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Reproductive performance in East Greenland polar bears (Ursus maritimus) maybe affected by organohalogen contaminants as shown by physiologically-basedpharmacokinetic (PBPK) modelling

Christian Sonne a,*, Kim Gustavson b, Frank F. Rigét a, Rune Dietz a, Morten Birkved b,c, Robert J. Letcher d,Rossana Bossi e, Katrin Vorkamp f, Erik W. Born g, Gitte Petersen b

a Section for Contaminants, Effects and Marine Mammals, Department of Arctic Environment, National Environmental Research Institute, Aarhus University, Frederiksborgvej 399,PO Box 358, DK-4000 Roskilde, Denmarkb DHI Water Environment and Health, Department of Environmental Risk Assessment, Agern Allé 5, DK-2970 Hørsholm, Denmarkc DTU Management Engineering, Department of Management Engineering, MAN-QSA, Technical University of Denmark, Produktionstorvet Building 426, Room 111,DK-2800 Kgs. Lyngby, Denmarkd Wildlife and Landscape Science Directorate, Science and Technology Branch, Environment Canada, National Wildlife Research Centre, Carleton University, Ottawa, ON,Canada K1A 0H3e Department of Atmospheric Environment, National Environmental Research Institute, Aarhus University, Frederiksborgvej 399, PO Box 358, DK-4000 Roskilde, Denmarkf Department of Environmental Chemistry and Microbiology, National Environmental Research Institute, Aarhus University, Frederiksborgvej 399, PO Box 358,DK-4000 Roskilde, Denmarkg Greenland Institute of Natural Resources, PO Box 570, DK-3900 Nuuk, Greenland, Denmark

a r t i c l e i n f o

Article history:Received 9 June 2009Received in revised form 1 September 2009Accepted 21 September 2009Available online 27 October 2009

Keywords:Polar bearPCBsPFCsReproductive toxicityPBPK modellingDieldrin

a b s t r a c t

Polar bears (Ursus maritimus) feed mainly on ringed seal (Phoca hispida) and consume large quantities ofblubber and consequently have one of the highest tissue concentrations of organohalogen contaminants(OHCs) worldwide. In East Greenland, studies of OHC time trends and organ system health effects, includ-ing reproductive, were conducted during 1990–2006. However, it has been difficult to determine the nat-ure of the effects induced by OHC exposures on wild caught polar bears using body burden data andassociated changes in reproductive organs and systems. We therefore conducted a risk quotient (RQ)evaluation to more quantitatively evaluate the effect risk on reproduction (embryotoxicity and teratoge-nicity) based on the critical body residue (CBR) concept and using a physiologically-based pharmacoki-netic (PBPK) model. We applied modelling approaches to PCBs, p,p0-DDE, dieldrin, oxychlordane, HCHs,HCB, PBDEs and PFOS in East Greenland polar bears based on known OHC pharmacokinetics and dynam-ics in laboratory rats (Rattus rattus). The results showed that subcutaneous adipose tissue concentrationsof dieldrin (range: 79–1271 ng g�1 lw) and PCBs (range: 4128–53 923 ng g�1 lw) reported in bears in theyear 1990 were in the range to elicit possible adverse health effects on reproduction in polar bears in EastGreenland (all RQs P 1). Similar results were found for PCBs (range: 1928–17 376 ng g�1 lw) and PFOS(range: 104–2840 ng g�1 ww) in the year 2000 and for dieldrin (range: 43–640 ng g�1 lw), PCBs (range:3491–13 243 ng g�1 lw) and PFOS (range: 1332–6160 ng g�1 ww) in the year 2006. The concentrationsof oxychlordane, DDTs, HCB and HCHs in polar bears resulted in RQs < 1 and thus appear less likely tobe linked to reproductive effects. Furthermore, sumRQs above 1 suggested risk for OHC additive effects.Thus, previous suggestions of possible adverse health effects in polar bears correlated to OHC exposureare supported by the present study. This study also indicates that PBPK models may be a supportive toolin the evaluation of possible OHC-mediated health effects for Arctic wildlife.

� 2009 Elsevier Ltd. All rights reserved.

1. Introduction

Assessing the deleterious and potentially toxic effects of OHCs(organohalogen contaminants) exposure in wildlife is difficult. Of-ten, the individual life history of the study animal is unknown andusually only a ‘‘snap-shot” of the OHC tissue or blood plasma con-centrations at a given time in the individual’s life is obtained. Fur-thermore, it is difficult to conduct follow-up studies and to study

0045-6535/$ - see front matter � 2009 Elsevier Ltd. All rights reserved.doi:10.1016/j.chemosphere.2009.09.044

* Corresponding author. Tel.: +45 4630 1200 (switch board), +45 4630 1954(direct), mobile: +45 2521 4686; fax: +45 4630 1914.

E-mail addresses: csh@dmu.dk (C. Sonne), kig@dhigroup.com (K. Gustavson),ffr@dmu.dk (F.F. Rigét), rdi@dmu.dk (R. Dietz), birk@man.dtu.dk (M. Birkved),Robert.Letcher@ec.gc.ca (R.J. Letcher), rbo@dmu.dk (R. Bossi), kvo@dmu.dk (K.Vorkamp), ewb@mail.ghsdh.dk (E.W. Born), gip@dhigroup.com (G. Petersen).

URL: http://www.neri.dk (C. Sonne).

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clinical endpoints in order to evaluate the exposure scenario andpotential adverse health effects on the individual. Furthermore,the disruptive properties of exposure to the ‘‘cocktail” of persistentand bioaccumulative chemicals on endocrine systems are difficultto interpret due to the unknown agonistic and antagonistic interac-tions of the various pollutants that free roaming animals are ex-posed to. Recent studies on polar bears (Ursus maritimus) haveshown that various OHCs may affect vitamin concentrations (Ska-are et al., 2001; Braathen et al., 2004) as well as hormones (Skaareet al., 2001; Haave et al., 2003; Oskam et al., 2003, 2004; Braathenet al., 2004) all of which are important for the reproductive system.Furthermore, high OHC concentrations also appear to be impli-cated in the reduction in the size of the sexual organs of polar bears(Sonne et al., 2006a, 2007a) and as a consequence could disrupttheir reproductive capacity.

It has been a challenge for years to extrapolate OHC effects ob-served in laboratory mammals to wildlife. However, the ArcticMonitoring and Assessment Programme (AMAP, 1998, 2004)extrapolated toxic effects of OHC body residues in laboratory ani-mals to Arctic marine mammals in order to evaluate sub-lethalchronic OHC toxicity. This has been comprehensively assessed ina very recent AMAP review of OHC exposure and effects in Arcticwildlife and fish (Letcher et al., 2009) showing that extrapolationof effects seen in laboratory experiments to wildlife has someinherent weaknesses. Among these are differences in species-spe-cific sensitivity to OHCs, metabolism, life and reproductive cycle.Furthermore, laboratory animals are most often exposed to singleOHCs or technical products at high doses for short periods of time.In contrast, wildlife experience long-term low dosage exposure of avariety of OHCs. Arctic species are known to be more sensitive tothe reproductive effects of contaminants than laboratory animals(Sandell, 1990). On the other hand, an understanding of the link-ages between contaminants and potential health effects is mostlikely to come from the studies on laboratory animals. Once theselinkages are known in established laboratory animal trials they caninform our understanding of the linkages between contaminantsand health effects of wild animals.

A weight of evidence of correlative associations of more recentOHC exposure levels and various physiological and biochemicalendpoints (e.g. endocrine and immune) in polar bears from ‘‘hotspot” populations including those from East Greenland, supportsthe case that some populations of polar bears may be at greaterhealth risk. However, very recently a comprehensive review ofOHC exposure and effects in Arctic wildlife, including polar bears,concluded that there remains a virtual absence of data demonstrat-ing direct OHC-mediated cause-effect (Letcher et al., 2009). Tryingto address the concerns outlined by Letcher et al. (2009), the pres-ent study evaluates the possible linkages of OHC exposure and ef-fects on the reproductive system in polar bears from EastGreenland. To achieve this, we conducted ‘‘risk quotient” (RQ) cal-culations based on the ‘‘critical body residue” (CBR) concept usinga physiologically-based pharmacokinetic (PBPK) model developedby Cahill et al. (2003) and used for various contaminant purposesin, e.g. humans (Redding et al., 2008). Data on polar bears from EastGreenland were chosen as a lot of information on various OHCcompounds and biological parameters was available from this sub-population, covering the period 1990–2006 (Dietz et al., 2004,2007, 2008a; Sonne, 2004; Smithwick et al., 2005; Verreaultet al., 2005; Muir et al., 2006; Dietz, 2008; Gebbink et al., 2008;Letcher et al., 2009a,b). This allowed for an estimation of the riskof possible adverse effects on the reproductive outcome (embryo-toxicity and teratogenicity) in polar bears by comparing polar bearOHC body residues with determined CBR values obtained fromstudies of primarily orally OHC-exposed laboratory rats (Rattus rat-tus) (http://www.inchem.org/; Thomas and Colborn, 1992; Seed,2000; Viberg et al., 2001a,b, 2002). This cross-comparison between

species, however, necessarily assumes that differences in sensi-tively and modes of actions by which OHCs disrupt the endocrinesystems are negligible between species. Indeed, such assumptionsare routinely made in areas such as pharmacology where drugs areinitially tested on model organisms, or in toxicology where safedrinking water standards are set based on the effects found in lab-oratory models. The results of the PBPK model exercise were sub-sequently validated by using the data on OC (organochlorine)concentrations in ringed seals (Phoca hispida) and estimates of dai-ly consumption by East Greenland polar bears of ringed sealblubber.

2. Materials and methods

2.1. Chemical analyses

Polar bear subcutaneous adipose and liver tissues were col-lected from 84 subadults, 31 adult females and 36 adult males dur-ing 1990–2006 by local subsistence hunters in Scoresby Sound,East Greenland (69�N–74�N). In the field, sub samples of both tis-sues were randomly selected from the organs briefly after the killand stored in separate plastic bags. All samples were kept at out-door temperature (�5 to �20 �C) until frozen at �20 �C prior tochemical OHC analyses. All chemical analyses were performed atthe National Environmental Research Institute (NERI, Denmark),and the Letcher Labs previously at the Great Lakes EnvironmentalResearch Institute, University of Windsor (Ontario, Canada) andthe National Wildlife Research Centre (NWRC, Environment Can-ada, Carleton University (Ottawa, Canada), or the National WaterResearch Institute, Environment Canada (Ontario, Canada). TheOHC chemicals monitored included PCBs, DDTs, dieldrin, chlord-anes, HCHs, HCB, PBDEs and PFCs. Details of analyses are givenby Bossi et al. (2005), Dietz et al. (2004, 2007, 2008a), Gebbinket al. (2008), Muir et al. (2006), Verreault et al. (2005). The ex-tracted polar bear and ringed seal contaminant data are given inTable 1.

Ringed seal blubber samples were likewise obtained from 30individuals during 2006 by local subsistence hunters in ScoresbySound, East Greenland and analysed at NERI and at the NWRC.For technical details consult Bossi et al. (2005), Letcher et al.(2009b), Rigét et al. (2006), Vorkamp et al. (2004, 2008). The ex-tracted polar bear and ringed seal contaminant data are given inTable 2.

2.2. Age estimation

Polar bear age was estimated by counting the cementumGrowth Layer Groups (GLGs) of the lower I3 tooth after decalcifica-tion, thin sectioning (14 lm) and staining (toluidine blue) usingthe methods described by, e.g. Hensel and Sorensen (1980), Dietzet al. (1991) with few modifications (Kirkegaard, 2003). Ringed sealages were determined from canine tooth GLGs following the meth-ods given in Hensel and Sorensen (1980), Dietz et al. (1991). If malepolar bears were 6 years of age or older they were categorized asadults. This was also the case with female bears 5 years or olderwhile all other bears were categorized as juveniles.

2.3. OHC critical body residue (CBR) and effect data

In ‘‘classical” laboratory mammals toxicity studies, toxicity iscommunicated as critical daily doses (e.g. mg kg�1 bw d�1). Basedon the critical oral doses determined in the studies of laboratoryrats, corresponding critical body residues (CBRs) of the selectedOHCs were estimated using the PBPK (physiologically-based phar-macokinetic) model. CBR relates the effect of a chemical to its

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internal effect concentration in the body (e.g. lg kg�1 bw) in con-trast to an external effect concentration or oral intake dose. Hence,CBR is a more precise basis for toxicity determination of OHCs asthe tissue concentration more accurately reflects the fraction ofOHCs that has been assimilated by the animal and therefore maybe biologically active. The CBR concept thus includes the integratedresults of bioavailability, bioaccumulation, biomagnification andbiotransformation (Cahill et al., 2003).

Earlier assessments performed by AMAP of threshold levels re-lated to internal effects considered only few OHC substances (PCBs

and dioxin) and for the other OCs in question critical oral doseswere used (AMAP, 1998, 2004). Such a threshold exposure ap-proach was considered too uncertain in the most recent AMAP ef-fect review exercise (Letcher et al., 2009a). However, the presentCBR risk analyses are based on multiple chemicals and their accu-mulative effects and assess the impact on reproduction (embryo-toxicity and teratogenicity) which is a sensitive and importanteffect parameter at the population level. Critical body residueswere estimated based on the data from WHO EnvironmentalHealth Criteria Monographs (http://www.inchem.org/), supplied

Table 1Monitoring data on polar bears sampled in the Scoresby Sound/Ittoqqortoormiit area (East Greenland) in 1990, 1999–2001 and 2006. Monitoring data originally based on wetweight are transformed to lipid weight using 10% lipids in liver and 70% lipids in adipose tissue (fat). All data are given as lg kg�1 lipid weight except PFOS which is given inlg kg�1 ww. Data from Norstrom et al. (1998), Dietz et al. (2004, 2008a) and Dietz (Unpubl. data).

Gender Substance Tissue Year Mean Median 25%-Fractile 75%-Fractile Minimum Maximum No. of samples

Females Dieldrin Adipose 1990 444 420 345 494 175 835 71999–2001 180 155 139 226 110 286 212006 125 120 82 167 43 213 3

HCB Adipose 1990 173 95 57 116 25 749 71999–2001 57 45 32 64 17 229 212006 59 62 44 75 26 89 3

Oxychlordane Adipose 1990 2854 2490 1781 3986 1606 4345 71999–2001 1274 1037 753 1429 433 6022 212006 924 823 663 1135 502 1448 3

PCBs Adipose 1990 19 724 18 069 14 153 24 378 12 558 30 380 71999–2001 5050 4811 3178 5521 1928 17 035 212006 5209 4790 4141 6069 3491 7347 3

PFOS Liver 1999–2001 1251 1350 852 1652 237 1890 192006 3100 3122 2841 3381 2700 3456 4

ppDDE Adipose 1990 267 216 205 346 147 405 71999–2001 322 347 216 399 67 857 212006 128 124 121 134 118 143 3

HCHs Adipose 1990 367 289 256 444 175 705 71999–2001 188 144 127 201 81 817 212006 103 94 78 123 62 153 3

Males Dieldrin Adipose 1990 237 211 125 306 79 505 71999–2001 222 180 97 247 60 1020 182006 249 201 175 256 68 640 8

HCB Adipose 1990 88 52 38 128 35 263 101999–2001 102 59 29 81 2 921 182006 235 152 91 246 21 795 8

Oxychlordane Adipose 1990 1161 923 321 1343 118 4377 101999–2001 780 757 608 1006 202 1702 182006 1148 876 776 1272 580 2671 8

PCBs Adipose 1990 15 552 15 380 8893 18 322 6317 33 790 101999–2001 7070 6659 5042 7684 3580 13 452 182006 12 385 12 078 10 716 13 243 8725 13 243 8

PFOS Liver 1999–2001 1137 1036 904 1344 797 1660 152006 3108 2594 2263 3510 1332 6160 6

ppDDE Adipose 1990 339 219 152 456 44 1009 101999–2001 467 426 170 604 94 1411 182006 268 214 134 335 84 638 8

HCHs Adipose 1990 301 297 261 339 172 436 81999–2001 264 223 171 258 91 729 182006 118 120 86 153 11 204 8

Juveniles Dieldrin Adipose 1990 397 338 236 502 128 1271 271999–2001 214 199 147 245 93 547 512006 143 138 112 160 77 236 6

HCB Adipose 1990 168 61 38 254 1 817 261999–2001 115 76 54 110 29 553 512006 88 64 62 102 31 192 6

Oxychlordane Adipose 1990 2341 1951 1208 2670 436 8510 281999–2001 1179 1009 689 1569 322 4533 512006 731 586 484 871 343 1459 6

PCBs Adipose 1990 20 193 17 436 13 536 21 970 4128 53 923 281999–2001 5321 4775 3941 6248 2092 17 376 512006 7387 7556 7230 8114 3896 9895 6

PFOS Liver 1999–2001 962 836 595 1130 104 2840 452006 3003 3056 2270 3726 2108 3868 7

ppDDE Adipose 1990 529 439 303 604 142 1582 281999–2001 399 375 273 495 76 1147 512006 152 134 90 217 36 286 6

HCHs Adipose 1990 390 338 276 517 113 654 251999–2001 191 171 138 228 117 400 512006 93 90 85 99 76 113 6

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with more recent literature values and combined with the use of aPBPK model (Thomas and Colborn, 1992; Seed, 2000; Viberg et al.,2001a,b, 2002). Critical oral dose for rat reproduction (embryotox-icity and teratogenicity) obtained from the literature, and the cor-responding CBR values estimated from the physiologically-basedpharmacokinetic (PBPK) model are shown in Table 3. For risk eval-uations the estimated CBR values are directly comparable withOHC concentrations measured in polar bears from East Greenland(Tables 1 and 2). The CBR data estimates for rats are presentlyscaled to evaluate the possible effects in East Greenland polarbears, despite the obvious inter-species sensitivity differences to-wards OHCs between rats and polar bears.

For improving the assessment of the potential effects related toOHCs in polar bears, effect values on reproduction (embryotoxicityand teratogenicity) were generated from methodically uniformlaboratory mammal test (primary on rats) where critical oral doseswere determined. The WHO and recent literature data evaluatedwere all chronic and sub-chronic effect studies mainly performedwith rats and in which the animals were dosed orally(mg kg�1 bw d�1) for a long period (e.g. 30 d) and up to exposurefor three generations.

2.4. Physiologically-based pharmacokinetic (PBPK) model

Based on the validated effects data, the internal effect dose(CBR) for blood, adipose and liver tissues was estimated usingthe PBPK model developed at the Canadian Environmental Model-ling Centre at Trent University and described in Cahill et al. (2003).The PBPK modelling uses a special type of pharmacokinetics wherephysiology and anatomy of the mammalian body, and the bio-chemistry of the chemical or chemicals of interest are incorporatedinto a conceptual model for computer simulation. Unlike classicalpharmacokinetics, PBPK modelling is a powerful tool for manytypes of extrapolation, including species-to-species, route-to-route, and dose-to-dose. A complete model description, including

model equations and default parameterization, is presented onthe Environmental Toxicology and Chemistry web site (SETACSupplemental Data Archive, Item ETC 22-01-001; http://etc.allenpress.com).

Besides simulation time and number of iterations, the modelincorporates data for chemical exposure (oral or inhalation), bio-metric data (e.g. mass of animal, volume of intestine and organs),rates of metabolism and release and different physical–chemicalproperties (partition coefficients). For evaluation of the contribu-tion of each OHC compound to the potential environmental riskof the species or the group of species in question, a generic riskquotient (RQi) was calculated as:

RQ i ¼BRi

CBRi

where BRi is the mean body residue monitored in the organism orthe group of organisms and CBRi is the critical body residue, whichrelates the specific effect of a chemical to its molar concentration inthe body. The combined effect of the monitored OHCs or the poten-tial environmental generic risk to the organism from several sub-stances is determined by adding the RQi for each of thecomponents assuming additivity of the potential toxic effect (bio-chemical response, reproduction and mortality). When the sum ofRQ is <1, no generic risk of ecotoxic effects is assumed. When RQis P1, it should be assumed that there is a generic risk of ecotoxiceffects (Pedersen and Petersen, 1996):

RQ ¼Xn

i¼1

BRi

CBRi

where RQ P 1 indicates potential risk of the evaluated effect; andRQ < 1 indicates no potential risk of the evaluated effect.

In the present PKPB modelling, a combined excretion/metabolicrate constant on 10 h�1 has been used. A detailed description of thepresent PBPK model and an example on the CBR assessment forHCH are given in Gustavson et al. (2008).

Table 2Monitoring OC data in ringed seal blubber sampled in the Scoresby Sound/Ittoqqortoormiit area (East Greenland). All data are given as lg kg�1 lipid weight. Data from Rigét et al.(2006) and Rigét (Unpubl. data).

OC Year Mean Median 25%-Fractile 75%-Fractile Minimum Maximum No. of samples

PCBs 2006 822 658 446 922 151 3098 30ppDDE 2006 776 535 379 954 112 3060 30HCHs 2006 45.2 41.4 36.6 48.7 19.8 105 30HCB 2006 9.8 9.0 7.6 11.5 3.4 22.5 30Oxychlordane 2006 167 139 87 193 33.6 706 30

Table 3Critical oral dose for rat reproduction (embryotoxicity and teratogenicity) obtained from the literature and corresponding CBR values estimated for polar bears using thephysiologically-based pharmacokinetic (PBPK) model. CBR values are estimated for adipose tissue or liver tissue. All CBR values are in lg kg�1 lipid weight with exception forPFOS which is in wet weight.

Group Substance Critical daily dose (NOEC) (mg kg�1 bw) Reference Tissue CBR

Dieldrin Dieldrin 0.015 Thomas and Colborn (1992) Adipose 1571HCB HCB 1.0 WHO Adipose 18 141Oxychlordane Chlordane 0.25 WHO Adipose 4912PCBs Arochlor 1260 0.32 WHO Adipose 1825ppDDE DDT 10 WHO Adipose 168 571HCHs Lindane 5.0 WHO Adipose 73 143PBBs Polybrominated biphenyls 6.0 WHO Adipose 12 486Dioxins TCDD 0.005 WHO Adipose 89Furans TBDF 0.01 WHO Adipose 141Toxaphene/camphechlor Toxaphene/camphechlor 6.5 WHO Adipose 120 571PFOS Perfluorooctane sulfonates 1.6 Seed (2000) Liver 1600Dioxins TCDD 0.005 WHO Liver 34Furans TBDF 0.01 WHO Liver 55PBDEs RPBDEs 10.5 Viberg et al. (2001a,b, 2002) Liver 24 107

WHO: WHO Environmental Health Criteria Monographs (http://www.inchem.org/).

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The principles for the above risk characterisation are describedin the technical guidance document on risk assessment based onPredicted Environmental Concentrations (PECs) and external NoObserved Effect Concentrations (NOECs) (EEC, 2003). When usingthese principles for risk characterisation, assumptions regardingmonitored OHC data on Arctic animals equalled direct measureof the internal body residues (BRs).

2.5. Model validation

The PBPK model was evaluated for polar bears by estimating thetheoretical intake of OCs via ringed seal blubber using the ringedseal data for the year 2006 (Table 2) and estimated blubber con-sumption estimated by Stirling and McEwan (1975), Kingsley(1998). The CBR estimations were compared with actual concen-trations in adipose and liver tissues in ringed seals and polar bears.

3. Results and discussion

Concentrations of OHCs in East Greenland polar bears are givenin Table 1. It is seen that PFOS was highest in 2006 for all age/sexgroups while all organochlorine compounds have either decreasedor stabilized in the period 1990–2006 (Dietz et al., 2004, 2008a;Dietz, Unpubl. data; Norstrom et al., 1998). The concentrations ofOHC compounds in the polar bears increased for nearly all groupsin the order HCBs � HCHs < p,p0-DDE < dieldrin < oxychlor-dane < PFOS < PCBs. Only for p,p0-DDE and dieldrin in juvenilesthe order was reversed probably due to the lower metabolism inthis age cohort (AMAP, 1998, 2004). OC concentrations were, ex-cept for p,p0-DDE, significantly lower in ringed seals than in polarbears (Table 2) and in ringed seals the OCs increased in the orderHCB < HCHs < oxychlordane < p,p0-DDE < PCBs which reflect thebiomagnification in the Arctic marine food web and the high DDTmetabolic capacity of polar bears (AMAP, 1998, 2004; Vorkampet al., 2004, 2008; Rigét et al., 2006; Letcher et al., 2009b).

3.1. Critical body residues (CBRs) and risk quotients (RQs)

Critical body residues were estimated using the PBPK modeland the oral doses for rats obtained from the scientific literature.In the present PBPK modelling, a combined excretion/metabolicrate constant on 10 h�1 has been used as this was regarded as arealistic conservative value for most of the OHCs evaluated. Fur-thermore, the combined effects of different OHCs were estimatedon the basis of the added risk quotients for the single monitoredOHCs based on a specific effect parameter such as reproduction.This assumption might underestimate the risk as it is known thatsome substances and mixtures of different PCB congeners havebeen found to cause synergistic estrogenic effects (Crews et al.,1995).

RQs for East Greenland polar bears were calculated from meanbody residues (Tables 1 and 2) in the periods 1990, 1999–2001 and2006 and estimated CBR values for each OHC group (PCBs, p,p0-DDE, dieldrin, oxychlordane, HCHs, HCB, PBDEs and PFOS) as wellas the added risk for all substances (Figs. 1–3). Fig. 1 shows thatthe RQ values were P1 for dieldrin and nearly 10 for PCB body res-idues around 1990 indicating potential risk for adverse effects onreproduction in both adult male and adult female and in juvenilepolar bears. Furthermore, the RQ for combined OHC effects wasP10 indicating that the synergistic effect from all OHC body resi-dues presents an even greater potential risk to the health of thereproductive performance (embryotoxicity and teratogenicity) inpolar bears.

Between 1990 and 2000 (represented by data from 1999 to2001), RQs decreased for PCB body residues although they are still

above 1 indicating reproductive toxic effects (Fig. 2). Dieldrin RQvalues are nearly the same as those for PCBs although just below1 suggesting a lower risk for reproductive effects. However, RQfor combined OHC body residues was still around 5 in 2000, there-by suggesting negative effects on reproduction despite the decline.It should be noticed that RQ for PFOS is 1 for adult females (Fig. 2)sampled in 2000.

In 2006, the body residues of three OHC compounds haveRQs P 1 (Fig. 3). These are increasing in the order diel-drin > PFOS > PCBs. Except for dieldrin, all adult females, adultmales and juveniles had RQs P 1, which means that all thesegroups of individuals could be susceptible to reproductive effects(note that only three adult females were analysed in 2006). ForPCBs in adult males, the concentrations in subcutaneous adiposetissue had increased from 1999 to 2001 to a level in 2006 similarto that in 1990. We suggest that this could be a result of accessto open water-related seal species (i.e. harp [Phoca groenlandica],bearded [Erignathus barbatus] and hooded [Cystophora cristata]),changes in biomagnification processes in the East Greenland mar-ine food webs related to reduced ice extent, sea ice break-up, tem-perature increase and NAO (North Atlantic Oscillation) (Macdonaldet al., 2003, 2005; Dietz et al., 2008b–d; McKinney et al., 2009).

3.2. PCBs

PCBs and their OH-metabolites are known to exhibit variousendocrine properties such as estrogenic and thyroid-like (Colbornand Clement, 1992; Colborn et al., 1993; Kramer et al., 1997;Moore et al., 1997; Andersson et al., 1999; Matthews and Zac-harewski, 2000). Wildlife studies and controlled laboratory studiesof rodent and primates have shown that these compounds inter-fere with male and female reproductive performances such as tes-tes and endometrial pathology, sperm quality, foetal development,growth and litter size at food concentrations of 8–250 lg kg�1 d�1

and adipose tissue concentrations of 7.5–77 lg g�1 lw (Golub et al.,1991; WHO, 1992; AMAP, 1998, 2004). Likewise, numerous studieshave shown that PCBs induce negative reproductive effects in mink(Mustela vison) similar to those in rats. This is important to noticeas mink has a reproductive cycle that includes delayed implanta-tion similar to that of polar bears. However, mink is rather suscep-tible to PCB exposure as a species, and it is therefore difficult toextrapolate to Arctic species and the possible effects may beover-interpreted. Studies of mink feeding on environmentally rele-vant concentrations of MeSO2–PCB metabolites for 1 year showedlower birth weights and reduced kit survival at muscle concentra-tions of 18 000 ng g�1 lw and 21 000 ng g�1 lw (Lund et al., 1999).It should be noted that polar bears, including those from EastGreenland, have been reported to contain considerable levels ofMeSO2–PCB metabolites in adipose and liver, and thus carry bodyburdens similar to PCBs (Gebbink et al., 2008; Letcher et al.,2009b). Furthermore, in vitro studies showed increased livermetabolism of progesterone. In mink, continuous oral exposureto 250 ng PCB g�1 food of contaminated fish induced oestrus delayand lowered whelping rates, while dietary oral exposure to a con-centration of 500 ng PCB g�1 increased litter mortality and bodyweight of kits (Restum et al., 1998).

East Greenland polar bears contain the highest blood levels ofPCBs and OH-metabolites for any (Arctic) species (Gebbink et al.,2008; Letcher et al., 2009a). Only little research on these highPCB concentrations on reproductive and ontogenic developmentaleffects has been conducted in polar bears due to extreme logisticand financial challenges (Letcher et al., 2009a). Sonne et al.(2006a, 2007a) and Sonne (submitted for publication) reported in-verse correlations between RPCB concentrations and size of maleand female sexual organs at adipose tissue and food concentrationsof 0.8–20 lg g�1 lw and 16 lg kg�1 d�1, respectively, suggesting a

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pre- and postnatal neuroendocrine disruption of the thalamic–hypothalamic–pituitary–target organ axis (Colborn and Clement,1992; Colborn et al., 1993; Morse et al., 1993; Tryphonas, 1994;Morse, 1995; Faroon et al., 2001; Zoeller et al., 2002; Tung and Iq-bal, 2007; Hertz-Picciotto et al., 2008; Klecha et al., 2008; Langer,2008). These concentrations and relationships are similar to thosefrom controlled laboratory studies and may in worst case lead toreduced sperm/egg and baculum quality/quantity resulting in low-er reproductive rates and population declines as reflected in the

high PCB RQs (Bergeron et al., 1999; Rune et al., 1991a,b; Sharpeand Skakkebæk, 1993, 2003; Facemire et al., 1995; Gray, 1998;Weidner et al., 1998; James, 1999; Pfleiger-Bruss et al., 1999; Hosieet al., 2000; Baskin et al., 2001; Fielden et al., 2001; Skakkebæket al., 2001; Jarrell et al., 2002). The RQs P 1 are strong suggestionsthat East Greenland polar bear reproductive health is at risk. Der-ocher et al. (2003) discussed whether OHCs could have affectedthe demography of Svalbard polar bears. As East Greenland andSvalbard polar bears are polluted at the same magnitude (Smith-

Fig. 1. Risk quotients based on monitoring data (1990 mean concentrations) and estimated CBR values for females, males and juvenile polar bears, respectively.

Fig. 2. Risk quotients based on monitoring data (1999–2001 mean concentrations) and estimated CBR values for females, males and juvenile polar bears, respectively.

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wick et al., 2005; Verreault et al., 2005; Muir et al., 2006) it maysupport the present findings of RQs P 1.

A few East Greenland and Svalbard polar bear studies have re-ported on the morphological changes in sexual organs (enlargedand penis-formed clitoris) that could be theoretically related toPCB exposure but no solid conclusions were made (Wiig et al.,1998; Sonne et al., 2005a). Furthermore, a single case of hypospa-dias in an East Greenland sledge dog has been reported (Sonneet al., 2008a). Such malformations totally prevent the individualfrom reproducing. However, such effects have not yet been re-ported on polar bears. In addition due to high TEQs (dioxin equiv-alents) in the Greenland sledge dogs – among others – teratogenicchanges from endocrine-disrupting chemicals in the neonatal envi-ronment could not be excluded.

3.3. PFOS

The oral daily exposure to PFOS could not be evaluated for EastGreenland polar bears. In other studies, PFOS effect dose reducingpup survival and weight of rats was 1.6–3.2 mg kg�1 d�1 (�internaldose of 58 lg g�1 liver ww) while LOAEL was 0.4 mg kg�1 d�1 andNOAEL was 0.1 mg kg�1 d�1 (�internal dose of 15 lg g�1 liver ww)(Seed, 2000). In rabbits, PFOS caused maternal toxicity (decreasedbody weight gain) at 1 mg kg�1 d�1 or higher (Case et al., 2001) andcaused abortions and reduced foetal weights. According to Seacatet al. (2003), PFOS liver residue NOAEL was set to 358 ppm for liverlesions in male rats and 370 ppm for female rats, which is approx-imately 300 times higher than the concentrations found in thepresent East Greenland polar bear liver tissue (Dietz et al.,2008a; Sonne et al., 2008b). However, Beach et al. (2006) expectpopulation reproductive and immune effects at 5000 ng g�1 wwwhich is only five times above the mean PFOS concentrations re-ported in East Greenland polar bears. Following the considerableincreasing trends (up to 2005) of perfluorinated sulfonates (PFSAs;and essentially all PFOS) and several longer chain perfluorinatedcarboxylates (PFCAs) in East Greenland polar bears (Dietz et al.,2008a), these PFCs are projected to increase further and estimated

to reach the levels that could be deleterious to polar bears by 2012.However, PFOS and C8–C13 chain length PFCAs increases are unli-kely to continue beyond 2012 as these longer chain PFCs have orwill be phased out in production (Houde et al., 2006).

3.4. Dieldrin

Ecologically relevant concentrations of dieldrin have been asso-ciated with reduced fertility in rats (15 ng g�1 bw d�1) (Thomasand Colborn, 1992) and increased pup mortality in mice, rats anddogs at doses above 25 ng g�1 bw d�1 (WHO, 1989). Estimated dai-ly oral exposure of East Greenland polar bears is 1 ng g�1 bw (Son-ne, in press) which is only slightly lower than the concentrationsthat induce effects in other species. Indeed, Sonne et al. (2006a)showed an inverse relationship between the adipose tissue diel-drin concentration and sexual organ size in East Greenland malepolar bears. These data support the observation we made in thepresent study that dieldrin levels, especially in 1990 with RQ > 1,have likely affected polar bear reproductive success and providedirection for future studies that would evaluate the mechanismsthrough which dieldrin could induce these effects.

3.5. Chlordanes

For oxychlordane, earlier rat studies have shown that a short-term exposure of ca. 25 lg g�1 bw d�1 impairs fertility and repro-ductive outcome (Ambrose et al., 1953; Keplinger et al., 1968;Welch et al., 1971; Balash et al., 1987; WHO, 1989). Comparing thisdaily short-term toxicity dose in mice and rats with the estimatedlong-term exposure for polar bears at 3.3 lg kg�1 bw d�1, there isno risk of acute toxic effects while the chronic long-term impacton polar bear reproductive success is unknown. However, the polarbear adipose concentrations of oxychlordane were high and de-spite the RQs for all three periods being below 1, this compoundmust still be considered a problem for polar bear reproductivity.This is supported by the high sumRQs and the fact that adipose tis-sue chlordane concentrations were also associated with reduced

Fig. 3. Risk quotients based on monitoring data (2006 mean concentrations) and estimated CBR values for females, males and juvenile polar bears, respectively.

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size of sexual organ in East Greenland male and female polar bearsand thereby potentially reproductive irreversible toxic impact(Sonne et al., 2006a).

3.6. Evaluation of the PBPK utility for polar bears

In the present study a PBPK model has been used to estimateCBR values from the data for critical oral doses in rats. In continu-ation of this we tested the utility of the PBPK model for estimationof body residual concentrations in polar bears by using the datafrom 2006 of OHC concentrations in ringed seals (Table 2). In cen-tral East Greenland ringed seals are assumed to constitute the maindiet for polar bears. The assumptions taken for evaluation of theutility of the PBPK model are (Stirling and McEwan, 1975; Kingsley,1998): (1) The weight of an average polar bear is set to 200 kg andit is assumed that the bear is reproductively inactive and not grow-ing. (2) The yearly intake of ringed seal blubber is set to 1000 kg.(3) The model was run for a time period of 2 years. In Table 4,the monitored and modelled concentrations of selected OHCs insubcutaneous adipose tissue of East Greenland polar bears arecompared. In general, the model fits very well with actual concen-trations based on monitoring data. For PCB, a relatively high devi-ation was found probably due to the fact that the PBPK modelvalidation, for practical reasons, only covered one congener ofPCB while the monitoring data cover a wide range of PCB congen-ers. For the other OHCs in question (p,p0-DDE, HCB, HCHs, dieldrinand oxychlordane) the PBPK model reflects the monitored data inEast Greenland polar bears very well meaning that the PBPK modelis applicable to the physiology and feeding behaviour of EastGreenland polar bears.

3.7. Interpretations

Based on the present calculations, other Arctic mammals suchas the Arctic fox (Vulpes lagopus) and killer whale (Orcinus orca)(AMAP, 1998, 2004; Fuglei et al., 2007; Wolkers et al., 2007; Let-cher et al., 2009a) may be at risk due to high OHC exposure andsubsequent body burdens. In addition the East Greenland Inuits –who consume the seals from the same area as the East Greenlandpolar bears – are also likely to be at greater risk. Although adverse

effects in the reproduction in the human population of East Green-land to our knowledge have not been demonstrated despite long-term medical surveillance. This was also the conclusion of a riskassessment by Nielsen et al. (2006). In addition to the risk forreproductive success calculated in the present study, as reviewedand summarized in Letcher et al. (2009a), other possible health ef-fects on other key parameters such as vitamin concentrations (Ska-are et al., 2001; Braathen et al., 2004), hormone concentrations(Skaare et al., 2001; Haave et al., 2003; Oskam et al., 2003, 2004;Braathen et al., 2004), CYP450 (Bandiera et al., 1995; Letcheret al., 1996), bone (Sonne et al., 2004, 2006a), organ (Sonne et al.,2005b, 2006b, 2007b, 2008b) and immune effects (Lie et al.,2004, 2005) should be added in order to complete the list of possi-ble OHC effects in polar bears. Polar bears are a typical k-selected(long living and low reproductive competitor species in crowdedniches) species which makes them vulnerable to impacts on theirreproductive system and succeed. Interpretive caution is empha-sized for the present study as it is based on the assumption thatrat data for OHC reproductive sensitivity can be extrapolated to po-lar bears. Finally, similar PBPK modelling should be conducted onimmune and thyroid toxicities as these organ systems are likelyto be negatively influenced by OHCs in East Greenland polar bearsas well (Letcher et al., 2009a).

4. Conclusions

Based on the critical oral doses, CBR values for adipose and livertissue OHC concentrations were estimated using the PBPK model.OHC concentrations measured in polar bears from East Greenlandcollected in 1990, 1999–2001 and 2006 had RQs P 1 for PCBs,PFOS and dieldrin for most periods which suggested increased riskfor adverse reproductive effects. Furthermore, all sumRQs were P1showing synergistic OHC toxicity which is also supported by previ-ous laboratory-controlled studies and field data from East Green-land and Svalbard bears. RQs above 10 were observed for PCBs in1990 suggesting greater risk for reproductive performance in bothadult male and adult female and in juvenile polar bears. SumRQsclose to 10 were also observed in 2006 due to PCB increase after2000 and the general increase of PFCs. These findings increasethe health concern for OHC-exposed East Greenland polar bears

Table 4Monitored and PBPK model-estimated OC concentrations of contaminants in ringed seal blubber and polar bear adipose tissue from central East Greenland, 2006. All values are inlg kg�1 lipid weight. Data from Dietz (Unpubl. data) and Rigét (Unpubl. data).

Monitored data PBPK modelled data

OC group Ringed seal Polar bear females Polar bear males Polar bear juveniles

p,p0-DDEAvg. 776 128 268 152 147Max 3060 143 638 286 580Min 112 118 84 36 21

HCBAvg. 10 59 235 88 54Max 23 89 795 192 124Min 3 26 21 31 19

HCHsAvg. 45 103 118 93 81Max 105 153 204 113 190Min 20 62 11 76 36

OxychlordaneAvg. 167 924 1148 731 680Max 707 1448 2671 1459 2800Min 33 502 580 343 135

PCBsAvg. 822 5209 12 385 7387 4500Max 3098 7347 13 243 9895 16 960Min 151 3491 8725 3896 827

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as they are among the most polluted wildlife species and due to theslow reproductive rates of this k-selected species. Furthermore, theGreenland Inuit population, sledge dogs, killer whales and Arcticfoxes may also be of risk with respect to toxic reproductivity. Wesuggest that similar evaluations are conducted for Inuits and forimmunopathy and thyroid function as well.

Conflict of interest

No conflicts of interest were reported.

Acknowledgements

The Danish Cooperation for Environment in the Arctic isacknowledged for financial support, and H. & B. Sandell, JonasBrønlund and local hunters are acknowledged for organizing thesampling in East Greenland.

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