persistent organic pollutants in marine birds, arctic hare and ringed seals near qikiqtarjuaq,...
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Baseline
Persistent organic pollutants in marine birds, arctic hare andringed seals near Qikiqtarjuaq, Nunavut, Canada
Mark L. Mallory a,*, Birgit M. Braune b, Mark Wayland c, Ken G. Drouillard d
a Canadian Wildlife Service, P.O. Box 1714, Iqaluit, NU, Canada X0A 0H0b Canadian Wildlife Service, NWRC, Raven Road, Carleton University, Ottawa, ON, Canada K1A 0H3
c Canadian Wildlife Service, 115 Perimeter Road, Saskatoon, SK, Canada S7N 0X4d Great Lakes Institute for Environmental Research, University of Windsor, 401 Sunset Avenue, Windsor, ON, Canada N9B 3P4
The contamination of Canadian Arctic marine wild-
life by persistent organic pollutants (POPs) has beenknown since the early 1970s (reviewed in Muir et al.,
1992, 1999; Jensen et al., 1997; AMAP, 1998). These
compounds typically arrive via long range transport
from source areas in tropical or temperate regions,
and are deposited in the Arctic by condensing, cool,
polar air (Bright et al., 1995; Jensen et al., 1997). Once
assimilated into the lower parts of marine food chains,
these lipophilic compounds biomagnify with increasingtrophic level, and can reach concentrations in top pred-
ators that can be deleterious to reproduction or survival
(Jensen et al., 1997; Dietz et al., 2000; Borga et al., 2001;
Bustnes et al., 2003).
After the initial discovery of POPs in Arctic wildlife
(once thought to be too distant from sources to be pol-
luted), monitoring programs were established to track
contaminant levels in various wildlife groups. Becausethey are long-lived, forage over large areas, and are
often colonial and thus easy to sample, marine birds
integrate many aspects of local marine ecosystem condi-
tions and have proven to be particularly effective bio-
indicators of marine pollution (e.g. Braune et al., 2002).In the Canadian Arctic, concentrations of many POPs
in marine birds decreased from the 1970s through the
1980s, and then levelled off by the 1990s (Muir et al.,
1999; Braune et al., 2001). Despite a general pattern of
decline, however, there remain clear regional differences,
with some sites continuing to support wildlife with con-
siderably higher contaminant concentrations than other
locales (Muir et al., 1999; Braune et al., 2002). Manygaps remain in our geographic knowledge of contami-
nant levels in Arctic wildlife tissues.
Contamination of Arctic wildlife has been a key con-
cern to Inuit (the aboriginal residents of the Canadian
Arctic), because harvest of wild animals is the principal
source of meat and an important cultural activity (Jen-
sen et al., 1997). During our seabird research near two
large colonies southeast of Qikiqtarjuaq (formerlyBroughton Island), Nunavut (Fig. 1) on eastern Baffin
Island, the community expressed concern over possible
contamination of wildlife near the colonies, which were
located close to two former military sites (Durban and
Padloping islands). To address community concerns,
evaluate the possibility of point-source contamination
in the local marine food chain, and fill in a large geo-
graphic gap for contaminant data in Nunavut, we
Edited by Bruce J. Richardson
The objective of BASELINE is to publish short communications on different aspects of pollution of the marineenvironment. Only those papers which clearly identify the quality of the data will be considered for publication.
Contributors to Baseline should refer to ‘Baseline—The New Format and Content’ (Mar. Pollut. Bull. 42, 703–704).
* Corresponding author. Tel.: +1 867 975 4637; fax: +1 867 975
4645.
E-mail address: [email protected] (M.L. Mallory).
www.elsevier.com/locate/marpolbul
Marine Pollution Bulletin 50 (2005) 95–104
sampled marine birds, arctic hare (Lepus arctica) and
ringed seals (Pusa hispida) near Qikiqtarjuaq to deter-mine their POP levels.
Tissue samples were collected 100km southeast of
Qikiqtarjuaq during 10–14 September 2001 (with the
exception of seals, n = 2, collected from local hunters
in October 2001). Target wildlife species were collected
using a 12 gauge shotgun, firing 70mm, #4 steel shot-
shells. The species were chosen to represent a range of
trophic levels or food webs, with glaucous gulls (Larushyperboreus; n = 1) and northern fulmars (Fulmarus gla-
cialis; n = 2) representing top marine predators and
scavengers, black guillemots (Cepphus grylle; n = 8) rep-
resenting nearshore piscivores, common eiders (Somate-
ria mollissima borealis; n = 8) representing benthic
molluscivores, and arctic hare (n = 2) representing
terrestrial herbivores. Most bird samples were of adults,
although one eider and six guillemots were juveniles.Whole carcasses were placed in individual bags, labelled
with the time and location of the collection, and shipped
frozen to Iqaluit where carcasses were partially thawed
to permit measurement and dissection of �5g breast
muscle and liver samples which were wrapped in hex-
ane-rinsed foil, refrozen and shipped to the Great Lakes
Institute for Environmental Research (GLIER) for POP
analyses.POP analyses of sample tissues included determina-
tion of chlorobenzenes (P
CBz = 1,2,4,5-tetrachloroben-
zene, 1,2,3,4-tetrachlorobenzene, pentachlorobenzene
and hexachlorobenzene), hexachlorocyclohexanes
(P
HCH = a-, b-, and c-hexachlorocyclohexane),chlordane-related compounds (
PCHLOR = oxychlor-
dane, trans-chlordane, cis-chlordane, trans-nonachlor,
cis-nonachlor and heptachlor epoxide), DDT and itsmetabolites (
PDDT = p,p0-DDE, p,p0-DDD and
p,p0-DDT), mirex (P
MIREX = photomirex and mirex),
dieldrin, and PCBs (P
PCB = 78 congeners identified
according to IUPAC numbers (Ballschmiter and Zell,
1980): 16/32, 17, 18, 19, 20/33, 22, 24/27, 25, 26, 28, 31,
40, 42, 44, 45, 49, 52, 60, 64, 66/95, 70, 74, 85, 87, 91,
92, 95, 97, 99, 101, 105, 110, 118, 129, 130, 132, 134,
135/144, 136, 137, 138, 141, 146, 149, 151, 153, 156,157, 158, 170/190, 171, 172, 174, 176, 177, 178, 179,
180, 182/187, 183, 185, 194, 195, 199, 200, 201, 202,
203, 206, 207, and 208).
All samples were analyzed according to CAEAL-
accredited standard operating procedures (Environment
Canada, 1989). Chemical extraction and cleanup of
PCBs and organochlorine pesticides followed the proce-
dures of Lazar et al. (1992). Briefly, tissue homogenateswere ground and spiked with 1,3,5-tribromobenzene as a
surrogate recovery standard and extracted with 350ml
of dichloromethane: hexane (50:50% v/v, OmniSolve-
Grade, VWR, ON, Canada). Cleanup of the sample
was performed by gel permeation chromatography fol-
lowed by activated Florisil (VWR, ON, Canada) chro-
matography. Chemical analysis was performed using a
Hewlett-Packard 5890 gas chromatograph with 5973mass selective detector (GC-MSD) and 7673 autosam-
pler. The column was a 60m · 0.250mm · 0.1lm DB-
5 (Chromatographic Specialties, Brockville, ON). For
every batch of five samples injected, the surrogate stand-
ard, PCB standard mixture, organochlorine standard
mixture, method blank and in-house reference tissue
also were analyzed. PCBs were quantified according to
the method described by Drouillard and Norstrom(2003). Detection limits ranged from 0.01 to 0.08ng/g
wet weight depending on the chemical of study. Blanks
and reference tissues, quantified during each batch of
sample extractions, were in compliance with the normal
quality assurance procedures instituted by GLIER’s
CAEAL certified organic analytical laboratory. Sample
recoveries for the surrogate standard averaged 89 ± 2%
(mean ± SE). Chemical concentrations were not recov-ery corrected.
We loge-transformed data to approximate normal
distributions, and then used t-tests with Bonferroni cor-
rection to compare concentrations of POPs in tissues of
various species. The following PCB congeners were not
detected in any bird or hare samples: PCB 19, 24/27, 22,
25, 45, 40, 42, 199. In seals, PCBs 19, 40, 132, 176, 199,
200, 207 were not detected; seal fat samples had detect-able levels of all 71 other congeners. Dieldrin was not
detected in eider, hare or seal muscle, nor in seal liver.
Mirex was not detected in eider or seal liver, nor hare
muscle. All fulmar and gull tissues, as well as seal fat,
had detectable levels of all POP residues examined.
Compared to the marine species, concentrations of most
POPs in arctic hare were low. In guillemots, eiders and
seals, POP concentrations were typically higher in livertissue than in muscle, but this pattern was reversed in
the fulmars and glaucous gull (Table 1).
Fig. 1. Location of the study area in eastern Arctic Canada (black
circles denote former military sites, star denotes community of
Qikiqtarjuaq). Gray areas on the Cumberland Peninsula represent
glaciers.
96 Baseline / Marine Pollution Bulletin 50 (2005) 95–104
Table 1
Geometric means ± SD and ranges (in brackets) in ng/g wet wt. as well as lipid-normalized means ± SD (ng/g) of POPs in muscle and liver tissues of wildlife from the Qikiqtarjuaq area
Mean ± SD (Range) Species tissue (n)
Black guillemot (8) Common eider (8) Northern fulmar (2) Glaucous gull
(1)
Arctic hare (2) Ringed seal (2)
Muscle Liver Muscle Liver Muscle Liver Muscle Liver Muscle Muscle Liver Fat
% Lipid 2.46 ± 0.41 2.84 ± 1.64 1.75 ± 0.67 1.99 ± 0.46 8.10 6.12 5.43 2.74 2.36 2.76 3.10 97.9
(1.88–3.01) (1.27–6.18) (0.78–2.48) (1.29–2.44) (5.25, 10.95) (3.39, 8.85) – – (2.32, 2.40) (1.82, 3.67) (2.81, 3.38) (95.3, 100.7)P
PCB 27.00 ± 36.34 34.71 ± 56.33 2.73 ± 1.64 5.32 ± 6.46 38.21 30.12 560.4 219.6 7.95 7.58 14.51 804.6
(8.36–100.93) (5.90–172.52) (0.58–5.70) (0.43–19.33) (20.21, 56.21) (14.00, 46.25) – – (0.11, 15.80) (6.08, 9.08) (6.70, 22.33) (302.8, 1306)
975 ± 1793 1494 ± 2508 152 ± 65 276 ± 316 449 468 10,320 8013 343 327 449 808P
DDT 3.32 ± 2.52 6.51 ± 4.16 1.36 ± 0.83 1.40 ± 0.71 29.19 19.60 291.8 116.8 0.05 4.52 4.66 491.4
(0–6.93) (2.38–13.53) (0.14–2.72) (0.28–2.31) (17.81, 40.56) (9.30, 29.90) – – (0, 0.10) (3.01, 6.04) (1.79, 7.54) (236.9, 745.8)
143 ± 120 264 ± 176 79 ± 35 77 ± 37 355 306 5374 4263 2.1 203 143 495P
CHLOR 2.89 ± 2.63 7.00 ± 4.38 0.99 ± 0.82 4.50 ± 3.80 17.12 20.95 66.29 39.76 0.12 2.77 8.31 509.2
(0–7.34) (3.36–15.18) (0–2.64) (0.68–11.37) (9.57, 24.66) (15.33, 26.57) – – (0, 0.23) (2.68, 2.85) (3.89, 12.72) (250.9, 767.4)
126 ± 131 285 ± 184 60 ± 39 227 ± 150 204 376 1221 1451 5.0 113 257 513
Dieldrin 0.23 ± 0.64 2.71 ± 1.03 0 3.30 ± 1.87 6.53 11.26 3.28 1.99 0 0 0 58.03
(0–1.82) (1.72–4.05) – (1.90–7.00) (3.76, 9.29) (8.20, 14.32) – – – – – (57.68, 58.39)
12 ± 34 51 ± 70 0 133 ± 102 78 202 60 73 0 0 0 59P
MIREX 0.16 ± 0.28 0.44 ± 0.47 0.01 ± 0.02 0 1.01 0.85 15.33 8.88 0 0.63 0 13.86
(0–0.70) (0–1.37) (0–0.04) – (0.51, 1.50) (0.17, 1.53) – – – (0.54, 0.72) – (3.44, 24.28)
8 ± 14 18 ± 21 0.2 ± 0.6 0 12 11 282 324 0 27 0 14P
CBz 3.06 ± 2.20 6.90 ± 3.67 1.37 ± 0.92 1.56 ± 0.85 14.26 10.84 22.62 13.57 0.06 0.59 0.56 33.43
(0–5.35) (3.66–14.34) (0.10–3.17) (0.64–3.27) (9.62, 18.90) (6.35, 15.33) – – (0.02, 0.10) (0.48, 0.70) (0.44, 0.67) (28.65, 38.21)
130 ± 103 270 ± 113 80 ± 41 84 ± 35 178 180 417 495 2.6 22 18 34P
HCH 1.41 ± 1.11 2.58 ± 1.59 0.42 ± 0.44 1.63 ± 1.67 0.70 0.52 2.67 1.25 0 0.85 0.91 32.61
(0–2.61) (0.98–5.09) (0–1.06) (0.44–4.93) (0.46, 0.94) (0.28, 0.76) – – – (0.52, 1.18) (0.72, 1.10) (30.62, 34.60)
60 ± 51 105 ± 65 26 ± 28 69 ± 64 9 8 49 45 0 30 29 33
Baselin
e/Marin
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50(2005)95–104
97
Among all wildlife, contaminant concentrations (ex-
pressed in ng/g wet weight) were considerably higher
in seal fat than in all other tissues (Table 1). Considering
only muscle and liver tissues, the glaucous gull had the
highest concentration of all POPs examined, except for
dieldrin (highest in fulmar tissues) andP
HCH (highestin guillemot livers). Concentrations of
PPCB,
PDDT,P
CHLOR andP
MIREX increased following this pat-
tern: eider < guillemot < fulmar < gull. Fulmars, how-
ever, had higher dieldrin and lowerP
HCH levels than
the glaucous gull. The glaucous gull had the highest con-
taminant burdens for all residues except dieldrin (high-
est in fulmars) andP
HCH (highest in guillemots)
when comparing lipid-normalized values of POPs.One outlier was found among the samples, an adult
common eider, which was excluded from the analyses
above and Table 1. For this bird,P
PCB was 5487ng/g
wet wt. in muscle and 4173ng/g wet wt. in liver (approx-
imately 1000 times higher than the next closest eider).
Most of the detectable PCB congeners were elevated in
this specimen, but PCBs 55, 105, 118, 156, 157, 183,
187, and particularly 138, 146, 153, 170/190, 180, 187and 194 were detected at much higher levels than in
the other eiders. Other POPs, however, were in the same
range as the rest of the eiders.
Concentrations of POPs in muscle tissue of black
guillemots and common eiders were similar (all
P > 0.05; Table 1). In liver tissues, concentrations ofPCBz and
PDDT were significantly higher in guille-
mots (P < 0.01; Table 1), while other POP levels didnot differ significantly between the species. The propor-
tional contribution of various PCB congeners to thePPCB also differed among species (Fig. 2).
PPCB in
black guillemot had relatively similar contributions
from congeners 105, 118, 138 and 153 whereas, in the
other birds, PCB 153 was the dominant congener fol-
lowed by PCB 138, with lower contributions (<7%) from
PCB 105. Ringed seals had higher proportions of lower-chlorinated PCB congeners than the other species, while
arctic hare had higher proportions of the hepta- and
octachlorobiphenyls (Table 2), but also from specific
PCBs 87, 110 and 149. Guillemots also differed from
the other birds by having high proportions of both
penta- and hexachlorobiphenyls.
For the suite of POPs examined, almost all residue
concentrations in muscle and livers of wildlife fromQikiqtarjuaq were below levels thought to be of concern
for wildlife health (Kamrin and Ringer, 1996; Weime-
yer, 1996; Braune et al., 1999a; Muir et al., 1999). Ex-
cept for the possibility of the single common eider
outlier identified above, there was no evidence of
point-source contamination for POPs in the local mar-
ine environment, nor of trace elements (Mallory et al.,
2004), and therefore local breeding marine birds donot appear to be contaminated from wastes at the for-
mer local military sites.
Seabird concentrations of PCBs were lower than
most species studied in the North Pacific and Southern
Oceans (Guruge et al., 2001), and most of the POPs
examined in this study were lower than other sites in
the eastern Arctic (Braune et al., 1999a,b; Buckman
et al., 2004), and were much lower than those measuredat sites known to be affected by point-source contamina-
tion (e.g. Connell et al., 2003) or at levels known to be
lethal or deleterious to birds (Weimeyer, 1996).
Glaucous gulls are a top predator in Arctic marine
ecosystems, and typically accumulate high contaminant
burdens (e.g. Savinova et al., 1995; Muir et al., 1999;
Skaare et al., 2000; Braune et al., 2002). The gull sam-
pled at Qikiqtarjuaq hadP
PCB,P
CHLOR,P
DDTand
PMIREX concentrations in breast muscle that
were approximately ten times lower than values found
at some other sites in the eastern Arctic (Braune et al.,
1999b). It also had POP concentrations similar to lower
levels found for glaucous gulls at Ny-Alesund and Bjør-
nøya (Savinova et al., 1995), particularly for PCBs. In
general, concentrations of POPs are much higher in gulls
of the Barents Sea than in the Canadian Arctic (e.g.,Henriksen et al., 2000; Braune et al., 2002; Buckman
et al., 2004).
For guillemots, samples from Qikiqtarjuaq also had
POP concentrations about 10 times lower than found
elsewhere in eastern Arctic (Braune et al., 1999b), and
similar to values from Greenland birds (Vorkamp
et al., 2004), butP
PCBs andP
CHLOR in young guil-
lemots were similar to those in nestling black guillemotsfrom an uncontaminated site in Labrador (Kuzyk et al.,
2003). Hexachlorocyclohexanes (P
HCH) were higher in
guillemots from Qikiqtarjuaq compared to those from
Iceland (Olafsdottir et al., 2001), and similarlyP
PCB,
DDT and dieldrin were higher in Qikiqtarjuaq guille-
mots than those collected in western Hudson Bay (John-
stone et al., 1996). This suggests that Qikiqtarjuaq
guillemots, or perhaps guillemots inhabiting the easternArctic coastline along Baffin Island and northern Labra-
dor, may be relying on different foods than those from
the other regions, and may be wintering in areas exposed
to higher concentrations of these contaminants.
Eiders at Qikiqtarjuaq had POP concentrations lower
than those typically found in Maritime Canada, but
consistent with values from across the eastern Arctic
(Braune et al., 1999b). Compared to the European Arc-tic, concentrations of
PCBz were similar, but Qikiqtar-
juaq eiders had higher concentrations ofP
HCH, lower
concentrations ofP
PCB andP
CHLOR, and much
lower concentrations ofP
DDT (Savinova et al.,
1995). Concentrations ofP
PCB and DDT in eiders
from Iceland (Olafsdottir et al., 2001) were higher than
those found in the Qikiqtarjuaq birds. The exception
to these patterns is the outlier eider, which hadP
PCBconcentrations higher than any reported in the studies
above. The liverP
PCB concentration in this eider
98 Baseline / Marine Pollution Bulletin 50 (2005) 95–104
was in the range of values found for guillemots from a
contaminated former military site in Newfoundland,
Canada (Kuzyk et al., 2003) where negative physiologi-
cal responses were observed. The effect of PCB concen-
trations on birds varies by species (Hoffman et al., 1996),
and it is not known what levels may be harmful to
Fig. 2. Distribution of PCB congeners representing at least 2% of the Total PCBs in muscle tissue of at least one species examined. Species were as
follows: BLGU––black guillemot; NOFU––northern fulmar; GLGU––glaucous gull; COEI––common eider; ARHA––arctic hare; RISE––ringed
seal.
Table 2
Proportional contribution of PCB congeners by chlorination group in muscle tissue of examined species
Chlorination group Proportion of total PCB burden (%)
Black guillemot Common eider Northern fulmar Glaucous gull Arctic hare Ringed seal
Trichlorobiphenyls 0.5 1.1 1.6 0.1 0.2 3.3
Tetrachlorobiphenyls 6.4 7.5 12.7 3.1 3.9 17.5
Pentachlorobiphenyls 41.1 22.3 27.6 17.5 26.6 40.5
Hexachlorobiphenyls 41.3 53.5 41.5 53.3 40.3 32.5
Heptachlorobiphenyls 9.7 15.1 14.0 22.1 24.1 6.2
Octachlorobiphenyls 1 0.5 2.1 3.5 4.5 0
Nonachlorobiphenyls 0 0 0.5 0.4 0.4 0
Baseline / Marine Pollution Bulletin 50 (2005) 95–104 99
eiders. It is possible that high PCB levels in this bird
resulted from it feeding in a very localized area of
PCB pollution at one of the former military sites (e.g.,
Kuzyk et al., 2003). However, the fact that this was an
adult bird, and that none of the other local wildlife
had concentrations within an order of magnitude of thisspecimen suggests that the contaminants were acquired
during migration or wintering.
The ringed seals sampled at the site had POP concen-
trations within ranges reported elsewhere in the Cana-
dian Arctic; somewhat lower than reported for western
Hudson Bay and southwestern Ellesmere Island, but
higher than Amundsen Gulf and Barrow Strait (Muir
et al., 1999). PCB, DDT and CBz concentrations weresimilar in Qikiqtarjuaq seals to ringed seals sampled in
Alaska, but Alaskan seals had much higher levels of
HCH (Kucklick et al., 2002). However, Hoekstra et al.
(2003) found very similar lipid-normalized concentra-
tions of PCB, DDT, CHLOR and HCH in 20 seals in
the Beaufort-Chukchi Seas as we observed near Qikiq-
tarjuaq. The two seals sampled in this study differed in
PCB concentrations by a factor of four; male seals areknown to accumulate considerably higher POP concen-
trations than females (Muir et al., 1999), and that may
explain the differences at Qikiqtarjuaq.
PCB congeners 153 and 138 dominated the PCB con-
gener pattern in all species at Qikiqtarjuaq (Fig. 2). In
the eiders, fulmars and the gull, PCBs 153, 138, 180
and 118 were found in the highest concentrations and
therefore contributed the highest proportions toP
PCB(Fig. 2). A similar pattern has been found in other sea-
birds from disparate regions around the world (e.g.,
Focardi et al., 1989; Klasson-Wehler et al., 1998; Gur-
uge et al., 2001; Buckman et al., 2004). The proportional
contribution of each congener to the overall PCB burden
clearly differed across species at Qikiqtarjuaq (Fig. 2),
which has also been shown in Iceland (Olafsdottir et al.,
2001). Differences in the proportional distribution ofPCB congeners among species may be attributable in
part to exposure and uptake, but also to differential
metabolic capabilities for dealing with these chemicals,
particularly between birds and mammals (e.g., Fisk
et al., 2001; Ruus et al., 2002a,b). The major PCB con-
geners measured in birds are characteristically blocked
at meta–para-sites along both phenyl rings and thus
are not expected to undergo significant biotransforma-tion (Drouillard and Norstrom, 2003).
Organochlorine contaminants biomagnify up food
chains, and particularly marine food chains, because
they are longer than terrestrial chains and because mar-
ine animals are relatively high in lipids, a tissue to
which organochlorines readily bind (Muir et al., 1992;
Dietz et al., 2000; Ruus et al., 2002b). Near Qikiqtarjuaq,
POP concentrations generally increased from eiders toguillemots to fulmars to gulls, notably in the lipid-nor-
malized values, and representing as much as a 100-fold
amplification. Similar patterns in POP concentrations
among marine bird trophic webs have been observed
in other studies (Savinova et al., 1995; Borga et al.,
2001; Braune et al., 2002). The main exception to this
pattern was the result forP
HCH, which did not bio-
magnify appreciably at this site. This is consistent withresults from Iceland (Olafsdottir et al., 2001) and else-
where (Muir et al., 1999), and may be attributable to
the ability of birds to metabolize these substances as well
as the physical and chemical properties of HCHs, which
do not favour biomagnification (e.g., relatively low
octanol–water partition coefficient compared to other
organochlorines).
Most marine birds collected near Qikiqtarjuaq, east-ern Baffin Island, Canada, showed no evidence of high,
point source POP contamination; rather most measured
concentrations were lower than or consistent with back-
ground levels over much of the Canadian Arctic. Even
the one eider with higher PCB concentrations was well
below levels associated with health concerns. With the
exception of glaucous gulls, these levels appear to pose
little threat for the long-term health of these species.However, recent evidence suggests that contaminants
not addressed in this study, such as brominated flame
retardants, may be increasing (Braune, 2003), and thus
monitoring levels in birds of this region may prove
warranted.
Acknowledgments
Many thanks to J. Noble Jr., J. Aliqatuqtuq, J.
Nookiguak, D. Pickle, J. Akearok, and A. Fontaine
for assistance with aspects of this study. Financial sup-
port was provided by Environment Canada (CWS)
and Indian and Northern Affairs Canada, Northern
Contaminants Program (Local Contaminant Concerns).
This study was conducted under permits NUN-SCI-01-02, WRP 000140, and SRL 0100401N-A.
References
AMAP, 1998. AMAP assessment report: arctic pollution issues. Arctic
Monitoring and Assessment Programme (AMAP), Oslo.
Ballschmiter, K., Zell, M., 1980. Analysis of polychlorinated biphenyls
(PCB) by glass capillary gas chromatography. Composition of
technical Arochlor- and Clophen-PCB mixtures. Fresenius Z
Analytica Chemica 302, 20–31.
Borga, K., Gabrielsen, G.W., Skaare, J.U., 2001. Biomagnification of
organochlorines along a Barents Sea food chain. Environmental
Pollution 113, 187–198.
Braune, B.M., 2003. Retrospective survey of dioxins, furans, coplanar
PCBs and polybrominated diphenyl ethers in Arctic seabird eggs.
In: Northern Science and Contaminants and Research Directorate
(Ed.), Synopsis of Research Conducted Under the 2001–2003
100 Baseline / Marine Pollution Bulletin 50 (2005) 95–104
Northern Contaminants Program. Department of Indian Affairs
and Northern Development, Ottawa, pp. 251–254.
Braune, B., Muir, D., DeMarch, B., Gamberg, M., Poole, K., Currie,
R., Dodd, M., Duschenko, W., Eamer, J., Elkin, B., Evans, M.,
Grundy, S., Hebert, C., Johnstone, R., Kidd, K., Koenig, B.,
Lockhart, L., Marshall, H., Reimer, K., Sanderson, J., Shutt, L.,
1999a. Spatial and temporal trends of contaminants in Canadian
Arctic freshwater and terrestrial ecosystems: A review. The Science
of the Total Environment 230, 145–207.
Braune, B.M., Malone, B.J., Burgess, N.M., Elliot, J.E., Garrity, N.,
et al., 1999b. Chemical residues in waterfowl and gamebirds
harvested in Canada, 1987–95. Canadian Wildlife Service, Techni-
cal Report Series no. 326, Ottawa.
Braune, B.M., Donaldson, G.M., Hobson, K.A., 2001. Contaminant
residues in seabird eggs from the Canadian Arctic. I. Temporal
trends 1975–1998. Environmental Pollution 114, 39–54.
Braune, B.M., Donaldson, G.M., Hobson, K.A., 2002. Contaminant
residues in seabird eggs from the Canadian Arctic. II. Spatial
trends and evidence from stable isotopes for intercolony differ-
ences. Environmental Pollution 117, 133–145.
Bright, D.A., Dushenko, D.W., Grundy, S.L., Reimer, K.J., 1995.
Effects of local and distant contaminant sources: polychlorinated
biphenyls and other organochlorines in bottom-dwelling animals
from an Arctic estuary. The Science of the Total Environment 160/
161, 265–283.
Buckman, A.H., Norstrom, R.J., Hobson, K.A., Karnovsky, N.J.,
Duffe, J., Fisk, A.T., 2004. Organochlorine contaminants in seven
species of Arctic seabirds from northern Baffin Bay. Environmental
Pollution 128, 327–338.
Bustnes, J.O., Erikstad, K.E., Skaare, J.U., Bakken, V., Mehlum, F.,
2003. Ecological effects of organochlorine pollutants in the Arctic:
A study of the glaucous gull. Ecological Monographs 13, 504–515.
Connell, D.W., Fung, C.N., Minh, T.B., Tanabe, S., Lam, P.K.S.,
Wong, B.S.F., Lam, M.H.W., Wong, L.C., Wu, R.S.S., Richard-
son, B.J., 2003. Risk to breeding success of fish-eating ardeids due
to persistent organic contaminants in Hong Kong: Evidence from
organochlorine compounds in eggs. Water Research 37, 459–467.
Dietz, R., Riget, F., Cleemann, M., Aarkrog, A., Johansen, P.,
Hansen, J.C., 2000. Comparison of contaminants from different
trophic levels and ecosystems. The Science of the Total Environ-
ment 245, 221–231.
Drouillard, K.G., Norstrom, R.J., 2003. The influence of diet
properties and feeding rates on PCB toxicokinetics in the ring
dove. Archives of Environmental Contamination and Toxicology
44, 97–106.
Environment Canada, 1989. Analytical Methods Manual. Water
Quality Branch, Environment Canada, Ottawa.
Fisk, A.T., Hobson, K.A., Norstrom, R.J., 2001. Influence of chemical
and biological factors on trophic transfer of persistent organic
pollutants in the Northwater Polynya marine food web. Environ-
mental Science and Technology 35, 732–738.
Focardi, S., Leonzio, C., Fossi, M.C., 1989. Variations in polychlo-
rinated biphenyl congener composition in eggs of Mediterranean
water birds in relation to their position in the food chain.
Environmental Pollution 52, 243–255.
Guruge, K.S., Tanaka, H., Tanabe, S., 2001. Concentration and toxic
potential of polychlorinated biphenyl congeners in migratory
oceanic birds from the North Pacific and the Southern Ocean.
Marine Environmental Research 52, 271–288.
Henriksen, E.O., Gabrielsen, G.W., Trudeau, S., Wolkers, J., Sagerup,
K., Skaare, J.U., 2000. Organochlorines and possible biochemical
effects in glaucous gulls (Larus hyperboreus) from Bjørnøya, the
Barents Sea. Archives of Environmental Contamination and
Toxicology 38, 234–243.
Hoekstra, P.F., O’Hara, T.M., Fisk, A.T., Borga, K., Solomon, K.R.,
Muir, D.C.G., 2003. Trophic transfer of persistent organochlorine
contaminants (OCs) within an Arctic marine food web from the
southern Beaufort-Chukchi Seas. Environmental Pollution 124,
509–522.
Hoffman, D.J., Rice, C.P., Kubiak, T.J., 1996. PCBs and dioxins in
birds. In: Beyer, W.N., Heinz, G.H., Redmon-Norwood, A.W.
(Eds.), Environmental contaminants in wildlife: Interpreting tissue
concentrations. Lewis Publishers, CRC Press, Boca Raton, pp.
165–208.
Jensen, J., Adare, K., Shearer, R., 1997. Canadian Arctic assessment
report. Indian and Northern Affairs Canada, Ottawa, ON. 460 pp.
Johnstone, R.M., Court, G.S., Fesser, A.C., Bradley, D.M., Oliphant,
L.W., MacNeil, J.D., 1996. Long-term trends and sources of
organochlorine contamination inCanadian tundra peregrine falcons,
Falco peregrinus tundrius. Environmental Pollution 93, 109–120.
Kamrin, M.A., Ringer, R.K., 1996. Toxicological implications of PCB
residues in mammals. In: Beyer, W.N., Heinz, G.H., Redmon-
Norwood, A.W. (Eds.), Environmental contaminants in wildlife:
Interpreting tissue concentrations. Lewis Publishers, New York,
pp. 153–163.
Klasson-Wehler, E., Bergman, A., Athanasiadou, M., Ludwig, J.P.,
Auman, H.J., Kannan, K., Van denberg, M., Murk, A.J., Feyk,
L.A., Giesy, J.P., 1998. Hydroxylated and methylsulfonyl poly-
chlorinated biphenyl metabolites in albatrosses from Midway
Atoll, North Pacific Ocean. Environmental Toxicology and
Chemistry 17, 1620–1625.
Kucklick, J.R., Struntz, W.D., Becker, P.R., York, G.W., O’Hara,
T.M., Bohonowych, J.E., 2002. Persistent organochlorine pollut-
ants in ringed seals and polar bears collected from northern Alaska.
The Science of the Total Environment 287, 45–59.
Kuzyk, Z.Z.A., Burgess, N.M., Stow, J.P., Fox, G.A., 2003. Biological
effects of marine PCB contamination on black guillemot nestlings
at Saglek, Labrador: Liver biomarkers. Ecotoxicology 12, 183–197.
Lazar, R., Edwards, R.C., Metcalfe, C.D., Metcalfe, T., Gobas,
F.A.P.C., Haffner, G.D., 1992. A simple, novel method for the
quantitative analysis of coplanar (Non-ortho substituted) poly-
chlorinated biphenyls in environmental samples. Chemosphere 25,
493–504.
Mallory, M.L., Wayland, M., Braune, B.M., Drouillard, K.G., 2004.
Trace elements in marine birds, arctic hare and ringed seals
breeding near Qikiqtarjuaq, Nunavut, Canada. Marine Pollution
Bulletin 49, 136–141.
Muir, D.C.G., Wagemann, R., Hargrave, B.T., Thomas, D.J., Peakall,
D.B., Norstrom, R.J., 1992. Arctic marine ecosystem contamina-
tion. The Science of the Total Environment 122, 75–134.
Muir, D., Braune, B., DeMarch, B., Norstrom, R., Wagemann, R.,
Lockhart, L., Hargrave, B., Bright, D., Addison, R., Payne, J.,
Reimer, K., 1999. Spatial and temporal trends and effects of
contaminants in the Canadian Arctic marine ecosystem: A review.
The Science of the Total Environment 230, 83–144.
Olafsdottir, K, Petersen, Æ, Magnusdottir, E.V., Bjornsson, T.,
Johannesson, T., 2001. Persistent organochlorine levels in six prey
species of the gyrfalcon Falco rusticolus in Iceland. Environmental
Pollution 112, 245–251.
Ruus, A., Sandvik, M., Ugland, K.I., Skaare, J.U., 2002a. Factors
influencing activities of biotransformation enzymes, concentrations
and compositional patterns of organochlorine contaminants in
members of a marine food web. Aquatic Toxicology 61, 73–87.
Ruus, A., Ugland, K.I., Skaare, J.U., 2002b. Influence of trophic
position on organochlorine concentrations and compositional
patterns in a marine food web. Environmental Toxicology and
Chemistry 21, 2356–2364.
Savinova, T.N., Polder, A., Gabrielsen, G.W., Skaare, J.U., 1995.
Chlorinated hydrocarbons in seabirds from the Barents Sea area.
The Science of the Total Environment 160/161, 497–504.
Skaare, J.U, Bernhoft, A., Derocher, A.E., Gabrielsen, G.W., Gok-
soyr, A., Henriksen, E.O., Larsen, H.J., Wiig, O., 2000. Organo-
chlorines in top predators at Svalbard: Occurrence, levels and
effects. Toxicology Letters 112–113, 103–109.
Baseline / Marine Pollution Bulletin 50 (2005) 95–104 101
Vorkamp, K., Christensen, J.H., Glasius, M., Riget, F.F., 2004.
Persistent halogenated compounds in black guillemots (Cepphus
grylle) from Greenland–levels, compound patterns and spatial
trends. Marine Pollution Bulletin 48, 111–121.
Weimeyer, S.N., 1996. Other organochlorine pesticides in birds. In:
Beyer, W.N., Heinz, G.H., Redmon-Norwood, A.W. (Eds.),
Environmental contaminants in wildlife: interpreting tissue con-
centrations. Lewis Publishers, New York, pp. 99–115.
Polycyclic aromatic hydrocarbons in capelin (Mallotus villosus) in theBarents Sea by use of fixed wavelength fluorescence measurements
of bile samples
C. Haugland a,*, K.I. Ugland a, J.F. Børseth b, E. Aas b
a Section of Marine Biology and Limnology, University of Oslo, P.O. Box 10634, Blindern, 0316 Oslo, Norwayb RF-Rogaland Research, P.O. Box 2503, Ullandhaug, N-4091 Stavanger, Norway
Polycyclic aromatic hydrocarbons (PAHs) are pres-
ent in most marine and terrestrial habitats. Some PAH
molecules have been shown to possess mutagenic and
carcinogenic effects on organisms (Varanasi et al.,
1989; Kurelec, 1993). Other damaging effects includeskin lesions, skeletal deformities and tumours (Krahn
et al., 1986; Malins et al., 1988; Collier and Varanasi,
1991; Hose et al., 1996; Baumann, 1998). PAHs in the
environment derive primarily from two different
sources: (1) pyrogenic molecules that are formed during
incomplete combustion of organic material, and (2)
petrogenic molecules that are components of fossil fuels
(Varanasi et al., 1989). Petrogenic PAHs are of increas-ing concern in the marine environment due to oil explo-
ration and transport. The Barents Sea is a productive
ocean in the Arctic climate zone (Fig. 1) located between
northern Norway, Svalbard and Novaja Zemlja (Loeng,
1991). In the southeast Barents Sea, the Norwegian oil
company Statoil is developing a large gas field called
Snøhvit.
The field was discovered in 1984 and is planned to beoperational in 2006. Statoil conducts several environ-
mental baseline investigations in the southern Barents
Sea. The purpose of this study was to determine the lev-
els of PAH in capelin by use of fixed wavelength fluores-
cence (FF). Capelin (Mallotus villosus Muller) is a small
coldwater species with a northern circumpolar distribu-
tion. Capelin is a key species in the Barents Sea since it is
the main food source for marine mammals, seabirds andlarger fish species like cod (Gjøsæter, 1998). This species
is therefore a natural choice in a monitoring program.
PAH compounds can be absorbed by organisms by
ingestion, diffusion through surfaces or by gas exchange
(Varanasi et al., 1989). After the molecules are taken up
by fish they are biotransformed into polar metabolites,
and thereby enhance the efficiency of excretion (Lawand Klungsøyr, 2000). The metabolites are concentrated
in the gall bladder. Sensitive measurements of PAHs
may therefore be based on samples from the bile (Aas
et al., 2000; Jonsson et al., 2004).
Sampling of capelin was carried out in 2002 at two
different locations. On April 13th, 50 individuals were
taken (15 males and 35 females) at location 1
(70�530N; 29�240E). On April 17th, 18 individuals, allmales, were caught at location 2 (70�520N; 29�290E)(Fig. 1). The fish were kept alive in tanks with running
water. Dead fish were not used for further analysis.
From each individual, the gall bladder was removed
and stored in cryo-tubes frozen in liquid nitrogen for
shipment.
The fluorescence analyses were conducted according
to Aas et al. (2000) using a Perkin Elmer LS50B lumi-nescence spectrofluorometer. This method is considered
to be semi-quantitative and semi-qualitative due to lack
of absolute values of the different PAH metabolites.
Thus the measurements only provide an indication of
exposure, but the advantages are that relatively large
samples can be analysed at a low cost (Aas et al.,
2000). The method uses wavelength pairs for excited
and emitted light as a result of the ability of PAH mol-ecules to absorb ultraviolet light followed by emission of
light of a longer wavelength.
The bile samples were analysed at the wavelength
pairs 290/335, 341/383 and 380/430 nm, optimised for
the detection of naphthalene, pyrene and benzo[a]pyrene
type metabolites (Krahn et al., 1987; Lin et al., 1996;
0025-326X/$ - see front matter Crown Copyright � 2004 Published by Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2004.10.029
* Corresponding author. Tel.: +47 986 38 786; fax: +47 22854438.
E-mail addresses: [email protected], [email protected]
(C. Hauglnad).
102 Baseline / Marine Pollution Bulletin 50 (2005) 95–104