ent 378 (2007) 403–417www.elsevier.com/locate/scitotenv

Science of the Total Environm

Levels and trends of organochlorines and brominated flameretardants in Ivory Gull eggs from the Canadian Arctic, 1976 to 2004

Birgit M. Braune a,⁎, Mark L. Mallory b, H. Grant Gilchrist a,Robert J. Letcher c,1, Ken G. Drouillard c

a Environment Canada, National Wildlife Research Centre, Carleton University, Raven Road, Ottawa, Ontario, Canada K1A 0H3b Environment Canada, Box 1714, Iqaluit, Nunavut, Canada X0A 0H0

c Great Lakes Institute for Environmental Research, University of Windsor, Windsor, Ontario, Canada N9B 3P4

Received 21 September 2006; received in revised form 19 February 2007; accepted 2 March 2007Available online 5 April 2007

Abstract

The ivory gull (Pagophila eburnea) is a circumpolar marine bird which has recently been listed as an endangered species inCanada. To determine whether contaminants may be playing a role in the population decline of this species, ivory gull eggscollected in 1976, 1987 and 2004 from Seymour Island in the Canadian Arctic were analyzed for organochlorines, polychlorinateddibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs) and non-ortho PCBs. This study also provides the first account ofpolybrominated diphenyl ethers (PBDEs), hexabromocyclododecanes (HBCDs) and polybrominated biphenyls (PBBs) in ivorygulls. The most quantitatively abundant legacy organochlorines found in the ivory gull eggs were p,p′-DDE, ΣPCB andoxychlordane. Concentrations of the organochlorines analyzed either decreased or showed little change between 1976 and 2004.Concentrations of ΣPCDD in ivory gull eggs were greater than ΣPCDF, and the non-ortho PCBs (primarily PCB-126) contributedthe largest fraction to the total TEQ value in all years sampled. Concentrations of PCDDs, PCDFs and ΣTEQ decreased from 1976to 2004. In contrast, concentrations of the PBDEs steadily increased between 1976 and 2004 driven primarily by increases in BDE-47. Although concentrations of the persistent chlorinated compounds (i.e. organochlorine pesticides, PCBs, PCDDs, PCDFs)reported in this study were below published toxicological threshold values for eggs of wild birds, we cannot rule out the possibilityof synergistic/additive, sublethal effects. Very few studies have been carried out to evaluate the exposure-effect relationship of thepersistent brominated compounds in avian species. Given the scarcity of information on toxicity threshold levels for PBBs andPBDEs in avian species, coupled with the trend toward increasing concentrations in ivory gulls, continued monitoring and furthertoxicological studies of these compounds are warranted.© 2007 Elsevier B.V. All rights reserved.

Keywords: Ivory gull; Canadian Arctic; Organochlorines; Dioxins; Brominated flame retardants

⁎ Corresponding author. Tel.: +1 613 998 6694; fax: +1 613 9980458.

E-mail address: birgit.braune@ec.gc.ca (B.M. Braune).1 Current address: Environment Canada, National Wildlife Research

Centre, Carleton University, Raven Road, Ottawa, Ontario, CanadaK1A 0H3.

0048-9697/$ - see front matter © 2007 Elsevier B.V. All rights reserved.doi:10.1016/j.scitotenv.2007.03.003

1. Introduction

It has been known for several decades that biota atnorthern latitudes are exposed to a variety of chemicalresidues that are, for the most part, transported there byair and ocean currents (Macdonald et al., 2000). More

Fig. 1. Location of sampling site (Seymour Island) in the CanadianArctic.

404 B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

recently, brominated organic compounds such aspolybrominated diphenyl ethers (PBDEs), hexabromo-cyclododecanes (HBCDs) and polybrominated biphe-nyls (PBBs) have been detected in a wide variety ofbiota including arctic wildlife (de Wit, 2002; Law et al.,2003). Wania and Dugani (2003) demonstrated that thelower-brominated PBDE congeners have an Arctic con-tamination potential comparable to those PCBs knownto undergo significant long-range transport.

Persistent chlorinated and brominated organo-con-taminants biomagnify up the food chain making thosespecies feeding at high trophic positions more vulner-able to contaminant exposure via their diet (Hop et al.,2002; Borgå et al., 2004; Vorkamp et al., 2004; Wolkerset al., 2004). One top predator, the ivory gull (Pagophilaeburnea), is an arctic marine bird which is experiencingserious population declines (Gilchrist and Mallory,2005) and has recently been listed as “Endangered” inCanada (COSEWIC, 2006). Ivory gulls feed largely onfish and invertebrates, but also scavenge carrion ofmarine mammals (Haney and MacDonald, 1995). Giventheir relatively high trophic position in marine foodwebs (Hobson et al., 2002), contaminants have beenproposed to be among the health stressors that could beaffecting this species.

Elliott et al. (1992) reported that most legacy organo-chlorines (e.g. PCBs, p,p′-DDE, chlordanes, hexachlor-ocyclohexanes, mirex) as well as 2,3,7,8-TCDD hadremained the same or increased in ivory gulls from theCanadian Arctic between 1976 and 1987, and Brauneet al. (2006) found that mercury levels in eggs of ivorygulls were among the highest ever reported for seabirdeggs from the arctic marine environment. There isstrong evidence from other data sets that the majority oflegacy persistent organic pollutants (e.g. PCBs, DDTs)have significantly declined in Canadian Arctic biotaover the last several decades whereas contaminants suchas the hexachlorocyclohexanes (HCHs) and Hg haveeither remained relatively constant or increased in anumber of arctic species (Braune et al., 2005).

Since contaminant burdens in the egg reflect residuesassimilated by the female, seabird eggs have been usedto monitor trends and patterns of contaminants in theCanadian Arctic marine environment (Braune et al.,2001, 2002; Braune, in press). In this study, we re-analyzed archived samples of ivory gull eggs collectedfrom Seymour Island in the western Canadian Arctic(Fig. 1) in 1976 and 1987 for the legacy organochlorinesusing standardized laboratory protocols, and we extend-ed the time series by analyzing eggs collected from thatsame location in 2004. We also provide new informationfor ivory gulls on polychlorinated dibenzo-p-dioxins

(PCDDs), dibenzofurans (PCDFs), non-ortho PCBs,and polybrominated diphenyl ethers (PBDEs) as well ashexabromocyclododecanes (HBCDs) and some poly-brominated biphenyls (PBBs). We include data for sta-ble isotopes of nitrogen (15N/14N), as a reflection of thediet of the female prior to or during egg-laying (Hobson,1995; Hebert et al., 1999), to monitor possible temporalchanges in trophic position that could affect the con-centrations of contaminants found in the eggs. Theresidue data were compared with published toxicitythreshold levels to determine whether persistent chlori-nated or brominated pollutants may be playing a role inthe population decline of this species.

2. Methods

2.1. Sample collection and storage

Ivory gull eggs were collected by hand from SeymourIsland, Nunavut, Canada (Fig. 1) during 1976 (n=9),1987 (n=9) and 2004 (n=6) under appropriate CanadianWildlife Service (CWS) collection permits. Each eggwas removed from a different nest to maintain inde-pendence among samples. Eggs collected in 1987 and2004 were kept cool in the field and shipped to theNational Wildlife Research Centre (NWRC) for proces-sing and archival. Eggs collected in 1976 were preservedin 10% formalin in the field prior to shipment to NWRCwhere they were stored frozen. Egg contents from allyears were homogenized and stored frozen (−40 °C) inacetone–hexane rinsed glass vials. Because the period ofstorage in formalin had dehydrated the 1976 samples, aquantity of 10 to 20% of RO-water (water purified using

405B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

reverse osmosis; BioLabWater Purification System) wasadded to each egg sample from 1976 to aid in homog-enization of those samples. Numerous studies utilizingtissues stored in formalin for periods up to several de-cades have shown little to no effects of formalin storageon organochlorine concentrations (French and Jefferies,1971; MacGregor, 1974; Neidermyer and Hickey, 1976;Wiemeyer et al., 1984; Loganathan et al., 1993; Renaudet al., 1995, 1999, 2004) or δ15N (Kaehler andPakhomov, 2001; Sarakinos et al., 2002) in tissues, andthus we did not consider any special treatment of thesamples from 1976.

2.2. Organochlorine analysis

Egg homogenates were analyzed as pooled (compos-ite) samples with each pool consisting of three individualegg samples. The number of eggs collected in each yeardictated the number of pools available for analysis (seeTable 1). A sample of the RO-water used to rehydrate the

Table 1Mean concentrations of organochlorines (ng g−1 lipidweight±standarderror based on pooled analyses) and stable nitrogen isotope ratios (δ15Nas o/oo±standard error based on individual analyses) in eggs of ivorygulls collected between 1976 and 2004

Year 1976 1987 2004

n (# pools) 9 (3) 9 (3) 6 (2)δ15N 15.6±0.7 18.4±0.3 17.3±0.2% lipid 4.8±0.6 8.7±0.4 10.3

1245-TeCB 48.6±4.8 72.7±19.3 19.81234-TeCB 22.1±2.3 22.5±1.1 13.2PnCB 93.4±6.3 92.3±8.4 38.0HCB 1463±184 1244±56.2 515ΣCBz 1627±197 1432±61.5 586α-HCH 71.3±19.6 113±27.7 15.1β-HCH 70.8±9.2 188±26.4 160ΣHCH 142±20.8 301±54.1 175cis-Chlordane 70.5±12.0 102±19.5 58.2trans-Nonachlor 341±169 1255±149 838cis-Nonachlor 132±41.9 259±36.7 140Oxychlordane 1782±129 2967±285 1655HE 531±89.0 1036±112 545ΣCHLOR 2855±175 5627±574 3240p,p′-DDE 25 874±4795 25 614±2910 10 624p,p′-DDD 110±35.3 63.9±26.1 9.3p,p′-DDT 1183±187 359±115 111ΣDDT 27 168±4902 26 037±2925 10 744OCS 28.9±3.7 27.1±1.4 7.4Dieldrin 904±145 896±122 511Mirex 226±56.5 379±31.3 164ΣPCBa 9654±1303 11 848±1287 4877

The number of eggs (n) is given with the number of egg pools shown inbrackets.a Does not include non-ortho PCB congeners 77, 81, 126 and 169.

1976 egg samples was also analyzed. Organochlorineanalyses of the pooled samples included determination ofchlorobenzenes (ΣCBz=1,2,4,5-tetrachlorobenzene,1,2,3,4-tetrachlorobenzene pentachlorobenzene and hex-achlorobenzene), hexachlorocyclohexanes (ΣHCH=α-,β- and γ-hexachlorocyclohexane), chlordane-relatedcompounds (ΣCHLOR=oxychlordane, trans-chlordane,cis-chlordane, trans-nonachlor, cis-nonachlor and hepta-chlor epoxide), DDT and its metabolites (ΣDDT= p,p′-DDE, p,p′-DDD and p,p′-DDT), octachlorostyrene(OCS), mirex, dieldrin and PCB congeners (ΣPCB).ΣPCB consisted of 85 congeners identified according toIUPAC numbers (Ballschmiter et al., 1992) and includedcongener numbers 16/32, 17, 18, 19, 20/33, 22, 24/27, 25,26, 28/31, 40, 42, 44, 45, 47/48, 49, 52, 56/60, 64/41/71,66, 70/76, 74, 84, 85, 87, 91, 92, 95, 97, 99, 101, 105, 110,118, 128, 130, 134, 136, 137, 138, 141, 144/133, 146,149, 151, 153/132, 156, 157, 158, 170/190, 171, 172, 174,176, 177, 178, 179, 180, 183, 187/182, 185, 194, 195,196/203, 200, 199, 201, 202, 206, 207, and 208. Of these,PCB congeners 105, 118, 156 and 157 were mono-orthosubstituted PCBs (MO-PCBs). Congeners separated by aslash chromatographically co-eluted during the separationprocess and are therefore reported together.

Samples were analyzed for organochlorines at theGreat Lakes Institute for Environmental Research(GLIER) at the University of Windsor, Windsor, ON,according to CAEAL-accredited standard operatingprocedures (Environment Canada, 1989). Lipids weredetermined by gravimetric methods. Chemical extrac-tion and clean-up of PCBs and organochlorine pesticidesfollowed the procedures of Lazar et al. (1992). Briefly,tissue homogenates were ground, spiked with recoverysurrogate standards and extracted with dichloromethane:hexane (50:50% v/v). Sample clean-up was performedby gel permeation chromatography followed by activat-ed Florisil chromatography. Chemical analysis wasperformed using a Hewlett-Packard 5890 gas chro-matograph with 5973mass selective detector (GC-MSD)operated in the electron impact (EI) mode and usingselected ion monitoring (SIM). The column was a60 m×0.250 mm×0.1 μm DB-5 column. For everybatch of five samples injected, a MSD-PCB standardprepared from Aroclor 1242, 1254 and 1260 mixtures(Accu Standards, CT), two organochlorine pesticidestandards (Supelco, PA), a method blank and two in-house reference tissues (CWS-DCCO Reference EggPool; GLIER Detroit River fish homogenate) were alsoanalyzed. PCBs and organochlorine pesticides weredetermined using an external quantification approach asdescribed by Drouillard and Norstrom (2003), i.e., usingan average response factor for equivalent compounds in

Table 2Concentrations (pg g−1 lipid weight) and TEQs for PCDDs, PCDFs,and dioxin-like non-ortho PCBs (NO-PCBs) and mono-ortho PCBs(MO-PCBs) in eggs of ivory gulls collected between 1976 and 2004

Year 1976 1987 2004

n (# pools) 9 (1) 9 (1) 6 (1)% lipid 5.0 9.3 9.6

PCDDs2378-TCDD 97 97 2112378-PnCDD 27 25 7.2123478-HxCDD 3.6 2.8 1.4123678-HxCDD 23 18 5.4123789-HxCDD 4.3 b2.2 b1.01234678-HpCDD 13 b5.4 5.0OCDD 39 b2.7 16ΣPCDD 207 143 55

PCDFs2378-TCDF 1.5 0.8 1.012378-PnCDF 1.4 1.2 0.823478-PnCDF 33 38 13123478-HxCDF 4.8 4.5 2.3123678-HxCDF 9.3 7.3 2.2234678-HxCDF 5.9 4.6 2.4123789-HxCDF b2.3 1.5 b1.11234678-HpCDF 5.3 b2.0 2.3ΣPCDF 61 58 24

NO-PCBsPCB-77 339 178 195PCB-81 482 287 102PCB-126 4346 3376 1165PCB-169 1806 1694 760ΣNO-PCB 6973 5535 2221

TEQsPCDD-TEQ 125 122 28PCDF-TEQ 37 41 14NO-PCB-TEQ 502 377 137MO-PCB-TEQa 34 36 13ΣTEQ 697 576 193ΣTEQ (pg g−1 ww) 34 54 19

The number of eggs (n) is given for each pool.a PCB congeners 105, 118, 156 & 157.

406 B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

PCB or organochlorine secondary standard mixtures.The nominal detection limit was 0.3 ng g−1 wet weight(ww). Blanks and reference tissues, quantified duringeach batch of sample extractions, were in compliancewith the normal quality assurance procedures institutedby GLIER's CAEAL certified organic analytical labo-ratory. PCB and organochlorine standard recoveriesfrom samples and reference materials averaged 103±4.2% (mean±SE). The recovery efficiency was highand, therefore, residue concentrations were not recovery-corrected.

2.3. Analysis of PCDDs, PCDFs and non-ortho PCBs

Egg homogenates were analyzed as three annualpooled (composite) samples, i.e. one pool each for eggscollected in 1976, 1987 and 2004. Numbers of samplesin each pool varied according to availability (Table 2).Samples were analyzed at the Research and ProductivityCouncil (RPC) Laboratory, Fredericton, NB, for poly-chorinated dibenzo-p-dioxins (PCDDs), dibenzofurans(PCDFs) and non-ortho substituted polychlorinated bi-phenyls (NO-PCBs) using a method based on US EPAMethods 1613B and 8290A. Briefly, samples wereground with anhydrous sodium sulfate, transferred to achromatographic column and eluted with solvent. Ex-tracts were cleaned up and fractionated by gel perme-ation chromatography and alumina column clean-upfollowed by carbon/glass fibre column separation. Fur-ther separation of the PCDDs and PCDFs from the NO-PCBs was achieved using Florisil column chromatogra-phy. All samples were spiked with a solution containingspecific amounts of each of 15 isotopically-labelledPCDD/PCDF surrogates and four isotopically-labellednon-ortho PCB surrogates prior to extraction. Samplerecoveries for the surrogate standard averaged 89±1.2%(mean±SE). Samples were analyzed using a VGAutoSpec high-resolution mass spectrometer (HRMS)linked to a HP 5890 Series II high-resolution gas chro-matograph (HRGC) computerized data system run inSelected Ion Recording (SIR) mode. Matrix blanks, anin-house herring gull egg reference material, HGQA(Wakeford and Turle, 1997), provided by the CanadianWildlife Service, and a certified reference material,WMF-01 (freeze-dried fish), were run for quality con-trol. The results for both reference materials were withinacceptable ranges. All reported residue levels were cor-rected for surrogate standard recoveries. Minimumdetection limits were assessed for each sample and arereported in the results where relevant. Although1,2,3,4,7,8,9-HpCDF and OCDF were assayed,they were not detected in any of the samples (highest

minimum detection limits: 0.155 pg g−1 ww for1,2,3,4,7,8,9-HpCDF, 0.315 pg g−1 ww for OCDF)and are, therefore, not reported. The non-ortho PCBcongeners identified according to IUPAC numbers(Ballschmiter et al., 1992) were 77, 81, 126 and 169.

2.4. Analysis of brominated contaminants

Egg homogenates were analyzed as pooled sampleswith each pool consisting of three individual egg samples.Sample sizes for each year are given in Table 3. Sampleswere analyzed at the Great Lakes Institute for

Table 3Mean concentrations (ng g−1 lipid weight±standard error) of HBCD,PBDE and PBB congeners in eggs of ivory gulls collected between1976 and 2004

Year 1976 1987 2004

n (# pools) 9 (3) 9 (3) 6 (2)% lipid 4.8±0.6 8.7±0.4 10.3

BDE-17 0.34±0.34 0.48±0.25 1.8BDE-28 NQa 0.73±0.37 2.9BDE-30 0.48±0.08 0.52±0.06 0.78BDE-47 8.7±0.7 15.4±1.3 27.1BDE-49 NQ 0.12±0.12 0.56BDE-66 NQ NQ 0.59BDE-85 NQ NQ 0.39BDE-99 3.9±0.04 3.6±0.6 4.3BDE-100 2.4±0.8 1.8±0.2 2.6BDE-138 2.5±1.3 3.1±1.7 2.6BDE-153 0.09±0.05 0.26±0.06 0.81BDE-183 NQ NQ NQBDE-190 NQ NQ NQBDE-209 NQ NQ NQΣBDE 18.3±2.3 26.0±4.3 44.5Total-(α)-HBCDb 3.8±0.8 3.0±0.3 2.1BB-101 5.6±0.5 9.3±1.0 6.6BDE154/BB-153 c 12.2±0.9 20.5±1.0 9.6

The number of eggs (n) is given with the number of egg pools shown inbrackets.a Not Quantified (b0.01 ng g−1 ww; MLOQ).b α-HBCD is representative of total-HBCD (see text).c Predominantly BB-153 (see text).

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Environmental Research (GLIER; laboratory of R.Letcher) at the University of Windsor for polybrominateddiphenyl ethers (PBDEs), polybrominated biphenyls(PBBs) and total (α) hexabromocyclododecane (HBCD)based on methods for bird eggs described in Verreault etal. (2005, 2006), but based on sample extraction andclean-up methods previously described in Letcher et al.(1995). Briefly, samples were ground with anhydroussodium sulfate, spiked with internal standards andtransferred to a glass extraction column and eluted with50:50 CH2Cl2 (DCM)/n-hexane. Bulk lipid removal fromthe sample extract was accomplished by gel permeationchromatography, with further fraction clean-up andanalyte separation by Florisil (1.2% H2O deactivated)column chromatography.

Chemical analysis for 14 BDE congeners (BDE-17,-28, -30, -47, -49, -66, -85, -99, -100, -138, -153, -154(co-elution with BB-153), -183 and -190), BB-101 andtotal-(α)-HBCD was performed using an Agilent gaschromatograph (GC) equipped with a 5973 quadrupolemass spectrometer (MS) detector run in electron capturenegative ion (ECNI) mode. The GC column was a30 m×0.25 mm×0.25 μm HP-5MS capillary column.BDE-209 was analyzed separately on a similar 15 m GC

column using the same GC-MS(ECNI) analysis para-meters as described for the other 14 BDE congeners andtotal-(α)-HBCD. A more rapid GC elution was requiredfor the determination of BDE-209 to minimize the ana-lytical degradation (debromination) that occurs at highertemperatures. The determination of α-HBCD by GC-MS is representative of total-HBCD as it is generallyacknowledged that, starting in the injection port, any β-and γ-HBCD residues are thermally isomerized to α-HBCD at temperatures exceeding 160 °C.

Brominated compounds were identified on the basisof their retention times on the HP-5MS GC columnsrelative to authentic standards. For every batch of fivesamples injected, method blanks and two in-house ref-erence tissues (CWS-DCCO Reference Egg Pool andHGQA provided by the Canadian Wildlife Service)were run for quality control. The results for both ref-erence materials were within acceptable ranges. Anexternal standard solution was also run for every fivesamples to generate a relative response factor (RRF).Quantification of the brominated compounds was per-formed using an internal standard method based on therelative ECNI response factor of the 79Br+ 81Br anionsof BDE-71 and to that of authentic congener standardsin the neutral fractions. The method limit of quantifi-cation (MLOQ) was 0.01 ng g−1 ww. Recovery of theinternal standards averaged 62±18% (mean±SE) forBDE-71. All reported residue levels were corrected forinternal standard recoveries.

2.5. Stable isotope analysis

Freeze-dried homogenates for individual ivory gullegg samples were analyzed for stable nitrogen isotopes(15N/14N) by the Department of Earth Sciences at theUniversity of Ottawa, Ottawa, ON. Stable-nitrogenisotope assays were performed on 1 mg subsamples ofhomogenized materials by loading them into tin cupsand combusting them at 1800 °C in a Robo-Prep ele-mental analyzer. Resultant N2 gas was then analyzedusing an interfaced Europa 20:20 continuous-flow iso-tope ratio mass spectrometer (CFIRMS) with one labo-ratory standard run for every 8 samples analyzed.Stable-nitrogen abundance was expressed in δ notationas the deviation from standards in parts per thousand(‰) according to the following equation:

dX ¼ Rsample=Rstandard

� �−1

� �� 1000

where X is 15N and R is the corresponding ratio15N/14N. The Rstandard values were measured against theVienna PeeDee Belemnite (V-PDB) scale. Analytical

Fig. 2. Mean concentrations of ΣDDT, ΣPCB, ΣCBz and ΣCHL(ng g−1 lipid weight±standard error) in eggs of ivory gulls collectedbetween 1976 and 2004.

408 B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

precision, as determined by replicate measurements ofthe internal laboratory standards, was ±0.2o/oo.

2.6. Data handling

All residue concentrations are expressed on a lipidweight basis because the % lipid in the eggs collected in1976 were lower than those collected in 1987 and 2004(Tables 1 and 2). As well, the moisture content of the1976 samples was manipulated to overcome moistureloss during original temporary storage in formalin ren-dering wet weight concentrations unreliable. Due to thelimited number of total samples (n=8) and time periods(n=3), the data set was not considered large enough forrobust statistical analyses (see van Belle, 2002). Forcalculation of the sum of congeners for PCBs (ΣPCB),BDEs (ΣBDE), PCDDs (ΣPCDD), PCDFs (ΣPCDF),non-ortho PCBs (ΣNO-PCB), mono-ortho PCBs(ΣMO-PCB), and sum of organochlorine groups suchas HCHs (ΣHCH), chlordanes (ΣCHL), etc, a value ofzero was used for values below detection/quantificationlimits.

BDE-153 and BDE-154 have very similar physical–chemical characteristics (Tittlemier et al., 2002) andpublished data for biota show that BDE-154 generallyoccurs at lower or similar concentrations to BDE-153(e.g., Elliott et al., 2005; Hites, 2004; Norstrom et al.,2002). Based on a comparison of the BDE-153 andBDE-154/BB153 data in Table 3, BDE-154/BB153 waslikely comprised of over 90%BB153 and, therefore, wasnot included in the calculation of ΣBDE concentrations.

Toxic equivalents (TEQs) for non-ortho and mono-ortho PCBs, PCDDs and PCDFs were calculated bymultiplying congener concentrations by a congener-specific avian toxic equivalency factor (TEF) as re-commended by the World Health Organization (Van denBerg et al., 1998). The resulting TEQs were then summedto provide an indication of the toxic potential of thesample expressed as 2,3,7,8-TCDD toxic equivalents.

3. Results and discussion

3.1. Trophic level and migration

Stable isotopes of nitrogen (δ15N) can be used as anindicator of trophic position inmarine foodwebs (Hobsonet al., 2002) which allows us to evaluate trophic shifts indiet over time. Although the mean δ15N values in ivorygull eggs varied among years (Table 1), they did notchange in a consistent directional pattern suggesting thatany trends in residue concentrations over the time periodsampled were not directly attributable to changes in diet.

Unlike many other seabirds which breed in the Arctic andoverwinter in more temperate latitudes, the ivory gull israrely found far from drifting pack ice at any time of year.After the breeding season, the birds move offshore fromthe Canadian breeding colonies, eventually moving east-ward to overwinter in large areas of open water along theice edge from the Labrador Sea to Davis Strait in thenorthwest Atlantic (Haney and MacDonald, 1995).Therefore, the chemical residue patterns found in theivory gull reflect only the environmental contaminantspresent in arctic marine food chains.

3.2. Organochlorines

The major organochlorine groups found in the ivorygull eggs were ΣDDT (predominantly p,p′-DDE) fol-lowed by ΣPCB, ΣCHL (mainly oxychlordane) andΣCBz (mainly HCB) (Table 1). Although statistical an-alyses were not performed, it can be seen from Table 1and Fig. 2 that there have been decreases in concentra-tions of ΣDDT (and all three measured metabolites),ΣPCB, ΣCBz (and 1,2,3,4-TeCB, PnCB and HCB) aswell as OCS, dieldrin, and possibly mirex, from 1976and 1987 to 2004, whereas there was no discernibletrend in concentrations ofΣCHL orΣHCH. It is interest-ing to note that many compounds actually increased orchanged only minimally in concentration between 1976and 1987 (Table 1), an observation also made by Elliottet al. (1992). The reasons for this are unclear. The d15Nvalues did increase between 1976 and 1987 but ourlimited sample size did not allow us to correct for thispossible effect. The pattern of decline or no change formost organochlorines in ivory gull eggs between 1976and 2004 is very similar to that observed for other seabirdspecies breeding in the Canadian Arctic over that same

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time period (Braune, in press), as well as for other marinebiota from the Canadian Arctic such as ringed seals(Phoca hispida), beluga (Delphinapterus leucas) andpolar bears (Ursus maritimus) (Braune et al., 2005).

As a percentage of ΣCHL, oxychlordane decreased(from 63% to 51%) and trans-nonachlor increased (from12% to 26%) between 1976 and 2004 (Fig. 3). As well,α-HCH decreased from 49% to 9% and β-HCH increasedfrom 51% to 91% ofΣHCHover that time period (Fig. 3).γ-HCH was not detected in any of the samples. It shouldalso be noted that, proportionately, p,p′-DDE increasedfrom 95% to 99% between 1976 and 2004 as theproportions of p,p′-DDT and p,p′-DDD progressivelydecreased. The major PCB congeners found in the eggs,that is, those congeners averagingmore than 5%ofΣPCBoverall, were, in order of decreasing contribution: 153/132, 138, 180, 118 and 99. Cumulatively, these congenersaccounted for 69% (in 1976) to 74% (in 2004) of ΣPCB.

The proportional increase of p,p′-DDE and decreaseof p,p′-DDDand p,p′-DDTover the study period parallelsthat seen in northern fulmar eggs (Braune, in press) andsuggests greater exposure to DDT in the 1970s. Althougha ban onDDTusewent into effect in the U.S., Canada andmost western European countries in the early 1970s,Russia and other eastern European countries continuedtheir use of DDT during the 1970s and 1980s (AMAP,2004) whichmay have been the source of some continuedexposure to DDT delivered to the Canadian Arctic viaatmospheric transport (Macdonald et al., 2005).

Fig. 3. Mean percent contribution HCH isomers to ΣHCH, and of chlordane-1976 and 2004.

Oxychlordane was the major chlordane-related com-pound found in the ivory gull eggs which agrees withthe findings of Fisk et al. (2001b) for chlordane com-ponents in liver and fat of ivory gulls from northernBaffin Bay. However, given that oxychlordane is ametabolite of cis- and trans-chlordane, and thebiomagnification factor is generally considerablyhigher for oxychlordane than trans-nonachlor (Fisket al., 2001a; Hop et al., 2002), the decrease in oxy-chlordane and increase in trans-nonachlor as apercentage of ΣCHL between 1976 and 2004 maysuggest an increased exposure to chlordane. Theconcentration data (Table 1) support this speculation.Although chlordane use patterns were restricted in the1980s and finally banned world-wide in 1997 (AMAP,2004), chlordane is persistent in the environment andsome chlordane compounds (i.e. cis- and trans-chlordane) are highly volatile (AMAP, 1998) facilitat-ing atmospheric transport to the Arctic. Based ontemporal trends of chlordane isomers in arctic air,Bidleman et al. (2002) suggested that sources of chlor-dane have shifted from the freshly applied pesticide toa higher proportion of weathered residues, possiblyrecycled from soils, in recent years.

TheΣHCH composition in the ivory gull eggs shiftedfrom an almost equal contribution of α-HCH (49%) andβ-HCH (51%) in 1976 to predominantly β-HCH (91%)in 2004, a pattern which is consistent with other arcticseabird species (Braune, in press). HCH isomers are

related compounds to ΣCHL, in eggs of ivory gulls collected between

Fig. 4. Concentrations (pg g−1 lipid weight) of (A) ΣPCDD and ΣPCDF, and (B) TEQ contributions from PCDDs, PCDFs, non-ortho PCBs (NO-PCB) and mono-ortho PCBs (MO-PCB), in eggs of ivory gulls collected between 1976 and 2004.

Fig. 5. Concentrations (pg g−1 lipid weight) of PCDD and PCDFhomologs in eggs of ivory gulls collected between 1976 and 2004.

410 B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

subject to biotransformation, and seabirds appear toreadily metabolize the γ and α isomers whereas β-HCHis recalcitrant and biomagnifies in the food web (Borgået al., 2004; Hop et al., 2002; Moisey et al., 2001).

Overall, concentrations of the major organochlorinegroups in ivory gull eggs were somewhat higher (muchhigher in the case of ΣDDT) or similar to levels mea-sured over the same time period in eggs of thick-billedmurres (Uria lomvia), northern fulmars (Fulmarusglacialis) and black-legged kittiwakes (Rissa tridactyla)breeding on Prince Leopold Island in the CanadianArctic (Braune, in press). ΣDDT concentrations in theivory gull eggs from 1987 were similar to levels foundin eggs of glaucous gulls (Larus hyperboreus) collectedfrom Prince Leopold Island in the Canadian Arctic in1993 (Braune et al., 2002). Hepatic concentrations oforganochlorines measured by Buckman et al. (2004) in avariety of seabird species from northern Baffin Bay in1998 showed that levels in ivory gulls were most similarto glaucous gulls and northern fulmars even though, inthat study, glaucous gulls were shown to be feeding onetrophic level higher than ivory gulls and northernfulmars. Braune et al. (2006), however, showed thativory gulls from Seymour Island had a relatively highmean δ15N value compared with eggs of other seabirdscollected from the Canadian Arctic during 2003–2004.Ivory gulls, glaucous gulls, and northern fulmars will allopportunistically scavenge carrion, including marinemammals (Haney and MacDonald, 1995; Hatch andNettleship, 1998; Gilchrist, 2001), which could period-ically affect their contaminant intake. There is also someevidence that ivory gulls have a higher metabolic ratethan most other gulls as well as fulmars (Gabrielsen andMehlum, 1989) and, therefore, require a higher caloricintake. This could be achieved through consumption ofmore prey which could also increase contaminant

intake. Fisk et al. (2001a) demonstrated that the bio-magnification factors for organochlorines in the North-water Polynya (northern Baffin Bay) from arctic cod(Boreogadus saida) to ivory gulls were the highest (e.g.for ΣDDT, ΣPCB, ΣHCH, dieldrin, mirex) or secondhighest compared with other seabird species.

3.3. PCDDs and PCDFs

Concentrations of ΣPCDD in ivory gull eggs weregreater than ΣPCDF in all years sampled but bothPCDDs and PCDFs decreased from 1976 to 2004(Table 2, Fig. 4A). ΣTCDD (2,3,7,8-TCDD) dominatedthe PCDD pattern although OCDD was prominent in1976 and present in concentrations comparable to2,3,7,8-TCDD in 2004 (Fig. 5). ΣPnCDF (primarily2,3,4,7,8-PnCDF) followed by ΣHxCDF dominated thePCDF pattern (Fig. 5). OCDF and 1,2,3,4,7,8,9-HpCDFwere not detected in any of the samples. PCB-126 wasthe major non-ortho PCB congener in all years (Table 2)

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making up 52–62% of ΣNO-PCB followed by PCB-169 contributing 26–34% to ΣNO-PCB. PCB-118 wasthe predominant mono-ortho PCB congener in all years(73–75% of MO-PCB) followed by PCB-105 (19–21%of ΣMO-PCB). The non-ortho PCBs (primarily PCB-126) were the largest contributor (65–72%) to ΣTEQfollowed by the PCDDs (15–21%) in all years (Table 2,Fig. 4B).ΣTEQ concentrations decreased between 1976and 2004 without any major shift in the contributingfractions (Fig. 4B).

Decreases in ΣPCDD, ΣPCDF and ΣNO-PCB havealso been observed in other species of seabirds in theCanadian Arctic (Braune and Simon, 2003a) as well asfrom the Baltic Sea (de Wit et al., 1994). Addison et al.(2005), however, found no significant changes in con-centrations of PCDDs, PCDFs, NO-PCBs or MO-PCBsin ringed seals from the Canadian Arctic between 1981and 2000. Concentrations of ΣPCDD were 2–3 timeshigher (or more in 1976) thanΣPCDF in ivory gull eggswhereas ΣPCDF was higher than ΣPCDD in otherseabirds breeding in the Canadian Arctic (Braune andSimon, 2003a,b). Although the PCDF congener patternfound in the ivory gull eggs was similar to that found inarctic-breeding northern fulmars and thick-billed murreswith the predominance of the 2,3,4,7,8-PnCDF conge-ner (Braune and Simon, 2003b), PCDF concentrationsin ivory gull eggs were about an order of magnitudelower. Concentrations of ΣPCDD in the ivory gull eggsfrom 1976 and 1987 were comparable to concentrationsfound in eggs of arctic-breeding northern fulmars andblack-legged kittiwakes collected in 1993 (Braune andSimon, 2003b). The PCDD congener pattern in ivorygulls was, however, quite different from that seen in thefulmar and murre eggs where the 1,2,3,7,8-PnCDD and1,2,3,6,7,8-HxCDD congeners dominated the pattern(Braune and Simon, 2003b). The predominance of the2,3,7,8-TCDD and OCDD congeners in ivory gull eggsin 1976 and 2004 (Fig. 5) was more similar to the PCDDpattern found in marine mammals in the CanadianArctic (Norstrom et al., 1990; Addison et al., 2005). Asnoted by Norstrom et al. (1990), atmospheric depositionis the most likely source of TCDD and OCDD in theCanadian Arctic. 2,3,7,8-TCDD was the major PCDD/PCDF contaminant found in the herbicide 2,4,5-Twhichwas extensively used in the United States and otherindustrialized countries in the 1960s, and TCDD wasalso produced by pulp and paper mills using a chlorinebleaching process (Rice et al., 2003). A comparison ofsource to sink PCDD/PCDF congener profiles generatedby combustion sources identified OCDD as the domi-nant congener in the sink profile (Baker and Hites,2000) suggesting combustion processes as a source of

OCDD to the Arctic. The high biomagnification factorfor TCDD relative to other PCDD congeners (Brauneand Norstrom, 1989) would have enhanced the presenceof TCDD in the congener pattern found in the ivory gulleggs. Decreasing concentrations of PCDDs and PCDFsin the ivory gull eggs reflect legislated reductions inemissions/use. The relative increase of OCDD to TCDDin the 2004 ivory gull eggs (Fig. 5) may be related tothe continued photochemical generation in condensedatmospheric water of OCDD (and to a lesser extent,HpCDD) from pentachlorophenol (PCP), a widely usedwood preservative (Baker and Hites, 2000) in additionto combustion sources.

PCB-126 was the predominant non-ortho PCBcongener and PCB-118, the predominant mono-orthoPCB congener in ivory gulls. This same pattern for thecoplanar PCBs has been reported for glaucous gulls andblack guillemots (Cepphus grylle) from the Svalbardarea (Daelemans et al., 1992), black guillemots from theBaltic Sea (Koistinen et al., 1995), two species of cor-morants from the Canadian Pacific coast (Harris et al.,2003), five albatross species from the North Pacific andSouthern Oceans (Tanabe et al., 2004) as well as double-crested cormorants (Phalacrocorax auritus) and herringgulls (Larus argentatus) from the Great Lakes (Kannanet al., 2001). PCB-126 was also the predominant non-ortho PCB congener in northern fulmars, thick-billedmurres and black-legged kittiwakes from the CanadianArctic (Braune and Simon, 2003b). Concentrations ofPCB-126 in the ivory gull eggs from 1987 were verysimilar to those found in fulmar eggs collected from theCanadian Arctic in 1993 (Braune and Simon, 2003b).However, whereas the NO-PCBs provided the dominantcontribution to ΣTEQ consistently over all years in theivory gull eggs (Fig. 4B), the PCDF-TEQ had equal and,in the case of the fulmars, greater prominance in theΣTEQ of kittiwakes, fulmars and murres from theCanadian Arctic (Braune and Simon, 2003b). The pre-dominance of the NO-PCB-TEQ in ivory gull eggs,however, agrees well with the findings for double-crested cormorants and herring gulls from the GreatLakes (Kannan et al., 2001) and five albatross speciesfrom the North Pacific and Southern Oceans (Tanabeet al., 2004). Coplanar PCBs also contributed the maintoxic load in black guillemots in the Baltic Sea althoughPCB-118 was the major PCB-TEQ contributor in thatstudy (Koistinen et al., 1995).

3.4. Brominated compounds

Concentrations of ΣBDE steadily increased between1976 and 2004 driven primarily by increases in BDE-

Fig. 6. (A) Concentrations (ng g−1 lipid weight±standard error) of total-(α)-HBCD and ΣBDE in eggs of ivory gulls collected between 1976 and2004. (B) Percent contributions of BDE congeners to ΣBDE in eggs of ivory gulls collected between 1976 and 2004.

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47, whereas concentrations of total-(α)-HBCD de-creased (Table 3, Figure 6A). Except for a peak in1987, concentrations of BDE-154/BB-153 and BB-101did not change greatly between 1976 and 2004 (Table 3).Concentrations of the higher chlorinated congeners,BDE-183, -190 and -209, were b0.01 ng g−1 ww(MLOQ) in all samples. BDE-47 was the predominantBDE congener in all years followed by BDE-99, BDE-138 and BDE-100 (Table 3, Fig. 6B). However, thepercent contribution of BDE-47, as well as a number ofthe minor congeners (e.g. BDE-28, -153, -17, -49, -66,-85), to ΣBDE increased between 1976 and 2004 whileproportional contributions of BDE-99, BDE-138 andBDE-100 decreased (Fig. 6B).

The BDE data for ivory gulls agrees well with thepattern of significant increase seen during the past twodecades in North America (Hale et al., 2003; Law et al.,2003). Dramatic increases in concentrations of PBDEshave been documented for a variety of wildlife speciesincluding northern fulmars and thick-billed murres(Braune et al., 2005), ringed seals (Ikonomou et al.,2002) and beluga whales (Delphinapterus leucas) (Sternand Ikonomou, 2000) from the Canadian Arctic, double-crested cormorants and great blue herons (Ardeaherodias) from the Canadian Pacific coast (Elliottet al., 2005), herring gulls from the Great Lakes(Norstrom et al., 2002) and belugas from the St.Lawrence Estuary (Lebeuf et al., 2004), as well asperegrine falcons (Falco peregrinus) from South Green-land (Vorkamp et al., 2005).

Concentrations of ΣBDE in yolk sacs of kittiwakehatchlings sampled from Svalbard in the NorwegianArctic in 2002 (Murvoll et al., 2006a) were about 10times higher, and HBCD, 56 times higher, comparedwith ivory gull eggs from 2004. Likewise, concentra-

tions of ΣBDE in yolk sacs of European shag(Phalacrocorax aristotelis) hatchlings sampled fromthe Norwegian coast in 2002 (Murvoll et al., 2006b)were about 6 times higher, and HBCD, almost 200 timeshigher, compared with ivory gull eggs from 2004. Incontrast, ΣBDE was less than 2 times higher in blackguillemot eggs collected in 2001 from East Greenland(Vorkamp et al., 2004). ΣBDE concentrations in ivorygull eggs from 1976 and 1987 were about 4 to 9 timeshigher than those in thick-billed murre and northernfulmars eggs from the Canadian Arctic in 1975 and 1987(Braune et al., 2005).

Levels of BDE-47 in herring gull eggs from LakeOntario (Norstrom et al., 2002) were about 11 timeshigher than in ivory gull eggs in 1987 but had increasedto 84 times higher in 2000 compared with the 2004ivory gull data. However, concentrations of BDE-47 inivory gull eggs from 1976 were actually about 4 timeshigher than levels in double-crested cormorant eggsfrom 1979 from the Canadian Pacific coast (Elliott et al.,2005) but, by 2002, BDE-47 levels in cormorant eggswere about 9 times higher than those found in ivory gulleggs from 2004. These data suggest that BDE-47concentrations are increasing at a faster rate in avianwildlife at temperate North American latitudes com-pared with the Canadian Arctic. In contrast, the patternof change for BDE-47 and BDE-99 in guillemots/com-mon murres (Uria aalge) from the Baltic Sea showed aperiod of increase during the 1970s and into the late1980s followed by a rapid decrease during the 1990s(Sellström et al., 2003) which follows the general trendsdocumented for both Europe and Japan (Hale et al.,2003; Law et al., 2006). The pattern of change forHBCD was likewise different in the Baltic guillemotsshowing an increase during that same time period

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(Sellström et al., 2003). However, declines in HBCDwere observed in peregrine falcons from South Green-land (Vorkamp et al., 2005) which agree with our ob-servations for ivory gulls. This may be related to the factthat demand for HBCD is relatively low in NorthAmerica compared with Europe (de Wit, 2002).

Deca-BDE, Penta-BDE and Octa-BDE are the threemajor PBDE commercial mixtures commonly used inNorth America (Hale et al., 2003) and the pattern ofBDE congeners found in North American biota is areflection of source inputs and metabolism of the var-ious congeners found in these products. Although theDeca-BDE product (mainly BDE-209) is widely used,BDE-209 is less environmentally mobile and bioavail-able due to its physical–chemical properties (Boonet al., 2002; Wania and Dugani, 2003), which isprobably why it was not quantified in the ivory gulleggs. BDE-47 and BDE-99 are important componentsof the commercial Penta-BDE product used in polyure-thane foam, for which North America accounts for 98%of global demand (Hale et al., 2003). The large NorthAmerican reservoir of this product, in conjunction withthe relatively high arctic contamination potential ofthese congeners (Wania and Dugani, 2003), wouldaccount for the dominant presence of these BDEcongeners in the ivory gull egg pattern.

3.5. Toxicological significance

Concentrations of most of the organochlorine com-pounds in ivory gull eggs collected from SeymourIsland between 1976 and 2004 were below toxicologicalthreshold levels for eggs of wild birds (see reviews byBlus 1996, 2003; Hoffman et al., 1996; Peakall, 1996;Wiemeyer, 1996). Although concentrations of p,p′-DDE in the ivory gull eggs were relatively high (rangingfrom 774 to 2426 ng g−1 ww), those values were stillbelow the threshold level (3000 ng g−1 ww) associatedwith reproductive success in one of the more sensitiveavian species, the brown pelican (Pelecanus occidenta-lis) (Blus 1996, 2003). As well, the higher p,p′-DDEvalues in the ivory gulls were measured in samples from1987 and 1976, both on a wet weight and lipid weightbasis, whereas the more recent samples from 2004 hadconcentrations at the lower end of the range. Concen-trations of β-HCH in ivory gull eggs increased between1976 and 2004, and β-HCH is thought to have estro-genic effects (Willett et al., 1998). However, studiesmeasuring levels of HCH in eggs have not found anyassociated toxic effects (Wiemeyer, 1996) and our dataare about three orders of magnitude below those re-ported levels. Since most of the contaminants measured

in this study were either decreasing or not showingmuch change over time, those concentrations are notcurrently of toxicological concern although continuedmonitoring of some compounds such as DDE and β-HCH may be warranted.

A wide range of TEQ effects thresholds based onconcentrations of PCDDs, PCDFs and coplanar PCBshave been presented in the literature reflecting thevariation in species sensitivities and study criteria. Ivorygull eggs from all the years sampled exceeded the no-observed-effect-level (NOEL) of 7 pg g−1 TEQ ww ineggs for Great Lakes bald eagles (Giesy et al., 1995)which was, however, a conservative estimate and anorder of magnitude lower than the NOEL (100 pg g−1

TEQ ww in eggs) and LOEL (210 pg g−1 TEQ ww ineggs) suggested by Elliott et al. (1996) for bald eaglechicks. The ivory gull eggs from 1987 exceeded theNOEL (37 pg g−1 TEQ ww in eggs) but not the LOEL(130 pg g−1 TEQ ww in eggs) for CYP1A in Pacificcoast osprey (Pandion haliaetus) chicks (Elliott et al.,2001). Although all of the ivory gull ΣTEQs exceededthe reproductive NOELs for Caspian terns (Sternacaspia) (7.5 pg g−1 ww), herring gulls (10 pg g−1 ww)and double-crested cormorants (4.6 pg g−1 ww) reportedby Giesy et al. (1994) for the Great Lakes, the ivory gullΣTEQs were orders of magnitude lower than the TEQlevel (2175 pg g−1 ww) associated with effects onhatching success in Forster's tern (Sterna forsteri)(Kubiak et al., 1989) and TEQ LD50's for embryos ofCaspian terns, herring gulls and double-crested cormor-ants (750, 2000 and 460 pg g−1 ww, respectively) (Giesyet al., 1994). As well, wet weight concentrations of2,3,7,8-TCDD and PCB-126 in all of the ivory gull eggswere orders of magnitude lower than the experimentally-determined values for TCDD LD50 (4.0 ng g

−1 ww) andPCB-126 LD50 (177 ng g

−1 ww) for embryo mortality ineggs of double-crested cormorants (Powell et al., 1998).Given that the TEQs calculated for the ivory gulls do notappear critical compared with literature values, in con-junction with the trend towards declining TEQ levels, itis unlikely that current levels of PCDDs, PCDFs orcoplanar PCBs are affecting the reproductive success ofivory gulls in the Canadian Arctic.

The toxicological effects of PBBs and PBDEs arehypothesized to be similar to those of the structurally-related PCBs and dibenzodioxins (Pijnenburg et al., 1995)although Darnerud et al. (2001) suggest that the primarymechanisms of toxicity of PBDEs are probably differentthan those of dioxins. Very few studies, however, havebeen carried out to evaluate the exposure-effect relation-ship of these compounds in avian species. Several recentstudies (Fernie et al., 2005; Murvoll et al., 2006a,b) have

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begun to examine potential biochemical/physiologicaleffects in birds but the PBDE levels reported for thehatchlings from those studies were greater than the con-centrations found in the ivory gull eggs. Earlier studiesinvestigating the effects of PBB in chickens (Polin andRinger, 1978) and Japanese quail (Babish et al., 1975)found reproductive effects at egg concentrations whichwere several orders of magnitude higher than the levels ofPBBs found in the ivory gull eggs. However, given thescarcity of information on toxicity threshold levels forPBBs and PBDEs in avian species, coupled with the trendtoward increasing concentrations in Canadian Arcticspecies, continued monitoring and further toxicologicalstudies are warranted.

The levels of chlorinated and brominated compoundsfound in this study do not suggest that ivory gulls areexperiencing lethal exposure to these contaminants, butwe cannot rule out the possibility of synergistic/additive,sublethal effects as observed in other species (Sagerupet al., 2000; De Roode et al., 2002; Helberg et al., 2005).

Acknowledgements

We are grateful to S. MacDonald and D. Gill,formerly with the National Museum of Natural Sciencesin Ottawa, for making the egg collections in 1976 and1987, respectively. Sample preparation and archivalwere carried out by the Laboratory Services personnel atthe National Wildlife Research Centre (NWRC) inOttawa. We thank the Great Lakes Institute forEnvironmental Research (GLIER) at the University ofWindsor, Windsor, ON, for the organochlorine analyses,the RPC Laboratory in Fredicton, NB, for the PCDD/PCDF analyses, and S.-G. Chu and S. Shahmiri at theUniversity of Windsor and NWRC, respectively, for theanalyses of the brominated compounds. Stable isotopeanalyses for ivory gull eggs were carried out by P.Middlestead of the Department of Earth Sciences at theUniversity of Ottawa. Funding was provided by theCanadian Wildlife Service (Environment Canada) andthe Northern Ecosystem Initiative. Logistical supportout of Resolute Bay was provided by the PolarContinental Shelf Project, Natural Resources Canada.

References

Addison RF, Ikonomou MG, Fernandez MP, Smith TG. PCDD/F andPCB concentrations in Arctic ringed seals (Phoca hispida) have notchanged between 1981 and 2000. Sci Total Environ 2005;351–352:301–11.

AMAP. AMAP Assessment Report: Arctic Pollution Issues. ArcticMonitoring and Assessment Programme (AMAP), Oslo, Norway;1998. 859 pp.

AMAP. AMAP Assessment 2002: persistent organic pollutants in theArctic. Arctic Monitoring and Assessment Programme (AMAP),Oslo, Norway; 2004. 309 pp.

Babish JG, Gutenmann WH, Stoewsand GS. Polybrominatedbiphenyls: tissue distribution and effect on hepatic microsomalenzymes in Japanese quail. J Agric Food Chem 1975;23:879–82.

Baker JI, Hites RA. Is combustion the major source of polychlorinateddibenzo-p-dioxins and dibenzofurans to the environment? A massbalance investigation. Environ Sci Technol 2000;34:2879–86.

Ballschmiter K, Bacher R, Mennel A, Fischer R, Riehle U, Swerev M.The determination of chlorinated biphenyls, chlorinated dibenzo-dioxins, and chlorinated dibenzofurans by GC-MS. J High ResolChromatogr 1992;15:260–70.

Bidleman TF, Jantunen LMM, Helm PA, Brorström-Lunden E, JunttoS. Chlordane enantiomers and temporal trends of chlordaneisomers in arctic air. Environ Sci Technol 2002;36:539–44.

Blus LJ. DDT, DDD, and DDE in birds. In: Beyer WN, Heinz GH,Redmon-Norwood AW, editors. Environmental Contaminants inWildlife: Interpreting Tissue Concentrations. SETAC SpecialPublication Series. Boca Raton, FL: CRC Press; 1996. p. 49–71.

Blus LJ. Organochlorine pesticides. In: Hoffman DJ, Rattner BA,Burton Jr GA, Cairns Jr J, editors. Handbook of Ecotoxicology.Second edition. Boca Raton, FL: CRC Press; 2003. p. 313–39.

Boon JP, Lewis WE, Tjoen-a-Choy MR, Allchin CR, Law RJ, de BoerJ, et al. Levels of polybrominated diphenyl ether (PBDE) flameretardants in animals representing different trophic levels of theNorth Sea food web. Environ Sci Technol 2002;36:4025–32.

Borgå K, Fisk AT, Hoekstra PF, Muir DCG. Biological and chemicalfactors of importance in the bioaccumulation and trophic transferof persistent organochlorine contaminants in Arctic marine foodwebs. Environ Toxicol Chem 2004;23:2367–85.

Braune B.M. Temporal trends of organochlorines and mercury inseabird eggs from the Canadian Arctic, 1975 to 2003. EnvironPollut; in press.

Braune BM, Norstrom RJ. Dynamics of organochlorine compounds inherring gulls: III. Tissue distribution and bioaccumulation in LakeOntario gulls. Environ Toxicol Chem 1989;8:957–68.

Braune BM, Simon M. Dioxins and furans in Canadian Arcticseabirds. Organohalog Compd 2003a;61:317–20.

Braune BM, Simon M. Dioxins, furans, and non-ortho PCBs inCanadian Arctic seabirds. Environ Sci Technol 2003b;37:3071–7.

Braune BM, Donaldson GM, Hobson KA. Contaminant residues inseabird eggs from theCanadianArctic. I. Temporal trends 1975–1998.Environ Pollut 2001;114:39–54.

Braune BM, Donaldson GM, Hobson KA. Contaminant residues inseabird eggs from the Canadian Arctic. II. Spatial trends andevidence from stable isotopes for intercolony differences. EnvironPollut 2002;117:133–45.

BrauneBM,Outridge PM, FiskAT,MuirDCG,HelmPA,HobbsK, et al.Persistent organic pollutants and mercury in marine biota of theCanadian Arctic: an overview of spatial and temporal trends. SciTotal Environ 2005;351–352:4–56.

Braune BM, Mallory ML, Gilchrist HG. Elevated mercury levels in adeclining population of ivory gulls in the Canadian Arctic. MarPollut Bull 2006;52:969–87.

Buckman AH, Norstrom RJ, Hobson KA, Karnovsky NJ, Duffe J, FiskAT. Organochlorine contaminants in seven species of Arcticseabirds from northern Baffin Bay. Environ Pollut2004;128:327–38.

COSEWIC. Press Release. Cypress Hills Interprovincial Park, Saskatch-ewan,May 1, 2006, Committee on the Status of EndangeredWildlifein Canada (COSEWIC), 2006; www.cosewic.gc.ca.

415B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

Daelemans FF, Mehlum F, Schepens PJC. Polychlorinated biphenylsin two species of arctic seabirds from the Svalbard area. BullEnviron Contam Toxicol 1992;48:828–34.

Darnerud PO, Eriksen GS, Johannesson T, Larsen PB, Viluksela M.Polybrominated diphenyl ethers: occurrence, dietary exposure, andtoxicology. Environ Health Perspect 2001;109:49–68.

DeRoodeDF,GustavssonMB,RantalainenAL,KlompAV,Koeman JH,Bosveld ATC. Embryotoxic potential of persistent organic pollutantsextracted from tissues of guillemots (Uria aalge) from the Baltic Seaand the Atlantic Ocean. Environ Toxicol Chem 2002;21:2401–11.

de Wit CA. An overview of brominated flame retardants in theenvironment. Chemosphere 2002;46:583–624.

de Wit CA, Järnberg UG, Asplund LT, Jansson B, Olsson M, Odsjö T,et al. The Swedish Dioxin Survey: summary of results fromPCDD/F and coplanar PCB analyses in biota. Organohalog Compd1994;20:47–50.

Drouillard KG, Norstrom RJ. The influence of diet properties andfeeding rates on PCB toxicokinetics in the ring dove. Arch EnvironContam Toxicol 2003;44:97–106.

Elliott JE, NobleDG, NorstromRJ,Whitehead PE, SimonM, Pearce PA,et al. Patterns and trends of organic contaminants in Canadian seabirdeggs, 1968–90. In: Walker CH, Livingstone DR, editors. PersistentPollutants inMarine Ecosystems. SETACSpecial Publication Series.Oxford, U.K.: Pergamon Press; 1992. p. 181–94.

Elliott JE, Norstrom RJ, Lorenzen A, Hart LE, Philibert H, KennedySW, et al. Biological effects of polychlorinated dibenzo-p-dioxins,dibenzofurans, and biphenyls in bald eagle (Haliaeetus leucoce-phalus) chicks. Environ Toxicol Chem 1996;15:782–93.

Elliott JE, Wilson LK, Henny CJ, Trudeau SF, Leighton FA, KennedySW, et al. Assessment of biological effects of chlorinated hydro-carbons in osprey chicks. Environ Toxicol Chem 2001;20:866–79.

Elliott JE, Wilson LK, Wakeford B. Polybrominated diphenyl ethertrends in eggs of marine and freshwater birds from BritishColumbia, Canada, 1979–2002. Environ Sci Technol 2005;39:5584–91.

Environment Canada. Analytical methods manual. Water QualityBranch, Environment Canada, Ottawa; 1989.

Fernie KJ,MayneG, Shutt JL, PekarikC,GrasmanKA, Letcher RJ, et al.Evidence of immunomodulation in nestling American kestrels(Falco sparverius) exposed to environmentally relevant PBDEs.Environ Pollut 2005;138:485–93.

Fisk AT, Hobson KA, Norstrom RJ. Influence of chemical andbiological factors on trophic transfer of persistent organicpollutants in the Northwater Polynya marine food web. EnvironSci Technol 2001a;35:732–8.

Fisk AT, Moisey J, Hobson KA, Karnovsky NJ, Norstrom RJ.Chlordane components and metabolites in seven species of Arcticseabirds from the Northwater Polynya: relationships with stableisotopes of nitrogen and enantiomeric fractions of chiralcomponents. Environ Pollut 2001b;113:225–38.

French MC, Jefferies DJ. The preservation of biological tissue fororganochlorine insecticide analysis. Bull Environ Contam Toxicol1971;6:460–3.

Gabrielsen GW, Mehlum F. Thermoregulation and energetics of arcticseabirds. In: Bech C, Reinertsen RE, editors. Physiology of ColdAdaptation in Birds. New York: Pergamon Press; 1989. p. 137–45.

Giesy JP, Ludwig JP, Tillitt DE. Dioxins, dibenzofurans, PCBs andcolonial, fish-eating water birds. In: Schecter A, editor. Dioxinsand Health. New York: Plenum Publishers; 1994. p. 249–307.

Giesy JP, Bowerman WW, Mora MA, Verbrugge DA, Othoudt RA,Newsted JL, et al. Contaminants in fishes from Great Lakes-influenced sections and above dams of three Michigan Rivers: III.

Implications for health of bald eagles. Arch Environ ContamToxicol 1995;29:309–21.

Gilchrist HG. Glaucous Gull (Larus hyperboreus). In: Poole A, Gill F,editors. The Birds of North America, No. 573. Philadelphia, PA:The Birds of North America, Inc.; 2001. 32 pp.

Gilchrist HG, Mallory ML. Declines in abundance and distribution ofthe Ivory Gull (Pagophila eburnea) in Arctic Canada. BiolConserv 2005;121:303–9.

Hale RC, Alaee M, Manchester-Neesvig JB, Stapleton HM, IkonomouMG. Polybrominated diphenyl ether flame retardants in the NorthAmerican environment. Environ Int 2003;29:771–9.

Haney JC, MacDonald SD. Ivory Gull (Pagophila eburnea). In: PooleA, Gill F, editors. The Birds of North America, No. 175. TheAcademy of Natural Sciences, Philadelphia, PA, and TheAmerican Ornithologists’ Union, Washington, D.C., 1995, 24 pp.

Harris ML, Wilson LK, Norstrom RJ, Elliott JE. Egg concentations ofpolychlorinated dibenzo-p-dioxins and dibenzofurans in double-crested (Phalacrocorax auritus) and pelagic (P. pelagicus)cormorants from the Strait of Georgia, Canada, 1973–1998.Environ Sci Technol 2003;37:822–31.

Hatch SA, Nettleship DN. Northern Fulmar (Fulmarus glacialis). In:Poole A, Gill F, editors. The Birds of North America, No. 361.Philadelphia, PA: The Birds of North America Inc.; 1998. 32 pp.

Hebert CE, Shutt JL, Hobson KA,Weseloh DVC. Spatial and temporaldifferences in the diet of Great Lakes herring gulls (Larusargentatus): evidence from stable isotope analysis. Can J FishAquat Sci 1999;56:323–38.

Helberg M, Bustnes JO, Erikstad KE, Kristiansen KO, Skaare JU.Relationships between reproductive performance and organochlo-rine contaminants in great black-backed gulls (Larus marinus).Environ Pollut 2005;134:475–83.

Hites RA. Polybrominated diphenyl ethers in the environment and inpeople: a meta-analysis of concentrations. Environ Sci Technol2004;38:945–56.

Hobson KA. Reconstructing avian diets using stable-carbon andnitrogen isotope analysis of egg components: patterns of isotopicfractionation and turnover. Condor 1995;97:752–62.

Hobson KA, Fisk A, Karnovsky N, Holst M, Gagnon JM, Fortier M. Astable isotope (δ13C, δ15N) model for the North Water food web:implications for evaluating trophodynamics and the flow of energyand contaminants. Deep-Sea Res Part II 2002;49:5131–50.

Hoffman DJ, Rice CP, Kubiak TJ. PCBs and dioxins in birds. In:Beyer WN, Heinz GH, Redmon-Norwood AW, editors. Environ-mental Contaminants in Wildlife: Interpreting Tissue Concentra-tions. SETAC Spec Publ Ser. Boca Raton, FL: CRC Press; 1996.p. 165–207.

Hop H, Borgå K, Gabrielson GW, Kleivane L, Skaare JU. Food webmagnification of persistent organic pollutants in poikilotherms andhomeotherms from the Barents Sea. Environ Sci Technol 2002;36:2589–97.

Ikonomou MG, Rayne S, Addison RF. Exponential increases of thebrominated flame retardants, polybrominated diphenyl ethers, inthe Canadian Arctic from 1981 to 2000. Environ Sci Technol2002;36:1886–92.

Kaehler S, Pakhomov EA. Effects of storage and preservation on theδ13C and δ15N signatures of selected marine organisms. Mar EcolProgr Ser 2001;219:299–304.

Kannan K, Hilscherova K, Imagawa T, Yamashita N, Williams LL,Giesy JP. Polychlorinated naphthalenes, -biphenyls, -dibenzo-p-dioxins, and –dibenzofurans in double-crested cormorants andherring gulls from Michigan waters of the Great Lakes. EnvironSci Technol 2001;35:441–7.

416 B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

Koistinen J, Koivusaari J, Nuuja I, Paasivirta J. PCDEs, PCBs, PCDDsand PCDFs in black guillemots and white-tailed sea eagles fromthe Baltic Sea. Chemosphere 1995;30:1671–84.

Kubiak TJ, Harris HJ, Smith LM, Schwartz TR, Stalling DL, Trick JA,et al. Microcontaminants and reproductive impairment of theForster's tern on Green Bay, Lake Michigan — 1983. ArchEnviron Contam Toxicol 1989;18:706–27.

Law RJ, Alaee M, Allchin CR, Boon JP, Lebeuf M, Lepom P, et al.Levels and trends of polybrominated diphenyl ethers and otherbrominated flame retardants in wildlife. Environ Int 2003;29:757–70.

Law RJ, Allchin CR, de Boer J, Covaci A, Herzke D, Lepom P, et al.Levels and trends of brominated flame retardants in the Europeanenvironment. Chemosphere 2006;64:187–208.

Lazar R, Edwards RC, Metcalfe CD, Metcalfe T, Gobas FAPC,Haffner GD. A simple, novel method for the quantitative analysisof coplanar (non-ortho substituted) polychlorinated biphenyls inenvironmental samples. Chemosphere 1992;25:493–504.

Lebeuf M, Gouteux B, Measures L, Trottier S. Levels and temporaltrends (1988–1999) of polybrominated diphenyl ethers in belugawhales (Delphinapterus leucas) from the St. Lawrence Estuary,Canada. Environ Sci Technol 2004;38:2971–7.

Letcher RJ, Norstrom RJ, Bergman A. An integrated analytical methodfor determination of polychlorinated aryl methyl sulfone metabo-lites and polychlorinated hydrocarbon contaminants in biologicalmatrices. Anal Chem 1995;67:4155–63.

Loganathan BG, Tanabe S, Hidaka Y, Kawano M, Hidaka H, TatsukawaR. Temporal trends of persistent organochlorine residues in humanadipose tissue from Japan, 1928–1985. Environ Pollut 1993;81:31–9.

Macdonald RW, Barrie LA, Bidleman TF, Diamond ML, Gregor DJ,Semkin RG, et al. Contaminants in the Canadian Arctic: 5 years ofprogress in understanding sources, occurrence and pathways. SciTotal Environ 2000;254:93–234.

Macdonald RW, Harner T, Fyfe J. Recent climate change in the Arcticand its impact on contaminant pathways and interpretation oftemporal trend data. Sci Total Environ 2005;342:5–86.

MacGregor JS. Changes in the amount and proportions of DDT and itsmetabolites, DDE and DDD, in the marine environment offsouthern California, 1949–72. Fish Bull 1974;72:275–93.

Moisey J, Fisk AT, Hobson KA, Norstrom RJ. Hexachlorocyclohexane(HCH) isomers and chiral signatures of α-HCH in the Arcticmarine food web of the Northwater Polynya. Environ Sci Technol2001;35:1920–7.

Murvoll KM, Skaare JU, Moe B, Anderssen E, Jenssen BM. Spatialtrends and associated biological responses of organochlorines andbrominated flame retardants in hatchlings ofNorthAtlantic kittiwakes(Larus tridactyla). Environ Toxicol Chem 2006a;25:1648–56.

Murvoll KM, Skaare JU, Anderssen E, Jenssen BM. Exposure andeffects of persistent organic pollutants in European shag(Phalacrocorax aristotelis) hatchlings from the coast of Norway.Environ Toxicol Chem 2006b;25:190–8.

Neidermyer WJ, Hickey JJ. Chronology of organochlorine compoundsin Lake Michigan fish, 1929–66. Pest Monit J 1976;10:92–5.

Norstrom RJ, Simon M, Muir DCG. Polychlorinated dibenzo-p-dioxins and dibenzofurans in marine mammals in the CanadianNorth. Environ Pollut 1990;66:1–19.

Norstrom RJ, Simon M, Moisey J, Wakeford B, Weseloh DV.Geographical distribution (2000) and temporal trends (1981–2000)of brominated diphenyl ethers in Great Lakes herring gull eggs.Environ Sci Technol 2002;36:4783–9.

Peakall DB. Dieldrin and other cyclodiene pesticides in wildlife. In:Beyer WN, Heinz GH, Redmon-Norwood AW, editors. Envi-

ronmental Contaminants in Wildlife: Interpreting Tissue Con-centrations. SETAC Spec Publ Ser. Boca Raton, FL: CRC Press;1996. p. 73–97.

Pijnenburg AMCM, Everts JW, de Boer J, Boon JP. Polybrominatedbiphenyl and diphenylether flame retardants: analysis, toxicity, andenvironmental occurrence. Rev Environ Contam Toxicol 1995;141:1–26.

Polin D, Ringer RK. PBB fed to adult female chickens: its effect onegg production, reproduction, viability of offspring, and residuesin tissues and eggs. Environ Health Perspect 1978;23:283–90.

Powell DC, Aulerich RJ, Meadows JC, Tillitt DE, Kelly ME,Stromborg KL, et al. Effects of 3,3′,4,4′,5-pentachlorobiphenyland 2,3,7,8-tetrachlorodibenzo-p-dioxin injected into the yolks ofdouble-crested cormorant (Phalacrocorax auritus) eggs prior toincubation. Environ Toxicol Chem 1998;17:2035–40.

Renaud CB, Kaiser KLE, Comba ME. Historical versus recent levelsof organochlorine contaminants in lamprey larvae of the St.Lawrence River basin, Québec. Can J Fish Aquat Sci 1995;52:268–75.

Renaud CB, Comba ME, Kaiser KLE. Temporal trend of organochlo-rine contaminant levels in the northeastern part of Lake Superiorbasin based on lamprey larvae lipid burdens. J Great Lakes Res1999;25:918–29.

Renaud CB, Alfonso N, Kaiser KLE, Comba ME. A comparison oforganochlorine contaminant levels, on a lipid basis, in sea lampreylarvae fromMad River, Lake Huron basin and Michipicoten River,Lake Superior basin (1960–1976). Water Qual Res J Can2004;39:75–82.

Rice CP, O'Keefe PW, Kubiak TJ. Sources, pathways, and effects ofPCBs, dioxins, and dibenzofurans. In: Hoffman DJ, Rattner BA,Burton Jr GA, Cairns Jr J, editors. Handbook of Ecotoxicology.Second Edition. Boca Raton, FL: CRC Press; 2003. p. 501–73.

Sagerup K, Henriksen EO, Skorping A, Skaare JU, Gabrielsen GW.Intensity of parasitic nematodes increases with organochlorinelevels in the glaucous gull. J Appl Ecol 2000;37:532–9.

Sarakinos HC, Johnson ML, Vander Zanden MJ. A synthesis of tissuepreservation effects on carbon and nitrogen stable isotopesignatures. Can J Zool 2002;80:381–7.

Sellström U, Bignert A, Kierkegaard A, Häggberg L, de Wit CA,Olsson M, et al. Temporal trend studies on tetra- and pentabro-minated diphenyl ethers and hexabromocyclododecane in guille-mot egg from the Baltic Sea. Environ Sci Technol 2003;37:5496–501.

Stern GA, Ikonomou M. Temporal trends of polybrominatedbiphenyls and polybrominated and polychlorinated diphenylethers in southeast Baffin beluga. In: Kalhok S, editor. Synopsisof Research Conducted Under the 1999–2000 Northern Con-taminants Program. Indian and Northern Affairs Canada, Ottawa;2000. p. 227–32.

Tanabe S,Watanabe M, Minh TB, Kunisue T, Nakanishi S, Ono H, et al.PCDDs, PCDFs, and coplanar PCBs in albatross from the NorthPacific and Southern Oceans: levels, patterns, and toxicologicalimplications. Environ Sci Technol 2004;38:403–13.

Tittlemier SA, Halldorson T, Stern GA, Tomy GT. Vapor pressures,aqueous solubilities, and Henry's law constants of some brominat-ed flame retardants. Environ Toxicol Chem 2002;21: 1804–10.

van Belle G. Statistical Rules of Thumb. New York: John Wiley &Sons; 2002. 221 pp.

Van den Berg M, Birnbaum L, Bosveld ATC, Brunström B, Cook P,Feeley M, et al. Toxic equivalency factors (TEFs) for PCBs,PCDDs, PCDFs for humans and wildlife. Environ Health Perspect1998;106:775–92.

417B.M. Braune et al. / Science of the Total Environment 378 (2007) 403–417

Verreault J, Letcher RJ, Muir DCG, Chu S, Gebbink WA, GabrielsenGW. New organochlorine contaminants and metabolites in plasmaand eggs of glaucous gulls (Larus hyperboreus) from theNorwegian Arctic. Environ Toxicol Chem 2005;24:2486–99.

Verreault J, Villa RA, Gabrielsen GW, Skaare JU, Letcher RJ. Maternaltransfer of organohalogen contaminants and metabolites to eggsof Arctic-breeding glaucous gulls. Environ Pollut 2006;144:1053–60.

Vorkamp K, Christensen JH, Riget F. Polybrominated diphenyl ethersand organochlorine compounds in biota from the marine environ-ment of East Greenland. Sci Total Environ 2004;331: 143–55.

Vorkamp K, Thomsen M, Falk K, Leslie H, Møller S, Sørensen PB.Temporal development of brominated flame retardants in peregrinefalcon (Falco peregrinus) eggs from South Greenland (1986–2003).Environ Sci Technol 2005;39:8199–206.

Wakeford B, Turle R. In-house reference materials as a means toquality assurance: the Canadian Wildlife Service experience. In:Clement RE, Keith LH, Siu KWM, editors. Reference Materialsfor Environmental Analysis. Boca Raton, FL: CRC Press; 1997.p. 205–31.

Wania F, Dugani CB. Assessing the long-range transport potential ofpolybrominated diphenyl ethers: a comparison of four multimediamodels. Environ Toxicol Chem 2003;22:1252–61.

Wiemeyer SN. Other organochlorine pesticides in birds. In: BeyerWN, Heinz GH, Redmon-Norwood AW, editors. Environmen-tal Contaminants in Wildlife: Interpreting Tissue Concentra-tions. SETAC Spec Publ Ser. Boca Raton, FL: CRC Press;1996. p. 99–115.

Wiemeyer SN, Moore JF, Mulhern BM. Formalin preservation ofavian blood for metal and DDE analysis. Bull Environ ContamToxicol 1984;33:525–32.

Willett KL, Urlrich EM, Hites RA. Differential toxicity andenvironmental fates of hexachlorocyclohexane isomers. EnvironSci Technol 1998;32:2197–207.

Wolkers H, van Bavel B, Derocher AE,Wiig Ø, Kovacs KM, LydersenC, et al. Congener-specific accumulation and food chain transfer ofpolybrominated diphenyl ethers in two arctic food chains. EnvironSci Technol 2004;38:1667–74.