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Biological effects and environmental contaminants in herring and Baltic Sea top predators
Suzanne Faxneld1, Björn Helander1, Britt-Marie Bäcklin1, Charlotta Moraeus1, Anna Roos1, Urs Berger2, Anna-Lena Egebäck2, Anna Strid3, Amelie Kierkegaard2, Anders Bignert1
___________________________________________
Naturhistoriska Riksmuseet
Enheten för miljöforskning och övervakning Box 50 007
104 05 Stockholm
Rapport nr 6:2014
2 Affiliations:
1. Department of Environmental Research and Monitoring, Swedish Museum of Natural History
2. Department of Applied Environmental Science, Stockholm University
3. Department of Materials and Environmental Chemistry, Stockholm University
Chemical analysis:
Organochlorines/bromines, perfluoroalkylated chemicals, siloxanes Department of Applied Environmental Science, Stockholm University Phenols
Department of Materials and Environmental Chemistry, Stockholm University PCDD/PCDF
Department of Chemistry, Umeå University
3 Summary
The present study has increased our knowledge of the contaminant situation in the Baltic and the contaminant stress to top-predators through several new time-series of less known
chemicals (cyclic siloxanes, halogenated phenols). It has also given us additional analysis of previously analysed contaminants (chlorinated, brominated and perfluorinated chemicals).
This report comprises the results of observed biological effects together with the studied contaminants collected during a period of more than 40 years of monitoring.
Overall, concentrations of phenols and siloxanes increased in herring. PFOS concentrations increased from the start of the time series, but after 2000 decreasing trends were indicated for grey seals, guillemots and white-tailed sea eagles (WTSEs). In contrast, concentrations of perfluoro carboxylic acids (PFCAs), increased in herring, grey seals and WTSEs from most of the collection areas, and in addition, some of these PFCAs also increased during the most recent ten years in WTSEs. Polybrominated flame retardants (PBDEs) increased from the start of the time series until 1990s and decreased after that for herring, seals and WTSE. But
concentrations of HBCDD are increasing in guillemots, seals and WTSEs. Concentrations of DDE and PCBs decreased from the start of the sampling period until today. The same trend was also seen for PCDD/F in guillemot from Stora Karlsö and in herring from
Ängskärsklubb.
Blubber thickness in seals decreased over time, but no consistent explanation was found.
Liver/heart index in adult seals decreased, which might be due to decreasing trends of pollution.
Intestinal ulcers increased from the 1980s until the middle of the 1990s thereafter the
prevalence of intestinal ulcer decreased. HBCDD and sPCDD/F showed positive correlations with intestinal ulcers.
Pregnancy rate in seals increased from the 1970s until today. Negative correlations with pregnancy rate were found for sPCB, sDDT, and sPCDD/F.
In WTSE, productivity increased from the mid-1980s to 2000 where it levelled off near the background level. Mean productivity was negatively correlated with DDE in egg content and with sPCB, sDDT, and sPCDD/F as exposure index.
Eggshell thickness in WTSE increased from the 1970s until today at all sites except in the inland samples. However, still the shell thickness has not reached the confidence interval for the reference eggs in the Baltic Sea. Also in guillemot, eggshell thickness increased from the 1970s until today.
The desiccation index of WTSE eggs showed improvement in health from the 1970s until today. Egg desiccation was negatively correlated with DDE and PCB in egg content.
A present target level of DDE was checked in one example with WTSE eggshell thickness. It was shown that the target value (EAC, Environmental Assessment Criteria provided by OSPAR) was too high to protect the avian fauna from adverse effects and needs to be revised.
This example also indicates that the target values of today, needs to be evaluated also in the environment, for potentially more sensitive species at the top of the food web impossible to test in a laboratory.
To study the possible causal relationships between biological effects and various
contaminants analysed from environmental samples is of course very difficult since, unlike
the situation in a laboratory, we cannot control factors like the mixture of contaminants and
also several biological confounders like age and condition. Despite the large number of
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analysed specimens the sample sizes with contrasting, low and high concentrations of
combinations of various contaminants becomes yet too small. The multiple regression analyses show in general too much multicollinearity to become meaningful. Some of the biological effects show clear statistical correlation with the analysed contaminants which implies that we cannot rule them out as the possible cause of that effect, but since we cannot guarantee that all relevant contaminants (or other relevant factors) are included in the
analysis, we cannot establish causal link between a certain effect and a specific chemical or a combination of chemicals. However, the advantages with environmental studies are obviously that they reflect a realistic exposure situation and can give indications that can be tested in the laboratory. The long term monitoring is in this respect invaluable.
Acknowledgements
Several persons have been involved at various stages in this long term project. We would especially want thank: Elin Boalt, Henrik Dahlgren, Mats Hjelmberg, Anh Le Van, Aroha Miller from the Swedish Museum of Natural History.
Petra Wallberg at the former “Havsmiljöanslaget” for fruitful discussions at the initiation of the project. Tove Lundeberg (Swedish EPA) and Ann-Sofie Wernersson (Swedish Agency for Marine and Water Management) as always supporting contact persons.
Swedish EPA for funding the national monitoring programs that has delivered the core of the
data that was not financed directly by the project.
5 Introduction
The Baltic Sea has had problems with contaminants for several decades. During and after the 1950s there was a large decrease in seal and eagle populations due to the usage of DDT and PCB. However, since the 1980s the populations have increased due to the ban of these compounds. Unfortunately, other contaminants have been discovered in the ecosystems that potentially might affect the organisms. White-tailed sea eagles (WTSE) still show signs of depressed reproduction regionally, with lower nestling brood size and a higher frequency of dead eggs in the southern Bothnian Sea. There is also a current occurrence of desiccated eggs and total breeding failure among females on the northern coast of the Bothnian Sea. Earlier, a decline in WTSE productivity and population size was strongly related to elevated levels of DDE and PCB (Helander et al. 1982, 2002; Helander 1994b). However, present differences in reproduction between the southern Bothnian Sea and the Baltic Proper cannot be explained by DDE or PCB alone. Between 1990s and 2005 the grey seal population in the Baltic Sea increased with approximately 8% per year. However, between 1987 and 1996 the prevalence of intestinal ulcers increased in 1-3 years old grey seals. In older seals, 4-20 years, the
increase was seen from 1997 to 2007, indicating that the intestinal lesions started in the young seals, and followed that generation. Also, decreases in blubber thickness have been observed during the 2000s. If continued, negative effects on the grey seal population and their
reproduction will certainly be seen. Moreover, it has been shown that concentrations of the brominated flame retardant HBCDD have increased in young grey seals. In addition, this increase of brominated flame retardants occurred at the same time as the increase of intestinal ulcer. Additionally, the fat content in herring has decreased at several sites in the Baltic. Thus, all this might indicate that new compounds have entered the Baltic Sea and are affecting the organisms living there.
Several hypotheses have been suggested for this general decrease in condition and fat content of different organisms. This includes a change in the abundance and species composition in the Baltic ecosystem, but influences from contaminants cannot be excluded, e.g. halogenated phenols are believed to affect the environment. One known effect of exposure of halogenated phenols is the inhibition of ATP in the mitochondria, which leads to loss of energy in terms of forming heat instead of storing fat (Wallace and Starkov 2000).
Means to evaluate potential harmful effects of hazardous substances in the environment have been proposed, representing a threshold below which concentrations are believed not to cause harmful effects on the environment. The assessment criteria that are used have been
developed by the EC, Environmental Quality Standards (EQSs) and the OSPAR organisation Environmental Assessment Criteria (EAC). EQS and EAC are based on controlled,
toxicological lab studies of model organisms in enclosed systems. The results from these experiments are then used as target levels, where it is assumed that no harmful effects will occur on the most sensitive organism. A shortcoming of EQS and EAC are that they do not consider the potential of combination effects of chemical contaminants. Combinations of several compounds, with the same or similar toxicological endpoints (Backhaus et al. 2010) but also dissimilar endpoints (Kortenkamp et al. 2009) have been shown to be more toxic than the compounds one by one. Already available compound specific target levels would
therefore benefit from additional recommendations with the ambition to include combination
effects.
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Few species inhabit the Baltic Sea and hence, the food-chain is rather short.
The figure below (Figure 1) illustrates a typical food-chain in the Baltic Sea, where herring is a typical zooplankton feeder, the grey seal represents a fish eater and the white-tailed sea eagle is on the top of the food-chain, eating fish, seals and fish-eating birds. Trophic position might be important for the accumulation of several contaminants. Most persistent organic pollutants (POPs) are fat soluble and persistent in the environment and, as a result, tend to bioaccumulate in organisms (Newman and Unger 2003). The dominant pathway for uptake of POPs in larger marine organisms is via food ingestion. This pathway together with a slow excretion rate and metabolism, leads to biomagnification, i.e. reach the highest concentrations in top predators (Fisk et al. 2001; Braune et al. 2005). Also, the top predators are more long lived compared to their prey.
Figure 1. A schematic food web in the Baltic Sea. Illustrated by S. Faxneld & B. Helander with images taken from ian.umces.edu/imagelibrary.
7 Aim of study
More than 140.000 chemicals are pre-registered by ECHA (not including e.g. pesticides, pharmaceuticals, PCPs (Personal Care Products), by-products, metabolites etc. Only a small fraction of these are analysed and monitored. The potential to reveal new causal relationships between found biological effects and exposure of chemicals are thus limited. Nevertheless, one of the important aims with effect monitoring is to disclose changes in wildlife health as an early warning system, in order to initiate research to find plausible explanations. Despite these limitations, combining data on contaminant concentrations in herring, grey seal and white- tailed sea eagle from the national marine monitoring programme and data on biological effects on these marine top predators has given us possibility to a) relate biological effects in the marine top predators to certain contaminants or concentrations of contaminants b)
calculate life-time exposures and relate biological effects to these indices c) supply information that can be used in a future revision of the present Quality Standards.
Background Chemicals:
Cyclic siloxanes
Dimethylsiloxanes are the building blocks for many PDMS (polydimethylsiloxane) based silicone polymers. All dimethylsiloxanes are characterised by –Si(CH
3)
2-O- repeating units.
There are three major cyclic methylsiloxanes used in commercial production:
octamethylcyclotetrasiloxane (D4), decamethylcyclopentasiloxane (D5) and
dodecamethylcyclohexasiloxane (D6). D4 contains four repeating units, D5 and D6 contain 5 and 6 repeating units, respectively.
D4, D5 and D6 concentrations in herring and grey seal from the Baltic was reported by Kierkegaard et al. (2013). The concentrations of D5 and D6 were similar in herring caught in the highly populated Baltic Proper and in the less populated Bothnian Sea and Bothnian Bay.
Herring from the North Sea had lower levels of all three chemicals. The concentrations of D4, D5 and D6 in grey seal blubber were lower than the lipid normalized concentrations in
herring, indicating that they do not biomagnify in grey seals (Kierkegaard et al. 2013) Usage
The three major commercially produced cyclic methylsiloxanes are commonly used in cosmetics, conditioners, soaps, lip gloss, deodorants (Stevens 1998; CES 2007; SCCS 2010) and as dry cleaning solvents (CES 2007), often as a blend of all three, and sometimes alone.
D4 and D5 evaporate readily from formulations applied in thin layers, thus the environment is probably their main sink (CES 2007). Personal care products are believed to be one primary source of emissions to the environment, accounting for an estimated 539, 16500 and 1899 tonnes/year in Europe for D4, D5, and D6, respectively. The second major source is emission of residuals from PDMS, accounting for an estimated 944, 787, and 462 tonnes/year in Europe (Brooke et al. 2009a-c).
Biological effects
Very few studies have investigated toxic effects of siloxane exposure. Most studies to date
have been performed on rats. Burns-Naas et al. (1998) found that rats exposed to 160 ppm of
D5 for two weeks showed increased liver weight and liver-to-body-weight ratios. There was
also a slight difference between females and males, where the increase was larger for females.
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This might indicate that females are more sensitive. Moreover, no treatment-related pathology in the livers was observed. Klykken et al. (submitted; cf: Burns-Naas et al. 1998) found similar results in rats after exposure to D4 (i.e. increases in liver weight and liver-to-body- weight ratios).
Conventions, Aims and Restrictions
Thus far, cyclic dimethylsiloxanes have been subjected to intense regulatory scrutiny in Canada and the EU. In Canada, D4 and D5 were identified as harmful to the environment in January 2009, while no concerns were identified for D6. The Canadian government proposed that D4 and D5 be added to the Toxic Substances List, a decision, which means that measures to reduce their release to the environment must be developed within 2 years. This decision for D5 was appealed by the producing industry. In response, the Canadian government appointed a Board of Review to examine the scientific evidence indicating that D5 is harmful to the environment. The report of the Board of Review concluded that D5 is not harmful to the environment, and the Canadian government withdrew their proposal.
Within the EU, the United Kingdom took responsibility for the risk assessment of D4, D5, and D6. In their submission to ECHA, they concluded that D4 was both a PBT (persistent, bioaccumulative and toxic) chemical and a vPvB (very persistent and very bioaccumulative) chemical. If accepted, this would lead to D4 being classified as a SVHC (Substance of Very High Concern) under REACH. For D5 the conclusion was that it was a vPvB chemical, but they were unsure whether the evidence was sufficient to justify extra risk management. The vPvB designation is an exposure hazard based classification designed to prevent substances being released to the environment that – due to their persistence and bioaccumulation – would be very difficult to manage, should harmful effects be identified in the future. No concerns were identified for D6. ECHA has not yet made a decision based on these submissions prepared by the United Kingdom.
Target levels
The regulatory concern in Europe is focused on exposure hazard, not risk. Therefore there are no target levels.
Halogenated phenols
Phenols are a class of aromatic organic compounds that consist of one or more hydroxyl groups (–OH) bonded directly to an aromatic hydrocarbon group (Barlow and Johnson 2007), while halogenation is a chemical reaction that sees a halogen atom incorporated into a
molecule as a substitution of a hydrogen atom. There are four types of halogenation – chlorination, fluorination, bromination and iodination (Gribble 2003).
Usage
There are many different types of halogenated organic compounds found in the environment.
One well-known example is bisphenol-A, commonly found in polycarbonate containers, and 2,5-dichlorphenol, otherwise known as mothballs (Barlow and Johnson 2007). The
chlorophenols are believed to be the result of human activity (Wannstedt et al. 1990). Some halogenated phenols have been used as external and internal disinfectants (Zondek and
Shapiro 1943) e.g., trichlorophenol, better known as TCP, is a mild antiseptic used for treating sore throats, mouth ulcers, cuts, bites, pimples, boils, and as a general disinfectant.
Halogenated phenols can also be used as pesticides and solvents (Häggblom and Young
1995), while some chlorinated phenols e.g. tetrachlorophenols and pentachlorophenols, are
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used as fungicides (The Dow Chemical Company 1985). Chlorine is widely used to disinfect drinking water, and from this, halogenated by-products can be formed (Cancho et al. 1999).
Toxic chlorophenols are generated by the bleaching process used in pulp and paper mills (Wannstedt et al. 1990). Sources of halogenated phenols include direct discharge, metabolism and/or environmental transformation from other pesticides, chlorination of water and natural production by bacteria, plants and animals (Wannstedt et al. 1990). Hydroxylated
polybrominated diphenyl ethers (OH-PBDEs) are another group of halogenated organic compounds and they are known to be metabolites of PBDEs (Örn and Klasson-Wehler 1998;
Marsh et al. 2006). These can both occur naturally (Gribble 2003; Malmvärn 2007) in the environment but can also originate from anthropogenic compounds.
Biological effects
Very few studies have investigated biological effects of OH-PBDEs. In a study by Meerts et al. (2001) some OH-PBDEs were shown to induce the estrogen receptor signal transduction pathway in vitro. Thus, OH-PBDEs have estrogenic potencies, and Meerts et al. (2001) suggests that they are in the same range as for Bisphenol A.
Conventions, Aims and Restrictions
Pentachlorphenol (PCP) is under review for being included in The Stockholm Convention on POPs.
Target levels
There are no target levels for halogenated phenols.
Brominated Flame Retardants
Polybrominated diphenyl ethers (PBDEs) and hexabromocyclododecane (HBCDD) are both groups of brominated flame retardants (BFRs) that have an inhibitory effect on combustible materials. There are a possible 209 PBDE congeners, and a possible 16 HBCDD
stereoisomers. PBDEs are produced as three different technical products: Penta-, Octa- and DecaBDE. Each of these products includes a few major congeners. For PentaBDE these are BDE-47, -99, -100, 153 and -154. OctaBDE contains mainly BDE-183, while DecaBDE includes almost exclusively BDE-209 (LaGuardia et al. 2006). HBCDD is produced as a mixture of three stereoisomers: α-, β- and γ-HBCDD (Covaci et al. 2006).
Usage
Both PBDEs and HBCDD are added as flame retardants into building materials, thermal insulation, furnishings, motor vehicles, plastics, textiles and other products that should not catch fire. Leakage of these substances to the environment occurs from their production and use in consumer products, and long-range transport via both gas phase and air borne particles.
The PBDE congeners that are most commonly found in fish are BDE-47, -99 and -100, while PBDE congeners with a higher degree of bromination are more common in soil and
organisms in the terrestrial environment.
Biological effects
Several PBDE congeners and HBCDD have been shown to cause neurotoxic effects in rats and mice. Animals exposed to PBDEs and HBCDD during a sensitive stage of brain
development have later shown impaired motor behaviour (Eriksson et al. 2002; Viberg et al.
2003a) and reduced memory and learning abilities (Viberg et al. 2003b). BFRs are also
considered to be endocrine disruptors, and in particular, effects on the thyroid hormone
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system have been seen (Darnerud 2008). In addition, chronic exposure to HBCDD can lead to increased liver weights and thyroid hyperplasia (Darnerud 2003; Stapleton 2006). A field study on ospreys (Pandion haliaetus ) showed depressed reproduction at high concentrations of PBDEs (Henny et al. 2009).
Conventions, Aims and Restrictions
The Penta- and OctaBDE products were banned within the EU in 2004 (Cox and Efthymiou 2003) and both were added in 2009 to the UNEP Stockholm Convention on Persistent Organic Pollutants for global bans (UNEP 2009). A Swedish ban of DecaBDE was
established in 2007, but this ban was withdrawn when DecaBDE was included in the RoHS directive in 2008. Recently, major producers in the US have announced that they will discontinue DecaBDE production and use by 2013 (Albemarle 2009; Chemtura 2009; Hess 2009). PBDEs are also on the list of prioritized substances within the Water Framework Directive.
HBCDD is under review by the Persistent Organic Pollutants Review Committee (POPRC) as a proposed substance to be listed under the Stockholm Convention (Arnot et al. 2009).
Target levels
For PBDEs, there is an EQS
biotafor the sum of BDE congeners, and it is set at 0.0085 ug/kg wet weight
The EQS
biotafor HBCDD is based on secondary poisoning of predators and set at 167 ug/kg fresh weight.
For further information about target levels for PBDEs see Bignert et al. (2014).
DDTs (Dichlorodiphenylethanes)
Usage
DDT is a persistent synthetic pesticide and it was first synthesised in the early 1870s,
however, its insecticidal properties were not discovered until 1939 (Metcalf 1973). DDT was introduced in the beginning of the 1940s and later on used worldwide as an insecticide, until it was recognised that it bioaccumulates and had detrimental effects. Currently, DDT is still used but under strict conditions for the control of malarial vectors in some areas (WHO 2007).
DDE and DDD are formed by the degradation of DDT.
Biological effects
DDE, which is a degradation product of DDT, is very persistent and has a high fat solubility.
This leads to bioaccumulation in the food web. One of the most known effects of DDT is eggshell thinning in predatory bird and some other birds´ eggs (Ratcliffe 1967; Blus et al.
1971, 1974; Peakall and Lincer 1996; Lundholm 1997), even at low doses. The white-tailed sea eagle became critically endangered in the Baltic range during the 1960s and 1970s largely due to DDT and the reproductive impairments caused by its breakdown product DDE
(Helander et al. 2002).
Conventions, Aims and Restrictions
DDT is listed among the initial 12 Persistent Organic Pollutants (POPs) included in The
Stockholm Convention on POPs, an international agreement requiring measures for reducing
or preventing the release of dangerous substances into the environment.
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In Sweden, DDT was partially banned as a pesticide in 1970, and completely banned in 1975 due to its persistence and environmental impact.
Target levels
There are no EQS or EC foodstuff regulation developed for any of the DDTs. The current, recommended EAC for DDE is set to 5 µg/kg wet weight.
PCB
Polychlorinated biphenyls (PCBs) are constructed of two benzene rings, with between one to ten chlorine atoms attached, making 209 PCB congeners possible. Twenty of these congeners have non-ortho chlorine substitutions, and so can attain a planar structure similar to the highly toxic polychlorinated dibenzo-p-dioxins and dibenzofurans (McKinney et al. 1985).
Usage
PCBs have been used in a wide variety of manufacturing processes, especially as plasticizers in paints and cements, insulators, and as fire retardant fabric treatments. They are widely distributed in the environment due to inappropriate handling of waste material e.g., leakage from large condensers, hydraulic systems, and buildings. They were also found in e.g.
lubricating oils, sealants, and adhesives.
Biological effects
PCBs degrade very slowly and they are fat and oil soluble, which leads to bioaccumulation in biota to high concentrations (Newman and Unger, 2003). The toxicological effect of PCBs on reproduction in mink, for example, is well documented (Aulerich et al. 1977; Jensen et al.
1977).
Conventions, Aims and Restrictions
PCBs are one of the original 12 Persistent Organic Pollutants (POPs) included in The
Stockholm Convention on POPs, an international agreement requiring measures for reducing or preventing the release of dangerous substances into the environment.
In 1973, the use of PCBs was banned in Sweden, except for in enclosed systems. In 1978, all new use of PCBs was forbidden.
The Minister Declaration from 1996, within HELCOM, and the declaration in Esbjerg 1995, called for measures for the use of toxic, persistent and bioaccumulating substances to have ceased completely by 2020.
Target levels
Several target levels have been developed for PCBs, including Environmental Quality Standards (EQSs), Environmental Assessment Criteria (EACs), and recommendations for foodstuffs. Currently, recommending a target level for environmental status assessments for PCBs is problematic. In this report we use the EAC: CB-118=24 ug/kg lipid weight, CB- 153=1600 ug/kg lipid weight. See Bignert et al. (2014) for more information.
Dioxins
“Dioxins” refer to polychlorinated dibenzo-p-dioxin (PCDD) and dibenzofuran (PCDF)
compounds. Seventeen (10 furans, 7 dioxins) of the 210 possible congeners, substituted in the
positions 2,3,7,8, are considered to be of toxicological importance. Some polychlorinated
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biphenyls (PCBs) are called dioxin-like PCBs (dl-PCBs) because they have a structure similar to that of dioxins and have dioxin-like effects. They are, however, not included in this report.
PCDD/Fs are characterized by low water solubility and low vapor pressure. In the
environment, they can undergo photolysis, but they are generally very resistant to chemical and biological degradation. Due to their persistent and hydrophobic properties, PCDD/Fs accumulate in sediments and organisms in the aquatic environment.
Usage
PCDD/Fs are not produced intentionally. They are formed as by-products in several industrial processes and from most combustion processes, such as municipal waste incineration and small-scale burning under poorly controlled conditions. They are also minor impurities in several chlorinated chemical products (e.g. PCBs, chlorophenols, hexachlorophene etc.).
Formerly, pulp bleaching using chlorine gas was an important source of PCDD/Fs.
Biological effects
PCDD/Fs can cause a variety of biological and toxicological effects in animals and humans.
The most relevant toxic effects are developmental toxicity, carcinogenity and
immunotoxicity. Most toxic effects are explained by the binding of PCDD/Fs to the aryl hydrocarbon (Ah) receptor. The sensitivity of various species to the toxic effects of PCDD/Fs varies significantly. 2,3,7,8-TCDD is the most toxic and well-studied congener and is used as a reference for all other related chemicals.
Each of the 17 relevant congeners is assigned a toxic equivalency factor (TEF), where 2,3,7,8- TCDD equals 1 (Van den Berg et al. 1998; Van den Berg et al. 2006). Dioxin concentrations are here reported as TCDD-equivalents (TEQ), which is the sum of the individual congener concentrations multiplied with its specific TEF.
Conventions, Aims and Restrictions
Dioxins are included in several international agreements, of which the Stockholm Convention and the Convention on Long Range Transboundary Air are among the most important for the control and reduction of sources to the environment. Several EU legislations regulate dioxins, e.g. the plan for integrated pollution prevention and control (IPPC) and directives on waste incineration (EC 2000, 2008). The EU has also adopted a Community Strategy for dioxins, furans and PCBs (EC 2001). PCDD/Fs are currently not included in the Water Framework Directive but are on the list of substances to be revised for adoption in the near future.
HELCOM has listed PCDD/Fs and dl-PCBs as prioritized hazardous substances of specific concern for the Baltic Sea (HELCOM 2010), like OSPAR on the List of Chemicals for Priority Action (OSPAR 2010).
WHO and FAO have jointly established a maximum tolerable human intake level of dioxins via food, and within the EU there are maximum allowable levels of dioxins in food and feed stuff (EC 2006). The EQS
biotafor concentrations of dioxins is based on human health and set at Σ PCDD/Fs+dioxin like PCBs: 0.0065 ug WHO05-TEQ/kg ww. PCDD/F levels in fat fish, mainly herring and salmon, from the Baltic Sea often exceed this limit. Sweden and Finland have since 2002 been authorised a derogation from this directive, allowing to sell on the domestic market or to non-member states (EC 2375/2001, EC 201/2002, EC 199/2006, EC 1881/2006).
Target levels
The EQS
biotafor concentrations of dioxins and dioxin-like PCBs (DL- PCBs) is based on
human health and set at Σ PCDD/Fs: 0.0035µgWHO05-TEQ /kg wet weight. For more
information regarding target levels for dioxins, see Bignert et al. (2014).
13 PFASs
Perfluoroalkyl substances (PFASs) are organofluorine compounds, where all hydrogens have been replaced by fluorine on a carbon chain. There are a number of different PFASs, but the two most well-known are perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS).
Usage
Perfluoroalkyl chemicals have been used industrially and commercially since the early 1950s.
They are found in a wide range of products e.g. grease proof packaging such as food boxes (microwave food), fire-fighting foams, outdoor clothing, Teflon and many cleaning and personal care products. PFOA is used in the production process of fluoropolymers such as PTFE. PFOS was used in Scotchguard products, which was sprayed on furniture and carpets to give a grease/stain proof layer. In 2000, the main producer of these products, 3M, started phasing out production of PFOS and PFOS-related chemicals (Buck et al. 2011).
Perfluoroalkyl acids (PFAAs) are a subgroup of PFASs containing a perfluorinated carbon chain and a hydrophilic acid group. PFAAs are strong surfactants with an extraordinary surface tension lowering potential. In the environment they can have two sources - direct from manufacturing and use, and indirect from degradation of volatile precursor compounds (Buck et al. 2011). PFOA and perfluorononanoic acid (PFNA) are intentionally produced
perfluoroalkyl carboxylic acids (PFCAs) and therefore a large portion of the PFOA and PFNA found in the environment probably originates from direct sources (mainly the production process of fluoropolymers, Prevedouros et al. 2006), waterborne and atmospheric transport to remote locations. Therefore, sewage treatment plant effluent from industry or larger cities could represent hot-spots. In contrast, longer-chain PFCAs than PFNA, e.g. PFUnDA (perfluoroundecanoic acid) and PFTrDA (perfluorotridecanoic acid) are unintentionally produced substances, and their presence in the environment is probably due to both direct sources (impurities in PFOA and PFNA production) and indirect sources (atmospheric transport and degradation of precursors, Buck et al. 2011). Also the role of PFOS derivatives for the distribution and accumulation of PFOS in the environment is currently under
investigation (Martin et al. 2010). Perfluorooctane sulfonamide (FOSA) is an intermediate product in the degradation of many PFOS precursors to PFOS, and is often analysed together with PFOS in the environment.
Biological effects
PFOA exposure to rat, mouse, rabbits and monkeys has shown to lead to reduced body weight
and increased liver weight. (Kennedy et al. 2004). The common carp (Cyprinus carpio)
exposed to different PFOS concentrations experienced decreases in glycogen, and declines in
condition factor and hepatosomatic index with increases in PFOS concentrations (Hagenaars
et al. 2008). In addition to animal studies, epidemiological studies on humans have during
recent years increased. For instance, concentrations of PFOA in maternal blood and PFOA
and PFOS in cord blood during pregnancy have been found to be negatively associated with
birth weight (Apelberg et al. 2007; Fei et al. 2007), ponderal index, head circumference
(Apelberg et al. 2007), and birth length (Fei et al. 2008a). In contrast, no associations between
concentrations of PFOA and PFOS in maternal plasma during pregnancy and developmental
milestones in early childhood have been found (Fei et al. 2008b).
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Conventions, Aims and Restrictions
PFOS, its salts, and perfluorooctane sulfonyl fluoride are among the nine new Persistent Organic Pollutants (POPs) included in Annex B in The Stockholm Convention on POPs, but it is listed with exemptions (UNEP 2009). Additionally, the use of PFOS and its derivatives is restricted in the EU by the Marketing and Use Directive 2006/122/EC.
Target levels
The EQS
biotafor PFOS concentration is based on human health and set at 9.1 ug/kg wet weight. There are no EAC or EC foodstuff regulation developed for PFOS.
Combination effects
Few studies have investigated combination effects of contaminants. However, it has been suggested that compounds with the same mode of action will most likely act additive
(Broderius et al. 1995; Broderius et al. 2005; Backhaus et al. 2010), but there is also evidence that compounds with dissimilar mode of action can act additive (Kortenkamp et al. 2009). In addition, endocrine disrupting chemicals (i.e. polychlorinated biphenyls and certain
pesticides) have potential to act synergistically when released together into the environment (Arnold et al. 1996). In a study by Evans and Nipper (2007) exposure to marine invertebrates to the combination of phenanthrene (polycyclic aromatic hydrocarbon) and lindane
(organochlorine pesticide) lead to an additive response for sea urchin fertilization while there was a synergistic effect for copepod reproduction. Additive or synergistic effects were also seen when aquatic midges were exposed to several combinations of organophosphate (OP) insecticides (Lydy and Austin 2004). Furthermore, Lydy and Austin (2004) did not observe any significant impact on midges after pre-exposure to DDE followed by exposures to different OPs.
Study species:
Herring (Clupea harengus)
Background/ biology
Herring is a pelagic species that feeds mainly on zooplankton. Herring becomes sexually mature at about 2 - 4 years of age in the Baltic and on the Swedish west coast. Herring can reach up to 25 years of age but a more common age is 10 years. West coast herring with a body length of 23-30 centimetres is a bit larger than herring populations in the Baltic at 15-24 cm. Normal weight varies between 40-200 grams on the west coast and a bit less for herring in the Baltic. Herring are important not only for human consumption but also for consumption for several other predators in the marine environment. In the Baltic Sea, herring is the main prey species for grey seals.
Past and current health status/population status
Generally, the condition i.e. weight versus length, of the Baltic herring has decreased since the mid-1980s (Cardinale and Arrhenius 2000; Bignert et al. 2014). Both condition and fat content are higher at the Swedish west coast compared to the Baltic (Bignert et al. 2014). The fat content (in muscle tissue) in autumn caught herring from Utlängan in the southern Baltic Proper has been less than 2% from 2004 until 2011 (Bignert et al. 2014), which is
exceptionally low.
15
Monitoring, why this species is useful within monitoring and status of monitoring network (geographical and temporal coverage).
Herring muscle tissue is fat and thus very appropriate for analysis of fat-soluble contaminants i.e. hydrocarbons. Herring is the most commonly used indicator species for monitoring contaminants in biota within the BMP (Baltic Monitoring Programme) in the HELCOM convention area, and is sampled by Finland, Estonia, Poland, Germany, Denmark, and Sweden.
Guillemot (Uria aalge)
Background/ biology
The common guillemot is abundant and widespread pelagic seabird found in the Baltic Sea.
They feed mainly on sprat (Sprattus sprattus) and herring (Clupea harengus) (Bignert et al.
1998). Guillemots breed for the first time at 4 - 5 years of age. Eggs hatch after about 32 days.
Normally the guillemot lay just a single egg but if this egg is lost, another may be laid. It has been shown that guillemot eggs that are laid late tend to contain significantly higher
concentrations of organochlorines compared to eggs laid early (Bignert et al. 1995).
Past and current health status/population status
The guillemot population in the Baltic was almost extinct at the end of the nineteenth century.
Stora Karlsö (Southern Baltic Sea) was the only known breeding locality, and in 1880 only 20 individuals were seen there (Hedgren 1975). In 1880s the birds were protected at Stora
Karlsö, and since then the population has increased to about 8000-10 000 breeding pairs (Österblom et al. 2002). There are today also other colonies in the Baltic, but Stora Karlsö is the largest.
Monitoring, why this species is useful within monitoring and status of monitoring network (geographical and temporal coverage).
Since 1912 guillemot chicks have been ringed on Stora Karlsö, and the population can therefore be followed. Guillemots are suitable for monitoring contaminants in the Baltic Sea as most do not migrate further than the southern parts of the Baltic Proper during the winter season (Österblom et al. 2002). The egg content is high in fat (11 - 13%), thus very
appropriate for analysis of fat-soluble contaminants i.e. hydrocarbons. Since the 1960s the species has been used in marine monitoring as an indicator of organic pollutants in the Baltic Sea (Bignert et al. 1998).
Grey seals (Halichoerus grypus)
Background/Life history
The grey seal is the largest seal species in the Baltic. The female can reach an age of more than 40 years. Cows usually produce their first pup when five or six years old. Bulls are sexually mature at about six years of age but most of the breeding bulls are between 12 and 18 years old. The period of birth occurs during the last three weeks of February to the first week of April with a peak in the first two weeks of March. Lactation lasts for 16-21 days and the mating takes place at the end of the lactation. After mating, the fertilized egg develops into a blastocyst within ten days and then remains dormant in the uterine horn for about 3.5 month.
The implantation of the blastocyst occurs in July/August. Females usually do not forage
during the reproductive period and the lactation. The pup is dependent on the blubber layer
that the females gain during the autumn and winter (King 1983). During lactation females
loose about 40 % of their body weight and most of the weight loss is blubber (Fedak and
16
Andersson 1982). The daily decrease in body weight during the reproductive period is less in adult grey seals males compared to females but the females gain weight again faster than the males (Andersen and Fedak 1985; Kastelein et al.1990). Moulting takes place in May-June and most of the seals then spend long periods hauled out on land or ice and that period is therefore the best time for population estimates (Helander 1998; Sjöberg 1999). The Baltic grey seals mostly feed on herring (Clupea harengus) which importance has increased since the 1970s. In an earlier study by Söderberg (1972), cod (Gadus morhua) was a larger part of the food compared to in a later study by Lundström et al. (2007).
Past and current health status/population status
During the 20
thcentury the number of Baltic seals was severely reduced, and in the 1970s there were only a few thousand Baltic grey seals left of the population (Almkvist et al. 1980;
Hårding and Härkönen 1999). The population of grey seals was estimated to around 100 000 in the beginning of the 1900s (Hårding and Härkönen 1999). In the late 1970s necropsies of by-caught and stranded Baltic seals, including analysis of persistent organic contaminants were initiated in Sweden. A high prevalence of uterine occlusions and stenosis was found in Baltic ringed and grey seals (Helle et al. 1976; Olsson 1977) as well as high levels of PCB and DDT in the Baltic biota including seals and other apex predators (Jensen et al. 1969). At this time, the necropsies of Baltic grey seals also revealed certain chronic organ lesions such as claw fold inflammation, loss of bone matrix, intestinal (mostly colonic) ulcers,
glomerulopathy and tubular cell proliferations in kidneys, cortical hyperplasia in adrenals, arteriosclerosis and besides the uterine occlusions and stenosis, uterine leiomyomas (Bergman and Olsson 1985, 1989; Bergman et al.1992; Bergman 1999; Bergman et al. 2001; Bäcklin et al. 2003; Bredhult et al. 2008).The ban of seal hunt in 1974, decreasing levels of DDT and PCB in the Baltic biota reported by Olsson and Reutergård (1986) and Bignert et al. (1995) and an improvement of the gynecological health conditions of grey seals resulted in an increasing Baltic grey seal population (Bergman 1999). The number of grey seals has been increasing with about 8 % per year between the early 1990s and the mid- 2000s. In 1987-1996 most of the earlier observed pathological changes decreased in prevalence in younger seals except intestinal ulcers. From the middle of 1980s the prevalence of intestinal ulcers, mostly localized in the ileum-caecum-colon region, in 1- to 3-year-old grey seals increased
significantly compared to the decade before. Thereafter (1997-2009), the prevalence of intestinal ulcers decreased. Several years after the increase in young seals, there was a significant increase of ulcers in 4-20 years old grey seals. This indicates that the ulcers observed in adult seals started to develop earlier in life. The high prevalence of intestinal ulcers seems unique for the Baltic population of grey seals. Examination of grey seal intestines from the Scottish east coast and Atlantic coast of Ireland, revealed no signs of ulcers (Bergman 1999; O´Neill and Whelan 2002). Only one case of intestinal ulcer in grey seals has been reported outside the Baltic Sea (Baker 1980; 1987). The high prevalence of ulcers of moderate to severe degree in the young Baltic grey seal indicates an impaired or delayed healing process, which may involve the immune- as well as the hormonal system.
The mean autumn/winter blubber thickness has decreased significantly in Baltic grey seals
since the beginning of 2000s, especially in 1-4 year-old seals from by-catch and hunt (Bäcklin
et al., 2011). There could be several reasons for a thinner blubber layer in the autumn/winter
season e.g. disease, contaminants, decreased fish stocks and change in diet, or a change in the
quality of the diet. The reason for the decreasing trend in blubber thickness in seals is
unknown but so far no correlations to disease have been found.
17
Monitoring, why this species is useful within monitoring and status of monitoring network (geographical and temporal coverage).
Being a top predator of the Baltic Sea, feeding on fish, the Baltic grey seal serves besides the white-tailed sea eagle and the guillemot, as a good indicator for the Baltic Sea ecosystem and for environmental contaminants in fish. Grey seals inhabit the whole eastern coastline of Sweden but also show high site fidelity, making them important as geographical markers (Karlsson 2003). Monitoring of Baltic grey seal health and population levels in Sweden has been included in the National Environment Monitoring Programme since 1989, although data and samples for the Environmental Specimen Bank have been collected since the 1970s. The increase in 1970s and 1980s and decrease in the 1990s in prevalence of female reproductive lesions followed the PCB trend and was a good indicator of the severity of that contaminant (Roos et al. 2012). The increase in prevalence of intestinal ulcers and lowered blubber thickness show that there are still environmental changes that need to be investigated more deeply.
White-tailed Sea Eagle (Haliaeetus albicilla)
Background/Life history
White-tailed sea eagles (WTSE) are large birds of prey, feeding mainly on fish and seabirds.
Females are usually larger (4.5 - 7 kg) than males (4 - 5.5 kg). The birds become sexually mature at 4-5 years of age and usually start their reproductive life from about 5 years old.
Adults have a mean lifespan of about 18 years (Helander 2003a), with a few exceeding 30 years. Mating pairs are generally faithful over their life-time, and remain at the same breeding site, with sites commonly used over many generations of eagles. An adult pair typically produces a single clutch of eggs annually (Helander 1985). Clutch size is 1-3 eggs with a mean weight of about 120 g per egg. The lipid content in 97 undeveloped sea eagle eggs from the Swedish Baltic coast averaged 5.08 % (Helander et al. 2002). The incubation period is 35 - 40 days and the nestling period is 70 - 85 days (usually longer for the bigger females than for the smaller males). The breeding season usually starts about one month prior to egg-laying and extends at least a month after the young fledge from the nest, thus comprising fully six months of the year.
In Sweden, juveniles move from their natal area during October to the southern Baltic Sea states. Adults in southern and central Sweden are mainly sedentary, whereas adults breeding at northern latitudes (Bothnian Bay, Lapland) are mainly migratory as a result of ice cover and the scarcity of food. Resightings of ringed breeding birds have shown a strong homing tendency in sea eagles in Sweden, with a mean distance of 90 km from birth place to settlement site for 35 males and 114 km for 37 females (Helander 2003a) and a very strong nest site fidelity among established breeders.
The WTSE is a top predator feeding on fish as well as on other fish- predators. Thus, the
eagle is at the very top of the food web. The eagles catch fish near the surface, and common
fish prey on the Baltic coast are pike Esox lucius, bream Abramis brama, ide Leuciscus idus,
roach Rutilus rutilus and perch Perca fluviatilis. These species are common among fish prey
also in fresh water systems in southern and central Sweden. In Lapland, where cyprinids are
not available, pike (dominating), burbot Lota lota, grayling Thymallus thymallus and perch
are taken. Bird prey taken on the coast includes eiders Somateria mollissima, mergansers
Mergus spp., cormorants Phalacrocorax spp., grebes Podiceps spp. and gulls Larus spp. Birdprey in Lapland includes mergansers, divers Gavia spp. and gulls. Mammalian prey is largely
retrieved as carrion, including seals Pinnipedia on the coast. Fish and birds make up most of
18
the food of sea eagles on the coast, whereas mammal prey (largely carcass remains of reindeer
Rangifer tarandus) make up some 10 % of prey items in nests in Lapland (Helander 1983).Past and current health status/population status
Over nearly 40 years, the reproduction of sea eagles around the Baltic Sea was seriously depressed, including related parameters: eggshell thinning, excessive rate of dehydration of eggs, low fertilization of eggs, and embryo mortality (Helander et al. 1982, 2002, 2008). The net effect of these anomalies was a strong (80 %) reduction in the number of young produced per pair and year, and a decreasing population. In the 1970s the WTSE was classified as critically endangered in the Baltic range, based on the negative population trends and very poor reproduction. The remnant populations were eventually saved through extensive conservation efforts, including feeding programs, protection of nest sites and education. An improvement in reproduction was first noticed in the mid-1980s, and since the mid-1990s the reproductive rate is almost back to normal within the Baltic range. In Sweden, the species is currently classified as near threatened (Gärdenfors 2010).
Monitoring, why this species is useful within monitoring and status of monitoring network (geographical and temporal coverage).
The white-tailed sea eagle was the first species that indicated there were deleterious effects from environmental pollutants in the Baltic Sea. Strong relationships have been indicated between reproductive impairments and high levels of DDTs and PCB in the eggs (Helander et al. 1982, 2002, Helander 1994b). If WTSE reproduction had been monitored in the Baltic Sea earlier during the 20th century, the negative impact of DDT may have been noticed as early as the 1950s. An improvement in Baltic WTSE reproduction was first noticed not earlier than 10 years after most countries around the Baltic had implemented bans of DDT and PCB. This serves as a clear reminder of the long-term effects of persistent pollutants. The subsequent recovery, from an 80% reduction in reproductive ability in the 1970s, is nevertheless important evidence of successful results from wise political decisions.
Feeding on fish, sea birds, and seals, the WTSE is strongly exposed to persistent chemicals that magnify in the food web. Territorial adults on the Baltic Sea coast are mainly sedentary and thus reflect the regional contaminant situation, providing good opportunities for long- term studies. A large portion of breeders in the Baltic region are currently ringed, improving possibilities for study of reproduction and contamination of individual birds over time.
Currently, eagles are breeding all along the coasts of the Baltic Sea, as well as in inland
freshwater systems, and are monitored in a network of national projects that use the same
methodology (Helander et al. 2011; Herrmann et al. 2011). Monitoring of sea eagle
reproduction in Sweden has been included in the National Environment Monitoring
Programme since 1989, as an indicator of effects from chemical pollutants. Pre-1954
background data on breeding success and pre-1950 background data on nestling brood size
are available from the Swedish Baltic coastline (Helander 1994a, 2003b). These data are used
as reference levels for evaluation of observations within the programme.
19 Materials & Methodology
Chemicals
Organochlorines and brominated flame retardants
The analyses of organochlorines and brominated flame retardants are carried out at the Institute of Applied Environmental Science (ITM) at Stockholm University. Before 1988, organochlorines were analysed by a packed column gas chromatography (GC). During 1988, analysis on a capillary column was introduced, allowing analysis of individual congeners (Eriksson et al. 1994). The extraction method originates from the method described by Jensen et al. (1983) where wet tissues are extracted with a mixture of polar and non-polar solvents.
The organochlorines are analysed on a gas chromatograph (GC) equipped with a μ-electron capture detector (Eriksson et al. 1994). The BFRs are analysed by a GC connected to a mass spectrometer operating in electron capture negative ionization mode (NICI) (Sellström et al.
1998).
Quality assurance
Quality control for organochlorines has continuously improved over the last 20 years, resulting in accreditation in 1999. Assessment is performed once a year by the accreditation body SWEDAC. The laboratory is fulfilling the obligations in SS-EN ICO/IEC 17025:2005.
The accreditation is valid for CB28, 52, 101, 118, 153, 138, 180, HCB, p,p'-DDE, p,p'-DDD,
p,p'-DDT and α, ß- and γ-HCH in biological tissues. So far the BFRs are not accredited butthe analysis of BDE-47, 99,100, 153, 154 and HBCDD are in many ways performed with the same quality aspects as the organochlorines.
Standards
The origin of all standards are well documented with known purity and certified concentration with uncertainty for the solutions.
Selectivity
To have the possibility to control impurities in solvents, equipments and glassware, one blank sample (a vegetable oil) is extracted together with each batch of environmental samples.
Coelution of PCB congeners and pesticides in GC analysis is dependent upon instrumental conditions such as column type, length, internal diameter, film thickness and oven
temperature. To minimize possible coelutions, two 60 m columns are used in parallel, the commonly used 5 % phenyl-methylsilicone phase and the more polar 14 %
cyanopropylphenyl-methylsilicone phase. The only remaining known coelution is for CB-138, which coelutes with CB-163 (Larsen et al. 1990). Therefore CB-138 is reported as
CB138+163. PBDE and HBCDD are analysed on a 30 m DB-5 MS column, monitoring m/z 79 and 81.
Reference Material
Two laboratory reference materials (LRM) are used as extraction controls, chosen with
respect to their lipid content and level of contaminants. The controls consist of herring
respectively salmon muscle, homogenised in a household mixer and stored in aliquots in
airtight bags of aluminium laminate at -80°C. At every extraction event one extraction control
is extracted as well.
20
Proficiency testing
Concerning PCBs and pesticides, the laboratory has participated in the periodic QUASIMEME proficiency testing since 1993, with two rounds every year, each one
containing two samples. Around 95% of all reported values have been satisfactory according to QUASIMEME, meaning they have been within +/- 2 standard deviations of the assigned value. In 2000, the laboratory participated in the first interlaboratory study ever performed for PBDEs and HBCDD, contaminants that since 2001 are incorporated in the QUASIMEME proficiency testing scheme. Around 80% of the values the laboratory has produced during the years have been satisfactory according to QUASIMEME.
Quantification limits and uncertainty in the measurements
Calculation of the uncertainty in the measurement is based on the Nordtest Report TR 537
“Handbook for calculation of measurement uncertainty in environmental laboratories”, where the within-laboratory reproducibility is combined with estimate of the method and laboratory bias. The within-laboratory reproducibility is calculated from LRM from more than 8000 PCB and pesticide values during a period of nearly 20 years and around 2000 BDE and HBCDD values during nearly 15 years. The bias is estimated from proficiency testing of more than 8 samples during at least 4 years. The bias for PBDE is used also for HBCDD since no reliable proficiency testing (or certified reference material) exists today. Finally, the expanded
uncertainty is calculated, using a coverage factor of 2 to reach approximately 95% confidence level (Table 5). The reproducibility for the PCBs and pesticides follows the theory stated by Horwitz where the relative standard deviation increase when the concentration levels decrease (Horwitz and Albert 2006). The reproducibility for the PBDEs and HBCDD follows a
function where the relative standard deviations increase first at the very lowest concentration.
Table 5. Expanded uncertainty (%) at different concentrations CB28,101,118,
153,138,180,HCB
CB52 HCH α, ß, γ
ppDDE ppDDD
ppDDT PBDEs HBCDD
ng/g lw % % % % ng/g lw % %
2-50 36 49 40 43 0.2-1 73
4-50 52 > 2 58
> 50 29 30 34 31 38 2-25 103
> 25 64
The quantification limit is estimated to approximately 2 ng/g fat weight for all analysed PCBs, α, ß, γ-HCH, HCB, pp-DDE and pp-DDD and 4 ng/g fat weight for pp-DDT. For all analysed PBDEs the quantification limit is estimated to approximately 0.2 ng/g fat weight and for HBCDD 2 ng/g.
Perfluoroalkyl substances
The analyses of perfluoroalkyl substances are carried out at the Department of Applied Environmental Science (ITM), Stockholm University.
Sample preparation and instrumental analysis
A sample aliquot of approximately 1.0 g (0.5 g for bird eggs) homogenized tissue in a
polypropylene (PP)-centrifuge tube was spiked with 1.0 ng (10 ng for bird eggs) each of a
suite of mass-labelled internal standards (
18O- or
13C-labelled perfluoroalkane sulfonates and
perfluoroalkyl carboxylic acids). The samples were extracted twice with 5 mL of acetonitrile
in an ultrasonic bath. Following centrifugation, the supernatant extract was removed and the
21
combined acetonitrile phases were concentrated to 1 mL under a stream of nitrogen. The concentrated extract underwent dispersive clean-up on graphitised carbon and acetic acid. A volume of 0.5 mL of the cleaned-up extract was added to 0.5 mL of aqueous ammonium acetate. Precipitation occurred and the extract was centrifuged before the clear supernatant was transferred to an autoinjector vial for instrumental analysis and the volume standards M8PFOA and M8PFOS were added.
Aliquots of the final extracts were injected automatically on an ultra performance liquid chromatography (UPLC) system (Acquity, Waters) coupled to a tandem mass spectrometer (MS-MS; Xevo TQS, Waters). Compound separation was achieved on a BEH C18 UPLC column (1.7 µm particles, 50 × 2.1 mm, Waters) with a binary gradient of ammonium acetate buffered methanol and water. The mass spectrometer was operated in negative electrospray ionisation mode. Quantification was performed in selected reaction monitoring
chromatograms using the internal standard method.
Quality control
The extraction method employed in the present study (with the exception of the concentration step) has previously been validated for biological matrices and showed excellent analyte recoveries ranging between 90 and 110% for PFCAs from C6 to C14 (Powley and Buck 2005). Including extract concentrations, we determined recoveries between 70 and 90% for C6- to C10-PFCAs and 65 70% for C11-C15 PFCAs. Extraction efficiencies for
perfluoroalkane sulfonates (PFSAs), including perfluorooctane sulfonamide (FOSA), were determined to 70 95%. Method quantification limits (MQLs) for all analytes were
determined on the basis of blank extraction experiments and ranged between 0.01 and 0.3 ng/g wet weight for the different compounds. A fish tissue sample used in an international inter-laboratory comparison (ILC) study in 2007 (van Leeuwen et al. 2009) was analysed as control sample along with all sample batches. The obtained concentrations were in good agreement with the mean concentrations from the ILC study for all seven compounds quantified in the ILC.
Halogenated phenols
The analyses were carried out at the Department of Materials and Environmental Chemistry (MMK) at Stockholm University.
Analytical information
Liver tissue was homogenized and extracted with isopropanol, cyclohexane and diethylether (Jensen et al. 2009). Separation of halogenated phenols from neutral compounds was done by partitioning with potassium hydroxide (0.5 M) in 50% ethanol, and the phenolic fraction was derivatized with diazomethane in diethylether (Hovander et al. 2000). Lipids were removed using concentrated sulfuric acid and further cleanup included activated silica gel impregnated with sulfuric acid columns. Chlorinated phenols were analysed using a GC equipped with an electron capture detector. The column used was a Varian CPsil 8 CB (25 m x 0.15 mm x 12 µm) column. Brominated phenols were analysed using a GC connected to a mass
spectrometer operating in electron capture negative ionization mode. The column used was a
DB5-HT (15 m x 0.25 mm x 0.1 µm).
22 Cyclic siloxanes
The analyses were carried out at the Department of Applied Environmental Science (ITM), Stockholm University and preparation of the samples was carried out at the Swedish Museum of Natural History.
Preparation of samples
Before each sampling occasion no products containing siloxanes (shampoo, soap, lotion etc.) were used by the laboratory personnel, and before the sampling started hands were washed with Dax soap to minimize contamination. Herring from every other year between 1988 and 2010 was sampled. From each year muscle from 12 individuals were divided into two pools.
From each individual 2 grams of muscle tissue was sampled, resulting in 12 grams per pool.
Between each sampling the pool was covered with aluminum foil to prevent contamination.
The samples were packed in aluminum foil and then marked and vacuumed packed in plastic bags. For every sampled year, 10 grams of perch muscle was placed next to the sample to serves as a sampling matrix blank given the risk of contamination from the siloxanes present in the room or on the person sampling. For each herring pool an additional pool of 2 g of muscle per individual was sampled and put in a glass jar for analysis of fat content.
Analytical Information
Extraction of the biological samples was performed with a purge and trap method described in Kierkegaard et al. (2010), with an extraction time of 40 h. Instrumental analysis was
performed on a Trace GC Ultra (Thermo Electron Corp.) equipped with a MD800 MS detector (Fisons Instruments SpA) using electron ionization (EI). Five µL of the extract was injected in a large volume splitless injector at a temperature of 220ºC. The GC temperature program and ions monitored are supplied in Kierkegaard et al. (2010). At least one procedural blank and a control sample were analyzed with every extraction round of 8 samples. The control sample was a herring homogenate stored frozen in portions of 10 g. The extraction and all handling of the ENV+ traps were performed in a clean air cabinet. Details of other measures taken to reduce contamination during sample preparation and instrumental analysis are described in Kierkegaard et al. (2010) and Kierkegaard and McLachlan (2010).
Dioxins, dibenzofurans and dioxin-like PCBs
The analyses of dioxins and dioxin-like PCBs are carried out at the Department of Chemistry, Umeå University. The extraction method is described by Wiberg et al. (1998), the clean-up method by Danielsson et al. (2005), and the instrumental analysis (GC-HRMS) by Liljelind et al. (2003). The laboratory participates in the annual FOOD intercalibration rounds, including laboratory reference material (salmon tissue) with each set of samples.
Study species:
Herring
Methods, sampling, and storage
Herring samples are collected each year from seventeen sites along the Swedish coasts:
Rånefjärden, Harufjärden, Kinnbäcksfjärden (Bothnian Bay), Holmöarna, Örefjärden,
Gaviksfjärden, Långvindsfjärden, Ängskärsklubb (Bothnian Sea), Lagnö, Landsort (northern
Baltic Proper), Byxelkrok, Abbekås, Hanöbukten, Utlängan (southern Baltic Proper), Kullen,
23
Fladen (Kattegatt) and at Väderöarna (Skagerrak). Herring are also collected from two sites in the open sea, the Baltic Proper and the Bothnian Sea (by the Swedish Board of Fisheries) (Bignert et al. 2014).
Sampling of the fish is carried out every autumn, outside the spawning season. However, from two sites, Ängskärsklubb and Utlängan, herring is also sampled in spring. The collected specimens are placed individually in polyethylene plastic bags, frozen as soon as possible, and transported to the sample preparation laboratory. The age of the herring specimens are
determined using their scales. The analysed specimens are females, between 2 - 5 years of age.
Biological effects/studied parameters
For each specimen, total body weight, total length, body length, sex, age, reproductive stage, state of nutrition, liver weight and sample weight are registered.
Condition factor was calculated for each fish:
The stoutness of fish, i.e. weight versus length, is a common measure of the ‘degree of well- being’ of an individual or a population.
In this report the commonly used ‘condition factor’, K, (Vibert and Lagler 1961) is used:
K = 100 W / L
3where weight (W) is given in grams and length (L) in centimetres.
Fat percentage:
Fat content is determined in herring samples that are analysed for organochlorines.
The sample fat content is determined after extraction with acetone and hexane with 10% ether without heating (Jensen et al. 1983) in the present investigation.
In general, an extremely low fat content due to, for example starvation, may cause elevated concentrations of organochlorines expressed on a fat weight basis.
In herring muscle tissue, the subcutaneous fat layer was removed before samples were
prepared. Analyses of fat content, including skin and subcutaneous fat, showed a fat content at least 1.5 times higher than samples without skin.
Guillemot
Methods, sampling and storage
Guillemot eggs are collected each year at Stora Karlsö (Figure 2) in the southern Baltic Proper. Ten guillemot eggs, collected between weeks 19-21 (22), are analysed each year.
Normally the guillemot only lay one single egg, but if this egg is lost another may be laid. In this report, only early laid eggs are included, except for dioxins, where all collected eggs are included.
Biological effects/studied parameters
Various shell parameters, such as dry shell weight, thickness and thickness index, are
recorded for each egg.
24 Grey Seal
Methods, sampling and storage
Grey seals (Halichoerus grypus) in this study were collected 1974-2012 all year around from the entire Baltic Sea. The by-caught grey seals (n=490) and seals found dead on beaches (n=210) were sent as whole bodies or as samples, fresh or frozen, by fishermen and collectors to the Swedish Museum of Natural History (SMNH) to be necropsied and sampled for the Swedish Environmental Specimen Bank. During 2001-2012, between the 16
thof April and 31
stof December every year, samples of (n=744) shot grey seals have been collected by seal hunters from Bothnian Bay, Bothnian Sea and northern Baltic Proper. The frozen internal organs, lower jaw (with teeth) and a piece on blubber together with information of blubber thickness, date of the hunt and position were sent to SMNH for examination and sample collection.
Sampled tissues (blubber, muscle, liver, kidney, adrenals, lung, spleen, brain and blood) for contaminant analyses were wrapped in aluminium foil and vacuum packed in plastic bags before stored in -25˚C.
Biological parameters in this study
The time trends of biological parameters in Baltic grey seals in this study are the prevalence of intestinal ulcers, pregnancy rate, the heart index of liver, adrenal and thyroid weights and blubber thickness. The condition of the intestinal mucosa was evaluated on a four-degree scale: <4 mm erosions (0), 4–10 mm erosions (1), >10 mm erosions or ulcers (2), and >10 mm ulcers also affecting the muscular layer (3). Only the prevalence of intestinal lesions of moderate (2) and severe (3) extent, i.e. lesions exceeding 10 mm in diameter, was considered to be pathological (Bergman 1999) and were used in statistical calculations.
The pregnancy rate is measures as proportion of pregnant 6 years or older grey seal females in the pregnancy period (August-February).
Heart, liver and adrenal gland weight were recorded from all seals as well as the weight of thyroid glands from whole body necropsied seals. The grey seal body weight varies over the year, therefore somatic index was not used, but heart index measuring the organ weights over time. The blubber thickness was measured at sternum between the muscle layer and the skin.
This was done by seal hunters in shot seals and at SMNH in whole body necropsies.
Age determination was performed by reading dental growth layers under polarized light in the cementum zones in longitudinal tooth sections (Hewer 1964).
White-tailed Sea Eagle
