In the recent decade
several monitoring studies have started to focus on not only "classic" persistent
organic pollutants (POPs), such as polychlorinated biphenyls
(PCBs), organochlorinated pesticides (OCPs), and/or
polychlorinated dibenzodioxins/polychlorinated dibenzofurans
(PCDDs/PCDFs) but also on other groups of halogenated
xenobiotics such as brominated flame retardantds (BFRs). These
chemicals are used mainly as additives in polymers to prevent
them from catching fire (de Wit 2002; Hale et al. 2003).
Generally, two types of BFRs can be distinguished: a) reactive
compounds, for instance tetrabromobisphenol A (TBBPA), are
incorporated by covalent binding into polymeric matrix and b) additive BFRs,
represented by polybrominated diphenyl ethers (PBDEs),
hexabromocyclododecane (HBCD), and/or polybrominated biphenyls
(PBBs), are merely dissolved in polymeric material. Although
TBBPA is used mainly in North America, the production of PBDEs
prevails in Europe (de Wit 2002; Rahman et al. 2001).
Products based on penta-, octa-, and
decabromodiphenyl ethers are currently the only commercially
interesting PBDEs (de Boer et al. 2000a). They are used in
the
housing and electronic parts of television sets or personal
computers and also in textiles [de Wit 2002; World Health
Organization (WHO) 1994, 1997]. PentaBDEs are mainly applied
in textiles and polyurethane foams, whereas decaBDEs are used
in
textile as well as in many other kinds of synthetic plastics
such as polyester used for electronic circuit boards (de Wit
2002; Petterson and Karlsson 2001). HBCD is used in foams and
expanded polystyrene and final products such as upholstered
furniture, interior textiles, and packaging material (de Wit
2002).
The occurrence of BFRs in various
environmental compartments is of great concern because of their
high lipophility (log Kow is
between 5 and 10) and/or high resistance to degradation
processes (Haglund et al. 1997). Although the first reports on
a presence of PBDEs in both abiotic and biotic matrices were
published as early as the late 1970s [see Zweidinger et al.
(1979) for early data on air particles] and the beginning of
the 1980s [see Andersson and Blomkvist (1981) concerning fish
from Swedish rivers], intensive investigation into their
occurrence in the environment started a decade later. Various
PBDE congeners were found in Dutch (de Boer et al. 2000b),
Swedish (Haglund et al. 1997; Sellström et al. 1993,
1998), Japanese (Ohta et al. 2002; Watanabe et al. 1987),
British (Allchin et al. 1999), and Canadian (Alaee et al. 1999)
fish samples. Similarly, these brominated POPs were also
detected in sediments, wastewaters, and air (Allchin et al.
1999; de Boer et al. 2003; Sellström et al. 1998). BFRs
may be released into the environment from many sources such
as a) landfills
(additive types may leach out); b) emissions originated during incineration
processes (brominated dioxins and furans may originate under
these conditions) (de Wit 2002); and/or c) effluents from sewage
treatment plants (STPs), and communal and industrial wastes.
Like other POPs, the BFR group is the
subject of a wide range of toxicologic and ecotoxicologic
studies. Some of these studies classified these chemicals as
endocrine disruptors. These substances may exhibit adverse
effects on the regulation of thyroid hormone and induce
immunotoxicity. They also induce neurotoxicity, causing
interferences at sensitive periods of brain development (de Boer
et al. 2000a; Rahman et al. 2001).
The goal of the present
study, which is the first conducted in the Czech Republic, was
to recognize the
extent of contamination by PBDEs and HBCD in aquatic
ecosystems. Several fish species common to the Czech rivers
Vltava, Elbe, and Tichá Orlice were used as biomonitors
for this purpose.
Sample collection. Five fish species—chub, barbel, bream, perch,
and trout—were caught at several sampling sites at three
Czech rivers (Vltava, Elbe, and Tichá Orlice) during
2001–2003 and were delivered to the laboratory in edible
form (fillets). Before storage at –18°C, fish samples
were pooled according to fish weight and length (parameters
correlating with age of the fish). Typically one pooled sample
was prepared from three to five individual fish. Lipid content
was determined in each composite sample using extraction by n-hexane:
dichlormethane (1:1, vol/vol). The characteristics of examined
samples are summarized in Tables 1 and 2; sampling sites are
shown in Figure 1.
Characterization of fish used as
biomonitors. The fish used as
biomonitors in the present study represented a spectrum of
freshwater species typically found in Czech aquatic ecosystems.
Chub (Leuciscus cephalus) is a relatively abundant fish found in Czech
rivers and is suitable as a bioindicator of contamination in
aquatic ecosystems. This omnivorous species grows slowly, and
its mean lipid content in muscle is about 2.5%. Barbel (Barbus barbus) and
bream (Abramis brama) have high lipid content in muscle (up to 7%);
for this reason they are able to bioaccumulate lipophilic
organic pollutants to a large degree. These fish live in close
contact with benthic sediments. Perch (Perca fluviatilis) and
trout (Salmo trutta) belong to a group of predators, and their
lipid content in fillet is relatively low (not more than 1%).
The fish were treated humanely and with regard for alleviation
of suffering.
Chemicals. Standard
solutions containing PBDE congeners (concentration 50 µg/mL
in nonane) are 2,4,4´-triBDE (BDE-28); 3,4,4´-BDE
(BDE-37); 2,2´,4,4´-tetraBDE (BDE-47); 2,2´,4,5´-tetraBDE
(BDE-49); 2,3´,4,4´-tetraBDE (BDE-66); 2,2´,3,4,4´-pentaBDE
(BDE-85); 2,2´,4,4´,5-pentaBDE (BDE-99); 2,2´,4,4´,6-pentaBDE
(BDE-100); 2,2´,4,4´,5,5´-hexaBDE (BDE-153); 2,2,4´,4,5´,6´-hexaBDE
(BDE-154); 2,2´,4,4´,5´,6-BDE (BDE-183) and
deca-BDE (BDE-209). All were obtained from Cambridge Isotope
Laboratories (≥ 98% pure; CIL, Andover, MA, USA). Working
standard solutions were prepared in isooctane and were stored
in a refrigerator (5°C). The α-HBCD standard (50 µg/mL in toluene) with
declared purity of 98% was supplied by CIL. A standard solution
of PCB-112 (10 µg/mL in isooctane) was purchased from Gr.
Ehrenstorfer GmBH (Augsburg, Germany).
The organic solvents (hexane,
cyclohexane, isooctane) declared as organic trace analysis
grade were supplied by Merck (Darmstadt, Germany). Ethylacetate
and dichloromethane were obtained from Scharlau (Barcelona,
Spain). Anhydrous sodium sulfate, supplied by Penta Chrudim
(Chrudim, Czech Republic), was heated at 600°C for 5 hr,
then stored in a desiccator before use.
Styrene–divinylbenzene gel (Bio Beads S-X3, 200–400
mesh) was purchased from Biorad Laboratories (Hercules, CA,
USA). Sulfuric acid (98%) was obtained from Merck.
Instruments. We used a homogenizer
(model 2094; Foss Tecator, Hilleroed, Denmark) to homogenize fish
samples. For the
extraction step, we used the Soxhlet extractor Gerhart 173200
EV (Gerhart, Königswinter, Germany) with a cellulose
extraction thimble (Whatman, Brentford, UK).
An automated gel permeation
chromatography (GPC) system consisting of 350 MASTER pump,
fraction collector, automatic regulator of loop XLI,
microcomputer (software 731 PC via RS32C), dilutor 402 (GILSON,
Villiers le Bel, France), and stainless steel column 500 x 8
mm inner diameter (i.d.) packed with Bio-Beads S-X3 (soft gel)
was used
for a cleanup of crude extracts.
A vacuum evaporator (Büchi Rotavapor
R-114) and water bath (B-480) (Büchi, Postfach,
Switzerland) were used for concentration of extracts.
We used an Agilent 6890 gas chromatograph
equipped with electronic pressure control (EPC),
split/splitless injector, and coupled to a mass selective
detector Agilent 5973 (Agilent Technologies, CA, USA).
Capillary columns used were the a) DB-XLB column (30 m
x 0.25 mm i.d. x 0.1-µm
film thickness), and b) BD-XLB (15 m x 0.25 mm i.d. x
0.1-µm film thickness (all from J&W
Scientific, Folsom, CA, USA) were employed for separation of
PBDEs and HBCD.
Extraction of fish samples. Thirty grams of homogenous fish muscle were
mixed with 120 g anhydrous sodium sulfate to form a flowing
powder. The sample was transferred into a cellulose extraction
thimble and stored in a desiccator for 12 hr to complete the
desiccation process, then inserted into a Soxhlet apparatus and
extracted for 8 hr (seven cycles per hour) with 340 mL solvent
mixture n-hexane:dichlormethane (1:1, vol/vol). The crude
extract was carefully evaporated by rotary vacuum evaporator,
and the residual solvents were removed by a gentle stream of
nitrogen. The lipid content was determined gravimetrically
using the analytical balance A&D MH–300 (A&D Co.,
Tokyo, Japan) with 0.001-g accuracy.
Cleanup. Extracted
lipids were dissolved in 10 mL of cyclohexane:ethylacetate
mixture (1:1, vol/vol) containing 5 ng/mL PCB-112 (this
congener is not present in commercial mixtures or environmental
samples); this was considered the recovery standard. Two
milliliters of this solution (corresponding to 6 g wet sample)
were loaded onto a GPC column. The mobile phase was cyclohexane:
ethylacetate (1:1, vol/vol) with a flow rate of 0.6 mL/min. The
fraction corresponding to the elution volume of 14–30 mL
was collected. The eluate was evaporated by rotary vacuum
evaporator, and the residual solvents were carefully eliminated
by a gentle stream of nitrogen to dryness. The residue was then
dissolved in 1 mL isooctane containing 1 ng/mL BDE-37
(3,4,4´-BDE) as syringe standard and treated with
concentrated sulfuric acid (approximately three drops) to
remove residual lipids. After 10 min of compete phase
separation, an aliquot of the upper organic (isooctane) layer
was taken and transferred into a glass vial for subsequent gas
chromatography (GC )analysis.
GC analysis. We used a high-resolution
GC (HRGC) unit resolution mass-selective detector (MSD) for analyses
of the PBDEs and
HBCD in purified extracts. The GC conditions (column 1) were
as follows: column temperature program, from 105°C (hold 2 min)
to 300°C at 20°C/min (hold 5 min); carrier gas, helium
(Linde, Prague, Czech Republic) with a constant flow of 1.5 mL/min;
injection temperature, 275°C; injection volume, 1 µL
using pulsed splitless injection mode (splitless time, 2 min).
An MSD with quadrupole analyzer was operated in a selective
ion-monitoring (SIM) mode in a negative chemical ionization
(NCI). Monitored ions (m/z) were 79, 81, 159, and 161 (PBDEs); 79, 81,
158, and 160 (HBCD); and 326 and 328 (PCB-112, internal
standard). Ion m/z 79 was used to quantify all target
analytes. Methane was used as a reagent gas (purity 99.995%,
Linde) and
was set at a pressure 2 x 10–4 mbar. Ion source temperature was
150°C and quadrupole temperature 105°C.
We monitored the presence of decaBDE
using the same GC coupled with negative chemical ionization
mass spectrometry (GC/MS-NCI) employing a shorter column
(column 2). The temperature program was
as follows: from
80°C (hold 2 min) to 280°C at 20°C/min and to
320°C at 5°C/min (hold 5 min); carrier gas, helium with
constant flow 3 mL/min; injection temperature, 285°C;
injection volume, 1 µL using pulsed splitless injection
mode (splitless time, 2 min). Monitored ions were m/z 485 and 487;
the ion at m/z 487 was used for quantification.
We identified the target analytes by
comparing their retention times with retention times of
standards and by MS confirmation. For quantification, a
multilevel calibration curve was used (at least 5 points for
each congener).
Quality assurance. For each extraction
batch (consisting of five fish samples), one procedure blank was
processed. The results were
corrected for blank interferences and for recovery (PCB-112 was
added as surrogate before GPC cleanup). Limit of detection
(LOD) was calculated as quantity of analyte that generates a
response 3 times greater than the noise level of the detection
system. Limits of quantification (LOQs) were the minimum
concentrations of analytes possible to quantify with acceptable
accuracy and precision. Under these conditions, the LOQ was the
lowest calibration level and corresponded for particular
analyte to 3 x LOD.
LOD values (nanograms per gram lipid
weight) for fish were BDE-28, 0.015; BDE-47, 0.015; BDE-49,
0.015; BDE-66, 0.015; BDE-85, 0.02; BDE-99, 0.015; BDE-100,
0.015; BDE-153, 0.02; BDE-154, 0.015; BDE-183, 0.015; BDE-209,
2.0; and HBCD, 0.1.
For recovery testing
of the overall analytical method, chub muscle was spiked at level
2 ng/g (of
each analyte) by 100 µL standard mixture (500 ng/mL) in
acetone. Real-life samples were also analyzed to obtain
background levels of analytes. PBDE recoveries ranged between
83–101%, and recovery of HBCD was 91%. Acceptable
recovery rate was 80–110%. We also determined the
precision of the analytical method (repeatability) by analyzing
six spiked fish samples; repeatability ranged from 4 to 12%
(expressed as relative SD). Recovery of BDE-209 was 78 ± 3%
(n
= 6). Chub muscle samples spiked at 20 ng/g wet weight and were
analyzed within the validation process. The method we used is
fully validated. The repeatability of our results is documented
by our participation in certification study BROC (biological
reference materials for organic contamination) (van Leeuwen et al.
2006).
As mentioned previously,
fish is widely used as a biomonitor of bioavailable POPs that
occur in aquatic
environments. However, interpretation of obtained data is not
simple. Both bioaccumulation and depuration processes may take
place in aquatic biota simultaneously, and the ratio of their
intensities may differ widely among the fish species. It should
be noted that the concentration of POPs measured in their
bodies is dependent on many factors such as age, sex, and/or
feeding habits of particular resident species. In practice it
is difficult to obtain homogenous sets of biomonitors from an
entire river. Differences exist among sampling localities in
terms of food availability, causes of variations of fat
content, and hence varying accumulation potential in fish.
Table 3 is a summary of fish characteristics and the results
(based on wet weight) of target PBDEs and HBCD (sum of isomers)
in collected samples. It should be noted that technical HBCD
mixtures consist of three diastereomers—α, β, and
γ—the
last being the typically dominating component (up to 80%) of
this primary polluting material. In other words,
biotransformation of γ-HBCD may occur in biota, resulting
in a changed contamination pattern, which may lead under certain
circumstances (e.g., biomagnification) to α-HBCD becoming
the dominant component in the diastereomers profile. One should
be
aware that under GC conditions (hot injection), thermal
conversion of γ-HBCD yielding α-diastereomer also may
occur. Therefore, in most studies using GC for quantification,
α-diastereomer is used as the calibration standard representing
HBCD groups.
For
determination of all individual HBCD diastereomers, LC/MS must
be used (Morris et al. 2004).
Table 3 shows that
in all examined fish samples, the major PBDE congener was BDE-47.
Levels of this
2,2´4,4´-tetrabromodiphenyl ether were
approximately one order of magnitude higher than those of other
monitored congeners. This was not surprising, as BDE-47 was a
main component in various kinds of technical mixtures (e.g.,
Bromkal 70-5DE) commonly used in industry. As in our samples,
this congener typically makes the major contribution to the
total PBDE content in the environmental samples collected in
Europe. Pentabromodiphenyl ether congeners BDE-99 and BDE-100,
and hexabromodiphenylether congeners BDE-153 and BDE-154 were
also present in most samples. The levels of these congeners
exceeded the LOD in 70% of fish, and at least one of these
PBDEs was detected. The presence of BDE-49 was confirmed in
only about 10% of the samples; BDE-66 and BDE-183 were not
detected in any sample. In accordance with similar studies (de Boer
et al. 2003; Eljarat et al. 2004, 2005), no detectable
decabromodiphenyl ether (congener 209) was present in any
examined fish sample. According to several authors (Geyer et al.
1999; Sellström et al. 1998), the superlipophilic nature
of this chemical (log Kow ~
10) might be responsible for the lack of detection. BDE-209 can
be
strongly bound to sediments, hence its actual dissolved
concentration in water is very low, and thus only a negligible
fraction of this BFR is expected to be bioavailable to fish.
As
discussed by Eljarat et al. (2005), the low bioaccumulation
potential of this chemical is due to its large molecular size
that hinders a passage over membranes (Andersson and Blomkvist
1981). The alternative explanation of minimal occurrence of the
deca-BDE congener in aquatic organisms is its rapid excretion
and/or biotransformation after entering their body (Eljarrat
et
al. 2005; Eriksson et al. 2004). Regardless, BDE-209 is a
relatively labile substance that easily decomposes under
environmental conditions in yielding a large range of lower
brominated congeners in addition to other bromine-containing
products when illuminated by sunlight (Eriksson at al. 2004;
Söderström et al. 2004).
The presence of HBCD in fish collected in
2002 and 2003 was detected in more than 80% of tested samples,
with the highest contamination found in fish species from
Srnojedy (Elbe River). Figure 2A,B shows examples of
concentrations of BDE-47, other ΣPBDEs (congeners 28, 49,
85, 99, 100, 153, and 154), and HBCD in chub from all sampling
localities. The
average concentration of BFRs in fish from Klecany at Vltava
River (aggregated data obtained within the monitoring period)
was almost 5 times that of samples obtained in Podolí
upstream from Prague. The data obtained by analysis of chub
from the Elbe River in 2001–2003 indicated that Srnojedy,
located downstream from Pardubice (a large industrial area) was
the most polluted locality along the Elbe River.
|
Figure 1. The
sampling sites on the Czech rivers.
|
|
Figure 2. Concentration of BDE-47, other ΣPBDEs
(BDE-28, -49, -66, -99, -100, -153, and -154), and HBCD in chub
samples from sampling sites (ng/g wet
weight).
Error bars represent mean ± SD. (A) River Vltava and (B) River Elbe.
|
|
Figure 3. Comparison of BDE-47, other ΣPBDEs
(BDE-28, -49, -66, -99, -100, -153, and -154), and HBCD levels
in tested fish species in Hrensko on the
Elbe river (ng/g wet weight) from 2001 to 2003. Error bars
represent mean ± SD.
|
|
Figure 4. Pattern
of PBDE congeners in various fish species in Elbe-Hrensko.
Error bars represent mean ± SD.
|
|
Figure 5. Comparison
of PBDE and PCB content in fish collected in six sampling
localities (aggregated data).
|
Table 1.
|
Table 2.
|
Table 3.
|
Table 4.
|
In Hluboká n/V (Vltava River) and
Hrensko (Elbe River), no significant variation among the
monitoring years was found, whereas the concentrations of BFRs
in the lower part of the Elbe River were largely variable
(Figure 2B). In addition to being caused by increasing
pollution, this trend might be attributed to differences of
seasonal flows in this part of Elbe River for individual years.
Extreme floods in 2002 were probably accompanied by the removal
of contaminated sediments from monitored localities in the
upper part of the Elbe River and the apparent drop of aquatic
ecosystem pollution, hence reduction of fish exposure to
bioaccumulatively chemicals. On the other hand, the total
rainfall in the upper part of the Elbe River basin in 2003 were
below the long-term average values (ELbe InformationsSystEm
2006) and the flow was low. Because of the existence of
permanent emission sources (industrial wastes) of PBDEs and
HBCD along the upper part of the Elbe River, the increases in
pollution at monitoring localities occurred again in the
following year.
Figure 3 shows large
differences in the extent of PBDE and HBCD bioaccumulation among
the examined fish
species. It is important to note that regardless of the
monitoring year and sampling place, the concentrations of these
BFRs (based on wet weight) were found in the following order:
barbel > bream > chub > perch. de Boer and Brinkman
(1994) and Geyer et al. (1999) have shown that the contribution
of persistent organohalogen compound buildup in the food chain
becomes relevant when log Kow values
are > 5–6.5. Because log Kow values
of major PBDE congeners monitored in our study range from 5.5
to 7, biomagnification of these
chemicals (i.e., their transfer within the trophic levels of
examined fish species that leads to a stepwise increase in
contamination) might be expected. Similarly, higher levels of
PBDEs in lipids were, for example, found in bass, a predator
fish (collected in Penobscot River in central Maine), compared
with those in white sucker, a benthic feeder, from the same
locality (Anderson and MacRae 2006). Conversely, the lowest
extent of contamination (regardless of the expression of BFR
content on wet weight or lipids) was found in perch, which
represent the highest trophic level among the fish examined in
our study. Similar trends were also reported in other studies
(Table 4). For example, Covaci et al. (2006) showed that the
concentration of PBDEs in benthic bream from the Danube delta
in Romania was about 50% higher than that in predator perch
from the same locality. This controversy could be attributed
to differing fat content in these two species, which is in
addition to other factors related to differences in their
feeding habits. Another reason for the lower concentration of
PBDEs in predator species such as perch might be the
fish's higher growth rate (the ratio between weight and
age), leading to "dilution" of accumulated
pollutants because of a rapid increase of fish muscle tissue.
Generally, slow-growing fish species are exposed to polluted
environments for a longer time. Moreover, in the case of
benthic species (represented in our study by barbel and bream),
intensive contact with highly contaminated sediments is also
considered a factor in the higher levels of their contamination
(Covaci et al. 2006).
Substantial differences
were observed between the contamination pattern of perch and
other fish
species. Figure 4 shows aggregated data obtained for four
experimental biomonitors collected in Hrensko (Elbe
River) in two monitoring years, 2001 (before floods) and 2003
(after floods, low rainfalls). As illustrated in the figure,
the spectrum of PBDEs (regardless of their total
concentrations, see Table 3) was almost identical to that found
in omnivorous and benthic fish with distinctly dominating BDE-47
(40–75% of the total PBDE content). In perch the content
of BDE-99 was equal to or even higher than that of BDE-47.
Their contribution to the total PBDEs ranged between 30 and 40%
and 25 and 45%, respectively. In most studies (Covaci et al.
2005; Luross et al. 2002; Sellström et al. 1998; de Boer
et al. 2003), the dominanting congeners were also BDE-47, BDE-99,
and BDE-100. On the other hand, in Spanish studies by Eljarat
et al. (2004 and 2005), hexa-BDEs (BDE-153 and BDE-154) as well
as hepta-BDE (183) were the most abundant BFRs occurring in
fish (barbel and bream) samples. Probably less common PBDE
technical mixtures were released into the aquatic environment.
In Figure 5 the mean values of PBDE
content (aggregated data) are compared with PCB levels (ΣPCB,
PCB-28, -52, -101, -118, -138, -153, and -180) determined in
the same
fish
samples in a parallel study concerned with chlorine-containing
POPs (Pulkrabová J, Suchan P, Kocourek V, Hajslová J,
Pudil F. Unpublished data 2006). Typically, the content of PCBs
in biomonitors was higher by one order of magnitude and
generally did not correlate with the extent of contamination
in by PBDEs in particular localities. Differences in pollution
sources were documented by large variations in the PCB/PBDE
ratios calculated for individual fish species. Barbel showed
the tendency of fish species to accumulate more PCBs than PBDEs
(in Klecany, this phenomenon was not pronounced, with the
PCB/PBDE ratio of 59). The correlation coefficient
characterizing the relationship between the 10 ΣPBDE
congeners and the 7 indicator ΣPCB congeners in all examined
fish species was 0.39 (p < 0.05). It is reasonable
to believe that such low correlation clearly indicates the independence
of PBDEs and
organochlorine compounds as sources of pollution.
In North-East Atlantic coastal ecosystems,
levels of BFR compounds generally decreased as a function of
increasing latitude (Figures 2 and 3). The obvious reason
for
this is that the use and leakage of BFRs into the environment
is higher in urbanized areas along the Norwegian coast than
in
the almost unpopulated Spitsbergen. High levels of BFRs have
been reported in sewage [see review by Law et al. (2006a)],
and
the source of BFRs in the southern part of the study area is
most likely local discharges from urban sewage and industrial
activity. Because of their semivolatile properties, POPs are
subject to long-range atmospheric transport (Wania and Mackay
1993, 1996), and this is thus most likely the origin of the
BFRs detected in endemic Arctic biota.
The finding herein—that levels of
BFRs are lower in marine Arctic ecosystems than in temperate
marine ecosystems—is also in accordance with previous
reports in marine mammals. Data compiled on PBDEs in marine
mammals from temperate environments and from the Canadian
Arctic showed that levels of PBDEs were about 1,000 times
higher in marine mammals from temperate marine ecosystems than
in ringed seals from the Arctic (Ikonomou et al. 2002).
Furthermore, levels of Σ;PBDEs (BDE-17, BDE-47, BDE-49, BDE-99,
BDE-100, BDE-119, BDE-140, BDE-153, BDE-154, BDE-183) in harbor
porpoises also decreased as a function of increasing latitude
along the Norwegian coast, and were lower in animals from
Iceland than from Norway (Thron et al. 2004). Concentrations
of polychlorinated biphenyls (PCBs) in seawater in the North-East
Atlantic have been shown to decrease as function of increasing
latitude (Sobek and Gustafsson 2004). This confirms that levels
of POPs in marine biota generally decrease as a function of the
distance from the release areas.
When we compare organisms that occupy
similar trophic levels, the Arctic is still a pristine
environment with respect to organohalogenated anthropogenic
compounds. This is contrary to the beliefs of many politicians,
governmental bureaucrats, and nongovernmental organizations,
who, because of the particularly high levels of PCBs reported
in polar bears (Ursus maritimus), seem to believe that
the Arctic is heavily polluted by POPs. The high levels of PCBs
in polar bears are
attributed to the fact that this species is an apex predator
that feeds almost exclusively on the blubber of seals (Derocher
et al. 2002). Thus, because of biomagnification of the most
persistent PCB congeners from seals to polar bears, levels of
Σ;PCB in polar bears become very high (Bernhoft et al. 1997;
Skaare
et al.
2002). Levels of all BFR compounds analyzed herein (except for
BDE-153) were lower in polar bears than in its main prey
species, the ringed seal (Sørmo et al. 2006), most
likely because the polar bear has a high ability to metabolize
POPs (Letcher et al. 1996).
The clear latitudinal decrease in levels
of BFRs was not that pronounced in the two tern species
compared with the other species included in the study (Figures
2 and 3). This is most likely linked to the fact that the terns
are migratory birds, whereas the other species are endemic
to
their regions. During their migration from Africa (common tern)
and Antarctica (arctic tern), they feed along the highly
urbanized and thus more polluted coasts of Europe. Therefore,
even though the terns may metabolize and excrete some of the
BFR compounds during their migration via urbanized and
industrialized polluted areas, levels still seem to be
relatively high when they reach their breeding areas. Because
migration is energetically costly, the birds will have to build
up lipid stores for egg laying when arriving at their breeding
sites. Thus, because levels of POPs are lower in prey at more
northern breeding sites, body burdens of POPs in female birds
will be diluted, resulting in the latitudinal decrease of BFR
compounds in their eggs reported herein. When the eggs are
laid, the lipophilic BFRs are transferred from the female to
her eggs.
The high levels of HBCD
reported in common tern eggs from the Netherlands [330–7,100
ng/g lw (Morris et al. 2004)] occur most likely because specimens
that breed
here are exposed to higher concentrations for a longer period
of time than specimens that only transiently pass the
Netherlands en route to Norway and the Arctic.
In another study on kittiwakes (Rissa tridactyla),
levels of the sum of 23 PCB congeners and the Σ;PBDEs
(BDE-28, BDE-47, BDE-99, BDE-100, BDE-153, BDE-154) in newly
hatched
chicks did not differ between the west coast of Norway (Runde,
62° N) and Spitsbergen (Kongsfjorden, 79° N) (Murvoll
et al. 2006b). The apparent lack of a latitudinal decrease in
PBDE levels in kittiwakes may be because they winter in the
North Atlantic, close to where these compounds are used and released.
In calanoid species, HBCD was detected
only in the Oslofjord (Figure 3A). In Atlantic cod, levels of
PBDEs (Figure 2B) and HBCD (Figure 3B) were somewhat higher
in
the Oslofjord than at Froan, and were lowest in polar cod from
Spitsbergen. Levels of PBDEs in the Oslofjord were considerably
lower than concentrations reported in cod from 16 different
locations in the North Sea and Skagerrak (Boon et al. 2002).
Because food webs are complex,
and because few species were studied, we acknowledge that it
is difficult
to estimate biomagnification rates in the different ecosystems
included in this study. However, our crude approach, assuming
a
simple cod–harbor seal food chain, will still give some
information on biomagnification processes of BFR compounds in
the three coastal ecosystems.
In the Oslofjord, the biomagnification of
BDE-99 from Atlantic cod to harbor seal was particularly high
(Table 2), perhaps because BDE-99 constitutes a large part
of
the technical penta-BDE mixture, and the releases of this
compound into the environment probably has been high. The high
level of BDE-99 reported in the nearby Drammens fjord (Zegers
et al. 2003) supports this. Furthermore, levels of BDE-99 were
quite high in calanoids from the Oslofjord (Table 1). Much
higher levels of BDE-99 than reported herein have been reported
in more pelagic stocks of Atlantic cod in the Skagerrak and
the
North Sea (Boon et al. 2002). It is also possible that harbor
seals in the Oslofjord prefer to prey on cod from the pelagic
stocks. The Atlantic cod sampled in this study may have
belonged to a more coastal bound stock which the harbor seal
does not prefer to prey on, and this may have resulted in
an
overestimation of the BMF for BDE-99. It should also be noted
that BDE-99 is meta-para-substituted and consequently not easily
metabolized (Veltman et al. 2005). These factors may help
explain the high BMF of BDE-99 from Atlantic cod to harbor
seals in the Oslofjord.
The BMF of BDE-153 was also high in harbor
seals from the Oslofjord, perhaps because BDE-153 has a
substitution pattern similar to that of PCB-153, which is the
most persistent PCB compound. It is therefore possible that
the
high BMF of BDE-153 in the Oslofjord is caused by its
persistence.
In most species from all locations, BDE-47
was the most abundant congener (Table 1). This congener was
also biomagnified throughout the three food chains (Table 2).
BDE-47 constitutes approximately 25% of the technical penta-BDE
mixture (Hites 2004) and has thus been released to the
environment in relatively large volumes. Furthermore, there are
indications that in fish, BDE-99 (which constitutes ~ 50% of
the penta-BDE mixture) is debrominated to BDE-47 (Stapleton et
al.
2004), and this may thus lead to a further biomagnification of
BDE-47 in marine food chains.
Even though BDE-209
often is the predominating PBDE congener in marine sediments
(de Boer et al.
2003), it has been reported to contribute very little to the
total PBDE burden in organisms (Law et al. 2006a). This is
believed to be caused by the large molecular size of the
compound and the resultant low transfer over cells and uptake
into the organisms (Stapleton et al. 2004). Recently, there has
been a growing body of evidence that suggests that BDE-209 is
bioaccumulated to a larger extent in terrestrial food chains
than in marine food chains (Law et al. 2006a). However, BDE-209
has been reported to account for > 50% of total BDE burden
in the detritus feeding ice-amphipod Gamarus wilkitzkii at
Spitsbergen (Sørmo et al. 2006). Herein, BDE-209 was
detected in animals from all the three ecosystems (Table 1).
Because BDE-209 is almost ubiquitous, all possible efforts were
made to avoid contamination of the samples during sampling,
storage, and analysis. During the analyses, blank samples were
run parallel to the samples to control for possible
contamination in the laboratory, and no such contamination
could be identified.
The highest BDE-209 levels were found in
arctic tern eggs from Spitsbergen (Table 1). Further, it should
be noted that the highest concentration of BDE-209 relative to
Σ;PBDEs was found in polar cod from Spitsbergen (ca. 16%
of Σ;PBDEs),
harbor seals from Spitsbergen (~ 3%), and arctic terns from Spitsbergen
(~ 2%).
BDE-209 has a strong affinity to
particles. It is therefore possible that the detected levels
in the calanoid species and in the two cod species are associated
with the cuticle/skin or sediment particles and/or prey species
in the intestines (Law et al. 2006a; Leonards et al. 2004).
However, the detection of BDE-209 in tern eggs and seal blubber
shows that it is accumulated also in marine food chains. This
is consistent with reports that BDE-209 was bioaccumulated
in
grey seal given a supplement of this congener in their diet
(Thomas et al. 2005). BDE-209 has recently also been reported
in adipose tissue and plasma from polar bears and glaucous
gulls (Larus hyperboreus) from Spitsbergen (Sørmo
et al. 2006; Verreault et al. 2005). The relatively high contribution
of
BDE-209 to Σ;PBDEs in animals from Spitsbergen demonstrates
that this congener is subject to long-range transport and dispersal.
Whereas the data herein indicate that
BDE-209 may be biomagnified from polar cod to harbor seals
(Table 2), this was not the case from Atlantic cod to seals at
Froan. Thus, the potential of BDE-209 to be transferred in
food
webs is unclear. The technical deca-BDE-mixture, in which
BDE-209 is the major congener, presently constitutes about 80%
of the world market demand of PBDEs (de Boer et al. 2003).
Thus, there is a clear need for more information on the ability
of BDE-209 to biomagnify and/or be debrominated in marine
ecosystems.
At Spitsbergen, levels of both PBDEs and
HBCD were higher in ringed seals than in harbor seals (Table
1). The most obvious differences between these two seal species
were related to the much higher levels of BDE-47, HBCD, and
BDE-100 in ringed seals. These differences are most likely
related to differences in species-specific differences in their
ability to metabolize and biotransform the BFR compounds, and
possibly also related to differences in prey preferences.
There are few reports concerning levels of
HBCD in marine ecosystems (Morris et al. 2004; Stapleton et al.
2006; Wolkers et al. 2004; Zegers et al. 2005). Herein, HBCD
were
found in animals from all trophic levels, except in calanoids
at Froan and at Spitsbergen. The commercial HBCD mixtures
mainly consist of the three stereoisomers γ-HBCD
(75–89%), α-HBCD (10–13%), and β-HBCD
(1–12%)
(Heeb et al. 2005). In biota, the HBCD isomer composition
changes, and α-HBCD dominates (Law et al. 2006a). In the
present study, we did not distinguish among the different
isomers of HBCD. However, in aquatic invertebrates, marine
fish, birds, and marine mammals HBCD is present predominantly
as α-HBCD (Covaci et al. 2006).
In cod, seals, and terns,
HBCD levels seemed to be similar in the Oslofjord and at Froan,
whereas
levels were much lower at Spitsbergen (Figure 3B–D),
except in ringed seals (Table 1). The particular high levels of
HBCD in the ringed seals indicate that the bioaccumulation
potential of HBCD in this species may be particularly high
(Sørmo et al. 2006). Previously, it has been reported
that HBCD does not seem to biomagnify from ringed seals to
polar bears (Sørmo et al. 2006) possibly because polar
bears generally have a large capacity to metabolize
organohalogenated compounds (Letcher et al. 1996).
In common dolphins (Delphinus delphis) from
the Central and South Atlantic coast of Europe (Scotland,
Ireland, the Netherlands, Spain), median concentrations of HBCD
ranged from 200 to 900 ng/g lw, whereas median concentrations
in harbor porpoises ranged from 100 to 5,100 ng/g lw (Zegers et
al. 2005). Levels were highest in the Irish Sea and in
North-East Scotland. In blubber of two harbor seals from the
western Wadden Sea, concentrations of HBCD ranged from 63 to
2,055 ng/g lw (Morris et al. 2004). In stranded and by-caught
harbor porpoises in the United Kingdom, HBCD levels ranged from
11 to 21,300 ng/g lw (Law et al. 2006b), and a time-trend
analysis of the data strongly indicated a sharp increase in
HBCD concentrations from about 2001 onward. The HBCD levels
reported in cetaceans from the South- and Central-East Atlantic
coast, and in the harbor seals from the Wadden Sea are much
higher than those found in harbor seals from the Oslofjord,
Froan, and Spitsbergen (Table 1).
Relatively high levels of HBCD have also
been reported in hatchlings of kittiwakes from the Norwegian
west coast (~ 260 ng/g lw; Runde, 62° N) and levels were
somewhat lower in hatchlings from Spitsbergen (~ 120
ng/g lw; Kongsfjorden 79° N) (Murvoll et al. 2006b). Furthermore,
even higher levels of HBCD were reported in hatchlings of
European shags (Phalacrocorax
aristotelis) from the western
Norwegian coast (~ 420 ng/g lw; Sklinna 65° N) (Murvoll et
al. 2006a). The much higher levels in the European shag and
kittiwake hatchlings may be related to differences in the
analytical matrix (whole egg herein vs. yolk sac in
hatchlings). However, the differences may also be related to
the fact that kittiwakes and European shags winter in the North
Sea, the Norwegian Sea, and along the Canadian east coast,
whereas common and arctic terns migrate to the more pristine
areas in Africa and Antarctica, respectively. In North Sea
estuaries (United Kingdom and the Netherlands), levels of HBCD
in cormorant livers (Phalacrocorax
carbo)
and common tern eggs were 330–7,100 and 138–1,320
ng/g lw, respectively (Morris et al. 2004). These concentrations
are higher than
those reported in the tern eggs herein.
Because HBCD has been reported to have
histopathologic and neurotoxic effects (Birnbaum and Staskal
2004; Darnerud 2003; Mariussen and Fonnum 2003), there is
cause
for concern about the spreading and uptake of this compound in
biota. In Californian sea lions (Zalophus
californianus) a significant
temporal increase in HBCD was reported from 1994 to 2004
(Stapleton et al. 2006), and in harbor porpoises from the
United Kingdom a sharp increase in HBCD concentrations was
found from about 2001 onward (Law et al. 2006b). Because there
currently are no restrictions on the use of HBCD (Stapleton et al.
2006), there are reasons to believe that the global spreading
of the compound will continue and that levels in Arctic biota
will increase with time.
As mentioned previously, this is the
first report on the occurrence of PBDEs and HBCD in freshwater
fish in Czech rivers. The results of the present study are
summarized as follows:
• Fish is
a suitable species for use as a bioindicator for monitoring BFRs
in aquatic ecosystem; the
greatest accumulation was measured in fatty benthic species
represented by barbel and bream. Conversely, the potential of
perch (predator fish) for bioaccumulation of these chemicals
was lower.
• Contamination
patterns and their extent in fish collected in the Elbe and Vltava
rivers are
comparable with those reported in other European studies
conducted in rivers in industrial areas. Technical mixtures
based on penta- congeners were probably the source of
pollution. When the entire data set generated in our present
study was compared with data sets in similar studies conducted
abroad, no extremely contaminated locality was found in Czech
rivers that we monitored; the extent of pollution was similar
to that in other industrial regions, for example, in Canada,
Sweden, and Spain.
• The levels of PBDEs were about
10–30 times lower than PCB levels determined in the same
fish samples. In barbel from Hrensko (Elbe River), the
level of PBDEs was 60 times lower compared with that of PCBs.
The concentrations of HBCD in fish were of the same order of
magnitude as those of the most abundant PBDE-47.
