Contaminant exposure and effects in Baltic ringed and grey seals as assessed by biomarkers

Marine Environmental Research 55 (2003) 73–99 www.elsevier.com/locate/marenvrev

Contaminant exposure and effects in Baltic ringed and grey seals as assessed by biomarkers Madeleine Nymana,*, Magnus Bergknutb, Marie Louise Fanta, Hannu Raunioc, Marika Jestoid, Charlotta Bengsa, Albertinka Murke, Jaana Koistinenf, Christina Ba¨ckmand, Olavi Pelkoneng, Mats Tysklindb, Timo Hirvid, Eero Hellea a

Finnish Game and Fisheries Research Institute, Box 6, 00721 Helsinki, Finland b Department of Chemistry, Umea˚ University, SE-90187 Umea˚, Sweden c Department of Pharmacology and Toxicology, University of Kuopio, Box 1627, 70211 Kuopio, Finland d National Veterinary and Food Research Institute, Box 368, 00231 Helsinki, Finland e Wageningen University, Division of Toxicology, Box 8000, 6400 EA Wageningen, The Netherlands f Division of Environmental Health, National Public Health Institute, Box 95, 70701 Kuopio, Finland g Department of Pharmacology and Toxicology, University of Oulu, Box 5000, 90401 Oulu, Finland Received 2 May 2002; received in revised form 27 June 2002; accepted 7 July 2002

Abstract The Baltic Sea ecosystem has suffered from a heavy pollutant load for more than three decades. Persistent organic pollutants (POPs) and heavy metals have been of most concern due to their persistence and toxic properties. Ringed seals (Phoca hispida baltica) and grey seals (Halichoerus grypus) living in the Baltic Sea have been suffering from pathological impairments, including reproductive disturbances, which have resulted in a depressed reproductive capacity. We investigated several biochemical parameters as potential biomarkers for exposure to and effects of the contaminant load in the Baltic seals. Seals from less polluted areas were used as reference material in terms of the pollution load. In both Baltic seal populations, the levels of some biochemical parameters diverged from those in the reference seals, and some of these showed a clear correlation with the individual contaminant load. Of the potential bioindicators, we propose cytochrome P4501A activity and vitamin E levels, in blubber or plasma, as exposure biomarkers for polychlorinated biphenyls (PCBs) and dichlorodiphenyltrichloroethane (DDT) in both species. The arylhydrocarbon receptor-mediated chemical-activated luciferase gene expression (CALUX) response reflects the whole PCB and DDT burden in ringed seals. Retinyl palmitate (vitamin A) levels showed a negative * Corresponding author. Tel.: +358-205-751274; fax: +358-205-751201. E-mail address: madeleine.nyman@rktl.fi (M. Nyman). 0141-1136/02/$ - see front matter # 2002 Elsevier Science Ltd. All rights reserved. PII: S0141-1136(02)00218-0

74

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

correlation with the individual POP load, and is proposed as potential effect biomarkers for the depletion of the vitamin A stores. As the nutritional levels of both vitamin A and E have an impact on the vitamin levels in the seals, more information on the dietary vitamin levels is needed before any conclusions can be drawn. As the relationship between biochemical parameters and contaminants varied between the two species, species-specific characteristics has to be considered when monitoring the health status and possible toxic effects of the contaminant load in ringed and grey seals. # 2002 Elsevier Science Ltd. All rights reserved. Keywords: Ringed seal; Grey seal; Baltic Sea; Biomarker; Vitamin A; Vitamin E; PCB; DDT; Metals

1. Introduction It was first recognised about three decades ago that the Baltic Sea had become heavily polluted (Jensen, 1972; Jensen, Johnels, Olsson, & Otterlind, 1969). Most of the biota, including the Baltic ringed seals (Phoca hispida baltica) and grey seals (Halichoerus grypus) were suffering from heavy contaminant loads. Exceptionally high mean concentrations of the sum of polychlorinated biphenyls (SPCB), of dichlorodiphenyltrichloroethane (DDT) and its metabolites and of mercury were reported in Baltic seals in the late 1960s and 1970s (Helle, 1981; Helle, Olsson, & Jenssen, 1976a, 1976b; Herva & Ha¨sa¨nen, 1972; Jensen et al., 1969; Kari & Kauranen, 1978). A complex of pathological disorders has been described in the Baltic seals, based on findings from histological samples obtained during the 1970s and 80s (Bergman & Olsson, 1986; Bergman, Olsson, & Reiland, 1992; Helle et al., 1976a, 1976b; Olsson, Karlsson, & Ahnland, 1994). Most of the pathological changes associated with the seals’ heavy pollution load have occurred less frequently as the contaminant burden has decreased (Bergman, 1999; Eero Helle unpublished data). The mechanisms linking the impairments and the pollutant load carried by the Baltic seals are not clear, although new information on toxic effects in experimental animals may elucidate this. In recent years, an number of persistent organic pollutants (POPs), in addition to PCBs and DDT, have been detected in tissue samples of the Baltic seals. However, PCBs and DDT and their metabolites are still thought to be toxicologically the most relevant, posing the biggest threat to the Baltic seals (Olsson, Andersson, Bergman, Blomkvist, Frank, & Rappe, 1992), but additive and synergistic effects of the other compounds are possible. Studies on metals in marine mammals have focused on mercury, cadmium and lead because of their known toxicity to humans and other mammals (AMAP, 1998; O’Shea, 1999), and because they are considered a potential threat to the Baltic Sea ecosystem (HELCOM, 1996). The level and type of contamination in an organism is dependent on the availability, persistence, accumulation and metabolism of the compound concerned. Many of the POPs are lipophilic and thus accumulate in lipid rich tissues. In aquatic mammals, these compounds tend to accumulate with age and be transferred to the offspring via milk (reviewed by Boon et al., 1992). Marine mammals at the top of the aquatic food chain are especially vulnerable to contaminant exposure as they have a

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

75

lower detoxification capacity than terrestrial animals and as their large lipid reserves serve as depositories for lipophilic compounds (Boon et al., 1992; Tanabe, Watanabe, Kan, & Tatsukawa, 1988; Tanabe, Iwata, & Tatsukawa, 1994). Although very little is known on the toxic effects of PCB residues in free ranging mammals (O’Shea, 1999), reproductive impairment has been suggested to occur when PCB blubber concentrations exceed 70 mg/kg lipid weight in ringed seals (Helle et al., 1976a, 1976b). As these levels correspond typically to the present contaminant load in the Baltic ringed seals but are much higher than in the Baltic grey seals, both physiological and pathological effects of the contaminant load could be expected especially in the ringed seals. There is a vast amount of information on contaminant residue levels in different marine mammals from different parts of the world. However, the fate and effects of all known and newly introduced chemicals is impossible to analyse due to time and funding restrictions. Alternatively, biological markers (biomarkers) and bioassays, which reflect the exposure levels, health status or specific toxic effects in wildlife populations, can be used to assess the combined impact of the pollution load. In this study, we tested the applicability of relatively established biomarkers [cytochrome P450 (CYP) and vitamin A], and of other more speculative biomarkers (vitamin E and clinical screening parameters) on samples obtained from Baltic ringed and grey seals. Levels of SPCB, SDDT, mercury, selenium, cadmium and lead, that are reported in detail elsewhere, were used in this study (Fant, Nyman, Halle, & Rudba¨ck, 2001; Nyman, Koistinen, Fant, Vartiainen, & Helle, 2002). A reporter gene assay (CALUX) was used to measure the total dioxin-like toxic potency in the seal samples. This exposure bioassay has been used for assaying sediments, pore water, blood and various tissues (Murk, Legler, Denison, Giesy, Van de Guchte, & Brouwer, 1996; Murk et al., 1997; Murk, Leonards, Van Hattum, Luit, Van der Weiden, & Smit, 1998). Samples from ringed seals from the European Arctic (Svalbard) and from grey seals from the Western Atlantic Ocean (Sable Island, Canada) were used as reference material in terms of the POP load.

2. Materials and methods 2.1. Sample collection and preparation Baltic ringed seals (29) and grey seals (30) were sampled on the ice in the Bothnian Bay in April and May 1996–1998. Reference samples were collected from 29 ringed seals (Svalbard, the Arctic) and 20 grey seals (Sable Island, Canada) in May and June 1996–1998. Seals were sampled according to a special permission granted to the Finnish Game and Fisheries Research Institute by the Ministry of Forestry and Agriculture in Finland. The samples were obtained from seals on Svalbard during the local hunting season according the local hunting law, and from seals on Sable Island with a the permission of the Department of Fisheries and Oceans, Canada. Liver, blubber and blood samples were obtained from 10 to 30 seals per population during their moulting season at approximately the same phase of their annual

76

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

reproductive cycle. Serum and plasma were separated from the blood in the field, and were stored together the with tissue samples for the vitamin analyses at
70  C. Sampling, preparation and storage of the liver samples for the chemical and CYP analyses have been described in detail elsewhere (Fant et al., 2001; Mattson, Raunio, Pelkonen, & Helle, 1998; Nyman et al., 2002). Age was estimated as described previously (Mattson et al., 1998). The condition of the seals was estimated by comparing the thickness of the blubber, measured at the sternum, to the body length. The condition index was modified from the methods developed by Read (1990) and Beck and Smith (1995). No geographical difference was observed in condition between the respective seal populations. Although the Canadian grey seals showed a higher mean age than the Baltic population, the age range was similar. As some of the biochemical parameters examined showed gender specific characteristics, sex was included as a covariant in the statistical analyses. 2.2. Reagents and standards TRIZMA1 Hydrochloride, TRIZMA1 Base, all-trans-retinol and all-trans-retinol palmitate were purchased from Sigma Chemical Co. (St. Louis, MO, USA). HPLC-grade methanol and ethyl acetate, d l (+) ascorbic acid and di-isopropyl ether, p.a. were obtained from J. T. Baker (Deventer, Holland). dl-a-tocopherol was purchased from E. Merck (Darmstadt, Germany), retinyl acetate from ICN Biomedicals Inc. (Aurora, Ohio, USA) and 5,7-dimethyltocol from Matreya Inc. (Pleasant Cap, PA, USA). 25% Glutaraldehyde was purchased from J. T. Baker (Holland). 2.3. Chemical analyses Extraction and analysis methods for DDT, PCB and metal determination have been described in detail by Nyman et al. (2002) and Fant et al. (2001). The extraction and analysis methods for the POPs in short: liver samples were freeze-dried and extracted in a Soxhlet apparatus with toluene. The extract cleanup included lipid removal in a multi-layer sulphuric acid-silica column, a further cleanup in a column of activated carbon mixed with celite, and an additional cleanup on alumnae. Coplanar PCBs were isolated from the other congeners on an activated carbon column. All analyses were performed by high-resolution gas chromatography–high resolution mass spectrometry (HRGC–HRMS) using selected ion recording (SIR) with 10,000 resolution. Of the 34 PCB congeners analysed, 24 (IUPAC nr 52, 77, 99, 101, 105, 110, 114, 118, 126, 128, 138, 141, 153, 156, 157, 167, 169, 170, 180, 187, 189, 194, 206 and 209) congeners were included in this study. The coplanar PCB congeners 77, 126 and 169 are presented as a sum of coplanar PCBs (Scoplanar PCB). The concentrations of the PCBs were converted to dioxin equivalents (PCBTEQ) and summed (SPCB-TEQ) using the toxic equivalent factors proposed for humans, fish and wildlife (Van den Berg et al. 1998). As no o,p0 -isomers of the DDT compounds were found, only the p,p0 -isomers of DDT, DDE and DDD are included, and presented as a sum (SDDT).

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

77

The metals were freeze-dried and mercury levels were determined by cold atomic vapour absorption, while cadmium and lead levels were determined using an atomic absorption spectrometer equipped with a graphite furnace (Fant et al., 2001). 2.4. In vitro reporter gene assay (CALUX) Chemical-activated luciferase gene expression (CALUX) is an in vitro assay for assessing the toxic potency of individual or of a mixture of compounds via the cellular arylhydrocarbon receptor (AhR), including PCBs, PCDD/Fs, PBDEs and PBBs (Aarts et al., 1995). It is based on a rat hepatoma cell line in which a dioxin responsive element and the luciferase reporter gene have been incorporated. Liver samples were extracted and assayed for total TEQs according to the methods of Murk et al. (1998). In short, liver samples were mixed with anhydrous sodium sulphate and extracted in 190 ml pentane/dichloromethane (1:1, v/v) over 6 h using the Soxhlet apparatus. The extract was cleaned in a multi-layer glass column and evaporated to almost 1 ml under a gentle filtered airflow. The cells used in the assay were cultured in 96-well culture plates and incubated with the extract for 24 h, after which the cells were frozen. The toxic potency was quantified by measuring the luciferase activity directly from the plates. The luciferase response was converted to a TEQ value using a TCDD standard curve developed for rats, and reflected the AhR-related toxic potency of the contaminant mixture. 2.5. Cytochrome P450 analyses The methods used for the CYP analyses were all end point assays, and have been reported elsewhere (Mattson et al., 1998; Nyman, Raunio, Taavitsainen, & Pelkonen, 2001). Of the six putative members of Cytochrome P450 enzymes analysed, we included two CYP forms (CYP1A and CYP3A) in this study,as only these showed a geographical difference between the Baltic and the reference seal populations. The means and ranges are presented in Table 2. 2.6. Vitamin analyses Liver, blubber and plasma samples were used for the vitamin analyses. A 1.0 g liver or blubber was homogenised in 3.0 ml 50 mM Tris–HCl-buffer (pH 7.5) with a rod homogeniser (Polytron PT 1200; Kinematiga Ag., Switzerland). Plasma and the homogenate were stored at
70  C if not analysed at once. The extraction method used was modified from Murk et al. (1996). During the whole extraction procedure, samples were protected from the light as much as possible. In short, 50 ml homogenate or 100 ml plasma was mixed with 50 ml methanol. The methanol fraction contained internal standards (retinyl acetate 4 mg/ml and 5,7-dimethyltocol 10 mg/ ml) and an antioxidant (0.1% ascorbic acid). One hundred microlitres of di-isopropyl ether (DIPE) was added and the samples were vortexed briefly, after which they were left at
20  C over night. After mixing and centrifugation (10 min at 5000 rpm) the organic layer was collected and washed with 100 ml DIPE. It was then filtered

78

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

and evaporated under a nitrogen flow until dry. The residue was dissolved in 100 ml methanol–ethyl acetate (3:1; v/v), containing 0.1% ascorbic acid. Both vitamins were assayed simultaneously by high-performance liquid chromatography (HPLC), using a Merck LiChroCART 125–4 LiChrospher 100 RP-18 (5 mm) reversed-phase column with a Merck LiChrospher 100 RP-18 (5 mm) guard column. The samples (20 ml) were eluted with 99% methanol at a flow rate of 1.5 ml/min. These isocratic conditions were developed experimentally by modifying the methods used in previous reports (Bieri, Tolliver, & Catignani, 1979; Catignani & Bieri, 1983; DeRuyter & De Leenheer, 1978; Got, Gousson, & Delacoux, 1995; Hendriks, Verhoofstad, Brouwer, de Leeuw, & Knook, 1985). Vitamin A compounds were detected using a Waters 486 tuneable absorbance UV-detector (detection wavelength 326 nm) and vitamin E compounds using a Waters 474 scanning fluorescence detector (excitation wavelength 295 nm; emission wavelength 325 nm). In this study, two compounds with vitamin A activity (retinol and retinyl palmitate) and one compound with vitamin E activity (a-tocopherol) were detected. Retinyl acetate and 5,7-dimethyltocol were used as internal standards for the vitamin A and E analyses, respectively. The absorbance of the vitamin standard solutions was measured with UV-spectrophotometer. The response of all the vitamins measured was linear to the internal standards used, therefore the seal samples could be quantified reliably. The extinction coefficients used for the vitamin A compounds, measured at 325 nm, were 1835 for retinol, 975 for retinyl palmitate and 1550 for retinyl acetate. For a-tocopherol and 5,7-dimethyltocol (measured at 292 nm), the extinction coefficients were 75.8 and 83.0, respectively. 2.7. Haematology and clinical chemistry For the haematological analyses, blood cells from 20 ml EDTA-blood were fixed in an 8% glutaraldehyde solution in the field and stored at room temperature until analysis. Cell subsets were analysed using an electronic counter (Sysmex F-800) that had been adjusted for seal blood cells. Differential leukocyte counts were microscopically evaluated from May–Grywald–Giemsa stained Blood smears (Dacie & Lewis, 1984). Haemoglobin was determined from EDTA-blood using spectrophotometry (Nuutinen, 1972). Haematocrit was obtained by centrifuging 10 ml EDTA-Blood in the field using a specific haematocrit centrifuge (ALC centrifugette 4206). The immunoglobulin (IgG) levels were determined from serum by immunoblotting using polyclonal antibodies raised against seal IgG. The protein bands were quantified using a standard curve of purified seal IgG concentrations. Clinical chemistry parameters were analysed from frozen serum using an automatic analyser (Olympus AU 5000). 2.8. Statistics For the statistical analyses, the data were divided into four groups according to species and geographical location. The material was LN-transformed to better fit the parametric tests. Geographical and species differences were tested by analysis of

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

79

variance (ANOVA) and Pearson’s correlation analysis. The level of significance was set at P < 0.05. Results are presented as mean values  SD and ranges. The whole data set was further evaluated by principal component analysis (PCA; Wold et al., 1984; Wold, Esbenson, & Geladi, 1987) using the SIMCA-P 8.0 software package (Umetrics AB, Umea˚, Sweden). In a PCA, dominant patterns (systematic variations) in a data table can be extracted and described by a few parameters that are simple to interpret. Theoretically PCA corresponds to a mathematical decomposition of the data matrix X into two parts, a score matrix T and a loading matrix P. The score plot shows the relation between objects (samples) and the loading plot shows the relationship between variables (data). Samples (here: seal samples) with similar data values will have points close to each other in the score plot whereas samples with dissimilar values will be located further apart. Sample groupings are thus easily identified. The loading plot shows the relationship between variables, and variables adjacent to each other are likely to be correlated. In this case, the variables consist of data describing organic pollutants or responses in different biomarkers. When superimposed over a respective score plot, the loading plot shows variables responsible for the class separation. PCA has been used previously in PCB congener composition evaluation and pattern recognition (Bauer, Cramer, Stanley, Fredette, & Giglinto, 1992; Boon et al., 1997; Macdonald, Metcalfe, Balch, & Metcalfe, 1993; Stalling, Norstrom, Smith, & Simon, 1985; Storr-Hansen, Spliid, & Boon, 1995).

3. Results 3.1. Chemical data Although detailed descriptions of the contaminant burden in all seal populations are presented elsewhere (Fant et al., 2001; Nyman et al., 2002), a summary of the geographical and species variations is presented in Table 1 as background information for the evaluation process of the potential biomarkers. The levels of SDDT and SPCB were clearly higher in both Baltic seal species compared to seals from the reference areas, while cadmium showed the opposite geographical trend. In ringed seals, mercury and selenium were 8–60 times higher in the Baltic seals than in the reference seals. As for the species comparison, POP levels were higher in the Baltic ringed seals, but lower in Svalbard ringed seals compared with the respective grey seal populations. Ten to 100-fold higher mercury and selenium levels were observed in the grey seals from Sable Island than in the ringed seals from Svalbard. A gender difference was only observed in the grey seals. Females showed higher cadmium levels in both areas, while Sable Island males had higher POP burden than Sable females. SDDT, SPCB, sum coplanar PCBs and SPCB-TEQ all showed a strong positive correlation with each other in both seal species. A significant age-dependent accumulation of SPCB and SDDT was observed only in a few seal populations, although the trend was increasing for most seal groups. Mercury, selenium and cadmium showed weak indications of an age-dependent accumulation. The POPs

80

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

Table 1 SPCB, SDDT, heavy metal and selenium concentrations in ringed and grey seals from different geographic areas. Results are presented as means (range) SPCBa (mg/g l.w.)

SDDTa (mg/g l.w.)

Mercurya,b (mg/g f.w.)

Seleniuma,b (mg/g f.w.)

Cadmiuma,b (mg/g f.w.)

Leada,b (mg/g f.w.)

Ringed seal Baltic Sea Svalbard

66 (17–152) 1.1 (0.3–8.8)

38 (10–101) 0.4 (0.2–1.7)

53 (6.4–124) 1.0 (0.4–2.6)

15 (2.6–35) 2.0 (1.0–3.2)

2.8 (0.8–6.7) 14 (1.4–51)

(n.d.–0.2) (n.d.–0.1)

Grey seal Baltic Sea Sable Island

28 (10–74) 8.2 (0.3–58)

7.6 (3.3–14) 1.8 (0.3–4.5)

78 (15–348) 109 (27–278)

20 (2.7–79) 28 (9.3–83)

3.4 (1.4–7.6) 10 (3.0–20)

(n.d.–0.1) (n.d.–0.1)

l.w., lipid weight; f.w., fresh weight; n.d., no detection. Cadmium is reported as kidney levels, while all other contaminants are reported as liver levels. a Detailed results are presented in Nyman et al. (2002). b Detailed results in Fant et al. (2001).

were negatively correlated with condition in most males, while no such trend was observed for any of the females. 3.2. CALUX The CALUX-TEQ-values were clearly higher in the Baltic ringed seals compared to the reference ringed seal population, but only slightly higher than the Baltic grey seal values (Table 2). No gender or species difference was seen for CALUX-TEQ values within or between any of the seal populations. Individual CALUX-TEQ values showed a strong positive correlation with SPCB-TEQ in ringed seals (r> 0.80, P < 0.001, Fig. 1A). In addition, a strong positive correlation was observed between CALUX-TEQ and SPCB, SDDT and sum coplanar PCBs in ringed seals (r> 0.80, P < 0.001). There were too few grey seal samples in the CALUX assay for any statistical comparisons. 3.3. Cytochrome P450 enzymes Detailed data on cytochrome P450 enzyme expression in grey and ringed seals have been presented elsewhere (Nyman, Raunio, & Pelkonen, 2000; Nyman et al., 2001). Of the different cytochrome P450 (CYP) forms characterised, putative members of CYP1A, CYP2D and CYP2E showed elevated enzyme activities in the Baltic seal populations. However, the results suggest that the assays used for the CYP2D and CYP2E activities are at least partly associated with CYP1A activity in these seals. Putative CYP3A members showed a slight opposite trend, being reduced in the Baltic seals, while CYP2A did not show any geographical trend.

81

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

Table 2 Haematological and clinical chemistry blood values in ringed and grey seals from different geographical areasa Ringed seal

Grey seal

Baltic Sea

Svalbard

Baltic Sea

Sable Island

CALUX-TEQ (ng/g l.w.) 0.2 (0.04–0.4)

0.04 (0.01–0.1)*** 0.1 (0.1–0.4)

Cytochromes P450 (pmol/mg prot. min) CYP1Ab CYP3Ac

322 (31–998)*** 1307 (371–1846)*** 259 (21–1082) 563 (129–1072)* 392 (181–654) 667 (343–1514)

1201 (374–2037) 441 (68–521)

Haematology Erythrocytes (1012/l) Haemoglobin (g/l) Haematocrit (%)

4.0 (3.1–5.6)### 203 (134–260)###d 52 (39–71)###

Leukocytes (109/l) Neutrophils (%) Eosinophils (%) Basophils (%) Lymphocytes (%) Monocytes (%) IgG (mg/l)

7.4 42 20 0.9 35 1.4 31

(2.6–17)# (26–64) (8–33) (0–4) (20–55) (0–9)##d (5–50)

Clinical chemistry Na (mmol/l) K (mmol/l) Cl (mmol/l) Ca (mmol/l) P (mmol/l) Total protein (g/l) Albumin (g/l) Creatinine (mmol/l) Uric acid (mmol/l) Urea (mmol/l) Bilirubin (total) (mmol/l) (conjugated) (mmol/l)

155 4.6 107 2.4 2.5 66 33 83 87 10 6.2 0.6

(149–164) ### (3.9–7.0) ### (101–113)## (1.9–2.6) (1.8–3.2)### (55–80)### (28–40)##d (53–136)# (43–158) (5.9–20) (2–13) (0–3)

Cholesterol total (mmol/l) HDL (mmol/l) LDL (mmol/l) Triglycerides (mmol/l) Fe (mmol/l) TIBC (mmol/l) Glucose (mmol/l) Fructosamine (mmol/l)

7.4 6.3 0.6 1.1 34 87 5.6 242

(4.5–10## (4.2–9.4)###d (0–2.9)# (0.5–1.8)### (9–70)#d (56–165)### (1.8–7.2) (121–304)##d

AFOS (U/l) ASAT (U/l) ALAT (U/l)

113 (55–231) 46 (12–88) 46 (15–141)

d

4.1 (2.4–5.3) 217 (177–268) 52 (45–68)

3.5 (0.5–5.3)* 177 (127–220)*d 47 (36–58)

4.2 (3.2–5.5) 151 (112–195)###d 45 (32–53)###d

6.5 42 23 1.9 30 1.4 10

(3.4–12)* (28–75) (3–39) (0–7)* (13–50)* (0–5) (5–50)**

6.2 47 15 1.5 32 2.6 5

(0.3–10.2)* (22–73) (8–27)** (0–5) (11–57) (1–7)** (5)*

12 51 11 2.0 34 1.2 97

152 4.8 104 2.3 2.3 73 29 77 98 13 4.6 0.5

(129–161)* (4.0–5.5) (87–113)* (2.0–2.7) (1.3–3.0) (53–88)***d (23–33)**d (54–114) (43–206)* (5.7–25)*** (1–11) (0–3)

151 5.1 106 2.4 2.8 74 29 88 75 9.7 4.6 1.0

(130–158) 153 (150–157) (4.3–6.4) 5.0 (4.3–7.1) (88–114)***d 116 (113–120)### (2.1–2.6)**d 2.3 (2.1–2.5) (2.0–3.8)***d 2.2 (1.2–2.9) (58–86) 76 (69–87) (26–35)** 27 (22–30)# (63–111) 93 (73–112)### (53–102)*** 61 (43–121)### (5.5–18)***d 14 (8.9–21.3) (2–8)* 3.6 (2–6) (0–3)*** 0.1 (0–1)#

7.7 5.8 1.3 1.4 36 74 5.7 277

(4.9–12.1) (3.5–8.4) (0.3–2.5)*** (0.6–3.7)*** (11–74) (53–113)**d (2.4–8.9) (172–589)

6.0 4.6 0.8 1.3 22 53 5.2 211

(4.6–7.3) (3.4–5.4)* (0.1–2.8) (0.8–2.0)* (5.3–47) (31–77) (2.3–7.3) (144–279)**

54 (0–148)*** 83 (35–174)*** 72 (24–236)***

141 (64–279)** 51 (25–114) 30 (11–81)

6.2 4.9 0.9 1.1 21 56 4.7 182

(8.6–15)### (36–71)## (1–33)### (0–6) (15–54) (0–3) (0.5–500)

(4.7–9.1)### (3.9–6.0)## (0.1–2.6)# (0.6–2.9) (6.2–43)### (41–75)### (2.5–6.1)# (134–222)###

89 (17–634) 49 (32–79)### 34 (10–69)### (continued)

82

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

Table 2 (continued) Ringed seal

gGT (U/l) CK (U/l) LD (U/l)

Grey seal

Baltic Sea

Svalbard

Baltic Sea

Sable Island

15 (5–51)# # 430 (83–1076)## 1344 (957–2137)

17 (6–54) 9.5 (4–19)** 11 (7–26)# 762 (104–3502)* 873 (197–5795) 629 (250–1495) 1591 (692–2790)** 1357 (662–3036)** 996 (849–1254)###

%, Percentage of total leukocytes; IgG, immunoglobulin G; HDL, high density lipoprotein; LDL, low density lipoprotein; TIBC, total iron binding capacity; AFOS, alkaline phosphatase; ASAT, aspartate aminotransferase; ALAT, alanine aminotransferase; gGT, gamma-glutinyl transferase; CK, creatine kinase; LD, lactate dehydrogenase. a Results are presented as means (ranges). The significance of geographical variation within each species presented as *P<0.05 or **P< 0.01, ***P< 0.001. Significant species differences within the Baltic Sea or between the two reference areas is shown as #P <0.05, ##P <0.01 or ###P <0.001. b Nyman et al., 2000. c Nyman et al., 2001. d Sexual difference influences the observed geographical or species variation.

3.4. Vitamins A and E Retinyl palmitate was the predominant form of vitamin A in liver and blubber (comprising 88–99% of the total vitamin A content), while only retinol was detected in plasma. Vitamin A levels were clearly highest in the liver, followed by the blubber and plasma. The plasma retinol levels showed no geographical trend in ringed seals. In grey seals, retinol concentrations in liver and blubber were significantly lower in the Baltic grey seals compared with the reference population, while the plasma retinol showed the opposite geographical trend (Fig. 2). Retinyl palmitate levels in the liver were lower in both Baltic seal populations compared with the reference seals (Fig. 3A). In grey seals, a corresponding geographical trend was observed also for retinyl palmitate in blubber (Fig. 3B). The vitamin A differences observed for plasma and blubber were associated with the sex-difference (tested with ANOVA), and thus probably due to a difference in gender rather than a geographical difference. A slight species variation in vitamin A levels were observed. Between the Baltic and between the reference seal populations retinyl palmitate concentration in liver were higher in ringed seals than in grey seals (P < 0.01), while retinol levels in plasma showed the opposite trend (P < 0.05). The tissue distribution of vitamin A in ringed seals differed from grey seals in that ringed seals had relatively higher vitamin A levels in the liver than in the other tissues. The a-tocopherol concentration represents the total vitamin E concentration in the seal tissues because it is the only form of vitamin E found in seals. Vitamin E levels were higher in all tissues in the Baltic seal populations than in the respective reference samples (Fig. 4). However, the geographical differences in liver and plasma were associated with the sex-difference (tested with ANOVA), and thus were probably due to a gender difference. In general, vitamin E levels were highest in the blubber, followed by the liver and plasma. The Canadian grey seals differed from the

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

83

Fig. 1. Correlation between (A) LN-transformed SPCB-TEQ and CALUX-TEQ levels in liver of ringed seals from the Baltic Sea and Svalbard, and between (B) CALUX-TEQ values and retinyl palmitate in liver of ringed and grey seals.

other seal groups by having similar vitamin E concentrations in liver and blubber. As for the species variation, vitamin E levels in plasma were higher in the Svalbard ringed seals than in the Sable Island grey seals (P < 0.001). In the Sable grey seals, hepatic vitamin E levels were positively correlated with condition (r> 0.70, P > 0.001), while retinol levels in plasma and blubber showed a negative condition dependence (r<
0.61, P < 0.01). None of the vitamins were age-related in any of the seal populations.

84

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

Fig. 2. Retinol concentrations in liver, blubber and plasma of ringed seals and grey seals from different geographical areas. Significant geographical differences are presented with asterisks. (*P <0.05, **P <0.01, ***P <0.001). N, number of animals.

3.5. Haematology and blood chemistry The haematological and clinical chemistry values for the seal populations, and their species or geographical variations, are presented in Table 2. The haematological parameters were very similar between the both ringed and grey seal populations, although some slight differences were observed. In all the seal populations studied, a negative relationship between neutrophils and lymphocytes was observed (r <
0.64, P < 0.001 for all seal populations), while a positive correlation between haemoglobin and haematocrit was observed (r > 0.63, P < 0.001 for all seal populations). A comparison of the haematology between the two species revealed higher haemoglobin and haematocrit levels in ringed seals (Table 2). Other species-specific variations differed between the Baltic and the references seal populations. Of the analysed blood chemistry parameters, both Baltic seal populations showed elevated albumin and alkaline phosphatase (AFOS) concentrations, while the urea levels were slightly reduced compared to the reference seal populations (Table 2). The geographical differences in blood parameters were all less than a factor of 2, except for the conjugated bilirubin that was eight-fold higher in the Baltic grey seals compared to the Sable Island grey seals. A species comparison showed higher levels of albumin, total cholesterol, high density lipoprotein (HDL), iron (Fe), total iron binding capacity (TIBC), fructosamine and gamma-glutinyl transferase (gGT) in ringed seals than in grey seals, while creatinine showed the opposite trend. Chemical blood parameters that showed a consistent positive correlation in all the seal populations

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

85

Fig. 3. Retinyl palmitate concentrations in (A) liver and (B) blubber of ringed and grey seals from the Baltic Sea and reference areas. Significant geographical differences are presented with asterisks (*P <0.05, **P <0.01, ***P <0.001). N, number of animals.

were total cholesterol with HDL and low density lipoprotein (LDL) (r> 0.51, P < 0.05), aspartate aminotransferase (ASAT) with alanine aminotransferase (ALAT) (r> 0.65, P < 0.01), and total bilirubin with its conjugated form (r> 0.53, P < 0.05). Many of the other blood parameters showed species specific correlation patterns that resulted in complex networks. The condition index was positively correlated to creatinine in Baltic ringed seals (r > 0.68, P < 0.001), and to erythrocytes, haemoglobin, haematocrit, ASAT and ALAT in the Sable grey seals (r> 0.66, P < 0.01). In the same seal population, phosphorus and potassium showed a negative

86

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

Fig 4. Vitamin E concentrations in liver, blubber and plasma of ringed and grey seals from different geographical areas. Significant geographical differences are presented with asterisks (*P<0.05, **P <0.01, ***P <0.001). N, number of animals.

condition dependency (r <
0.62, P < 0.01), and calcium and iron showed a negative age dependency (r <
0.60, P < 0.01). The relationship between the biochemical parameters differed in the two species. In ringed seals, vitamin E, CYP1A and AFOS levels showed a positive correlation with each other (Fig. 5A). In grey seals, vitamin E, CYP1A, calcium, phosphorus, albumin and uric acid levels were positively correlated, while an opposing coherent group comprised hepatic and blubber retinyl palmitate, chloride and urea (Fig. 5C). 3.6. Relationships between potential biomarkers and contaminants Of the biochemical parameters assayed, the ones showing a geographical difference that was statistically significant were included for the PCA analysis. When summarising the PCA results, the correlation pattern revealed was mostly consistent with the pattern described previously for the biochemical parameters; CYP1A and vitamin E in blubber being positively correlated with the POPs, while retinyl palmitate showed the opposite trend (Fig. 5). In addition, species specific correlation patterns were revealed. In grey seals, CYP1A, vitamin E in plasma and blubber, retinol and phosphate showed a positive correlation with the POPs, while chloride, urea, retinyl palmitate in blubber were negatively correlated with both the organic contaminants and with cadmium in liver and kidney. In ringed seals, CYP1A, CALUX and vitamin E in blubber showed a positive correlation to the POPs, while total protein, ASAT and retinyl palmitate in liver showed a negative correlation. For both species, principal component two (t[2]) only explains a fraction of the data, 13%, respectively, 7%, and hence the spread in y-axis shown in Fig. 5A and C

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

87

Fig. 5. PCA showing relationships between potential biomarkers and contaminants for grey seal (A and B) and ringed seal (C and D), respectively. Metals were not included in the ringed seal analysis, and CALUX was not included in the grey seal analysis. Vitamin A is presented as RP (retinyl palmitate) or retinol (in plasma), Vitamin E as Vit E, Chloride as Cl, phosphate as P, total protein in plasma as prot. and aspartate aminotransferase as ASAT.

should not be over-interpreted. The total amount of variance in the data explained by PCA analysis using two components is 49 and 50% for grey seals and ringed seals, respectively. Considering biological variation, these percentages are within an acceptable range. Based on the PCA findings further correlation analyses were conducted on the data, specifying the species specific correlative relationship between the contaminants and each potential biomarker. These results are presented in Table 3. In ringed seals, the CALUX-TEQ response did not show any linear relation to any other biochemical parameter than CYP1A throughout the CALUX-TEQ range (r> 0.70, P < 0.001). A trend of a negative correlation with retinyl palmitate (although not statistically significant) was observed in ringed seals up to CALUX-

88

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

Table 3 Correlation matrix (presented as r) between contaminant levels and biochemical parameters in (A) ringed seals and (B) grey seals. The significance of correlation is presented as *P<0.05, **P <0.01 and ***P <0.001. Sample size is marked as n

Ringed seals CYP1A-activity (n) Vitamin A Liver (RP) (n)

CALUXTEQ

Splanar PCBs

SPCBTEQ

SPCB

SDDT

Mercury

Selenium

0.71*** (29)

0.85*** (48)

0.77*** (48)

0.77*** (48)

0.77*** (48)

0.78*** (34)

0.72*** (34)


0.50** (28)


0.49** (28)


0.51** (28)


0.54** (28)

0.94** (6)

0.93** (6)


0.19 (9)

Vitamin E Liver (n) Blubber (n) Plasma (n)


0.84** (9) 0.46* (23) 0.46* (28)

0.50** (28) 0.66*** (42) 0.71*** (47)

0.53** (28) 0.66*** (42) 0.71*** (47)

0.55** (28) 0.71*** (42) 0.77*** (47)

0.48* (28) 0.69*** (42) 0.72*** (47)

0.44 (6) 0.50* (19) 0.78*** (25)

0.39 (6) 042 (19) 0.77*** (25)

Blood parameters IgG (n) Na (n) Cl (n) ASAT (n)

0.96*** (17) 0.26 (29) 0.36 (29)
0.58*** (29)

0.51** (34) 0.28 (47) 0.33 (47)
0.45** (48)

0.49** (34) 0.12 (47) 0.23 (47)
0.51*** (48)

0.51** (34) 0.17 (47) 0.27 (47)
0.52*** (48)

0.51** (34) 0.13 (47) 0.23 (47)
0.55*** (48)

n.d.

n.d.

0.55** (35) 0.54** (35)
0.69*** (35)

0.62*** (35) 0.62*** (35)
0.62*** (35)

Sample size

Splanar PCBs

STEQPCB

SPCB

SDDT

Cadmium Liver

(B) Grey seals CYP1A-activity (n) Vitamin A Plasma (RE) (n) Liver (RE) (n) Liver (RP) (n) Blubber (RP) (n)

0.75*** (35)

0.86*** (35)

0.80*** (35)

0.82*** (35)


0.80*** (36)

0.73*** (34)
0.58*** (35)
0.45** (35)
0.66*** (34)

0.81*** (34)
0.57*** (35)
0.41* (35)
0.61*** (34)

0.74*** (34)
0.46** (35)
0.27 (35)
0.48** (34)

0.81*** (34)
0.52** (35)
0.38* (35)
0.61*** (34)


0.57*** (39) 0.71*** (40) 0.55*** (40) 0.74*** (39)

Vitamin E Blubber (n) Plasma (n)

0.64*** (35) 0.82*** (34)

0.67*** (35) 0.78*** (34)

0.57*** (35) 0.65*** (34)

0.63*** (35) 0.78*** (34)


0.60*** (40)
0.73*** (39) (continued)

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

89

Table 3 (continued) CALUXTEQ Blood parameters Leukocytes (n) Cl (n) Ca (n) P (n) Urea (n) Uric acid (n)

Splanar PCBs

SPCBTEQ

SPCB

SDDT

Mercury


0.67** (15)
0.76*** (34) 0.53** (34) 0.77*** (34)
0.62*** (34) 0.56** (34)


0.61* (15)
0.78*** (34) 0.57*** (34) 0.86*** (34)
0.66*** (34) 0.61*** (34)


0.51 (15)
0.66***


0.51 (15)
0.81***

0.48** (34) 0.82*** (34)
0.53** (34) 0.58*** (34)

0.61*** (34) 0.88*** (34)
0.69*** (34) 0.55** (34)

0.50* (20) 0.80*** (39)
0.48** (39)
0.73*** (39) 0.68*** (39)
0.41** (39)

Selenium

RP, retinyl palmitate; RE, retinol; IgG, immunoglobulin G; ASAT, aspartate aminotransferase; n.d., no data.

TEQ levels of 40 pg TEQ/g lipid, after which the vitamin A levels did not decrease any further (Fig. 1B).

4. Discussion 4.1. CALUX The CALUX-TEQ-values in Baltic ringed seals reflect a higher burden of dioxinlike compounds in this population than in the Svalbard ringed seals. A strong positive relationship between the CALUX response and the SDDT and SPCB indicates that DDTs and non-planar PCBs are co-accumulated with the planar PCBs and other probably unknown dioxin-like compounds. The lack of a linear relationship between the CALUX response and most of the biochemical parameters in the ringed seals could be explained by the non-linear relationship throughout the CALUXTEQ range, and by the fact that the CALUX response is developed for rats and humans. The negative correlation pattern with retinyl palmitate, observed in ringed seals (Fig. 1B), is corresponding to what has been observed in otters previously (Murk et al., 1998). However, the seals seem to be more sensitive than otters. In otters the EC90 is 5 ng TEQ/g lipid and the dose-response curve is very steep. 4.2. Cytochrome P450 enzymes Cytochrome P450 enzymes are a large group of oxidating enzymes, of which a few are known to metabolise xenobiotic compounds. In addition to their xenobiotic-metabolising properties, some of these enzymes are induced by certain organochlorines.

90

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

CYP1, CYP2 and CYP3 associated enzymes have been reported in a number of different aquatic wildlife species, and especially the induction of CYP1A has been used as a biomarker for organochlorine exposure (review in Bainy, Woodin, & Stegeman, 1999; Letcher, Norstrom, Lin, Ramsey, & Bandiera, 1996; Marsili et al., 1998; Stegeman et al., 1992; Troisi & Mason, 1997). CYP1A is primarily induced through the Ah-receptor signal transduction pathway, which activates the production of CYP1A proteins (Safe, 1994). As only compounds with a planar configuration can bind to this receptor, the strong positive correlation between the CYP1A activity, expressed as ethoxyresorufin O-deethylase activity (EROD), and all DDT and PCB compounds in both species could be due to the co-accumulation of organic compound rather than all these compounds being associated with the induction of CYP1A. On the other hand, EROD is not a strict CYP1A marker as a strong induction occurs also in rats treated with phenobarbital, a typical CYP2B inducer (Nyman et al., 2000). The induction potency of EROD was clearly stronger in grey seals than in ringed seals, although the geographical difference in contaminant levels was much smaller. It seems that grey seals have a better metabolic capacity to react to increasing levels than the ringed seals. Further supporting this hypothesis is the higher total PCB and DDT load in the Baltic ringed seal although both species feed at approximately the same trophic level of the Baltic food web, and are therefore expected to be exposed to similar contaminant levels (So¨derberg, 1974; Tormosov & Rezvov, 1978). This warns for extrapolating between two species when using EROD activity as a biomarker. 4.3. Vitamins A and E Several POPs have been shown to interfere with the Vitamin A homeostasis by different mechanisms (Brouwer & Van den Berg, 1986; Brouwer, Blaner, Kukler, & van den Berg, 1988; Brouwer, Reijnders, & Koeman, 1989; Chen et al., 1992; Murk, Morse, Boon, & Brouwer, 1994c; Nilsson et al., 2000; Spear, Garcin, & Narbonne, 1988). Observed effects have been either reduced or increased plasma retinol levels, or a depletion of the vitamin A stores. The vitamin A homeostasis in the body has been used as a biomarker for PCB exposure in birds, otters and seals (Brouwer et al., 1989; Jenssen, Skaare, Wolstad, Nastad, Haugen, & Sørmo, 1995; Murk, Bosveld, Van den Berg, & Brouwer, 1994a; Murk et al., 1994b; Murck et al., 1998; Peakall, 1992; Ross et al., 1994). In this study, vitamins A and E showed geographical differences that could reflect differences in the seals’ general health status and functional response to environmental pollutant exposure, or reflect the dietary levels in each area. The vitamin A status of the Canadian grey seal liver and blubber was in accordance with previous reports from the same area, and plasma retinol levels in our study were in the same range as those reported previously for various pinniped species (Schweigert, Stobo, & Zucker, 1987; Schweigert, Ryder, Rambeck, & Zucker, 1990a). The correlation patterns of the vitamin A forms with the organic contaminants in the seals suggest an increased release of retinol from the vitamin A stores with increasing contaminant levels, leading to a depletion of the vitamin A stores.

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

91

Plasma retinol levels were slightly elevated in the Baltic grey seals compared to the reference seals. As plasma retinol concentrations in seals are tightly controlled over a wide range of dietary vitamin A availability (Mazzaro, Dunn, Furr, & Clark, 1995a), an extensive disturbance of the vitamin A balance would be needed before any changes occurred in the plasma retinol levels. In addition, plasma retinol levels have been shown to correlate poorly with the depletion of hepatic retinol concentrations (Got et al., 1995). The elevated plasma retinol levels in the Baltic grey seals further supports the hypothesis of a depletion of the vitamin A stores in the Baltic seals. Vitamin E is known as one of the major antioxidants, protecting the body against oxidative stress induced by both natural functions and xenobiotic metabolism (Slim, Toborek, Robertson, & Hennig, 1999; Yamamoto et al., 1994). It has been proposed as a bioindicator for contaminant induced oxygen radical stress (Stegeman et al., 1992), but its use as a biomarker has not received much attention. In both ringed and grey seals from Svalbard and Canada, liver and blubber vitamin E levels were similar to those previously reported in grey seals from Sable Island (Schweigert, Stobo, & Zucker, 1990b). Serum levels and the relative tissue distribution of vitamin E in all the seal populations studied were in the same range as has been reported for other pinniped species (Engelhardt, Geraci, & Walter, 1975; Schweigert, Ryder et al., 1990; Schweigert, Stobo et al., 1990). The two to three-fold elevation of vitamin E concentrations in blubber and plasma of both species of the Baltic seals could be due to higher vitamin E levels in the Baltic seal diet, or to a difference in the feeding activity of between seals in the Baltic Sea and those in the reference areas. On the other hand, as fish species caught in the wild and commonly fed to zoo animals meet the vitamin A and E requirements of marine mammals, it suggests that seals in the wild do not suffer from vitamin deprivation (Dierenfeld, Katz, Pearson, Murru, & Asper, 1991; Mazzaro, Dunn, Furr, & Clark, 1995b). Furthermore, all seals were sampled during moult, a period when food intake is greatly reduced (Castellini & Rea, 1992), and showed no difference in their condition. PCB exposure induces the production of free radicals as a by-product of oxidative metabolism (Stegeman & Hahn, 1994). Vitamin E has been shown to protects endothelial cells of the vasculature from oxidative stress caused by PCB in vitro (Slim et al., 1999). We have shown previously that cytochrome P4501A enzymes (CYP1A) are specifically induced in vascular endothelial cells in most non-hepatic tissues in both Baltic seal populations, indicating that the whole body is affected by the contaminant exposure (Hyyti, Nyman, Willis, Raunio, & Pelkonen, 2001). Similar enzymatic induction patterns in non-hepatic tissues have been observed in many animal species after treatment with or exposure to typical CYP1A inducers (Anderson et al., 1987; Dees, Master, Muller-Eberhard, & Johnson, 1982; Husøy, Myers, & Goksøyr, 1996; Thirman et al., 1994; Van Veld, Vogelbein, Cochran, Goksøyr, & Stegeman, 1997). The clear positive trend between vitamin E and both PCB and DDT in both seal species indicates that vitamin E could be involved in combating oxygen radical production caused by the heavy contaminant load. This trend being more obvious in ringed seals with higher contaminant levels further supports this hypothesis. The antioxidising function of vitamin E could also explain

92

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

the positive relation between this vitamin and mercury in ringed seals, since mercury can induce lipid perodixadion and radical production (Manca, Ricard, Trottier, & Chevalier, 1991; Stacey & Kappus, 1982). However, the negative trend for cadmium in grey seals is contradictory to this suggestion. 4.4. Haematology and clinical chemistry Clinical blood parameters are commonly used for screening the health of humans and domestic animals. In studies on the effects of exposure to POPs in experimental animals and humans, pathological changes have been more obvious than variations in biochemical parameters such as clinical blood values. Increased total serum lipids, cholesterol and triglycerides are the most commonly found changes in clinical chemistry parameters after PCB exposure (Allen, Carstens, & Abrahamson, 1976; Borlakoglu & Welch, 1992; Kreiss, Zack, Kimbrough, Needham, Smrek, & Jones, 1981; Yamamoto et al., 1994). Clinical blood parameters have been successfully used as biomarkers of toxic effects to some extent in fish and birds (reviewed by Peakall, 1992). Most parameters have been reported also in marine mammals as base line values for healthy captive seals. However, they have rarely been used as biomarkers for contaminant exposure or effects. The few reported observations of changes in clinical blood values in free-living and captive seals exposed to environmental contaminants have not been able to explain the observed changes by contaminant effects (De Swart et al., 1995; Hall, Pomeroy, Green, Jones, & Harwood, 1997; Reijnders, 1988; Schumacher, Heidemann, Skı´rnisson, Schumacher, & Pickering, 1995). In general, the blood chemistry parameters in our ringed and grey seal populations were within the ranges previously reported for wild and captive phocids (De Swart et al., 1995; Engelhardt, 1979; Geraci, Aubin, & Smith, 1979; Greenwood, 1971; McConnell & Vaughan, 1983). As reference values for many of the parameters analysed were not available for the species studied, the results were compared with the information available from other phocid species. The small geographical variations in the clinical screening values together with the lack of a clear relationship with the contaminant load for most parameters, suggests that the observed geographical differences may not have any toxicological relevance. The exception is the relationship between the POPs and phosphorus and calcium that could be of interest. A simultaneous elevation of phosphorus and calcium levels could be due to an increased binding of excess calcium to phosphorus in the blood, suggesting an increased mobilisation of calcium from the skeleton (Laurell, Lundh, & Nosslin, 1975). This hypothesis is supported by studies on experimental animals showing that POPs affect the bone structure by various mechanisms (Andrews, 1989; Andrews, Courtney, Stead, & Donaldson, 1989; Andrews, Jackson, Stead, & Donaldson, 1990; Lind, Eriksen, Sahlin, & O¨rberg, 1999). Further supporting the possible adverse effects of POPs on the bone structure of Baltic grey seals is the strong association between an elevated occurrence of skull bone lesions and the high POP load in this seal population (Bergman et al., 1992; Olsson et al., 1994).

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

93

4.5. Proposed biomarkers in seals When summarising the information on the relationships between potential biomarkers and contaminants and the geographical trends of the biochemical parameters, a clear and species- specific pattern was revealed (Fig. 5). In ringed seals, the POP parameters and metals showed a positive correlation to vitamin E and CYP1A, and a negative correlation to retinyl palmitate in liver, total protein and ASAT. The correlation pattern between CALUX-TEQ and the biochemical parameters showed similar trends, but were not consistent with the STEQ-PCB (Table 3). This suggests that CALUX-TEQ does not represent the complete SPCB-TEQ of the seals. It should be borne in mind that both the CALUX assay and the TEQ values are developed for humans and experimental animals, and extrapolation between species should be conducted with caution. In grey seals, vitamin E, CYP1A, plasma retinol and phosphorus showed a positive correlation with the POPs, while retinyl palmitate in blubber, chloride and urea were negatively correlated with the POPs. The biochemical parameters showed an opposite relationship to cadmium than to the organochlorines. The connection between POPs, CALUX, CYP1A and vitamin E could be explained by a dioxin like induction of CYP1A via the aryl hydrocarbon receptor, resulting in an increased need for radical scavengers such as vitamin E. The correlation pattern between the POPs and the vitamin A parameters indicate a toxic effect of these compounds depleting the vitamin A stores. The species-specific relationships between biochemical parameters and contaminants do not clearly suggest that either species is more vulnerable to the contaminant load in the Baltic Sea, but does indicate that each species reacts to the xenobiotic exposure in a different way. A better understanding of the underlying toxic mechanisms in needed before these assumptions can be tested. In addition, The observed network of relationships should be interpreted with caution as the complexity of patterns makes it difficult to discern the direct connections between parameters and contaminants from the indirect ones. Anyhow, it is important to be aware of the species differences when monitoring the health status and possible toxic effects of the contaminant load in ringed and grey seals. Currently, most ecotoxicological risk assessments are based on chemical residue analysis of abiotic matrices that are correlated with determinations of toxic effects in wildlife species. It is, however, often not known whether these correlation patterns are due to a causal relationship. The possible detection of direct effects in response to release of specific pollutants from point sources (such as industrial effluent or spills) is relatively easy. The detection, however, of chronic effects caused by mixtures such as POPs, which are released over long periods of time and from diffuse sources, is much more difficult. In addition, the response of populations to pollutants is often non-specific and hard to distinguish from responses to natural influences, such as nutritional status and physical stresses. Even if adverse effects observed under field conditions are found to be correlated with certain POP-concentrations, it remains unclear whether these contaminants play a role as etiological agents, since most study areas are polluted with a cocktail of contaminants. It can

94

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

never be excluded that the suggested causal relationship is due to co-correlation with other pollutants present in the exposure matrix that are not included in the exposure assessment or are below the detection limit.

5. Conclusions Most studies on the toxicity of antropogenic contaminants have focused on specific compounds, whereas wild animals are exposed to a mixture of compounds with varying toxicity. Although the chemical profile gives indications of biochemical responses to the contaminant load, possible additive, inhibiting, or exponential interactions between individual substances cannot be detected. With additional biological markers, biomarkers, it is possible to observe the subleathal responses of the total contaminant load, that ultimately may lead to pathological impairment and disease. Of the potential biomarkers and bioassays assessed in this study, CYP1A and vitamin E in blubber or plasma are proposed as potential exposure biomarkers for PCBs and DDTs in both species. The AhR-mediated CALUX response reflects the whole PCB and DDT burden in ringed seals. Used on blood samples it offers a non-destructive measure of the TEQ burden in seals. Retinyl palmitate levels are proposed as potential biomarkers for the depletion of the vitamin A stores. For both vitamins, more information on the dietary vitamin levels is needed before any conclusions can be drawn. Despite a strong positive correlation between plasma retinol and POPs in grey seals, plasma retinol cannot be recommended as a biomarker as POPs both increase and decrease the levels by different mechanisms. The clinical screening parameters show great individual variations and small geographical differences that do not make them suitable as biomarkers but rather as a tool for screening the general health of the individual.

Acknowledgements We thank Marcus Wikman, Bjørn Kraft, Christian Lydersen and the staff at the Polar Institute for assisting in the field sampling of ringed seals in Ny A˚lesund, in Svalbard. Richard Addison, Wayne Stobo and Chris Harvey-Clark we thank for coordinating and helping us with the grey seal sampling on Sable Island, Canada. We would like to thank Kalle Ja¨rvinen for the field assistance in the Bothnian Bay, and Sanna Sistonen, Soili Nikonen, Merja Luukkonen, Ritva Tauriainen and Pa¨ivi Tyni for their help in the laboratory. We are grateful to Esa Tulisalo for his assistance with the metal analyses, to Pa¨ivi Ekholm for conducting the Se analyses, and to Henrik Sundberg for performing the CALUX analyses. This study was supported by the Finnish Academy of Science and benefited from the method development work of the EUROCYP project.

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

95

References Aarts, J. J. M. J.G., Denison, M. S., Cox, M. A., Schalk, M. A. C., Garrison, P. A., & Tullis, K., et al. (1995). Species-specific antagonism of Ah receptor action by 2,20 ,5,50 -tetrachloro- and 2,20 ,3,30 ,4,40 -hexachlorobiphenyl. European Journal of Pharmacology and Environmental Toxicology, Section 293, 463–474. Allen, J. R., Carstens, L. A., & Abrahamson, L. J. (1976). Responses of rats exposed to polychlorinated biphenyls for fifty-two weeks I. Comparison of tissue levels of PCB and biological changes. Archives of Environmental Contamination and Toxicology, 4, 404–419. AMAP (1998). Assessment report: Arctic pollution issues. Chapter 7: Heavy metals (pp. 373–524). Oslo, Norway: Arctic Monitoring and Assessment Programme (AMAP). Anderson, L. M., Jerrold, J. M., Ward, M., Park, S. S., Jones, A. B., & Junker, J. L., et al. (1987). Immunohistochemical determination of inducibility phenotype with a monoclonal antibody to a methylchloranthrene-inducible isozyme of cytochrome P-450. Cancer Research, 47, 6079–6085. Andrews, J. E. (1989). Polychlorinated biphenyl (Arochlor 1254) induced changes in femur morphometry calcium metabolism and nephrotoxicity. Toxicology, 57, 83–96. Andrews, J. E., Courtney, K. D., Stead, A. G., & Donaldson, W. E. (1989). Hexachlorobenzeneinduced hyperparathyroidism and osteosclerosis in rats. Fundamental and Applied Toxicology, 12, 242–251. Andrews, J. E., Jackson, L. D., Stead, A. G., & Donaldson, W. E. (1990). morphometric analysis of osteosclerotic bone resulting from hexachlorobenzene exposure. Journal of Toxicology and Environmental Health, 31, 193–201. Bainy, A. C. D., Woodin, B. R., & Stegeman, J. J. (1999). Elevated levels of multiple cytochrome P450 forms in tilapia from Billings Reservoir-Sa˜o Paulo, Brazil. Aquatic Toxicology, 44, 289–305. Bauer, K. M., Cramer, P. H., Stanley, J. S., Fredette, C., & Giglinto, T. L. (1992). Multivariate statistical analyses of PCDD and PCDF levels in fish, sediment, and soil samples collected near resource recovery facilities. Chemosphere, 25, 1441–1447. Beck, G. G., & Smith, T. G. (1995). Distribution of blubber in the Northwest Atlantic harp seal, Phoca groenlandica. Canadian Journal of Zoology, 73, 1991–1998. Bergman, A. (1999). Health condition of the Baltic grey seal (Halichoerus grypus) during two decades. APMIS, 107, 270–282. Bergman, A., & Olsson, M. (1986). Pathology of the Baltic grey seal and ringed seal females with special reference to adrenocortical hyperplasia: is environmental pollution the cause of a widely distributed disease syndrome? Finnish Game Research, 44, 47–62. Bergman, A., Olsson, M., & Reiland, S. (1992). Skull-bone lesions in the Baltic grey seal (Halichoerus grypus). Ambio, 21, 517–519. Bieri, J. G., Tolliver, T. J., & Catignani, G. L. (1979). Simultaneous determination of a-tocopherol and retinol in plasma or red cells by high pressure liquid chromatography. The American Journal of Clinical Nutrition, 32, 2143–2149. Boon, J. P., van Arnhem, E., Jansen, S., Kannan, N., Petrick, G., & Schulz, D., et al. (1992). The toxicokinetics of PCBs in marine mammals with special reference to possible interactions of individual congeners with the cytochrome P450-dependent monooxygenase system—an overview. In C. H. Walker, & D. Livingstone (Eds.), Persistent pollutants in marine ecosystems (pp. 119–159). Oxford, UK: Pergamon Press. Boon, J. P., vanderMeer, J., Allchin, C. R., Law, R. J., Klunsoyr, J., & Leonards, P. E., et al. (1997). Concentration-dependent changes of PCB patterns in fish-eating mammals: Structural evidence for induction of cytochrome P450. Archives of Environmental Contamination and Toxicology, 33, 298–311. Borlakoglu, J. T., & Welch, V. A. (1992). Xenobiotic-induced aberrations of lipid metabolism. In C. K. Chow (Ed.), Fatty acids in foods and their health implementations (pp. 613–630). New York: Marcel Dekker Inc. Brouwer, A., & Van Den Berg, K. J. (1986). Binding of a metabolite of 3, 4,30 , 40 -tetrachlorobiphenyl to transthyretin reduces serum vitamin A transport by inhibiting the formation of the protein complex carrying both retinol and thyroxin. Toxicology & Applied Pharmacology, 85, 301–312.

96

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

Brouwer, A., Blaner, W., Kukler, A., & van den Berg, K. (1988). Study on the mechanism of interference of 3,4,30 ,40 -tetrachlorobiphenyl with the plasma retinol-binding proteins in rodents. Chem.–Biol. Interactions, 68, 203–217. Brouwer, A., Reijnders, P. J. H., & Koeman, J. H. (1989). Polychlorinated biphenyl (PCB)-contaminated fish induces vitamin A and thyroid hormone deficiency in the common seal Phoca vitulina. Aquatic Toxicology, 15, 99–106. Castellini, M. A., & Rea, L. D. (1992). The biochemistry of natural fasting at its limits. Experientia, 48, 575–582. Catignani, G. L., & Bieri, J. G. (1983). Simultaneous determination of retinol and a-tocopherol in serum or plasma by liquid chromatography. Clinical Chemistry, 29, 708–712. Chen, L.-C., Berberian, I., Koch, B., Mercerier, M., Azais-Braesco, V., & Glauert, H. P., et al. (1992). Polychlorinated and polybrominated biphenyl congeners and retinoid levels in rat tissues: structureactivity relationships. Toxicology & Applied Pharmacology, 114, 47–55. Dacie, J. V., & Lewis, S. M. (1984). Preparation and staining methods for blood and bone-marrow films. In J. V. Dacie, & S. M. Lewis (Eds.), Practical haematology (pp. 50–61). London, UK: Churchill Livingstone. Dees, J. H., Masters, B. S. S., Muller-Eberhard, U., & Johnson, E.F (1982). Effect of 2,3,7,8-tetrachlorodibenzo-p-dioxin and phenobarbital on the occurrence and distribution of four cytochrome P-450 isozymes in rabbit kidney, lung and liver. Cancer Research, 42, 1423–1432. DeRuyter, M. G. M., & DeLeenheer, A. P. (1978). Simultaneous determination of retinol and retinyl esters in serum or plasma by reversed-phase high-performance liquid chromatography. Clinical Chemistry, 24, 1920–1923. De Swart, R. L., Ross, P. S., Vedder, L. J., Boink, F. B., Reijnders, P. J., & Mulder, P. G., et al. (1995). Hematology and clinical chemistry values of harbor seals (Phoca vitulina) fed environmentally contaminated herring remain within normal ranges. Canadian Journal of Zoology, 73, 2035–2043. Dierenfeld, E. S., Katz, N., Pearson, J., Murru, F., & Asper, E. D. (1991). Retinol and a-tocopherol concentrations in whole fish commonly fed in zoos and aquariums. Zoo Biology, 10, 119–125. Engelhardt, F. R., Geraci, J. R., & Walker, B. L. (1975). Tocopherol distribution in the harp seal, Phagophilus groenlandicus. Comparative Biochemistry & Physiology, 52, 561–562. Engelhardt, F. R. (1979). Haematology and plasma chemistry of captive pinnipeds and cetaceans. Aquatic Mammals, 7, 11–24 Fant, M. L., Nyman, M., Helle, E., & Rudba¨ck, E. (2001). Mercury, cadmium lead and selenium in ringed seals (Phoca hispida) from the Baltic Sea and from Svalbard. Environmental Pollution, 111, 493–501. Geraci, J. R., Aubin, D. J., & Smith, T. G. (1979). Influence of age, condition, sampling time, and method on plasma chemical constituents in free-ranging ringed seals, Phoca hispida. J. Fish. Res. Board Can, 36, 1278–1282. Got, L., Gousson, T., & Delacoux, E. (1995). Simultaneous determination of retinyl esters and retinol in human livers by reversed-phase high-performance liquid chromatography. Journal of Chromatography B, 668, 233–239. Greenwood, A. (1971). Blood values in young gray seals. J.A.V.M.A, 159, 571–574. Hall, A., Pomeroy, P., Green, N., Jones, K., & Harwood, J. (1997). Infection, haematology and biochemistry in grey seal pups exposed to chlorinated biphenyls. Marine Environmental Research, 43, 81–98. HELCOM (1996). Third periodic assessment of the state of the marine environment of the Baltic Sea, 1989–1993; background document. Baltic Sea Environmental Proceedings, 64B, 144–153. Helle, E. (1981). Reproductive trends and occurrence of organochlorines and heavy metals in the Baltic seal populations. ICES (CM papers and reports) E: 37. Helle, E., Olsson, M., & Jenssen, S. (1976a). DDT and PCB levels and reproduction in ringed seal from the Bothnian Bay. Ambio, 5, 188–189. Helle, E., Olsson, M., & Jenssen, S. (1976b). PCB levels correlated with pathological changes in seal uteri. Ambio, 5, 261–263. Hendriks, H. F. J., Verhoofstad, W. A. M. M., Brouwer, A., de Leeuw, A. M., & Knook, D. L. (1985). Perisunoidal fat-storing cells are the main vitamin A storage sites in rat liver. Experimental Cell Research, 160, 138–149.

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

97

Herva, E., & Ha¨sa¨nen, E. (1972). Mercury in seals of the Gulf of Bothnia. Suomen Ela¨inla¨a¨ka¨rilehti, 78, 445–448 (in Finnish). Husøy, A.-M., Myers, M. S., & Goksøyr, A. (1996). Cellular localization of cytochrome P450 (CYP1A) induction and histology in Atlantic cod (Gadus morhua L) and European flounder (Platichthys flesus) after environmental exposure to contaminants by caging in Sorfjorden, Norway. Aquatic Toxicology, 36, 53–74. Hyyti, O., & Nyman, M., Willis, M.L., Raunio, H., & Pelkonen, O. (2001). Distribution of cytochrome P4501A (CYP1A) in the tissues of ringed and grey seals.. Marine Environmental Research, 51(5), 465–485. Jensen, S. (1972). The PCB story. Ambio, 1, 123–131. Jensen, S., Johnels, A. G., Olsson, M., & Otterlind, G. (1969). DDT and PCB in marine mammals from Swedish waters. Nature, 224, 247–250. Jenssen, B., Skaare, J., Wolstad, S., Nastad, A., Haugen, O., & Sørmo, E. (1995). Biomarkers in blood to assess effects of polychlorinated biphenyls in free-living grey seal pups. In A. S. Blix, L. Walløe, & O. Ulltang (Eds.), Whales, fish and man (pp. 607–615). Amsterdam: Elsevier. Kari, T., & Kauranen, P. (1978). Mercury and selenium contents of seals from fresh and brackish waters in Finland. Bulletin of Environmental Contamination & Toxicology, 19, 273–280. Kreiss, K., Zack, M., Kimbrough, R., Needham, L., Smrek, A., & Jones, B. (1981). Association of blood pressure and polychlorinated biphenyl levels. JAMA, 245, 2505–2509. Laurell, C.- B., Lundh, B., & Nosslin, B. (1975). Mineral—och skelettomsa¨ttning. In Klinisk kemi (pp. 85– 114). Lund, Sweden: Studentlitteratur. Letcher, R. J., Norstrom, R. J., Lin, S., Ramsay, M. A., & Bandiera, S. M. (1996). Immunoquantitation on microsomal monooxygenase activities of hepatic cytochromes P4501A and P4502B and chlorinated hydrocarbon contaminant levels in polar bears (Ursus maritimus). Toxicology and Applied Pharmacology, 137, 127–140. Lind, P. M., Eriksen, E. F., & Sahlin, L. & O¨rberg, J (1999). Effects of the anti-estrogenic environmental pollutant 3,30 ,4,40 ,5-pentachlorobiphenyl (PCB#126) in rat bone and uterus: diverging effects in ovariectomized and intact animals. Toxicology and Applied Pharmacology, 154, 236–244. Macdonald, C. R., Metcalfe, C. D., Balch, G. C., & Metcalfe, T. L. (1993). Distribution of PCB congeners in 7 lake systems—interactions between sediment and food-web transport. Environmental Toxicology and Chemistry, 12, 1991–2003. Manca, D., Ricard, A. C., Trottier, B., & Chevalier, G. (1991). Studies on lipid peroxidation in rat tissues following administration of low and moderate doses of cadmium chloride. Toxicology, 67, 303–323. Marsili, L., Fossi, M. C., Notarbartolo Di Sciara, G., Zanardelli, M., Nani, B., & Panigada, S., et al. (1998). Relationship between organochlorine contaminants and mixed function oxidase activity in skin biopsy specimens of Mediterranean fin whales (Balanoptera physalus). Chemosphere, 37, 1501–1510. Mattson, M., Raunio, H., Pelkonen, O., & Helle, E. (1998). Elevated levels of cytochrome P450 (CYP1A) in ringed seals from the Baltic Sea. Aquatic Toxicology, 43, 41–50. Mazzaro, L. M., Dunn, L. J., Furr, H., & Clark, R. M. (1995a). Vitamin A plasma kinetics in northern fur seals (Callorhinus ursinus), using vitamin A2 as a tracer. Canadian Journal of Zoology, 73, 10–14. Mazzaro, L. M., Dunn, J. L., Furr, H. C., & Clark, R. M. (1995b). Study of vitamin A supplementation in captive northern fur seals (Callorhinus ursinus) and its effect on serum vitamin E. Marine Mammal Science, 11, 545–553. McConnell, L. C., & Vaughan, R. W. (1983). Some blood values in captive and free-living common seals (Phoca vitulina). Aquatic Mammals, 10, 9–13. Murk, A. J., Bosveld, A. T. C., Van den Berg, M., & Brouwer, A. (1994a). Effects of polyhalogenated aromatic hydrocarbons (PHAHs) on biochemical parameters in chicks of the common tern (Sterna hirundo). Aquatic Toxicology, 30, 91–115. Murk, A. J., Van den Berg, J. H. J., Fellinger, M., Rozemeijer, M. J. C., Swennen, C., & Duiven, P., et al. (1994b). Toxic and biochemical effects of 3,30 ,4,40 -tetrachlorobiphenyl (CB77) and clophen A50 on eider ducklings (Somateria mollissima) in a semi-field experiment. Environmental Pollution, 86, 21–30. Murk, A., Morse, D., Boon, J., & Brouwer, A. (1994c). In vitro metabolism of 3,30 ,4,40 -tetrachlorobiphenyl in relation to ethoxyresorufin-O-deethylase activity in liver microsomes of some wildlife

98

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

species and rat. European Journal of Pharmacology and Environmental Toxicology, Pharmacology Section, 270, 253–261. Murk, A. J., Legler, J., Denison, M. S., Giesy, J. P., Van de Guchte, C., & Brouwer, A. (1996). Chemicalactivated luciferase gene expression (CALUX): a novel in vitro bioassay for Ah receptor active compounds in sediments and pore water. Fundamental and Applied Toxicology, 33, 149–160. Murk, A. J., Leonards, P. E. G., Bulder, A. S., Jonas, A., Rozemeijer, M. J. C., & Denison, M. S., et al. (1997). The CALUX (chemical activated luciferase expression) assay adapted and validated for measuring TCDD equivalents in blood plasma. Environmental Toxicology and Chemistry, 16, 1583–1589. Murk, A. J., Leonards, P. E. G., Van Hattum, B., Luit, R., Van der Weiden, M. E. J., & Smit, M. (1998). Application of biomarkers for exposure and effects of polyhalogenated aromatic hydrocarbons in naturally exposed European otters (Lutra lutra). Environmental Toxicology and Pharmacology, 6, 91– 102. Nilsson, C. B., Hoegeberg, P., Trossvik, C., Azais-Braesco, V., Blaner, W. S., & Fex, G., et al. (2000). 2,3,7,8-Tetrachlorodibenzo-p-dioxin increases serum and kidney retinoic acid levels and kidney retinol esterification in the rat. Toxicology and Applied Pharmacology, 169, 121–131. Nuutinen, M. (1972). Hemoglobiibin ma¨a¨ritys. In A. Hyva¨rinen, J. Jannes, E. Nikkila¨, N.-E. Saris, & P. Vuopio (Eds.), Kliiniset laboratoriotutkimukset (pp. 445–449). Helsinki, Finland: WSOY. Nyman, M., Raunio, H., & Pelkonen, O. (2000). Expression and inducibility of members in the cytochrome P4501 (CYP1) family in ringed and grey seals from polluted and less polluted waters. Environmental Toxicology and Pharmacology, 8, 217–225. Nyman, M., Raunio, H., Taavitsainen, P., & Pelkonen, O. (2001). Characterisation of xenobiotic-metabolising cytochrome P450 (CYP) forms in ringed and grey seals. Comparative Biochemistry and Physiology—Part C. Toxicology and pharmacology, 128, 99–112. Nyman, M., Koistinen, J., Fant, M. L., Vartiainen, T., & Helle, E. (2002). Current levels of DDT, PCBs and trace elements in the Baltic ringed seals (Phoca hispida baltica) and grey seals (Halichoerus grypus). Environmental Pollution. 119, 399–412. Olsson, M., Andersson, O¨, Bergman, A˚., Blomkvist, G., Frank, A., & Rappe, C. (2002). Contaminants and diseases in seals from Swedish waters. Ambio, 21, 561–562. Olsson, M., Karlsson, B., & Ahnland, E. (1994). Diseases and environmental contaminants in seals from the Baltic and the Swedish West Coast. Science of the Total Environment, 154, 217–227. O’Shea, T. (1999). Environmental contaminants and marine mammals. In J. E. Reynolds III, & S. A. Rommel (Eds.), Biology of marine mammals (pp. 485–536). Washington, USA: Smithsonian Institution Press. Peakall, D. (1992). Animal biomarkers as pollution indicators. London: Chapman & Hall. Read, A. J. (1990). Estimation of body condition in harbour porpoises (Phocaena phocaena). Canadian Journal of Zoology, 68, 1962–1966. Reijnders, P. J. H. (1988). Ecotoxicological perspectives in marine mammalogy: research principles and goals for a conservation policy. Marine Mammal Science, 4, 91–102. Ross, P., De Swart, R., Visser, I. K.G, Vedder, L. J., Murk, W., & Bowen, W. D., et al. (1994). Relative immunocompetence of the newborn harbour seal, Phoca vitulina. Veterinary Immunology and immunopathology, 42, 331–348. Safe, S. (1994). Polychlorinated biphenyls (PCBs) and polybrominated biphenyls (PBBs): Biochemistry, toxicology, and mechanisms of action. CRC Crit. Rev. Toxicol, 13, 319–393. Schumacher, U., Heidemann, G., Skı´rnisson, K., Schumacher, W., & Pickering, R. M. (1995). Impact of captivity and contamination level on blood parameters of harbour seals (Phoca vitulina). Comparative Biochemistry and Physiology, 112A, 455–462. Schweigert, F. J., Stobo, W. T., & Zucker, H. (1987). Vitamin A status in the grey seal (Halichoerus grypus) on Sable Island. Internat. J. Vit. Nutr. Res., 57, 239–245. Schweigert, F. J., Ryder, O. A., Rambeck, W. A., & Zucker, H. (1990a). The majority of vitamin A is transported as retinyl esters in the blood of most carnivores. Comparative Biochemistry and Physiology, 95A, 573–578. Schweigert, F. J., Stobo, W. T., & Zucker, H. (1990b). Vitamin E and fatty acids in the grey seal (Halichoerus grypus). Journal of Comparative Physiology, B159, 649–654.

M. Nyman et al. / Marine Environmental Research 55 (2003) 73–99

99

Slim, R., Toborek, M., Robertson, L. W., & Hennig, B. (1999). Antioxidant protection against PCBmediated endothelial cell activation. Toxicol. Sci., 52, 232–239. Spear, P. A., Garcin, H., & Narbonne, J. F. (1988). Increased retinoic acid metabolism following 3,30 ,4,40 5,50 -hexacbromobiphenyl injection. Canadian Journal of Physiology and Pharmacology, 66, 1181–1186. Stacey, N. H., & Kappus, H. (1982). Cellular toxicity and lipid peroxidation in response to mercury. Toxicology and Applied Pharmacology, 63, 29–35. Stalling, D. L., Norstrom, R. J., Smith, L. M., & Simon, M. (1985). Patterns of PCDD, PCDF, and PCB contamination in Great Lakes fish and birds and their characterization by principal components analysis. Chemosphere, 14, 627–643. Stegeman, J. J., Brouwer, M., Di Giulio, R. T., Fo¨rlin, L., Fowler, B. A., & Sanders, B. M., et al. (1992). Molecular responses to environmental contamination: enzyme and protein systems as indicators of chemical exposure and effect. In R. J. Hugget, R. A. Kimerle, P. M. Mehrle, & H. L. Bergman (Eds.), Biomarkers, biochemical, physiological, and histological markers of anthropogenic stress (pp. 235–335). USA: Lewis Publishers. Stegeman, J. J., & Hahn, M. E. (1994). Biochemistry and molecular biology of monooxygenases: current perspectives on forms, functions, and regulation of cytochrome P450 in aquatic species. In D. C. Malins, & G. K. Ostrander (Eds.), Aquatic toxicology: molecular, biochemical and cellular perspectives (pp. 87– 206). USA: Lewis Publishers. Storr-Hansen, E., Spliid, H., & Boon, J. P. (1995). Patterns of chlorinated biphenyl congeners in harbor seals (Phoca vitulina) and in their food: statistical analysis. Archives of Environmental Contamination and Toxicology, 28, 48–54. So¨derberg, S. (1974). O¨stersjo¨ns sa¨lpopulationer. SNV PM, 419, 3–29. Tanabe, S., Watanabe, S., Kan, H., & Tatsukawa, R. (1988). Capacity and mode of PCB metabolism in small cetaceans. Marine Mammal Science, 4, 103–124. Tanabe, S., Iwata, H., & Tatsukawa, R. (1994). Global contamination by persistent organochlorines and their ecotoxicological impact on marine mammals. Science of the Total Environment, 154, 163–177. Thirman, M. J., Albrecht, J. H., Krueger, M. A., Erickson, R. R., Cherwitz, D. L., & Park, S. S., et al. (1994). Induction of cytochrome CYP1A1 and formation of toxic metabolites of benzo (a) pyrene by rat aorta: a possible role in atherogenesis. Proceedings of the National Academy of Science, 91, 5397– 5401. Tormosov, D. D., & Rezvov, G. V. (1978). Information on the distribution, number and feeding habits of ringed and grey seals in the Gulfs of Finland and Riga in the Baltic Sea. Finnish Game Research, 37, 14–21. Troisi, G. M., & Mason, C. F. (1997). Cytochromes P450, P420 & mixed-function oxidases as biomarkers of polychlorinated biphenyl (PCB) exposure in harbour seals (Phoca vitulina). Chemosphere, 35, 1933– 1946. Van den Berg, M., Birnbaum, L., Bosveld, A. T. C., Brunstro¨m, B., Cook, P., & Feeley, M., et al. (1998). Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for human and wildlife. Environmental Health Perspectives, 106, 775–792. Van Veld, P., Vogelbein, K., Cochran, M. K., Goksøyr, A., & Stegeman, J. J. (1997). Route-specific cellular expression of cytochrome P4501A (CYP1A) in fish (Fundulus heteroclitus) following exposure to aqueous and dietary benzo[a]pyrene. Toxicology and Applied Pharmacology, 142, 348–359. Wold, S., Albano, C., Dunn, W.J., Edlund, U., Esbensen, K., Geladi, P., et al., (1984). Multivariate data analysis in chemistry. In Proceedings of the NATO advanced study on chemometrics. Mathematics and statistics in chemistry (pp. 1–79). Cosenza. Italy, D. Reidel Publishing Company, Dordrecht Holland. Wold, S., Esbensen, K., & Geladi, P. (1987). Principal component analysis. Chemometrics and Intelligent. Laboratory Systems, 2, 37–52. Yamamoto, K., Fukuda, N., Shiroi, S., Shiotsuki, Y., Nagata, Y., & Tani, T., et al. (1994). Ameliorative effect of dietary probucol on polychlorinated biphenyls-induced hypercholesterolemia and lipid peroxidation in the rat. Life Science, 54, 1019–1026.