause deficits in locomotor activity and the auditory system. OH-PCBs and OH-PBDEs have been shown to competitively bind with the thyroid hormone transport protein, transthyretin (TTR) from humans (Meerts et al., 2000; Ucán-Marín et al., 2009). Triclosan has been reported to interfere with thyroid hormone (thyroxine) metabolism in laboratory rodents (Crofton et al., 2007). Our study provides a glimpse at the accumulation, retention and metabolism of several classes of persistent contaminants in a

1082

Baseline / Marine Pollution Bulletin 58 (2009) 1078–1095

killer whale. While free-ranging resident killer whales in British Columbia consume a diet consisting largely of salmon rather than herring (Ford and Ellis, 2006), our results may have direct bearing on the types of metabolites found in this species. Nonetheless, the herring provided to Bjossa was of British Columbia origin, such that our results are relevant to marine mammals that consume this forage fish in coastal British Columbia. In addition, our results do reveal a species-specific capacity on the part of killer whales to metabolize POPs and generate certain metabolites. Since Bjossa’s diet was carefully controlled, the disadvantage of our small sample size is made up for by our ability to defensibly compare and contrast POP parent and POP metabolite profiles between predator and prey – something that is virtually impossible in free-ranging marine mammals. Acknowledgements Funding for this project was provided by the Species at Risk Act (SARA) program of Fisheries and Oceans Canada (PSR) and by the Natural Science and Engineering Research Council (NSERC) of Canada and the Canada Research Chairs (CRC) Program (RJL). We thank Ms. Ivy D’Sa (Canada Centre for Inland Waters, Environment Canada) for completing the GC/HRMS analysis. We also wish to thank Clint Wright of the Vancouver Aquarium Marine Science Center. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.marpolbul.2009.05.005. References Adolfsson-Erici, M., Petterson, M., Parkkonen, J., Sturve, J., 2002. Triclosan, a commonly used bactericide found in human milk and in the aquatic environment in Sweden. Chemosphere 46, 1485–1489. Brouwer, A., Morse, D.C., Lans, M.C., Schuur, A.G., Murk, A.J., Klasson-Wehler, E., Bergman, Å., Visser, T.J., 1998. Interaction of persistent environmental organohalogens with the thyroid hormone system: mechanisms and possible consequences for animal and human health. Toxicology and Industrial Health 14, 59–84. Clark, S.T., Odell, D.K., 1999. Nursing parameters in captive killer whales (Orcinus orca). Zoo Biology 18, 373–384. Crofton, K.M., Paul, K.B., DeVito, M.J., Hedge, J.M., 2007. Short-term in vivo exposure to the water contaminant triclosan: evidence for disruption of thyroxine. Environmental Toxicology and Pharmacology 24, 194–197. Ford, J.K.B., Ellis, G.M., 2006. Selective foraging by fish-eating killer whales Orcinus orca in British Columbia. Marine Ecology-Progress Series 316, 185–199. Ford, J.K.B., Ellis, G.M., Barrett-Lennard, L.G., Morton, A.B., Palm, R.S., Balcomb, K.C., 1998. Dietary specialization in two sympatric populations of killer whales (Orcinus orca) in coastal British Columbia and adjacent waters. Canadian Journal of Zoology 76, 1456–1471. Gebbink, W.A., Sonne, C., Dietz, R., Kirkegaard, M., Riget, F.F., Born, E.W., Muir, D.C.G., Letcher, R.J., 2008. Tissue-specific congener composition of organohalogen and metabolite contaminants in East Greenland polar bears (Ursus maritimus). Environmental Pollution 152, 621–629. Goksøyr, A., 1995. Whales, Seals, Fish and Man. Elsevier, Amsterdam, The Netherlands. Hakk, H., Letcher, R.J., 2003. Metabolism in the toxicokinetics and fate of brominated flame retardants (BFRs): a review. Environment International 29, 801–828. Hickie, B.E., Ross, P.S., Macdonald, R.W., Ford, J.K.B., 2007. Killer whales (Orcinus orca) face protracted health risks associated with lifetime exposure to PCBs. Environmental Science and Technology 41, 6613–6619. Hobbs, K.E., Muir, D.C.G., Michaud, R., Béland, P., Letcher, R.J., Norstrom, R.J., 2003. PCBs and organochlorine pesticides in blubber biopsies from free ranging St. Lawrence River Estuary beluga whales (Delphinapterus leucas), 1994–1998. Environmental Pollution 122, 291–302. Houde, M., Pacepavicius, G., Wells, R.S., Fair, P.A., Letcher, R.J., Alaee, M., Bossart, G.D., Hohn, A.A., Sweeney, J., Solomon, K.R., Muir, D.C.G., 2006. Polychlorinated biphenyls (PCBs) and hydroxylated polychlorinated biphenyls (OH-PCBs) in plasma of bottlenose dolphins (Tursiops truncatus) from the Western Atlantic and the Gulf of Mexico. Environmental Science and Technology 40, 5860– 5866.

Hua, W., Bennett, E.R., Letcher, R.J., 2005. Determination of triclosan in waste and surface waters from the Detroit River using high performance liquid chromatography–electrospray–tandem quadrupole mass spectrometry. Environment International 31 (5), 621–630. Kawano, M., Hasegawa, J., Enomoto, T., Onishi, H., Nishio, Y., Matsuda, M., Wakimoto, T., 2005. Hydroxylated polychlorinated biphenyls (OH-PCBs): recent advances in wildlife contamination study. Environmental Science 12, 315–324. Kelly, B.C., Ikonomou, M.G., Blair, J.D., Gobass, F.A.P.C., 2008. Hydroxylated and methoxylated polybrominated diphenyl ethers in a Canadian Arctic marine foods web. Environmental Science and Technology 42 (19), 7069–7077. Krahn, M.M., Hanson, M.B., Baird, R.W., Boyer, R.H., Burrows, D.G., Emmons, C.K., Ford, J.K.B., Jones, L.L., Noren, D.P., Ross, P.S., Schorr, G.S., Collier, T.K., 2007. Persistent organic pollutants and stable isotopes in biopsy samples (2004/2006) from southern resident killer whales. Marine Pollution Bulletin 54, 1903– 1911. Letcher, R.J., Norstrom, R.J., Muir, D.C.G., Sandau, C., Koczanski, K., Michaud, R., De Guise, S., Béland, P., 2000a. Methylsulfone polychlorinated biphenyl and 2, 2bis(chlorophenyl)-1, 1-dichloroethylene metabolites in beluga whale (Delphinapterus leucas) from the St. Lawrence River Estuary and western Hudson Bay, Canada. Environmental Toxicology and Chemistry 19, 1378–1388. Letcher, R.J., Klasson-Wehler, E., Bergman, Å., 2000b. The Handbook of Environmental Chemistry, New Types of Persistent Compounds, vol. 3. Springer-Verlag, Berlin/Heidelberg, Germany. Lewis, D.F.V., Eddershaw, P.J., Dickens, M., Tarbit, M.H., Goldfarb, P.S., 1998. Structural determinants of cytochrome P450 substrate specificity, binding affinity and catalytic rate. Chemical-Biological Interactions 115, 175–199. McKinney, M.A., Arukwe, A., Béland, P., Martineau, D., Dallaire, A., De Guise, S., Lebeuf, M., Letcher, R.J., 2004. Characterization and profiling of hepatic cytochromes p450 and Phase II xenobiotic-metabolizing enzymes beluga whales (Delphinapterus leucas) from the St. Lawrence River Estuary and the Canadian Arctic. Aquatic Toxicology 69, 35–49. McKinney, M.A., Martineau, D., Dallaire, A.D., Béland, P., De Guise, S., LeBeuf, M., Letcher, R.J., 2006. Organohalogen contaminants and metabolites in beluga whale (Delphinapterus leucas) liver from two Canadian populations. Environmental Toxicology and Chemistry 25 (5), 1246–1257. Meerts, I.A.T.M., van Zanden, J.J., Luijks, E.A.C., van Leeuwen-Bol, I., Marsh, G., Jakobsson, E., Bergman, A., Brouwer, Å., 2000. Potent competitive interactions of some brominated flame retardants and related compounds with human transthyretin in vitro. Toxicological Sciences 56, 95–104. Meerts, I.A.T.M., Lilienthal, H., Hoving, S., van den Berg, J.H., Weijers, B.M., Bergman, Å., Koeman, J.H., Brouwer, A., 2004. Developmental exposure to 4-hydroxy-2, 3, 3’, 4’, 5-pentachlorobiphenyl (4-OH-CB107): long-term effects on brain development, behavior, and brain stem auditory evoked potentials in rats. Toxicological Sciences 82, 207–218. Montie, E.W., Hahn, M.E., Reddy, C., Gebbink, W.A., Touhey, K.E., Letcher, R.J., 2009. Organohalogen contaminants and metabolites in cerebrospinal fluid and cerebellum gray matter in short-beaked common dolphins and Atlantic whitesided dolphins from the western North Atlantic. Environmental Pollution. Rayne, S., Ikonomou, M.G., Ross, P.S., Ellis, G., Barrett-Lennard, L.G., 2004. PBDEs, PBBs, and PCNs in three communities of free-ranging killer wales (Orcinus orca) from the northeastern Pacific Ocean. Environmental Science and Technology 38, 4293–4299. Reiss, R., Mackay, N., Habig, C., Griffin, J., 2002. An ecological risk assessment for triclosan in lotic systems following discharge from wastewater treatment plants in the United States. Environmental Toxicology and Chemistry 22, 2483– 2492. Ross, P.S., 2006. Fireproof killer whales: flame retardant chemicals and the conservation imperative in the charismatic icon of British Columbia. Canadian Journal of Fisheries and Aquatic Sciences 63, 224–234. Ross, P.S., Ellis, G.M., Ikonomou, M.G., Barrett-Lennard, L.G., Addison, R.F., 2000. High PCB concentrations in free-ranging Pacific killer whales, Orcinus orca: effects of age, sex and dietary preference. Marine Pollution Bulletin 40, 504– 515. Ross, P.S., Couillard, C.M., Ikonomou, M.G., Johannessen, S.C., Lebeuf, M., Macdonald, R.W., Tomy, G.T., 2009. Large and growing environmental reservoirs of DecaBDE present an emerging health risk for fish and marine mammals. Marine Bulletin Pollution 58 (1), 7–9. Sandala, G.M., Sonne-Hansen, C., Dietz, R., Muir, D.C.G., Valters, K., Bennett, E.R., Born, E.W., Letcher, R.J., 2004. Methyl sulfone and hydroxylated PCB metabolites in adipose and whole blood of polar bear (Ursus maritimus) from East Greenland. Science of the Total Environment 331, 125–141. Sandau, C.D., Meerts, I.A.T.M., Letcher, R.J., McAlees, A.J., Chittim, B., Brouwer, A., Norstrom, R.J., 2000. Environmental Science and Technology 34, 3871–3877. Teuten, E.L., Xu, L., Reddy, C.M., 2005. Two abundant bioaccumulated halogenated compounds are natural products. Science 307, 917–920. Ucán-Marín, F., Arukwe, A., Mortensen, A., Gabrielsen, G.W., Fox, G.A., Letcher, R.J., 2009. Recombinant transthyretin purification and competitive binding with organohalogen compounds in two gull species (Larus argentatus and Larus hyperboreus). Toxicological Sciences 107 (2), 440–450. Ueno, D., Darling, C., Alaee, M., Pacepavicius, G., Teixeira, C., Campbell, L., Letcher, R.J., Bergman, Å., Marsh, G., Muir, D.C.G., 2008. Hydroxylated polybrominated diphenyl ethers (OH-PBDEs) in the abiotic environment: surface water and precipitation from Ontario, Canada. Environmental Science and Technology 42, 657–1664.

Baseline / Marine Pollution Bulletin 58 (2009) 1078–1095 Valters, K., Li, H., Alaee, M., D’Sa, I., Marsh, G., Bergman, Å., Letcher, R.J., 2005. Polybrominated diphenyl ethers (PBDEs) and hydroxylated and methoxylated polybrominated and chlorinated analogues in the plasma of fish from the Detroit River. Environmental Science and Technology 39 (15), 5612–5619. Verreault, L., Dietz, R., Sonne, C., Gebbink, W.A., Shahmiri, S., Letcher, R.J., 2008. Comparative fate of organohalogen contaminants in two top carnivores in

1083

Greenland: captive sledge dogs and free-ranging polar bears. Comparative Biochemistry and Physiology Part C 147, 306–315. Ylitalo, G.M., Matkin, C.O., Buzitis, J., Krahn, M.M., Jones, L.L., Rowles, T., Stein, J.E., 2001. Influence of life-history parameters on organochlorine concentrations in free-ranging killer whales (Orcinus orca) from Prince William Sound, AK. Science of the Total Environment 281, 183–203.

0025-326X/$ - see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2009.05.005

Diuron increases spinal deformity in early-life-stage pink snapper Pagrus auratus Marthe Monique Gagnon *, Christopher Allan Rawson Department of Environmental and Aquatic Sciences, P.O. Box U987, Curtin University of Technology, Bentley Campus, Perth, Western Australia 6845, Australia

While the toxic effect of the anti-foulant herbicide diuron to photosynthetic aquatic biota has been widely studied, its sub lethal effects on different life stages of fish have been under-reported. Its main application since its initial commercial manufacture has been in agriculture for weed control in croplands. Diuron is one of a group of photosystem II inhibitors accounting for about 30% of herbicide use in Australia (Radcliffe, 2002). Its potency, persistence in soil and overall low mammalian toxicity have made it one of the most widely used emergence herbicides. In the aquatic environment the persistence of diuron has raised concern regarding the effect of agricultural runoff into freshwater systems and hence in the receiving marine environments (half-life of diuron > 100 days in seawater (Thomas et al., 2001)). In particular, its potential to cause damage to sensitive coral reefs in the North East of Australia has been the subject of great concern as these waters receive agricultural runoff from high diuron use areas (Haynes et al., 2000; Jones et al., 2003). Its addition as an anti-foulant to paints applied to marine vessel hulls is an additional source of diuron to the marine environment. Such usage increased through the 1980s until it was banned from commercial use in the UK in 2002 after high concentrations were measured in European marinas and ports (Chesworth et al., 2004). In Australia, use of diuron in both agriculture and marine vessels was reviewed in 2001 and some restrictions placed on its application in agriculture (APVMA, 2005). However, there were no restrictions placed on its use as an anti-foulant paint additive and its use is likely to increase as bans on the potent environmental androgen tributyltin (TBT) continue to take effect. Pink snapper (Pagrus auratus) are important commercial and recreational fish in Australia. They are widespread along the coastline but concern over their decline has been recently raised, particularly in Western Australia where catches are likely to be slowed by legislation in the near future. Previous data have suggested a possible effect of diuron on the hatch rate of pink snapper and the rate of spinal deformities in hatched larvae (unpubl. data). This study is a preliminary investigation into the toxicity of diuron to pink snapper in the early life stages. Specifically it aimed to investigate the effect of waterborne diuron exposure on the hatch-rate and spinal development of post-fertilisation pink snapper. Three year old pink snapper maintained within the Curtin University Aquatic Research Laboratory (CARL) were used in this study. These fish have been reared within this facility and are * Corresponding author. Tel.: +61 8 9266 3723; fax: +61 8 9266 2495. E-mail address: [email protected] (M.M. Gagnon).

maintained in a 10,000 l holding tank. Spawning of pink snapper in aquaculture occurs at dusk in late winter to early spring and is triggered by increased water temperature above 18 °C (Primary Industries and Resources South Australia, 2000). Fertilised pink snapper eggs are positively buoyant due to the presence of an oil globule while unviable and unfertilised eggs are negatively buoyant. The pink snapper at CARL spawned for the first time in September 2008 and viable, fertilised eggs were collected by automatic surface skimming of the holding tank and diverted to a large flow-through container with a mesh (100 lm) bag insert. Eggs were harvested from this container no more than 2 h postspawn. Two litres of filtered (0.22 lm) sea water was added to 5 l glass beakers which had been cut down to the 3 l mark to allow a greater surface area to volume ratio. These were placed in a large water bath heated to 24.5 °C to mimic the conditions in the pink snapper holding tank. One millilitre of diuron stock solutions (100, 10, 1 and 0.1 mg/l) made up in 100% ethanol was added to triplicate treatment chambers such that the nominal exposure concentrations were 50, 5, 0.5 and 0.05 lg diuron/l. These concentrations were chosen to cover the range of diuron concentrations estimated and measured in marine waters (Thomas et al., 2002). One millilitre of 100% ethanol was added to triplicate solvent control treatments, such that the ethanol concentration in each chamber (solvent control and diuron treatments) was 0.5 ppt and 1 ml of filtered sea water was added to two other test chambers as a procedural blank control. A gentle air flow was bubbled through each of the test chambers for 3 h prior to the addition of eggs and continued throughout the exposure. An estimated 200–300 buoyant pink snapper eggs were added to each of the test chambers (<0.5 h after collection) and observed to ensure mixing through the test fluid. The water bath was lit by a fluorescent lamp for the duration of the exposure. After 36 h the test chambers were removed from the water bath and the contents (hatched and unhatched pink snapper) were progressively sieved out using a 100 lm sieve. The sieve contents were preserved in 5% buffered formalin in specimen jars for 1 week after which they were transferred to 70% ethanol. The contents of individual specimen jars were transferred to a petri-dish with inscribed gridlines and inspected and counted under a dissection microscope. Unhatched eggs were classified as being more or less developed than stage 15 (Table 1). This was determined by the presence of a distinct tail margin. Hatched larvae were inspected for gross spinal deformities (spinal column or tail). Hatched larvae with straight tails were classified as normal