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Keywords:

  • Biotransformation;
  • Tissue distribution;
  • Fasting;
  • Cytochrome P450;
  • Seabird

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

The aim of the present study was to investigate how contaminant exposure and reduced food intake affect tissue distribution and biotransformation of halogenated organic contaminants (HOCs) in Arctic seabirds using herring gull (Larus argentatus) as a model species. Herring gull chicks were exposed for 44 d to cod liver oil containing a typical mixture of contaminants. Following exposure, food intake was reduced for a one-week period in a subgroup of the chicks. Polyclorinated biphenyls, organochlorine pesticides, and brominated flame retardants, as well as a wide range of hydroxy, methyl sulfone, and methoxy compounds were measured in liver, brain, and plasma samples. Additionally, phase I biotransformation enzyme activities and phase I and II messenger ribonucleic acid (mRNA) expression were investigated in the liver, brain, or both. Both contaminant exposure and reduced food intake had an increasing effect on the concentrations of HOCs and their metabolites. The HOC exposure and reduced food intake also led to increased 7-ethoxyresorufin-O-deethylation (EROD) activity, whereas mRNA expression of the biotransformation enzymes increased only following the reduced food intake. Tissue distribution of HOCs and their metabolites was not affected by either contaminant exposure or reduced food intake. In conclusion, the results indicate that biotransformation capacity and formation of HOC metabolites increase during reduced food intake. This finding supports the hypothesis that reduced food intake increases the susceptibility of Arctic animals to the effects of lipophilic HOCs. Environ. Toxicol. Chem. 2013;32:156–164. © 2012 SETAC


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

Arctic seabirds go through seasonal energy-demanding periods when breeding, migrating, and moulting. During these fasting periods energy is mobilized from adipose tissue, leading to significant loss in body mass. Mobilization of lipid stores leads to redistribution of lipophilic halogenated organic contaminants (HOCs) from relatively inactive adipose tissue to blood and vital organs, such as liver and brain 1. Seasonal fasting periods supposedly render Arctic animals more vulnerable to toxic effects of HOCs 2. Several studies have documented increased HOC concentrations in plasma, liver, and/or brain of Arctic animals during energy-demanding periods 3–8. In addition, HOCs in seabirds and fish have been reported to be redistributed to brain tissue more than liver and kidney during fasting 6, 7. When HOCs are redistributed from adipose tissue, they may induce phase I (cytochrome P450 [CYP]) and phase II (e.g., uridine diphosphate-glucuronosyltransferase [UDPGT] and glutathione S-transferase [GST]) enzymes in target tissue (e.g., liver and brain), leading to the formation of hydroxy (OH) and methylsulfone (MeSO2) metabolites of polychlorinated biphenyls (PCBs) 9. Some OH and MeSO2 metabolites are retained in the body, and their toxic potential toward the endocrine system is higher than that of the parent compounds 10, 11. For example, some OH-PCBs show structural similarities to thyroid hormones, and may have a higher affinity to thyroid hormone transport proteins in seabirds than thyroid hormones 12, 13.

The effect of fasting and lipid mobilization on contaminant biotransformation has received little attention in both experimental and wildlife studies. Enhanced biotransformation capacity and increased bioavailability of PCBs and dichlorodiphenyltrichloroethane (DDT) has been reported in experimentally fasted Arctic char (Salvelinus alpinus) and brown-headed cowbirds (Molothrus ater), respectively 7, 14, 15. Furthermore, concentrations of OH metabolites of HOCs were shown to be higher in fasting free-ranging ringed seals (Phoca hispida) compared to nonfasting animals 8. In addition, physiological changes associated with fasting have been shown to influence genes that encode key biotransformation enzymes in fish and rodents 16, 17.

The aim of the present study was to investigate how HOC exposure and reduced food intake affect accumulation, tissue distribution, and biotransformation responses of HOCs in Arctic seabirds using herring gull (chicks) as a model species.

MATERIALS AND METHODS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

Model species

The herring gull (Larus argentatus) breeds in several colonies along the coast of Norway and Bjørnøya (Svalbard, Norway) 18. This species is a generalist feeder, eating a variety of food, including other seabird species and fish 18. Due to its high trophic position in the food web, the herring gull accumulates high levels of HOCs 19.

Experimental design

The experimental design of the present study is described in detail by Hegseth et al. 20. Briefly, 40 herring gull chicks at the age of 2 to 5 d were collected at Sommarøya (Tromsø, Norway; 69.62°N, 18.04°E) at the end of June 2008 and transported to outdoor cage facilities located at Kårvika Marine Research Station. The chicks were fed herring (Clupea harengus) fillets and they received one food supplement tablet every fourth day (Fish eater tablets, Mazuri Zoo Foods). On day 21, the chicks were divided into two groups, control and exposed. The control and exposed groups were given clean filtered pharmaceutical cod (Gadus morhua) liver oil and contaminated cod liver oil/waste from the production of pharmaceutical oil, respectively. The filtered oil is produced by adding 1% (w/w) clay and 0.5% (w/w) active charcoal to cod liver oil to improve the visual quality of the oil and remove planar dioxin-like substances, respectively. The clay and charcoal are then removed by filtration. The filtered oil is distilled with increased temperature and low pressure. Distillation gives two products: clean cod liver oil (pharmaceutical product) and contaminated cod liver oil (waste containing ∼ 1,000 times more contaminants than the clean oil; see Supplemental Data, Table S1). The control and exposed groups were given daily 12 ml of clean cod liver oil and contaminated cod liver oil, respectively. After 45 d, blood was collected from half of each group (hereafter called nonfasted chicks) before the birds were euthanized and sacrificed for liver and brain tissue sampling. For the remaining individuals (hereafter called fasted chicks) of the exposed and control groups, food intake was reduced by 70% for one week, after which the chicks were sacrificed. Samples of liver and brain tissue were frozen in liquid nitrogen immediately after sampling and stored at −80°C. The samples for contaminant analysis were stored at −20°C until analysis. Due to self-induced accidents leading to injuries, three chicks were sacrificed before the end of the experiment. Further analyses were not conducted on these individuals. The experiment was approved by the Experimental Animal Welfare Committee of Norway (S-2008/69304).

Chemical analysis

Liver, plasma, and brain samples of the herring gull chicks were analyzed for a wide range of HOCs including legacy and emerging chlorinated and brominated compounds and their metabolites. The analytical procedures for determination of HOCs and their metabolites have been described elsewhere 21. Briefly, 2 ml of plasma sample was acidified, then [13C]-labeled internal standard mixture and a surrogate standard (3-MeSO2-4-Me-2′,3′,4′,5,5′-PentaCB) for the MeSO2-PCBs were added, and extraction was performed using a solid-phase extraction column (OASIS HLB, Waters). Lipid content (%, w/w) was determined enzymatically for plasma samples, and total amounts of lipids were calculated according to the following equation: TL = 1.677 × (TC – FC) + FC + TG + PL where TL = total lipids, TC = total cholesterol, FC = cholesterol, TG = triacylglycerol, and PL = phospholipids 22.

Liver and brain samples (2 g) were homogenized with dry Na2SO4, packed into columns, and extracted with acetone:cyclohexane (1:3,v/v). The extract was evaporated, and the lipid content (%, w/w) was determined. The extract was further cleaned up using gel permeation chromatography for lipid removal. All plasma, liver, and brain extracts were fractionated using Florisil. The collected Florisil fractions included fraction 1––dichloromethane:n-hexane (1:9,v/v) containing neutral compounds (PCBs, pesticides, polybrominated diphenyl ethers); fraction 2––acetone:n-hexane (1:9,v/v) containing MeSO2 compounds; and fraction 3––methanol:dichloromethane (1:5,v/v) containing OH compounds. The third fraction was gently evaporated to dryness and derivatized using diazomethane. A final clean-up of fractions 2 and 3 was performed on a solid-phase extraction column of 25% sulfuric acid silica with neutral silica on top, and extraction was performed with dichloromethane. All plasma, liver, and brain extracts were evaporated and transferred to gas chromatography (GC) vials with octachloronaphthalene added as volume corrector. The analysis of the compounds was performed on an Agilent 7890AGC with a 5975C mass spectrometer (MS) operated in single-ion monitoring mode. The MS was run in negative ion chemical ionization mode for the metabolites and pesticides and in electron impact mode for PCBs and DDTs. Polybrominated diphenyl ethers (PBDEs) were analyzed in the electron impact mode on an Autospec GC/high-resolution MS system (Waters). All sample fractions were analyzed separately. Fraction 1 (containing PCBs, pesticides, PBDEs, and MeO-PBDEs) was injected three times to the GC/MS or GC/high-resolution MS analysis instrument grouped into subgroups: PCBs/DDTs, pesticides/MeOBDEs, and PBDEs. Fraction 2 was analyzed for MeSO2-PCBs/dichlorodiphenyldichloroethylene (DDE), and fraction 3 was analyzed for the derivatized OH-PCBs, pentachlorophenol (PCP), 4-OH-heptachlorostyrene (4-OH-HpCS), tetrabromobisphenol-A (TBBP-A), and OH-PBDEs.

Polychlorinated biphenyls analyzed included the following congeners and metabolites: CB18, 28, 31, 33, 37, 47/49, 52, 99, 101, 105, 118, 123, 128, 138, 141, 149, 153, 156, 157, 167, 170, 180, 183, 187, 189, 194, 4-OH-CB120, 4-OH-CB107, 3-OH-CB153, 4-OH-CB146, 3-OH-CB138, 4-OH-CB130, 4-OH-CB163, 4-OH-CB187, 4-OH-CB172, and 4′-OH-CB193, as well as 3-MeSO2-CB52, 4-MeSO2-CB52, 3-MeSO2-CB49, 4-MeSO2-CB49, 3-MeSO2-CB91, 4-MeSO2-CB91, 3-MeSO2-CB101, 4-MeSO2-CB101, 3-MeSO2-CB87, 3-MeSO2-CB110, 4-MeSO2-CB110, 3-MeSO2-CB149, 4-MeSO2-CB149, 3-MeSO2-CB132, 4-MeSO2-CB132, 3-MeSO2-CB141, 4-MeSO2-CB141, 3-MeSO2-CB174, and 4-MeSO2-CB174. Quantified organochlorine pesticides and their metabolites included o,p′-DDT, o,p′-DDE, and o,p′-dichlorodiphenyldichlorethane (DDD), p,p′-DDT, p,p′-DDE, and p,p′-DDD, heptachlor, oxychlordane, trans-chlordane, cis-chlordane, trans-nonachlor, cis-nonachlor, PCP, 4-OH-HpCS, and MeSO2-DDE. The native [12C] PCBs and organochlorines (OCs) were purchased from Ultra Scientific Europe, and the isotopically labeled [13C] PCBs and OCs used as internal standards were purchased from Cambridge Isotope Laboratory. The native OH-PCBs and MeSO2-PCBs as well as PCP, 4-OH-HpCS, and MeSO2-DDE and the isotopically labeled OH-PCBs and PCP were kindly donated by the U.S. Centers for Disease Control and Prevention. In addition, a surrogate internal standard for the analysis of MeSO2-PCBs was used, 3-MeSO2-4-Me-2′,3′,4′,5,5′-PentaCB, which was also purchased from Cambridge Isotope Laboratory. The following brominated compounds were analyzed: BDE 28, 47, 66, 71/49, 77, 85, 99, 100, 119, 138, 153, 154, 183, 206, 207, and 209, 2,4,6,-tribromoanisole (TBA), TBBP-A, 2-OH-BDE68, 6-OH-BDE47/75, 5-OH-BDE47, 4-OH-BDE49, 5-OH-BDE100, 4-OH-BDE103, 5-OH-BDE99 4-OH-BDE101, 2-methoxy (MeO)-BDE68, 6-MeO-BDE47/2-MeO-BDE75, 5-MeO-BDE47, 4-MeO-BDE49, 5-MeO-BDE100, 4-MeO-BDE103, 5-MeO-BDE99, and 4-MeO-BDE101. Polychlorinated biphenyls 18, 33, and 37, o,p′-DDD, and TBA were not quantified in the brain samples, whereas 4-OH-HpCS was not quantified in the plasma samples. The native [12C] PBDEs, TBA, and TBBPA, MeO-BDEs, and isotopically labeled [13C] standard PBDEs were purchased from Wellington Laboratories.

All chemical analyses followed international requirements for quality assurance and control, for example, recommendations of the Arctic Monitoring and Assessment Program. In addition, the Norwegian Institute for Air Research participates in the Arctic Monitoring and Assessment Ring Test Program for persistent organic pollutants in human serum. Quality assurance and control of the sample preparation and analysis were ensured through the use of mass labeled internal standards for the PCBs, organochlorine pesticides, brominated compounds, and OH-PCBs, whereas a surrogate standard was used for the MeSO2 compounds in addition to sample preparation and analysis of certified reference materials and laboratory blanks. For each batch of 20 plasma samples, one standard reference (U.S. National Institute of Standards and Technology [NIST] human serum 1589a) and three blank samples were prepared, whereas for each batch of 10 brain and liver samples one standard reference (NIST cod oil 1588b) and one blank were prepared. The average recovery, determined from the added mass labeled internal standards ([13C]PCBs, pesticides, PBDEs, OH-PCBs) or surrogate standard (3-MeSO2-4-Me-2′,3′,4′,5,5′-PentaCB) for the MeSO2-PCBs was 76 ± 10% (standard deviation) for PCBs, 93 ± 13% for DDTs, 55 ± 15% for chlordanes (CHLs), 58 ± 18% for PBDEs, 60 ± 18% for OH compounds, 55 ± 18% for MeSO2 compounds, and 60 ± 16% for MeO-BDEs.

Gene expression

Expression of CYP1A4, CYP1A5, CYP3A, GST, and UDPGT mRNA was analyzed in the liver samples of the herring gull chicks. In addition, we analyzed CYP1A4, CYP1A5, and GST in the brain of the herring gull chicks. A detailed method for mRNA expression analysis has been described elsewhere 23.

Enzyme activity

Hepatic phase I enzyme activities were measured in liver microsomes of the herring gull chicks using methods described previously 24. Measurements of liver microsomal enzymes were based on the formation of resorufin from 7-ethoxyresorufin (EROD), 7-pentoxyresorufin (PROD), 7-benzyloxyresorufin (BROD), and 7-methoxyresorufin (MROD) in the presence of nicotinamide adenine dinucleotide phosphate. Fluorescence intensity was measured for 20 min in triplicate using a 96-well plate to determine enzyme activities. The excitation and emission wavelengths were 530 and 590 nm, respectively.

Data analysis

Statistical analysis was carried out using R Version 2.11.1 25. A compound was considered as present and was included in sum concentrations (on a wet-wt basis) of PCBs, PBDEs, DDTs, CHLs, OH-PCBs, and MeSO2 compounds if it was detected in >70% of the samples of a group (see Supplemental Data, Table S2). Samples with concentrations below the limit of detection were replaced by half of the limit of detection when included in the sum.

Linear mixed-effect models in the lme4 package 26 were used to analyze the effect of contaminant exposure and reduced food intake and lipid content on concentrations (ng/g wet wt) and tissue distribution of HOCs. Contaminant exposure (exposed and control), reduced food intake (fasted and nonfasted), and tissue (plasma, liver, brain) and lipid content were applied as fixed effects, whereas individual was included as a random effect. To investigate whether tissue distribution and lipid content were affected by the treatments, we included the following three-way interactions in the initial models: (1) contaminant exposure × reduced food intake × tissue; and (2) contaminant exposure × reduced food intake × lipid content. All mixed-effect models were fitted with random intercepts only. The most parsimonious models were manually selected using second-order Akaike's Information Criterion (AICc) from a pool of models, which all included the experimental treatments as predictor variables. The best model had the lowest AICc value. Maximum likelihood fitted models were used during model selection, whereas a restricted maximum likelihood fitted model was used for parameter estimation. Because concentrations of neutral HOCs (ΣPCBs, ΣDDTs, ΣCHLs, and ΣPBDEs) were strongly correlated (r > 0.93), sums of neutral HOCs (ΣHOCs) were used in the mixed models instead of individual HOC groups. The effect of contaminant exposure and reduced food intake on concentrations and tissue distribution of ΣOH and ΣMeSO2 metabolites, and lipid content was also tested using linear mixed-effect models as described above. The effect of reduced food intake on concentrations and tissue distribution of ΣOH and ΣMeSO2 metabolites was only tested in the exposed group, as these metabolites were not detected in all body compartments of the control (nonexposed) group. For the same reason, the effect of contaminant exposure on concentrations (and tissue distribution) was tested in brain and plasma for ΣOH metabolites, and in liver for ΣMeSO2 metabolites, respectively.

The effect of reduced food intake and contaminant exposure on mRNA expression (ng mRNA/µg total RNA) and activities of biotransformation enzymes (pmol/min/mg protein) was tested separately for liver and brain using linear models. Initial models included exposure, reduced food intake, and their interaction as response variables. The most parsimonious models were selected using AICc as described above.

As the present study was based on planned comparisons with well-defined control groups, we used the treatment contrast for all the models. Exposed chicks were compared to control chicks, and fasted chicks to the nonfasted ones, and plasma was set as a reference level for tissue comparisons. The null hypothesis was rejected at α = 0.05. Results based on parameter estimates (β) with 95% confidence intervals (CI) are given in the text. Diagnostic plots of residuals were used to verify that the model assumptions were met (most importantly constant variance between residuals).

RESULTS AND DISCUSSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

To our knowledge, no previous studies have investigated the effect of reduced food intake on biotransformation of HOCs in seabirds. Also, this is the first study to investigate the effect of reduced food intake on tissue distribution of metabolites of HOCs in seabirds.

Concentrations of HOCs

Concentrations of ΣHOCs in liver, plasma, and brain of the herring gull chicks were 21 (CI 15, 29) times higher in the exposed group compared to the control group (Fig. 1). The herring gull chicks were exposed to a total of approximately 16,700 ng/g body weight of ΣHOCs through the contaminated oil (ΣPCBs, ΣDDTs, ΣCHLs, ΣPBDEs; for congener-specific composition see Supplemental Data, Table S1). Halogenated organic contaminant concentrations in liver and plasma of the exposed herring gull chicks were similar to HOC concentrations reported in free-ranging glaucous gulls (Larus hyperboreus) from Bjørnøya, Svalbard 27, 28. Trace levels of HOCs observed in the control birds from the present study (Fig. 1) corresponded to a very low exposure of the control birds (total intake of ΣHOCs through the clean oil 18 ng/g body wt). The hepatic HOC concentrations in the control birds were similar to those reported in little auks (Alle alle) and Brünnich's guillemots (Uria lomvia) from northern Baffin Bay, Greenland 29. The relevance of the HOC exposure in the herring gull chicks used in the present experiment has been further discussed by Hegseth et al. 20.

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Figure 1. Concentrations of halogenated organic contaminants (ΣHOC) (polychlorinated biphenyls ΣPCB, organochlorine pesticides and polybrominated diphenyl ethers [ΣPBDE]), MeSO2 compounds (Σ MeSO2-PCBs, MeSO2-dichlorodiphenyldichloroethylene [DDE]), and hydroxy (OH)-PCBs in plasma, liver, and brain in control (C), exposed (E), and fasted (F) herring gull chicks. MeSO2 compounds were not detected in the plasma or brain of the control chicks. OH-PCBs were not detected in the liver of the control chicks.

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Neutral halogenated organic contaminant concentrations in liver, brain, and plasma were 1.8 (CI 1.2, 2.4) times higher in the fasted control chicks compared to the nonfasted control chicks (Fig. 1). The reduced food intake had a stronger effect on ΣHOC levels in the exposed chicks, among which ΣHOC concentrations were 2.9 (CI 2.1, 3.6) times higher in the chicks subjected to the reduced food intake compared to the nonfasted ones. During the period of reduced food intake, the animals lost on average 11% of their body weight (range, 7.0–18%), and the weight loss was similar in the exposed and the control group (β = −1.0 percentage unit [CI -4.2, 2.1]). As approximately 10% loss of body weight in male gulls has been related to loss of body fat 30, increased concentrations of ΣHOCs in the liver, plasma, and brain are most likely explained by mobilization of HOCs from the adipose tissue. Redistribution of HOCs from adipose tissue to blood and vital organs has been previously reported in seabirds, mammals, and fish. For example, increased plasma concentrations of HOCs were found following fasting in common eiders (Somateria mollissima) 4, harp seals (Phoca groendlica) 3, and northern elephant seals (Mirounga angustirostris) 31. Body condition (or body mass) has also been inversely associated with ΣPCB concentrations in the brain of black-legged kittiwakes (Rissa tridactyla) 6, liver and brain of glaucous gulls 5 and Arctic char 7, and liver of ringed seals 8. Lipid contents in plasma or liver were not affected by the experimental treatments or their interactions with tissue although lipid content in brain was slightly higher (0.46 percent units [CI 0.02, 0.90]) in the exposed chicks compared to the control chicks. Furthermore, the interaction terms between lipid content and the experimental treatments were excluded from the most parsimonious models explaining HOC concentrations (measured as wet wt), indicating that the increased concentrations of HOCs in the nonfasted herring gull chicks were not related to changes in lipid content due to the experimental treatments.

Concentrations of ΣHOCs in the herring gulls were 2.3 (CI 1.8, 5.5) times higher in liver compared to plasma, whereas ΣHOC concentrations were not significantly higher in brain compared to plasma (−40% [CI −57, 170]; Fig. 1). The interaction terms between tissue and the experimental treatments were excluded from the most parsimonious models, meaning that liver to plasma and brain to plasma concentration ratios of ΣHOCs were not influenced by contaminant exposure or reduced food intake. This finding is in contrast to previous reports on Arctic fish and seabirds. For example, in free-ranging black-legged kittiwakes, brain to liver ratios of PCBs (measured on wet-wt basis) were higher in fasting birds compared to nonfasting birds 6. Furthermore, in Arctic char dosed with Arochlor 1254, PCB concentrations increased more in brain tissue compared to liver tissue following fasting 7. The discrepancy in the effect of fasting on tissue distribution of ΣHOCs in herring gull chicks compared to black-legged kittiwakes and Arctic char could be linked to differences in the length of the fasting period. The reduced food intake in the herring gull chicks lasted only 7 d, producing an approximately 11% loss of body mass, whereas the black-legged kittiwakes and Arctic char followed their natural fasting period (several months) and lost on average 20% of their body mass 6, 7.

Concentrations of OH, MeO, and MeSO2 compounds

Nine different OH-PCBs were present in the herring gull chicks, whereas no OH-PBDEs, TBBPA, pentachlorophenol, 4-OH-HpCS, TBA, or MeO-PBDEs were detected (see Supplemental Data, Table S2). In total, 17 different MeSO2 metabolites of PCBs and DDE were present in plasma, liver, and/or brain samples of the exposed chicks, whereas MeSO2 metabolites were present only in liver samples of the control birds (see Supplemental Data, Table S2). Concentrations of ΣOH-PCBs and ΣMeSO2-compounds were 28 (CI 19, 41) and 21 (CI 13, 35) times higher, respectively, in the exposed group compared to the control group (Fig. 1). Reduced food intake increased the concentrations of ΣOH-PCBs and ΣMeSO2-compounds in liver, brain and plasma samples of the exposed chicks by 2.7 (CI 1.8, 4.1) and 2.6 (CI 1.9, 3.8) times, respectively (Fig. 1). An increase in plasma concentrations of OH-compounds following reduced food intake has previously been reported in free-ranging ringed seals 8.

The highest concentrations of ΣOH-PCBs were detected in plasma samples (Fig. 1). In comparison, brain concentrations of ΣOH-PCBs were only 3.2% (CI 2.2, 4.6) of those measured in plasma, whereas liver OH-PCBs were 13% (CI 8.9, 19) compared to concentrations in plasma. Concentrations of ΣMeSO2 compounds were 7.2 (CI 5.4, 9.7) and 2.9 (CI 2.3, 4.2) times higher in liver and brain, respectively, compared to plasma. The tissue distribution of ΣOH-PCBs and ΣMeSO2 compounds in the herring gull chicks is in accordance with previous studies on seals, polar bears (Ursus maritimus), and other seabird species 32–34. Exclusion of the interaction term of tissue and reduced food intake indicated that reduced food intake did not influence tissue distribution of HOC metabolites, that is, the liver to plasma and brain to plasma concentration ratios of ΣOH-PCBs or ΣMeSO2 compounds.

Uptake of HOCs, OH, and MeSO2 compounds

The PCB profile differed between the contaminated oil and the tissue samples of the herring gull chicks (Fig. 2). For example, the relative contribution of group IV and V PCBs (CB47/49, 52, 101, 141, and 149) to ΣPCBs was several times higher in the contaminated fish oil compared to the exposed chicks (Fig. 2), suggesting that these PCBs were readily biotransformed in the exposed herring gull chicks. This is in accordance with previous findings on PCB biotransformation in seabirds 35, 36 and American kestrels 37. Because group IV and V PCBs are precursors of MeSO2-PCB 9, MeSO2-PCBs detected in the exposed herring gull may result from biotransformation of group IV and V PCBs. However, MeSO2-PCBs in the exposed herring gull chicks may also originate from diet, because the intake of ΣMeSO2-PCBs from the contaminated oil was relatively high (63 ng/g body wt; for congener-specific composition, see Supplemental Data, Table S1). In contrast, the total intake of ΣOH-PCBs during the exposure period was minimal (0.10 ng/g body wt; for congener-specific composition, see Supplemental Data, Table S1). Exposed herring gull chicks with detected OH-PCBs are thus suggested to originate from biotransformation processes.

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Figure 2. Composition of polychlorinated biphenyls (PCBs), dichlorodiphenyltrichloroethane (DDT) compounds, chlordanes (CHLs), and polybrominated diphenyl ethers (PBDEs) in contaminated fish oil, and liver of the exposed and exposed and fasted birds. The PCB groups were divided based on the pattern of Cl substitution 48. CBI = no vicinal hydrogen (H)-atoms (CB153,167, 180, 183, 187, 189, 194); CBII = vicinal H-atoms only in ortho and meta positions in combination with = 2 ortho-Cl (CB99, 128, 138, 170); CBIII = vicinal ortho-meta H-atoms and = 1 ortho-Cl (CB28, 105, 118, 123, 156, 157); CBIV–V = vicinal meta-para H-atoms (CB47/49, 52, 101, 141, 149).

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The ΣDDT in the contaminated oil consisted of 26% p,p′-DDT and 63% p,p′-DDE, whereas in the herring gulls p,p′DDE comprised almost 100% of ΣDDT (Fig. 2). This indicates that p,p′-DDT was efficiently metabolized to p,p′-DDE in the herring gull chicks, which is in accordance with previous reports in seabirds 35, 38. Differences in chlordane patterns between the herring gull chicks and the contaminated oil suggest that cis-nonachlor and cis-chlordane were biotransformed to oxychlordane in the herring gull chicks (Fig. 2). Furthermore, the results of the present study suggest that food restriction increased the biotransformation of nonachlordanes to oxychlordane (Fig. 2). This corroborates a previous report suggesting efficient CHL biotransformation in seabirds including guillemots and black-legged kittiwakes 38.

The composition of ΣPBDE was similar in the contaminated oil and the herring gull chick livers, indicating similar accumulation of these compounds from diet in chicks. This is in contrast to previous findings on American kestrels showing different retention factors for individual PBDEs 37. MeO-PBDEs were not detected in the herring gull chicks, although the dietary intake was relatively high (total intake of ΣMeO-PBDEs through the contaminated oil 882 ng/g body wt). This suggests that either the absorption efficiency of MeO-PBDEs is very low or that these compounds are readily excreted via bile, urine, or feces. The results of the present study are similar to results found in field studies suggesting that MeO-PBDEs do not accumulate in seabirds at higher concentrations than in their prey 39.

Congener-specific patterns of ΣOH- or ΣMeSO2-PCBs

The ΣOH-PCBs consisted mainly of 4-OH-CB187 followed by 4-OH-CB146 (Fig. 3). These metabolites have also been reported to be the main OH-PCBs in free-ranging adult glaucous gulls, as well as chicks of northern fulmars and black-legged kittiwakes 28, 33. The 4-OH-CB187 may originate from hydroxylation of CB183 and/or CB187 9. The present study suggests that the formation of 4-OH-CB187 in the herring gull chicks is most likely a result of direct insertion of an OH-group in CB187, as the relative concentration of CB187 to ΣPCBs among the exposed herring gulls was significantly lower in the fasted chicks compared to the nonfasted chicks (p < 0.004, Wilcoxon rank sum test). No such difference was observed for relative concentrations of CB183.

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Figure 3. Concentrations of hydroxy polychlorinated biphenyls (OH-PCBs) in plasma (A) and MeSO2-PCBs in liver (B) in exposed (black circles), and exposed and fasted (white circles) herring gull chicks. Mean values are given with 95% confidence intervals.

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The metabolites 3-MeSO2-CB101, 4-MeSO2-CB101, 4-MeSO2-CB110, and 4-MeSO2-CB149 dominated the MeSO2-PCB pattern in the exposed herring gull chicks (Fig. 3). These metabolites may either be formed from their corresponding PCB precursors (i.e., CB101, CB110, and CB149) 9 or originate from the contaminated fish oil, as discussed earlier. The hypothesis that MeSO2-PCBs detected in the herring gull chicks originate from bioaccumulation is supported by the fact that the pattern of MeSO2-PCBs in the herring gulls was similar to that in the contaminated fish oil (Fig. 3, Supplemental Data, Table S1). The pattern of MeSO2-PCBs in the herring gull chicks was different from that of other seabird species 28, 33. For example, in adult glaucous gulls the main MeSO2-PCBs were 3-MeSO2-CB132 and 3- and 4-MeSO2-CB49 28, whereas 3-MeSO2-CB49, 4-MeSO2-CB101, and 3-MeSO2-CB87 were detected at highest concentrations in northern fulmar chicks, and 4-MeSO2-CB101 and 3-MeSO2-CB110 in black-legged kittiwake chicks 33. The different MeSO2-PCB profiles in different seabird species are likely due to species-specific differences in dietary accumulation, but may also be influenced by species-specific sulfone clearance and formation capacity 9.

Biotransformation enzymes

CYP1A4, CYP1A5, CYP3A, UDPGT, and GST mRNA were analyzed in the liver samples, whereas only CYP1A4, CYP1A5, and UDPGT were analyzed in the brain samples (Fig. 4). No significant differences were found in mRNA expression of the biotransformation enzymes between the control and the exposed herring gull chicks (Fig. 4). This was unexpected because concentrations of HOC metabolites were higher in the exposed herring gull chicks compared to the control birds, indicating that biotransformation enzymes had been active in the exposed chicks. For example, 4-OH-CB187 was detected at 33 (CI 25, 43) times higher concentrations in the exposed group compared to the control birds. The suggested precursor, CB187, is a bulky molecule, and thus a potential substrate of CYP3A 40. We would thus have expected that the mRNA levels of CYP3A would be higher in the exposed birds compared to the control birds. However, the HOC metabolites retained in the chicks are relatively persistent, and the concentrations of these metabolites reflect physiological processes over their whole life (∼ 2 months). In contrast, the measured mRNA expressions reflect transcript expression of biotransformation enzyme genes only during a transient time window (the moment of sampling). Therefore, the sampling for mRNA analysis might have missed this critical window of expression of these genes. Furthermore, HOCs having antagonistic effects on biotransformation enzymes 41 may have accumulated to a higher extent toward the end of the exposure period, which may explain the discrepancy between the results of HOC metabolite concentrations and mRNA expression. However, HOC metabolites detected in the herring gulls may also have been formed by CYPs, which were not investigated in the present study, for example, CYP2B-like enzymes.

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Figure 4. Messenger ribonucleic acid (mRNA) expression of CYP1A4, CYP1A5, and uridine diphosphate-glucuronosyltransferase (UDPGT), CYP3A, glutathione S-transferase (GST), and UDPGT in liver and/or brain, and activities of hepatic phase I enzymes measured as 7-ethoxyresorufin-O-deethylation (EROD), 7-methoxyresorufin (MROD), and 7-pentoxyresorufin (PROD) in control (C), exposed (E), and fasted (F) herring gull chicks.

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There was EROD, MROD, and PROD activity detected in liver microsomes of the herring gull chicks (Fig. 4). Results from previous avian or mammalian studies suggest that EROD, MROD, and PROD are catalyzed by CYP1A4, CYP1A5, and a wide range of other CYPs 42, 43. However, a recent study on seabirds suggests that catalytic activity and substrate specificities of CYP enzymes is species-specific between northern fulmars and black-legged kittiwakes 33. In the present study, BROD activity was not detected in the herring gull chicks, suggesting that BROD is a poor substrate for assessing CYP activity in herring gulls. Similarly, BROD activity was close to the detection limit in northern fulmar and black-legged kittiwake chicks 33. In contrast, BROD was suggested to be a suitable substrate for assessing CYP1A4 activity in common cormorants (Phalacrocorax carbo) 42.

The EROD activity was 2.1 times higher (CI 1.0, 4.4) in the herring gull chicks exposed to contaminated oil compared to the control ones, although mRNA expression of CYP enzymes were not affected by contaminant exposure (Fig. 4). The inconsistent results between mRNA expressions of CYPs and EROD activity suggest that there may be a time lag between mRNA expression of enzymes and detectible changes in enzyme activities; the half-life of mRNA is probably much lower compared to CYP enzymes. Contaminant exposure may also lead to posttranslational modification of proteins rather than transcriptional variation in genes 44. Alternatively, EROD formation of resorufin from 7-ethoxyresorufin may be catalyzed by other CYPs not identified in the present study. Contaminant exposure did not affect MROD or PROD activity (Fig. 4).

Reduced food intake increased liver CYP1A4, CYP1A5, and CYP3A mRNA expression by 47% (CI 1, 113), 132% (CI 57, 244), and 79% (CI 23, 159), respectively, in both the control and exposed groups (Fig. 4). Furthermore, reduced food intake led to three (CI 1.5, 6.2) times higher EROD activity (Fig. 4). The induction of CYP mRNA expressions and EROD activity following reduced food intake suggests that biotransformation enzymes may be induced by physiological effects of caloric restriction. This has previously been observed in experimental studies on rodents. For example, reduced food intake increased the expression of genes that encode key biotransformation enzymes in mouse 16 and hepatic CYP3A in rat 17. Similar to phase I enzymes, reduced food intake induced hepatic mRNA expression of GST in the herring gull chicks by 70% (CI 27, 128; Fig. 4). This finding is in contrast to previous studies showing that liver GST activity remained unchanged or decreased during food deprivation in rat and fish, respectively 45, 46. In the present study, reduced food intake also increased liver UDPGT mRNA expression by 99% (CI 5, 277) in the exposed herring gulls but not in the control group (Fig. 4). This finding indicates that the redistribution of HOCs to the liver induces hepatic UDPGT mRNA expression. A PCB-mediated induction of hepatic UDPGT has also been observed in Japanese quail (Coturnix japonica) 47.

CONCLUSIONS AND IMPLICATIONS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

Both contaminant exposure and reduced food intake resulted in increased concentrations of HOCs and their OH and MeSO2 metabolites in plasma, liver, and/or brain of the herring gull chicks. The enhanced biotransformation of HOCs as a result of reduced food intake is likely a consequence of remobilization of HOCs from fat storage compartments into the liver. However, biotransformation enzymes may also be induced by physiological processes due to caloric restriction. In conclusion, the results of the present study show that reduced food intake mobilizes a wide range of HOCs to organs of toxicological sensitivity such as brain and liver. Second, the results indicate that enzyme induction increases de novo synthesis of OH-PCBs (and possibly other contaminant metabolites), which have greater toxicological impacts compared to parent POPs, for example, toward the thyroid system 12, 13.

The results from the present study can be applied in studies investigating levels, temporal trends, and effects of HOCs in Arctic animals. As observed in the present study on herring gull chicks, loss of 10% body mass may lead to 80% higher concentrations of HOCs and over 2.5 times higher concentrations of HOC metabolites in plasma, liver, and brain. Spatiotemporal trends of HOC concentrations and correlative effect studies may thus be highly confounded by the effect of body condition. We recommend that the time window of sampling with respect to the fasting status should be carefully planned in studies that investigate levels and effects of HOCs in Arctic animals.

SUPPLEMENTAL DATA

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

Analytical methods for and congener-specific concentrations of HOCs and their metabolites in the contaminated and clean cod liver oil are presented in the Supplemental Data. In addition, compounds detected in >70% of the liver, brain, and plasma samples of the different treatment groups of herring gulls are presented. Concentrations of ΣPCB, ΣDDT, ΣCHL, ΣPBDE, ΣOH-PCB, ΣMeSO2-PCB, and ΣMeSO2-DDE in liver, brain, and plasma samples, and lipid % in liver and brain samples of the different treatment groups of herring gulls are also given. Results of most parsimonious linear statistical models explaining the effect of the experimental treatments on the concentrations and tissue distribution of HOCs and phase I and II enzymes are given in the supplemental information (426 KB PDF).

Acknowledgements

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

M. Hegseth, M.K. Jansen, T. Nøst, A. Sveistrup, and the staff at Kårvika Marine Research station are acknowledged for their technical assistance during the experiment. We thank H. Ludvigtsen for his help in catching the herring gull chicks. A. Sveistrup and T. Nøst kindly provided technical help during contaminant analysis. C. Bjørk is acknowledged for her technical assistance in the gene expression analysis and T. Nordstad for his help in organizing the data. We are grateful to N. Warner for his comments on the manuscript. The project was funded by the Research Council of Norway (BIOTRANS project 176073) and the Norwegian Ministry of Environment (Center funding). The Norwegian Institute for Air Research and the Norwegian Polar Institute also provided internal funding.

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  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information
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Supporting Information

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS AND DISCUSSION
  6. CONCLUSIONS AND IMPLICATIONS
  7. SUPPLEMENTAL DATA
  8. Acknowledgements
  9. REFERENCES
  10. Supporting Information

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