A Comparative Analysis of Polybrominated Diphenyl Ethers and Polychlorinated Biphenyls in Southern Sea Otters that Died of Infectious Diseases and Noninfectious Causes
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- Kannan, K., Perrotta, E., Thomas, N.J. et al. Arch Environ Contam Toxicol (2007) 53: 293. doi:10.1007/s00244-006-0251-8
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Southern sea otters (Enhydra lutris nereis) from the California coast continue to exhibit a slower population regrowth rate than the population in Alaska. Infectious diseases have been identified as a frequent cause of death. Infectious diseases caused by varied pathogens including bacteria, fungi, and parasites were suggestive of compromised immunological health of mature animals in this population. To test the hypothesis that elevated exposure to immunotoxic contaminants such as polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) contribute to disease susceptibility via immunosuppression, we determined concentrations of PBDEs and PCBs in livers of 80 adult female sea otters that died of infectious diseases, noninfectious causes, or emaciation. Concentrations of PBDEs and PCBs in sea otter livers varied widely (10–26,800 ng/g and 81–210,000 ng/g, lipid weight, respectively). Concentrations of PBDEs in sea otters were some of the highest values reported for marine mammals so far. Although PCB concentrations in sea otters have declined during 1992–2002, the mean concentration was at the threshold at which adverse health effects are elicited. Concentrations of PBDEs and PCBs were significantly correlated, suggesting co-exposure of these contaminants in sea otters. No significant association was found between the concentrations of PBDEs and the health status of sea otters. Concentrations of PCBs were significantly higher in otters in the infectious disease category than in the noninfectious category, suggesting an association between elevated PCB concentrations and infectious diseases in Southern sea otters.
The sea otter, once an abundant species with a range spanning from California in the east across the Pacific Rim to the Kuril Islands north of Japan in the west, was hunted to near-extinction in the 18th and 19th centuries for its fur (Estes 1990; Bacon et al. 1999). It was not until 1911 that the surviving population was legally protected by international law. The result was several remnant sea otter populations. Long-term studies exist for the sea otter subspecies of Attu Island of the Aleutian Archipelago, southeast Alaska, British Columbia, Washington, and California (Estes 1990; Bacon et al. 1999). Best estimates placed the California sea otter populations at 50 individuals in 1914. Prior to their exploitation, it was estimated that 16,000–18,000 sea otters inhabited the California coastline (DeMaster et al.1996; CDFG 2001). The California subspecies has exhibited a sluggish ∼5% per year population increase, whereas other sea otter populations have exhibited growth rates near species expansion limits, at 17–20% per year (Estes 1990). Additionally, the California sea otter has had two periods of population decline: one occurred during 1976–1984 and was attributed to mortalities from entanglement in near-shore fishing nets; the second took place from 1995 to 1999 and the cause remains unexplained (DeMaster et al.1996; CDFG 2001).
The California sea otter population has exhibited atypical mortality patterns. Approximately 40% of beached otters have died of infectious diseases, and mortality occurs throughout the year, with a seasonal peak in the spring through summer months (Thomas and Cole 1996; Estes et al. 2003). Additionally, the highest rate of mortality by age class was found among prime-aged adult otters (3–10 years), rather than among aged adults or pups (Otter Project 2005). It is believed that a high rate of infectious disease-related mortality, rather than low fecundity, was responsible for the stagnating population growth. Diseases caused by parasites, bacteria, or fungi, and diseases without a specified etiology, have been identified as the primary causes of death (Thomas and Cole 1996; Kreuder et al.2003). These findings suggest that the immunological health of mature animals in this population is compromised. In attempts to better understand the factors that contribute to high mortality and reduced immunocompetence in Southern sea otters, studies have examined residue levels of immunosuppressive chemical contaminants (Kannan et al. 1998; Nakata et al. 1998; Kannan et al. 2004a; Kannan et al. 2006a, 2006b), immunological parameters (Schwartz et al. 2005), and serology and clinical pathology (Hanni et al. 2003). However, because the contaminants exist in complex mixtures, and because interactions can potentially occur among contaminants, nutritional status, and other environmental factors, establishment of a link between contaminant exposure and disease mortality in marine mammals such as sea otters is a challenging task. Nevertheless, population-based epidemiologic studies to determine the risk of infection after exposure to potential contributing agents, such as polychlorinated biphenyls (PCBs) may help to discentangle this problem. (Hall et al.2006). Such case–control studies have been widely used in human and veterinary medicine (Shapiro 1989) but have received little attention in wildlife research, because of the logistical difficulties in obtaining disease and exposure data, and also because of the lack of necessary tools to investigate and understand the mechanisms of host disease resistance. A few earlier studies have compared residue levels of PCBs and butyltins between healthy and diseased sea otters (Kannan et al.1998; Nakata et al.1998) and other marine mammal populations (Kannan et al. 1993; Hall et al.2006), to examine whether high exposure levels were associated with the disease or mortality. Although the case–control studies still do not provide causal linkages, they can provide baseline information that allows us to develop a framework for future systematic investigations (Hall et al.2006). In this study, concentrations of polybrominated diphenyl ethers (PBDEs) and PCBs were measured in livers of sea otters, to compare exposure levels between individuals that died of infectious diseases and those that died of noninfectious causes.
Since the early 1990s, the putative link between exposure to PCBs and immunosuppression in marine mammals has received considerable attention (Lahvis et al.1995; De Swart et al. 1996; Jepson et al. 2005; Beineke et al. 2005). In addition to PCBs, several other halogenated contaminants such as DDTs, PBDEs, and dioxins can contribute to immunosuppression in marine mammals (Lahvis et al.1995; De Swart et al. 1996; Jepson et al. 2005; Beineke et al. 2005). Thymic atrophy and splenic depletion were significantly correlated with increased PCB and PBDE levels in harbor porpoises from the North Sea and the Baltic Sea (Beineke et al. 2005). PBDEs are flame retardants and are used in many household consumer products ranging from television sets and computers to sofas, blankets, and building materials (Hites 2004). Studies have shown that the concentrations of PBDEs in marine mammals have increased significantly over the last decade (She et al. 2002; Ikonomou et al. 2002; Kajiwara et al. 2004; Johnson-Restrepo et al. 2005a).
Although there are many known factors such as boat strike, entrapmentin fishing nets, human disturbances, and malnutrition that can contribute to the slow population growth rate of California sea otters, it remains unknown whether environmental contaminants contribute to the slowness of the recovery (Estes et al. 2003). In this study, we examined the association between exposure to PCBs and PBDEs and the disease status of the California sea otter. This was accomplished by determining and comparing hepatic concentrations of PCBs and PBDEs in sea otters that died of infectious diseases and those that died of noninfectious causes. The availability of large numbers of archived tissues of sea otters, which had been systematically examined for evidence of disease during postmortem investigations and necropsy, provided an opportunity to evaluate the association between PCB/PBDE concentrations and cause of mortality. Sea otters are unique and valuable sentinels that allow us to evaluate the health of coastal aquatic environments. The narrow home range, top-level predator status, and voracious appetite make the sea otter a suitable surrogate for human exposure to aquatic sources of environmental contaminants (Kajiwara et al. 2001; Kannan et al. 2004a).
Materials and Methods
Liver tissue, rather than adipose tissue, was chosen for analysis for the following reasons. First, this organ plays a significant role in toxicant processing. Second, the sea otter lacks blubber. Although PCBs and PBDEs are lipophilic by nature, high concentrations in fat tissue may not indicate direct toxicity associated with exposure. It is thought that the toxic effects are pronounced when contaminants are mobilized from the lipid pool and distributed to vital organs such as liver, heart, and kidney.
Analysis of PCBs and PBDEs
Liver samples were collected from the carcasses at the time of necropsy, wrapped in aluminum foil, enclosed in sterile sampling bags (TWIRL’EM, Fisher Scientific Inc., Hampton, NH), and stored at –20°C until analysis. Liver samples (∼10 g) were homogenized with anhydrous sodium sulfate (120 g), spiked with the internal standards PCB-30 and PCB-204 (25 ng), and extracted with mixed solvents dichloromethane (300 mL) and hexane (100 mL) using a Soxhlet apparatus for 20 h (Kannan et al. 2004a). The solvent was concentrated (11 mL), and an aliquot (1 mL) was taken for gravimetric determination of fat content. A second aliquot (5 mL) was spiked with 13C-labeled PCBs (10 ng each congener; MBP-MXF and MBP-CG; Wellington Laboratories, Guelph, Ontario, Canada) and 13C-labeled PBDEs (10 ng each congener; EO5100; Cambridge Isotope Laboratories, Andover, MA) and subjected to column chromatography for removal of lipids. Columns (10 mm i.d.) were prepared by layering silica gel (2 g, 100–200 mesh), 40% acidic silica gel (1 g), silica gel (1 g), 40% acidic silica gel (1 g), and silica gel (1 g). Sample extracts were eluted with 15% dichloromethane in hexane (150 mL). Samples were further subjected to lipid removal by treatment with concentrated sulfuric acid, if and when needed. Sample extracts (2 μL) were injected into a Hewlett-Packard 6890 gas chromatograph interfaced with a Hewlett-Packard 5973 mass spectrometer (GC-MS). Injections were made in the splitless mode, and samples were separated on a 30 m DB-5 (5% diphenyl/dimethylpolysiloxane) analytical capillary column with a 250 μm i.d. and a 0.25-μm film thickness. The oven temperature program was set to 100°C for 1 min, 10°C/min to 160°C, hold 3 min, and 2.5°C/min to 260°C, hold for 10 min. The inlet and interface temperatures were set to 270°C and 300°C, respectively. The MS was operated in an electron impact mode (70 eV) and selected ion monitoring mode (SIM). PBDE congeners were identified and quantified by SIM at m/z 406, 408; 486, 484; 564, 566; and 642, 644 for tri-; tetra-; penta-; and hexa-BDEs, respectively. Quantification was based on external calibration standard. PCB congeners were monitored at the two most intense ions of the molecular ion cluster. An equivalent mixture of Kanechlor (KC300, 400, 500, and 600), with known PCB composition, was used in the identification of PCB congeners. Quantification of PCB congeners was based on external calibration standards containing known concentration of di- through deca-CB congeners. Concentrations from individually resolved peaks of PCB isomers were summed, to obtain total PCB concentrations.
Quality Assurance and Quality Control
In order to assure the quality of our measurements, we have taken several steps. Extracts used for PBDE analysis were also analyzed for PCBs. Therefore, samples were spiked with internal standards PCB-30, PCB-204, and 13C-labeled PCBs and 13C-PBDEs, one from each homologue group. The recoveries of PCB-30 and PCB-204, spiked into samples prior to extraction, were 86 ± 15% and 91 ± 19%, respectively. Recovery of the 13C-labeled PCBs and 13C-PBDEs, spiked prior to lipid removal, ranged from 86% to 99%. The measured concentrations were not corrected for the recoveries of internal or surrogate standards. The limit of quantification (LOQ) was determined by several measures. The LOQ was considered to be two times greater than the signal determined in blank samples. If no signal was observed in the blank, the lowest point on the calibration curve that was within 30% of the back-calculated 1/x weighted calibration curve was used. Dilution factors, sample weight, and final extract volume were taken into account for calculating LOQ. A LOQ of 1 ng/g, wet weight, for total PBDEs and total PCBs was determined. Procedural blanks were analyzed simultaneously with every batch of five samples as a check for interferences or contamination arising from solvents and glassware. Total PBDEs represent the sum of BDE congeners 47, 99, 100, 153, and 154. Total PCBs represent the sum of all of the di- through decachlorobiphenyl congeners. Concentrations are reported on a lipid weight (wt) basis, unless reported otherwise.
All statistical analyses were performed using Statgraphics version 5 (Manugistics, Inc., Rockville, MD). The distribution of PBDE and PCB concentrations was skewed. Therefore, nonparametric tests were used when comparisons were made between mortality groups. The effects of location and biological parameters (e.g., nutritive status) on the concentrations of PBDEs and PCBs were tested using one-way ANOVA and the Kruskall-Wallis test or Fisher’s least significant difference test to determine whether median concentrations varied among sea otters by location or nutritive condition or through time. The level of significance used for all statistical tests was α ≤ 0.05.
Results and Discussion
Concentrations (ng/g) of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in livers of adult female sea otters from the California coast (1992–2002) stratified by mortality categories
PBDEs (lipid wt)
PBDEs (wet wt)
PCBs (lipid wt)
PCBs (wet wt)
Noninfectious causes (n = 27)
Mean ± SD
4.7 ± 2.7
2120 ± 2630
88.5 ± 107
25,600 ± 48200
970 ± 1520
Infectious disease (n = 26)
Mean ± SD
4.0 ± 1.6
2980 ± 5500
113 ± 191
15500 ± 32200
558 ± 1040
Emaciation (n = 27)
Mean ± SD
3.4 ± 1.5
1430 ± 1860
48.4 ± 79.2
10300 ± 23600
340 ± 940
Overall (n = 80)
Mean ± SD
4.0 ± 2.1
2170 ± 3710
82.8 ± 136
17100 ± 36700
624 ± 1220
High concentrations of PBDEs are known to occur in both inland and marine coastal regions of the United States. A study of the San Francisco Estuary, at the northern end of the California sea otters’ range, showed PBDE concentrations in water ranging from 3 to 513 pg/L; concentrations in sediment ranging from below detection limit to 212 ng/g, dry wt; and concentrations in clams (Corbicula fluminea) at 106 ng/g, dry wt (Oros et al. 2005). Water and sediment contained high proportions of octa- and deca- BDEs, whereas clams contained BDE-47 as the dominant congener. We did not analyze octa- or deca-BDE in sea otter tissues. Nevertheless, most studies have reported BDE-47, BDE-99, and BDE-100 as the most abundant congeners in biological tissues (Stapleton et al.2006). Sewage treatment plant effluents were suggested as a source of PBDEs for the San Francisco Estuary; an estimated 22.7 kg of PBDEs was discharged into this estuary every year (Oros et al. 2005). Mean concentration of PBDEs in edible portion of fish from San Francisco Bay in 2000 was 302 ng/g, lipid wt (Brown et al. 2006), approximately sevenfold lower than the level that we found in the sea otters.
Concentrations of PCBs in livers of sea otters ranged from 81 to 210,000 ng/g, lipid wt (mean ± SD: 17,100 ± 36,700; median: 3510) (Table 1). The highest concentration was found in an otter that died of undetermined causes in 1995. The next highest concentration, 143,000 ng/g, lipid wt, was found in an otter beached in 1994, and the COD of this otter was found to be an infectious neurological disease. An earlier study reported PCB concentrations as high as 266,000 ng/g, lipid wt, in livers of adult sea otters (Kannan et al.2004a), similar to that found in the present study.
The mean concentration of PCBs in livers of our sea otters was 2.5-fold lower than the concentration reported for California sea otters collected during 1992–1996 (Nakata et al. 1998). This suggests a reduction in PCB concentrations in sea otters in recent years. Similarly, concentrations of PCBs in California sea lions showed a trend of decline in the 1990s (Kannan et al.2004b). Despite the declining trend, concentrations of PCBs in sea otters and California sea lions were some of the highest concentrations (410,000 ng/g, lipid wt) found for marine mammals in recent years (Kajiwara et al. 2001; Kannan et al. 2004b). The mean concentration of PCBs in sea otters (17,100 ng/g, lipid wt) was at the threshold concentration (17,000 ng/g, lipid wt) that has been reported to elicit adverse physiological effects in marine mammals (Kannan et al.2000). A threshold PCB concentration of 45,000 ng/g, lipid wt, was reported for an increased risk of infection in harbor porpoises (Hall et al. 2006). A PCB concentration of 10,000–22,000 ng/g, lipid wt, was associated with suspected immunotoxic effects in harbor porpoises (Beineke et al. 2005). More than 25% of the sea otters analyzed in the present study contained PCB concentrations greater than 10,000 ng/g, lipid wt. These results suggest that individual sea otters may experience adverse effects due to exposure to high levels of PCBs.
Comparison of Concentrations by Location, Nutritive Condition, and Through Time
Comparison of Concentrations in the Infectious and Noninfectious Mortality Categories
Nine of the 20 individuals that contained PCB concentrations above 10,000 ng/g, lipid wt, were in either the subclass “undetermined causes” or “miscellaneous problems,” which were categorized as noninfectious COD in the present study (Figure 10). When the PCB concentrations for “undetermined” subclasses were removed (because of the lack of specific etiology) from the noninfectious COD category, concentrations of PCBs in the infectious COD otters were significantly greater than those in noninfectious-COD and emaciated otters (Mann-Whitney W test, p < 0.05) (Figure 9). Nevertheless, concentrations of PBDE did not differ significantly even after the undetermined and miscellaneous subclasses were removed from the noninfectious COD category.
In summary, these results suggest that elevated PCB concentrations in Southern sea otters are associated with infectious diseases. The sea otters analyzed in this study were also examined for the presence of a wide variety of contaminants including trace metals, butyltins, polycyclic aromatic hydrocarbons, pesticides, perfluorochemicals, and synthetic musks (Perrotta 2005; Kannan et al. 2005c). A recent study has reported the association between perfluorochemicals and infectious diseases (Kannan et al. 2006a) and between certain trace elements and disease (Kannan et al.2006b) in sea otters. Based on these results, PCBs and perfluorochemicals were found to be strongly associated with the disease status. Co-occurrence and exposure to multiple contaminants in sea otters may also contribute to immunosuppression. Further research is needed to establish the association between immunological parameters and contaminant levels as well as controlled-exposure studies to establish cause–effect linkages.
This study was funded by a grant from the Monterey Bay Sanctuary Foundation.