Science of the Total Environment 409 (2011) 2838–2844
Contents lists available at ScienceDirect
Science of the Total Environment
j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / s c i t o t e n v
Organohalogenated contaminants in eggs of rockhopper penguins (Eudyptes
chrysocome) and imperial shags (Phalacrocorax atriceps) from the Falkland Islands
Evi Van den Steen a,⁎, Maud Poisbleau a, Laurent Demongin a, Adrian Covaci b, Alin C. Dirtu b,c,
Rianne Pinxten a, Hendrika J. van Noordwijk d, Petra Quillfeldt d, Marcel Eens a
a
Laboratory of Ethology, Department of Biology, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium
Toxicological Centre, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk, Belgium
Department of Chemistry, “Al. I. Cuza” University of Iasi, Carol I Bvd, 700506, Iasi, Romania
d
Max Planck Institute for Ornithology, Vogelwarte Radolfzell, Schlossallee 2, 78315 Radolfzell, Germany
b
c
a r t i c l e
i n f o
Article history:
Received 7 February 2011
Received in revised form 1 April 2011
Accepted 1 April 2011
Available online 6 May 2011
Keywords:
Polychlorinated biphenyls
Organochlorine pesticides
Polybrominated diphenyl ethers
Methoxylated PBDEs
Bird eggs
Falkland Islands
a b s t r a c t
In this study, we evaluated the use of seabird eggs of two common bird species from the Falkland Islands as
bioindicators of contamination with organohalogenated contaminants (OHCs). We compared contamination
levels and profiles of different OHCs between eggs of the rockhopper penguin (Eudyptes chrysocome) and the
imperial shag (Phalacrocorax atriceps). In addition, laying order effects on OHC concentrations and profiles
were also investigated in both species. For polychlorinated biphenyls (PCBs), organochlorine pesticides
(OCPs) as well as polybrominated diphenyl ethers (PBDEs), concentrations were significantly lower in eggs of
rockhopper penguins (27.6 ± 0.70 ng/g lw, 56.5 ± 1.33 ng/g lw and 0.98 ± 0.04 ng/g lw, respectively)
compared to the imperial shags (140 ± 5.54 ng/g lw, 316 ± 11.53 ng/g lw, 1.92 ± 0.15 ng/g lw, respectively).
On the other hand, 2′MeO-BDE 68 and 6MeO-BDE 47, two brominated compounds of reported natural origin,
were significantly higher in the penguin eggs (0.55 ± 0.05 ng/g lw and 7.01 ± 0.64 ng/g lw, respectively)
compared to the shag eggs (0.17 ± 0.03 ng/g lw and 0.50 ± 0.06 ng/g lw, respectively). In addition, PCB, OCP
and PBDE contamination profiles differed markedly between the two species. Various factors, such as diet,
feeding behaviour, migratory behaviour and species-specific metabolism, may be responsible for the observed
results. For both rockhopper penguins and imperial shags, PCB, OCP and PBDE concentrations and profiles did
not significantly change in relation to the laying order. This suggests that, for both species, any egg of a clutch
is useful as a biomonitoring tool for OHCs. Although our results showed that OHCs have also reached the
Falkland Islands, concentrations were relatively low compared to other studies. However, future monitoring
may be warranted to assess temporal trends of different OHCs.
© 2011 Elsevier B.V. All rights reserved.
1. Introduction
The presence of organohalogenated contaminants (OHCs), such as
polychlorinated biphenyls (PCBs), organochlorine pesticides (OCPs)
and polybrominated diphenyl ethers (PBDEs), in the environment has
been a cause of concern because of their persistent character,
bioaccumulative potential and adverse effects on both humans and
wildlife (Vos et al., 2000; Birnbaum and Staskal, 2004). Different OHCs
have been shown to cause effects on reproduction in birds through
different mechanisms, such as eggshell thinning, embryotoxicity and
effects on reproductive behaviour (Gilbertson et al., 1991; Elliott and
Martin, 1994; McCarty and Secord, 1999; Fernie et al., 2008). There is
also evidence of long-range transport of these substances to regions
⁎ Corresponding author at: Laboratory of Ethology, Department of Biology,
University of Antwerp (Campus Drie Eiken), Universiteitsplein 1, 2610 Wilrijk, Belgium.
Tel.: + 32 3 265 22 85; fax: + 32 3 265 22 71.
E-mail address: [email protected] (E. Van den Steen).
0048-9697/$ – see front matter © 2011 Elsevier B.V. All rights reserved.
doi:10.1016/j.scitotenv.2011.04.002
where they have never been used or produced and, as a consequence,
these pollutants have been distributed worldwide and even remote
locations have been reached by OHCs (Braune, 2007). The transport of
these pollutants to remote areas is a direct consequence of their
persistence and ability to volatilise (Jones and De Voogt, 1999).
Deposition occurs mostly in colder regions, such as the polar and
subpolar regions (Wania and Mackay, 1996). The atmosphere and
ocean currents are the main transport routes by which OHCs enter
pristine regions. In addition, OHCs may also be transported to these
regions via pelagic organisms and migratory birds (Roosens et al.,
2007; Blais et al., 2007).
Different animal species are being used as sentinels of environmental pollution and human exposure (van der Schalie et al., 1999;
Van den Steen et al., 2009). Bird eggs have been successfully used to
monitor OHCs in numerous studies (Donaldson et al., 1999; Norstrom
et al., 2002; Elliott et al., 2005; Jaspers et al., 2005; Van den Steen et al.,
2006, 2009), as females can pass contaminants stored in their body
tissues into their eggs. For most bird species, eggs can be easily
E. Van den Steen et al. / Science of the Total Environment 409 (2011) 2838–2844
collected and the collection of a single egg from a clutch is expected to
have a minor effect on the population level (Furness, 1993). The
presence of laying order effects on the contaminant levels within a
clutch is of considerable interest for the purpose of biomonitoring.
Laying order effects of some OHCs, such as DDT, have been reported
for some avian species (Mineau, 1982; Nisbet, 1982). However, most
studies measuring OHCs showed that eggs within a clutch have
similar contaminant levels, and that a single egg statistically
represents the entire clutch (Newton and Bogan, 1978; Custer et al.,
1990; Van den Steen et al., 2006; Verreault et al., 2006).
Penguins have previously been used as biomonitors of pollution,
including OHCs, of pristine areas of the southern hemisphere, such as
the Falkland Islands (Bennington et al., 1975; Hoerschelmann et al.,
1979; de Boer and Wester, 1991). The most common penguin species
on the Falkland Islands is the southern rockhopper penguin (Eudyptes
chrysocome; Huin, 2007), but this population is in severe decline
(BirdLife International, 2009). During the 1930s, the Falkland Islands
were considered to hold one of the largest populations of the species
at 1,800,000 breeding pairs, but currently the total population is
210,000 breeding pairs (Pütz et al., 2003; Huin, 2007). Different
problems, such as overfishing, pollution and global warming may be
responsible for this severe decline (Cunningham and Moors, 1994;
Bingham, 2002). Another common bird species on the Falkland
Islands is the imperial shag (Phalacrocorax atriceps). They often share
colonies with rockhopper penguins (Bingham, 2001). Rockhopper
penguins and imperial shags differ in many aspects of their ecology
(Masello et al., 2010). Rockhopper penguins are opportunistic feeders
which generally rely on macrozooplankton, crustaceans and to a
lesser extent squid and fish (Stonehouse, 1975; Watson, 1975; Raya
and Schiavini, 2005), while imperial shags mainly feed on fish
(Masello et al., 2010; Michalik et al., 2010). Rockhopper penguins
have been shown to spend the winter between the Falkland Islands
and South America, up to 1400 km from the Falkland Islands (Pütz
et al., 2002) while imperial shags stay in the coastal region during this
period (del Hoyo et al., 1992).
The aim of the present study was to investigate the contamination
levels of OHCs, including PCBs, OCPs and PBDEs, in eggs of rockhopper
penguins and imperial shags from the Falkland Islands. In addition,
the presence of two methoxylated PBDE congeners (6MeO-BDE 47
and 2′MeO-BDE 68) was also assessed in the eggs of both study
species. Recent studies indicated that MeO-PBDEs found in wildlife
are mostly a consequence of accumulation via natural sources in
marine environments (e.g. via formation in sponges and green algae;
Marsh et al., 2004; Teuten et al., 2005). OHC concentrations and
profiles were compared between rockhopper penguins and imperial
shags, which have a different feeding ecology. Higher concentrations
of OHPs were expected in imperial shag eggs because of the higher
trophic position of this species compared to the rockhopper penguin.
However, rockhopper penguins may accumulate different pollutants
during the migration period. Other factors, such as maternal transfer
and feeding behaviour, may also contribute to differences between
the two species. In order to evaluate the eggs of both rockhopper
penguins and imperial shags as a biomonitoring tool for OHCs, laying
order effects on the OHC concentrations and profiles were also
investigated.
2. Materials and methods
2.1. Birds and study site
The study was carried out at the “Settlement colony” on New
Island, Falkland Islands (51°43′S, 61°17′W) from late October to
December 2008. This colony has approximately 5000 pairs of breeding
rockhopper penguins and 3000 pairs of breeding imperial shags. The
breeding biology of the rockhopper penguins at this colony has been
described by Poisbleau et al. (2008). Briefly, male penguins arrive at
2839
the colony first (early October) and establish nest sites. Female
penguins arrive few days later, for pairing and copulation. Laying
intervals are highly standardised in this species. Within clutches, the
second egg (B-egg) is generally laid 4 days after the first egg (A-egg)
(Poisbleau et al., 2008). Imperial shags arrive at the colony during
early October, when courtship and nest building commence. Egg
laying mainly takes place between early November and the end of
December. Female shags lay three eggs, several days apart (A-, B- and
C-egg, respectively). After the arrival of the first birds, we visited
study sites daily, initially to mark active nests and subsequently to
follow the egg laying.
2.2. Egg collection and preparation
When a new A-egg was detected in a penguin study nest, we
collected it. We replaced this egg with one egg found outside its own
nest that we considered as lost by their original parents. Afterwards,
we checked the nest daily until the laying of the B-egg. We then also
collected the B-egg as soon as it was detected in the study nest and
replaced it with one lost egg (see Poisbleau et al., 2009a, 2009b and
2011 for more details on the methods). As incubation in rockhopper
penguins typically does not start before clutch completion (Williams,
1995), the eggs were not incubated for longer than 24 h at collection.
In total, we collected 60 whole clutches. Seventeen A-eggs and 17 Beggs from the same clutch were used for the chemical analyses
(n = 34 eggs). For shag egg collection, eggs were collected at the day
of laying, in the morning before 11 AM, when only 44 males were on
the nest thus preventing the disturbance of females by our visits and
any potential influence of our nest checks on egg composition. We
collected a maximum of 2 out of 3 eggs in the nests. The collected egg
was replaced with a white chicken egg (similar in size and colour to
shag eggs). At clutch completion (3 eggs) we removed these chicken
eggs, leaving one shag egg per nest. Eight A-eggs and 8 B-egg from the
same clutch and 7 B-eggs and 7 C-eggs from the same clutch were
analysed (n = 30 eggs). After collection, we weighed the eggs to the
nearest 0.1 g using a digital balance and froze them whole at − 20 °C
for at least 4 days.
The same method was used to prepare all the frozen eggs
(Poisbleau et al. 2009a; 2009b; 2011). We first removed the shell
while the egg was still frozen. Then, we separated the yolk from the
albumen by taking advantage of the fact that albumen thaws more
quickly than yolk. We recorded the mass of the yolk to the nearest
0.1 g using a digital balance. We carefully homogenised the yolk by
swirling it with a mini-spatula in order to obtain a yolk sample
representative of the whole yolk. A small quantity of each homogenised yolk was transferred to a 1.5-mL Eppendorf tube and stored at
−20 °C until analysis.
2.3. Chemical analysis
A homogenised sample of approximately 0.5–1.0 g was weighed,
mixed with anhydrous Na2SO4 and spiked with internal standards (εHCH, CB 46 and CB 143, BDE 77 and BDE 128). Extraction was carried
out with 100 mL hexane/acetone (3:1, v/v) in an automated Soxhlet
extractor (Büchi, Flawil, Switzerland) in hot extraction mode for 2 h.
The lipid content was determined gravimetrically on an aliquot of the
extract (105 °C, 1 h), while the rest of the extract was cleaned up on a
column filled with ~ 8 g acidified silica and eluted with 15 mL hexane
and 10 mL dichloromethane. The eluate was concentrated to 100 μL
under a gentle nitrogen stream and transferred to an injection vial. In
all samples, concentrations of 22 PCB congeners (CB 28, 31, 74, 95, 99,
101, 105, 110, 118, 128, 138, 149, 153, 156, 163, 170, 180, 183, 187,
194, 196 and 199), 7 PBDE congeners (BDE 47, 49, 99, 100, 153, 154
and 183), 2 MeO-PBDE congeners (2′MeO-BDE 68 and 6MeO-BDE
47), dichlorodiphenyltrichloroethane (p,p′- and o,p′-DDT) and
metabolites (p,p′-DDE and p,p′-DDD), hexachlorocyclohexanes
2840
E. Van den Steen et al. / Science of the Total Environment 409 (2011) 2838–2844
(HCHs; α-, β- and γ-HCHs), chlordanes (CHLs; cis-chlordane (CC),
trans-chlordane (TC), trans-nonachlor (TN) and oxychlordane (OxC)),
and hexachlorobenzene (HCB) were determined.
blanks, LOQs were calculated for a signal-to-noise ratio equal to 10.
LOQs for the analysed compounds ranged between 0.1 and 0.5 ng/g
lipid weight (lw).
2.4. Instrumental analysis
2.6. Statistical analysis
For the PCB analysis, an Agilent 6890 gas chromatograph (GC)
connected to an Agilent 5973 mass spectrometer (MS) operated
in electron ionisation (EI) mode was equipped with a
25 m × 0.22 mm × 0.25 μm HT-8 capillary column (SGE, Zulte,
Belgium). The ion source, quadrupole and interface temperatures
were set at 230, 150 and 300 °C, respectively. The MS was used in
the selected ion-monitoring (SIM) mode with two ions monitored
for each PCB homologue group. One microlitre of the cleaned
extract was injected onto the column using the cold pulsed
splitless mode (injector temperature 90 °C (0.03 min) then to
300 °C at 700 °C/min), pressure pulse 25 psi, pulse time 1.5 min.
The splitless time was 1.5 min. Helium was used as carrier gas at
constant flow (1 mL/min). The temperature of the HT-8 column
was held at 90 °C for 1.5 min, then increased to 180 °C at a rate of
15 °C/min (held for 2.0 min), further increased to 280 °C at a rate
of 5 °C/min and finally raised to 300 °C at a rate of 40 °C/min, held
for 12 min.
For the analysis of OCPs and PBDEs, an Agilent 6890 GC connected
to an Agilent 5973 MS operated in electron capture negative ionisation
(ECNI) mode was equipped with a 25 m × 0.22 mm × 0.25 μm HT8 capillary column (SGE, Zulte, Belgium). Methane was used as
moderating gas and the ion source, quadrupole and interface
temperatures were set at 170, 150 and 300 °C, respectively. The MS
was used in the SIM mode with two ions monitored for each OCP in
specific windows, while ions m/z = 79 and 81 were monitored for
PBDEs during the entire run. One microlitre of the cleaned extract was
injected onto the column using the cold pulsed splitless mode, injector
temperature 90 °C (0.03 min) then to 300 °C at 720 °C/min, pressure
pulse 30 psi, pulse time 1.5 min. The splitless time was 1.5 min. Helium
was used as carrier gas at constant flow (1 mL/min). The temperature
of the HT-8 column was held at 90 °C for 1.5 min, then increased to
220 °C at a rate of 15 °C/min (held for 2.0 min), further increased to
242 °C at a rate of 3 °C/min and finally raised to 300 °C at a rate of
40 °C/min, held for 15 min.
In this study, data are expressed as mean ± standard error.
Statistical calculations were performed using Statistica for Windows
on lipid-normalised concentrations (Statsoft, 1997). The level of
significance was set at α = 0.05 throughout this study. Before data
analysis, samples with levels below LOQ were assigned a value of
½LOQ. Data were normally distributed (Kolmogorov–Smirnov test:
p N 0.05 for all cases). To compare contamination levels between both
species one-way ANOVAs were performed on the mean concentration
per clutch. The profile of PCBs, OCPs and PBDEs between eggs and
species was compared with principal component analysis on normalised concentrations. Principal components (PCs) with eigenvalues
above 1 were considered to account for a significant contribution to
the total variance according to the latent root criterion (Hair et al.,
1998). Factor loadings and factor scores were determined and used in
interpreting PC patterns. Compounds with factor loadings greater
than 0.65 on any PC were considered significant and included in the
figures. The first two PCs were used for the statistical analyses.
2.5. Quality control
Multi-level calibration curves in the linear response interval of the
detector were created for the quantification, and good correlation
(r2 N 0.999) was achieved. The identification of OHCs was based on the
relative retention times to the internal standard used for quantification, ion chromatograms and intensity ratios of the monitored ions. A
deviation of the ion intensity ratios within 20% of the mean values
obtained for calibration standards was considered acceptable. The
quality control was performed by regular analyses of procedural
blanks, by random injection of standards and solvent blanks. A
standard reference material SRM 1945 (PCBs, OCPs and PBDEs in
whale blubber) was used to test the method accuracy indicated that
the measured concentrations were within 10% of the certified values.
The quality control scheme was also assessed through regular
participation in inter-laboratory comparison exercises organised by
the Arctic Monitoring and Assessment Programme (AMAP) and the
National Institute of Standards and Technology (NIST). For each
analyte, the mean procedural blank value was used for subtraction.
BDE 47 and 99 had blank levels which were lower than 5% of the
values found in the samples. Nevertheless, the blank levels were
subtracted from the sample values. After blank subtraction, the limit
of quantification (LOQ) was set at 3 times the standard deviation of
the procedural blank and taking into account the amount of sample
used for analysis. For analytes that were not detected in procedural
3. Results
3.1. General contamination levels
Our results revealed that OCPs constituted the most abundant OHCs
in the eggs of both rockhopper penguins and imperial shags, with mean
sum concentrations of 56.5 ± 1.33 ng/g lw and 316 ± 11.53 ng/g lw,
respectively (Table 1, Fig. 1). The most abundant compounds among the
OCPs were HCB and p,p′-DDE (Table 1). Mean sum PCB concentrations
were 27.6 ± 0.70 ng/g lw and 140 ± 5.54 ng/g lw, respectively (Table 1,
Fig. 1). In eggs of both species, sum PBDE concentrations were much
lower compared to the sum OCPs and sum PCBs, with mean sum
concentrations of 0.98± 0.04 ng/g lw and 1.92 ± 0.15 ng/g lw, respectively (Table 1).
For both PCBs, OCPs as well as PBDEs, concentrations were
significantly lower in eggs of rockhopper penguins compared to
imperial shags (sum PCBs: F 1,30 = 248, p b 0.001; sum OCPs:
F1,30 = 335, p b 0.001; sum PBDEs: F1,30 = 43.3, p b 0.001; Figs. 1 and 2).
In eggs of both species, 6MeO-BDE 47 was more abundant than 2′
MeO-BDE 68 (Fig. 2). Concentrations of 2′MeO-BDE 68 and 6MeO-BDE
47 were significantly lower in the imperial shag eggs compared to the
penguin eggs (Fig. 2; 2′MeO-BDE 68: F1,30 = 21.3, p b 0.001; 6MeOBDE 47: F1,30 = 43.8, p b 0.001).
3.2. Contamination profiles
For PCBs, PCA revealed two PCs which accounted for 45.4% and
20.4% of the variance among the analysed PCB congeners, respectively
(Fig. 3a). CB 153, CB 138 and CB 118 were the most abundant PCB
congeners in both rockhopper penguins and imperial shags. However,
PC1 differed significantly between both species (F1,62 = 741.92,
p = 0.001), while there was no significant difference for PC2
(F1,62 = 0.53, p = 0.47). Shag eggs showed a higher contribution of
CB 101, CB 149, CB 146 and CB 156 and, on the other hand, a lower
contribution of CB 99, CB 105, CB 138, CB 187 and CB 128 compared to
the penguin eggs (Fig. 3a).
PCA revealed two PCs which accounted for 45.6% and 14.3% of the
variance among the OCPs in this study, respectively (Fig. 3b). In both
rockhopper penguins and imperial shags, HCB (55% and 65%,
respectively) and p,p′-DDE (34% and 33%, respectively) contributed
mainly to the sum OCPs. However, PC1 differed significantly between
both species (F1,63 = 300.89; p b 0.001), while there was no significant
E. Van den Steen et al. / Science of the Total Environment 409 (2011) 2838–2844
Compound
LOQ
CB 52
CB 101
CB 149
CB 146
CB 156
CB 99
CB 118
CB 105
CB 153
CB 138
CB 187
CB 183
CB 128
CB 180
CB 170
Sum PCBs
HCB
OxC
TN
CN
p,p-DDE
p,p-DDD
p,p-DDT
α-HCH
β-HCH
γ-HCH
Sum OCPs
BDE 28
BDE 49
BDE 47
BDE 100
BDE 99
BDE 154
BDE 153
BDE 183
Sum PBDEs
2′MeO-BDE 68
6MeO-BDE 47
0.50
0.40
0.20
0.20
0.20
0.40
0.30
0.20
0.20
0.20
0.20
0.20
0.20
0.20
0.20
Mean ± SD
0.10
0.10
0.10
0.10
0.25
0.25
0.25
0.20
0.20
0.20
0.10
0.10
0.10
0.10
0.10
0.20
0.20
0.20
0.20
0.20
Rockhopper penguin
Imperial shag
≤LOQ
≤LOQ
≤LOQ
≤LOQ
≤LOQ
1.27 ± 0.04
3.81 ± 0.10
1.30 ± 0.04
8.28 ± 0.22
4.91 ± 0.16
1.33 ± 0.03
0.64 ± 0.03
0.90 ± 0.03
3.10 ± 0.08
1.26 ± 0.03
27.55 ± 0.70
31.09 ± 1.11
1.48 ± 0.05
1.66 ± 0.05
0.37 ± 0.01
19.39 ± 0.49
≤LOQ
1.51 ± 0.11
0.15 ± 0.01
0.41 ± 0.06
0.28 ± 0.04
56.46 ± 1.33
≤LOQ
≤LOQ
0.13 ± 0.01
0.28 ± 0.02
0.09 ± 0.01
0.15 ± 0.01
0.13 ± 0.008
≤LOQ
0.98 ± 0.04
0.55 ± 0.05
7.01 ± 0.64
1.19 ± 0.21
1.22 ± 0.08
0.66 ± 0.07
7.09 ± 0.29
4.56 ± 0.24
5.61 ± 0.22
18.70 ± 0.78
5.82 ± 0.25
43.38 ± 1.93
18.68 ± 0.76
5.63 ± 0.23
3.54 ± 0.34
2.95 ± 0.14
14.29 ± 0.72
6.96 ± 0.33
140.28 ± 5.54
203.66 ± 7.49
3.29 ± 0.28
≤ LOQ
0.14 ± 0.02
105.27 ± 4.50
0.26 ± 0.08
0.15 ± 0.01
0.68 ± 0.06
1.14 ± 0.17
1.29 ± 0.20
315.93 ± 11.53
≤ LOQ
≤ LOQ
0.51 ± 0.07
0.14 ± 0.02
0.43 ± 0.05
0.26 ± 0.01
0.38 ± 0.03
≤ LOQ
1.92 ± 0.15
0.17 ± 0.03
0.50 ± 0.06
difference for PC2 (F1,63 = 0.08, p = 0.78). This significant difference is
due to a higher contribution of HCB and a lower contribution of OxC,
TN, CN and p,p′-DDT in the shag eggs compared to the penguin eggs
(Fig. 3b).
For PBDEs, PCA revealed two PCs which accounted for 42.3% and
30.2% of the variance among the analysed PBDE congeners, respectively (Fig. 3c). Both PC1 and PC2 differed significantly between
10
rockhopper penguin
imperial shag
8
ng/g lipid weight
Table 1
Limit of quantification (LOQ), mean concentration, and standard deviation (SD) of
different organohalogenated contaminants in eggs of rockhopper penguins (n = 34)
and imperial shags (n = 30) from the Falkland Islands (ng/g lipid weight). Mean % lipids
in the egg samples of rockhopper penguins and imperial shags was 31.39 ± 0.47 % and
28.09 ± 0.35 %, respectively.
2841
6
4
2
0
sum PBDEs 2'MeO-BDE 68 6MeO-BDE 47
Fig. 2. Mean concentrations (with standard errors) of sum PBDEs, 2′MeO-BDE 68,
6MeO-BDE47 and sum PBDEs in eggs of rockhopper penguins (n = 17) and imperial
shags (n = 15) from the Falkland Islands.
rockhopper penguins and imperial shags (PC1: F1,63 = 79.84;
p b 0.001; PC2: F1,63 = 20.89; p b 0.001). BDE 28, BDE 47, BDE 49, BDE
99 and BDE 183 contributed more to the PBDE profile of the imperial
shags, while BDE 100 and BDE 154 contributed less to the shag profile
(Fig. 3c).
3.3. Laying order effects
OHC concentrations in the A-eggs from the rockhopper penguins
were smaller compared to the B-eggs (Fig. 4), although we found no
significant differences in sum PCB, sum OCP and sum PBDE
concentrations between A-eggs and B-eggs (one-way ANOVA: Sum
PCBs: F1,32 = 2.16, p = 0.15; Sum OCPs: F1,32 = 2.31, p = 0.14; Sum
PBDEs: F1,32 = 1.87, p = 0.18).
For the imperial shags, OHC concentrations did not significantly
differ between the A-eggs and B-eggs (Fig. 5; One-way ANOVA: Sum
PCBs: F1,14 = 0.003, p = 0.96; Sum OCPs: F1,14 = 0.009, p = 0.93; Sum
PBDEs: F1,14 = 1.35, p = 0.27), between the B-eggs and C-eggs (Fig. 5;
One-way ANOVA: Sum PCBs: F1,12 = 0.60, p = 0.45; Sum OCPs:
F1,12 = 0.94, p = 0.35; Sum PBDEs: F1,12 = 1.56, p = 0.23).
Both for the rockhopper penguin and the imperial shag, profiles
did not significantly differ among different eggs from a clutch.
However, for the OCPs, there tended to be a difference in profile
between the A-eggs and C-egg from the cormorants (PC2:
F2,27 = 2.64; p = 0.09). In addition, PBDE profiles also tended to be
different between the A- and B-penguin eggs (PC1: F1,32 = 3.08,
p = 0.09).
4. Discussion
4.1. General contamination levels
350
ng/g lipid weight
300
rockhopper penguin
imperial shag
250
200
150
100
50
0
sum PCBs
sum OCPs
Fig. 1. Mean sum concentrations (with standard errors) of sum PCBs and sum OCPs in
eggs of rockhopper penguins (n = 17) and imperial shags (n = 15) from the Falkland
Islands.
The Falkland Islands are a remote location with no significant local
sources of OHC pollution. However, long-range transport may result
in toxicologically significant concentrations of these pollutants.
Concentrations in the present study were generally low compared
to other studies in which eggs of rockhopper penguins from the
Falklands were analysed (Hoerschelmann et al., 1979; Ballschmitter et
al., 1981). Few studies have been performed to investigate OHC
contamination on the Falkland Islands. Compared to our results,
Hoerschelmann et al. (1979) reported considerably (8 and 60 times,
respectively) higher concentrations of p,p′-DDT and p,p′-DDE in
rockhopper penguin eggs from the Falkland Islands. This decreasing
time trend of DDTs has also been observed in other bird species and
biological samples (Braune et al., 2005). de Boer and Wester (1991)
detected very low PCB concentrations in the tissues of two gentoo
penguins (Pygoscelis papua) from the Falkland Islands. They suggested
that a diet of squids may have reduced the total PCB load of the
2842
E. Van den Steen et al. / Science of the Total Environment 409 (2011) 2838–2844
a
70
1,5
A-egg
B-egg
60
1,0
PC2 (20.4%)
0,5
corm C
CB 138
CB 128
peng A
0,0
CB 149
CB 101
peng B
ng/g lipid weight
CB 180
CB 99
corm B
CB 146/156
CB 187
CB 105
-0,5
CB 118
corm A
-1,0
50
40
30
20
10
-1,5
0
-2,0
sum PCBs
-3
-2
-1
0
1
2
3
1,5
PC2 (14.3%)
1,0
α-HCH
corm A
γ-HCH
0,5
peng A
0,0
p,p’-DDT
HCB
corm B
TN
OxC
peng B
-0,5
corm C
-1,0
-3
-2
-1
0
1
2
3
PC1 (45.6%)
c
2,5
2,0
shag A
PC2 (30.2%)
1,5
shag C
1,0
BDE 47
BDE 99
0,5
shag B
0,0
-0,5
-1,0
-1,5
BDE 49/183
BDE 28
BDE 100
peng B
-2
BDE 154
peng A
-1
0
1
sum PBDEs
Fig. 4. Sum PCB, sum OCP and sum PBDE concentrations in A-eggs (n = 17) and B-eggs
(n = 17) from rockhopper penguins from the Falkland Islands.
PC1 (45.4%)
b
sum OCPs
4
2
3
PC1 (42.3%)
penguin and seabird species (Watanabe et al., 2004; Corsolini et al.,
2006; Yogui and Sericano, 2009).
The higher concentrations of PCBs, OCPs and PBDEs in the imperial
shags compared to the penguins may be due to the fact that imperial
shags are higher on the food chain (e.g. Weiss et al., 2009; Masello et
al., 2010). Southern rockhopper penguins are opportunistic feeders.
However, there is evidence that squid is of greater importance in the
diet of this species in waters around the Falkland Islands than
elsewhere (Clausen and Pütz, 2002). The main prey for imperial shags
breeding in the Falkland Islands is fish, although they also feed on
crustaceans (Michalik et al., 2010). Other species-specific factors, such
as the transfer of OHCs from the mothers to the eggs and differences in
metabolic capacities, may also have contributed to the observed
results.
Only few studies have investigated the presence of MeO-PBDEs in
eggs of birds. Verreault et al. (2007) reported a sum of 15 MeO-PBDEs
of 65 ng/g lw in egg yolk of glaucous gulls (Larus hyperboreus) from
the Norwegian Arctic. 2′MeO-BDE 68 and 6MeO-BDE47 were below
the detection limit in guillemot (Uria aalge) eggs from Faroe Islands,
Iceland, Norway and Sweden (Jörundsdóttir et al., 2009). In the
present study, rockhopper penguin eggs showed higher concentrations of MeO-PBDEs compared to the eggs of the imperial shag. This
again might be due to differences in diet, metabolic capacities and
feeding behaviour (pelagic versus benthic feeding). MeO-PBDE
concentrations were found to be higher in pelagic fish species from
the Southern North Sea compared to benthic species (Weijs et al.,
2009). Several studies have indicated that shags forage mostly on
benthic fish species (Blankley, 1981; Casaux et al., 1995), while
rockhopper penguins feed mainly on pelagic species (Cooper et al.,
1990).
Fig. 3. Plots of factor scores with standard errors and factor loadings from the principal
component analysis (PCA) for the (a) sum PCBs, (b) sum OCPs and (c) sum PBDEs.
Compounds with factor loadings greater than 0.65 on any PC were considered
significant. Peng A and peng B represent the first and second, respectively, egg from
clutches of rockhopper penguins from the Falkland Islands. Shag A, shag B and shag C
represent the first, second and third, respectively, egg from clutches of imperial shags
from the Falkland Islands.
A-egg
B-egg
C-egg
ng/g lipid weight
penguins (de Boer and Wester, 1991). Several studies in the
Netherlands and the Great Lakes (USA) have previously reported
PCB and OCP concentrations in other Phalacrocoracidae species (van
den Berg et al., 1994; Custer et al., 2001). Compared to these studies,
concentrations were relatively low in shag eggs from the present
study. This is probably because the Falkland islands are a remote
location and the presence of industrial activities is negligible. To the
best of our knowledge, the present study is the first to report
concentrations of PBDEs in birds from the Falkland Islands. PBDE
concentrations in our study were relatively low compared to other
400
300
200
100
0
Sum PCBs
Sum OCPs Sum PBDEs
Fig. 5. Sum PCB, sum OCP and sum PBDE concentrations in A-eggs (n = 8), B-eggs
(n = 15) and C-eggs (n = 7) in imperial shags from the Falkland Islands.
E. Van den Steen et al. / Science of the Total Environment 409 (2011) 2838–2844
2843
4.2. Contamination profiles
References
HCB and p,p′-DDE were the most abundant organohalogenated
compounds. A similar pattern has been shown in several fish and bird
species from Antarctic regions (Corsolini et al., 2007). The results of
the PCA showed marked differences for PCBs, OCPs and PBDEs
between the rockhopper penguin and the imperial shag. This might be
due to differences in diet between both species. In addition,
rockhopper penguins may be exposed to different pollutants during
their winter migration, which can last up to 6 months (Pütz et al.,
2002). Southern rockhopper penguins have been observed to migrate
in winter to coastal areas of the Falkland Islands and South America
(Pütz et al., 2002). Imperial shags, on the other hand, remain around
the Falklands throughout the year. Differences in metabolic capacity
may also be responsible for the observed results.
Bennington SL, Connors PG, Connors CW, Risebrough RW. Patterns of chlorinated
hydrocarbon contamination in New Zealand sub-Antarctic and coastal marine
birds. Environ Pollut 1975;8:135–47.
Bingham M. Penguins of the Falkland Islands and South America. 1st Books Library;
2001.
Bingham M. The decline of Falkland Islands penguins in the presence of a commercial
fishing industry. Rev Chil Hist Nat 2002;75:805–18.
Birdlife International. Rockhopper penguins. A plan for research and conservation
action to investigate and address population changes. Proceedings of the
international Rockhopper Penguin Workhop, BirdLife International 2009; 2009.
Ballschmitter K. Scholz Ch, Buchert H, Zell M, Figge K, Polzhofer K. Hoerschelmann H.
Studies of the global baseline pollution V. Monitoring the baseline pollution
of the Sub-Antarctic by penguins as bioinidicators. Fresenius Z Anal Chem
1981;309:1–7.
Birnbaum LS, Staskal DF. Brominated flame retardants: cause for concern? Environ
Health Perspect 2004;112:9-17.
Blais JM, Macdonald RW, Mackay D, Webster E, Harvey C, Smol JP. Biologically mediated
transport of contaminants in aquatic systems. “Critical review”. Environ Sci Technol
2007;41:1075–84.
Blankley WO. Marine food of Kelp Gulls, Lesser Sheathbills and Imperial Cormorants at
Marion Island (Subantarctic). Cormorant 1981;9:77–84.
Braune BM. Temporal trends of organochlorines and mercury in seabird eggs from the
Canadian Arctic, 1975–2003. Environ Pollut 2007;148:599–613.
Braune BM, Outridge PM, Fisk AT, Muir DC, Helm PA, Hobbs K, et al. Persistent organic
pollutants and mercury in marine biota of the Canadian Arctic: an overview of
spatial and temporal trends. Sci Total Environ 2005;351–352:4-56.
Casaux RJ, Favero M, Barrera-Oro ER, Silva P. Feeding trial on an Imperial Cormorant
Phalacrocorax atriceps: preliminary results of fish intake and otolith digestion.
Mar Ornithol 1995;23:101–6.
Clausen AP, Pütz K. Recent trends in diet composition and productivity of gentoo,
magellanic and rockhopper penguins in the Falkland Islands. Aquat Conserv: Mar
Freshwat Ecosyst 2002;12:51–61.
Cooper J, Brown CR, Gales RP, Hindell MA, Klages NTW, Moors PJ, et al. Diets and dietary
segregation of crested penguins (Eudyptes). In: Davis LS, Darby JT, editors. Penguin
biology. San Diego: Academic; 1990. p. 131–56.
Corsolini S, Berghesi N, Schiamone A, Focardi S. Polybrominated Diphenyl Ethers,
Polychlorinated Dibenzo-dioxins, -furans, and -biphenyls in three species of
Antarctic Penguins. Environ Sci Pollut R 2007;14:421–9.
Corsolini S, Covaci A, Ademollo N, Focardi S, Schepens P. Occurrence of organochlorine
pesticides (OCPs) and their enantiomeric signatures, and concentrations of
polybrominated diphenyl ethers (PBDEs) in the Adelie penguin food web,
Antarctica. Environ Pollut 2006;140:371–82.
Cunningham DM, Moors PJ. The decline of Rockhopper Penguins Eudyptes chrysocome
at Campbell Island, Southern Ocean and the influence of rising sea temperatures.
Emu 1994;94:27–36.
Custer TW, Custer CM, Hines RK, Stromborg KL, Allen PD, Melancon MJ, et al.
Organochlorine contaminants and biomarker response in double-crested cormorants nesting in Green Bay and Lake Michigan, Wisconsin, USA. Arch Environ
Contam Toxicol 2001;40:89-100.
Custer TW, Pendleton G, Ohlendorf HM. Within- and among-clutch variation of
organochlorine residues in eggs of black-crowned night-herons. Environ Monit
Assess 1990;15:83–9.
de Boer J, Wester P. Chlorobiphenyls and organochlorine pesticides in various subantarctic organisms. Mar Pollut Bull 1991;22:441–7.
del Hoyo J, Elliot A, Sargatal J. Handbook of the birds of the world, Vol. 1. Ostrich to
Ducks. Barcelona: Lynx Edicions; 1992.
Demongin L, Poisbleau M, Raya Rey A, Schiavini A, Quillfeldt P, Eens M, et al.
Geographical variation in egg size dimorphism in rockhopper penguins. Polar
Biol 2010;33:469–76.
Donaldson GM, Shutt JL, Hunter P. Organochlorine contamination in bald eagle eggs
and nestlings from the Canadian Great Lakes. Arch Environ Contam Toxicol
1999;36:70–80.
Elliott JE, Martin PA. Chlorinated hydrocarbons and shell thinning in eggs of (Accipiter)
hawks in Ontario, 1986–1989. Environ Pollut 1994;86:189–200.
Elliott JE, Wilson LK, Wakeford B. Polybrominated diphenyl ether trends in eggs of
marine and freshwater birds from British Columbia, Canada, 1979–2002. Environ
Sci Technol 2005;39:5584–91.
Fernie KJ, Shutt JL, Letcher RJ, Ritchie JI, Sullivan K, Bird DM. Changes in reproductive
courtship behaviors of adult American kestrels (Falco sparverius) exposed to
environmentally relevant levels of the polybrominated diphenyl ether mixture, DE71. Toxicol Sci 2008;102:171–8.
Furness RW. Birds as monitors of pollutants. In: Furness RW, Greenwood JJD, editors.
Birds as monitors of environmental change. London: Chapman and Hall; 1993.
p. 86-143.
Gilbertson M, Kubiak T, Ludwig J, Fox G. Great Lakes embryo mortality, edema, and
deformities syndrome (glemeds) in colonial fish-eating birds: similarity to chickedema disease. J Toxicol Environ Health 1991;23:455–520.
Hair JF, Anderson RE, Tatham RL, Black WC. Multivariate Data Analysis. 5th ed.
New-Jersey, U.S.A: Prentice Hall; 1998. p. 87-138.
Hoerschelmann H, Polzhofer K, Figge K, Ballschmitter K. Organochlorpestizide und
polychlorierte biphenyle in vogeleiern von den Falklandinseln und aus Norddeutschland. Environ Pollut 1979;20:247–69.
Huin N. Falkland Islands penguin census 2005/06. no. Stanley: Falklands Conservation;
2007.
4.3. Laying order effects
For the rockhopper penguins, concentrations were smaller in the
A-egg compared to the B-egg, although this difference was not
significant. Similarly, for the imperial shags, we also found no
significant difference in OHC concentrations between different eggs
from a clutch. In addition, both for the rockhopper penguin and the
imperial shag, contamination profiles did not significantly differ in
relation to the laying order. This suggests that, for both species, all
eggs of a clutch can be used for biomonitoring OHCs. However, for the
rockhopper penguin, we recommend using A-eggs for monitoring
purposes, because most of the A-eggs or newly hatched chicks from Aeggs are lost during the hatching period (Strange, 1982; Lamey, 1993;
Poisbleau et al., 2008; Demongin et al., 2010).
5. Conclusions
Although our results showed that the Falkland Islands have been
reached by OHCs, concentrations were relatively low compared to
other studies. Therefore, the study populations are not likely at risk for
adverse health effects. However, future monitoring may be warranted
to assess the presence and time trend of different OHCs. Different
factors, such as diet, feeding behaviour and migration may be
responsible for the observed differences in concentrations and
profiles between southern rockhopper penguins and imperial shags.
For both species, concentrations and profiles did not significantly
change in relation to the laying order. This suggests that, both species
are useful for biomonitoring purposes and any egg of a clutch can be
used to monitor OHCs.
Acknowledgements
We are grateful to the New Island Conservation Trust for the
permission to work on the island and for the logistic support. We wish
to thank Ian, Maria and Georgina Strange for their support during the
field season. All work was conducted under a research licence granted
by the Environmental Planning Department of the Falkland Islands
Government. Evi Van den Steen, Maud Poisbleau and Adrian Covaci
are postdoctoral fellows of the FWO (Fonds Wetenschappelijk
Onderzoek—Vlaanderen). Alin C. Dirtu acknowledges financial support from the University of Antwerp. Marcel Eens and Rianne Pinxten
are supported by the University of Antwerp and FWO. Fieldwork was
funded by a grant of the Deutsche Forschungsgemeinschaft (DFG) to
Petra Quillfeldt (Qu 148-1ff).
Appendix A. Supplementary material
Supplementary data to this article can be found online at
doi:10.1016/j.scitotenv.2011.04.002.
2844
E. Van den Steen et al. / Science of the Total Environment 409 (2011) 2838–2844
Jaspers V, Covaci A, Maervoet J, Dauwe T, Voorspoels S, Schepens P, et al. Brominated
flame retardants and organochlorine pollutants in eggs of little owls (Athene
noctua) from Belgium. Environ Pollut 2005;136:81–8.
Jones KC, De Voogt P. Persistent organic pollutants (POPs): state of science. Environ
Pollut 1999;100:209–21.
Jörundsdóttir H, Bignert A, Svavarsson J, Nygård T, Weihe P, Bergman Å. Assessment of
emerging and traditional halogenated contaminants in guillemot (Urua aalge) eggs
from North-Western Europe and the Baltic Sea. Sci Total Environ 2009;407:
4174–83.
Lamey TC. Territorial aggression, timing of egg loss, and egg-size differences in
rockhopper penguins, Eudyptes c. chrysocome, on New Islands, Falkland Islands.
Oikos 1993;66:293–7.
Marsh G, Athanasiadou M, Bergman Å, Asplund L. Identification of hydroxylated and
methoxylated polybrominated diphenyl ethers in Baltic Sea Salmon (Salmo salar)
blood. Environ Sci Technol 2004;38:10–8.
Masello JF, Mundry R, Poisbleau M, Demongin L, Voigt C, Wikelski M, et al. Diving
seabirds share foraging space and time within and among species. Ecosphere
2010;1(19) (28 pp).
McCarty JP, Secord AL. Nest-building behaviour in PCB-contaminated tree swallows.
Auk 1999;116:55–63.
Michalik A, van Noordwijk HJ, Brickle P, Eggers T, Quillfeldt P. The diet of the Imperial
Shag Phalacrocorax atriceps at a colony on New Island, Falkland/Malvinas Islands
combining different sampling techniques. Polar Biol 2010;33:1537–46.
Mineau P. Levels of major organochlorine contaminants in sequentially laid herring gull
eggs. Chemosphere 1982;11:679–85.
Newton I, Bogan J. The role of organo-chlorine compounds in breeding of British
sparrowhawks. J Appl Ecol 1978;15:105–16.
Norstrom RJ, Simon M, Moisey J, Wakeford B, Weseloh DVC. 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.
Nisbet ICT. Eggshell characteristics and organochlorine residues in common terns:
variation with egg sequence. Colonial Waterbirds 1982;5:139–43.
Poisbleau M, Demongin L, Angelier F, Dano S, Lacroix A, Quillfeldt P. What ecological
factors can affect albumen corticosterone levels in the clutches of seabirds? Timing
of breeding, disturbance and laying order in rockhopper penguins (Eudyptes
chrysocome chrysocome). Gen Comp Endocrinol 2009a;162:139–45.
Poisbleau M, Demongin L, Strange IJ, Otley H, Quillfeldt P. Aspects of breeding biology of
the southern rockhopper penguin Eudyptes c. chrysocome and new consideration
on the intrinsic capacity of the A-egg. Polar Biol 2008;31:925–32.
Poisbleau M, Demongin L, Trouvé C, Quillfeldt P. Maternal deposition of yolk
corticosterone in clutches of southern rockhopper penguins (Eudyptes chrysocome
chrysocome). Horm Behav 2009b;55:500–6.
Pütz K, Clausen AP, Huin N, Croxall JP. Re-evaluation of historical rockhopper penguin
population data in the Falkland Islands. Waterbirds 2003;26:169–75.
Pütz K, Ingham RJ, Smith JG, Lüthi BH. Winter dispersal of rockhopper penguins
Eudyptes chrysocome from the Falkland Islands and its implications for conservation. Mar Ecol Prog Ser 2002;240:273–84.
Raya Rey A, Schiavini A. Inter-annual variation in the diet of female southern rockhopper
penguin (Eudyptes chrysocome) at Tierra del Fuego. Polar Biol 2005;28:132–41.
Roosens L, Van Den Brink N, Riddle M, Blust R, Neels H, Covaci A. Penguin colonies as
secondary sources of contamination with persistent organic pollutants. J Environ
Monit 2007;9:822–5.
Statsoft. STATISTICA for Windows. Tulsa, Oklahama: Statsoft Inc; 1997.
Stonehouse B, editor. The biology of Penguins. London: Macmillan; 1975.
Strange IJ. Breeding ecology of the rockhopper penguin (Eudyptes crestatus) in the
Falkland Islands. Gerfaut 1982;72:137–88.
Teuten EL, Xu L, Reddy CM. Two abundant bioaccumulated halogenated compounds are
natural products. Science 2005;307:917–20.
van den Berg M, Craane BLHJ, Sinnige T, van Mourik S, Dirksen S, Boudewijn T, et al.
Biochemical and toxic effects of polychlorinated biphenyls (PCBs), dibenzo-pdioxins (PCDDs) and diben-zofurans (PCDFs) in the cormorant (Phalacrocorax
carbo) after in ovo exposure. Environ Toxicol Chem 1994;13:803–16.
Van den Steen E, Dauwe T, Covaci A, Jaspers VLB, Pinxten R, Eens M. Within- and
among-clutch variation of organohalogenated contaminants in eggs of great tits
(Parus major). Environ Pollut 2006;144:355–9.
Van den Steen E, Pinxten R, Jaspers VLB, Covaci A, Barba E, Carere C, et al. Brominated
flame retardants and organochlorine pollutants in the European environment
using great tit eggs as a biomonitoring tool. Environ Int 2009;35:310–7.
van der Schalie WH, Gardner HS, Bantle JA, De Rosa CT, Finch RA, Reif JS, et al. Animals
as sentinels of human health hazards of environmental chemicals. Environ Health
Perspect 1999;107:309–15.
Verreault J, Gebbink WA, Gauthier LT, Gabrielsen GW, Letcher RJ. Brominated flame
retardants in glaucous gulls from the Norwegian Arctic: more than just an issue of
polybrominated diphenyl ethers. Environ Sci Technol 2007;41:4925–31.
Verreault J, Villa RA, Gabrielsen GW, Skaare JU, Letcher RJ. Maternal transfer of
organohalogen contaminants and metabolites to eggs of Arctic-breeding glaucous
gulls. Environ Pollut 2006;144:1053–60.
Vos JG, Dybing E, Helmut AG, Ladefoged O, Lambré C, Tarazona JV, et al. Health effects of
endocrine disrupting chemicals on wildlife, with special reference to the European
situation. Crit Rev Toxicol 2000;30:71-133.
Wania F, Mackay D. Tracking the distribution of persistent organic pollutants. Environ
Sci Technol 1996;30:390A–6A.
Watanabe K, Senthilkumar K, Masunaga S, Takasuga T, Iseki N, Morita M. Brominated
organic contaminants in the liver and egg of the common cormorants (Phalacrocorax carbo) from Japan. Environ Sci Technol 2004;38:4071–7.
Watson GE. Birds of the Antarctic and sub-Antartic, Washington. Union: D.C. Amer.
Geogr; 1975.
Weijs L, Losada S, Das K, Roosens L, Reijnders PJ, Santos JF, et al. Biomagnification of
naturally-produced methoxylated polybrominated diphenyl ethers (MeO-PBDEs)
in harbour seals and harbour porpoises from the southern North Sea. Environ Int
2009;35:893–9.
Weiss F, Furness RW, McGill RAR, Strange IJ, Masello JF, Quillfeldt P. Trophic segregation
of Falkland Islands seabirds—insights from stable isotope analysis. Polar Biol
2009;32:1753–63.
Williams TD. The penguins. Oxford, UK: Oxford University Press; 1995.
Yogui GT, Sericano JL. Levels and patterns of polybrominated diphenyl ethers in eggs of
Antarctic seabirds: endemic versus migratory species. Environ Pollut 2009;157:975–80.