1
Contaminant profiles in female eelpout (Zoarces viviparus) and larvae from the
Baltic Sea area
Report for the BALCOFISH project
Jenny Hedman, Anh Le, Anders Bignert
___________________________________________
Swedish Museum of Natural History Department of Contaminant Research P.O.Box 50 007
SE-104 05 Stockholm Sweden
Report nr 13:2011
2 This report has been produced as a part of the BALCOFISH project.
October 2011
Analytical laboratories:
NERI National Environment Research Institute Aarhus University
Frederiksborgvej 399 PO box 358
DK-4000 Roskilde Denmark
IVL Swedish Environmental Research Institute P.O. Box 210 60
SE-100 31 Stockholm Sweden
Should be cited as:
Hedman, J.E., Le, V.A., Bignert, A. (2011). Contaminant profiles in female eelpout (Zoarces viviparus) and larvae from the Baltic Sea area. Report for the BALCOFISH project. Swedish Museum of Natural History, Sweden, report nr 13:2011.
3 SUMMARY
Here we investigated whether the eelpout (Zoarces viviparous) would be a useful indicator species for monitoring contaminants within the Baltic region, and if there were any spatial patterns in contaminant load. Eelpout adults and larvae were collected from a total of 17 sites, including both contaminated and reference sites, from 6 countries, and analysed for 7 different contaminant groups ranging from dioxins to phenolic compounds. Results showed no obvious spatial patterns between the areas investigated. However, more contaminant groups showed up in higher concentrations in Danish sites compared to the other sites. With the exception of PFASs, higher contaminant loads were usually seen in contaminated compared to reference sites, which was expected, although this was not apparent in the PCA analysis. Maternal transfer of contaminants from adult female eelpout to larvae was apparent. Because no obvious patterns were apparent across the sampled region, and as many contaminants were below LOQ in some sites, it is difficult to conclusively say whether eelpout would make the best indicator species for the Baltic region.
4 LIST OF ABBREVIATIONS
EU European union
µg Microgram
1234678-HpCDD 1,2,3,4,6,7,8-hepta chlorinated dibenzo-p-dioxins 1234678-HpCDF 1,2,3,4,6,7,8-hepta chlorinated dibenzo furan
1234789-HpCDF 1,2,3,4,7,8,9-hepta chlorinated dibenzo furan 123478-HxCDD 1,2,3,4,6,7,8-hexa chlorinated dibenzo-p-dioxins 123478-HxCDF 1,2,3,4,7,8-hexa chlorinated dibenzo furan
123678-HxCDD 1,2,3,6,7,8-hexa chlorinated dibenzo-p-dioxins 123678-HxCDF 1,2,3,6,7,8-hexa chlorinated dibenzo furan
123789-HxCDD 1,2,3,7,8,9-hexa chlorinated dibenzo-p-dioxins 123789-HxCDF 1,2,3,7,8,9-hexa chlorinated dibenzo furan
12378-PeCDD 1,2,3,7,8-penta chlorinated dibenzo-p-dioxins 12378-PeCDF 1,2,3,7,8-penta chlorinated dibenzo furan
2,4,6-TBP 2,4,6-tribromophenol 2,4-DBP 2,4-dibromophenol 234678-HxCDF 2,3,4,6,7,8-hexa chlorinated dibenzo furan 23478-PeCDF 2,3,4,7,8-penta chlorinated dibenzo furan 2378-TCDD 2,3,7,8-tetra chlorinated dibenzo-p-dioxins 2378-TCDF 2,3,7,8-tetra chlorinated dibenzo furan 4-NP 4-nonylphenol 4-t-OP 4-t-Octylphenol
AA Atomic absorption
BDE-100 2,2',4,4',6-Pentabromodiphenyl ether
BDE-153 2,2',4,4',5,5'-Hexabromodiphenyl ether
BDE-154 2,2',4,4',5,6'-Hexabromodiphenyl ether
BDE-17 2,2',4-Tribromodiphenyl ether
BDE-183 2,2',3,4,4',5',6-Heptabromodiphenyl ether
BDE-209 Decabromodiphenyl ether
BDE-28 2,4,4'-Tribromodiphenyl ether
BDE-47 2,2',4,4'-Tetrabromodiphenyl ether
BDE-49 2,2',4,5'-Tetrabromodiphenyl ether
BDE-66 2,3',4,4'-Tetrabromodiphenyl ether
BDE-85 2,2',3,4,4'-Pentabromodiphenyl ether
BDE-99 2,2',4,4',5-Pentabromodiphenyl ether
BPA Bisphenol-A Cd Cadmium Co Cobalt Cr Chromium
CRM Certified reference material
5 Cu Copper
DBT Dibutyltin
dl-PCBs Dioxin-like polychlorinated biphenyls
EQS Environmental quality standard
GC-HRMS Gas chromatography - High resolution mass spectrometry
HDPE High Density Polyethylene
Hg Mercury HPLC High pressure liquid chromatography
ICP-MS Inductively coupled plasma - Mass spectrometry
l.w. Lipid weight
LC-MS-MS Liquid chromatography - Mass spectrometry - Mass spectrometry
LOQ Limit of quantification
MBT Monobutyltin
MBTE Methyl tert-butyl ether
Me Metals Mn Manganese ng Nanogram Ni Nickel OCDD Octa chlorinated dibenzo-p-dioxins
OCDF Octachlorinated dibenzo furan
OSPAR The Oslo Paris convention
OTCs Organotin compounds
Pb Lead PBDEs Polybrominated diphenyl ethers
PBP Pentabromophenol
PCA Principal component analysis
PCB-105 2,3,3',4,4'-pentachlorobiphenyl PCB-114 2,3,4,4',5-pentachlorobiphenyl PCB-118 2,3',4,4',5-pentachlorobiphenyl PCB-123 2',3,4,4',5-pentachlorobiphenyl PCB-126 3,3',4,4',5-pentachlorobiphenyl PCB-156 2,3,3',4,4',5-hexachlorobiphenyl PCB-157 2,3,3',4,4',5'-hexachlorobiphenyl PCB-167 2,3',4,4',5,5'-hexachlorobiphenyl PCB-169 3,3',4,4',5,5'-hexachlorobiphenyl PCB-189 2,3,3',4,4',5,5'-heptachlorobiphenyl PCB-77 3,3',4,4'-tetrachlorobiphenyl PCB-81 3,4,4',5-tetrachlorobiphenyl
PCDDs Polychlorinated dibenzo-p-dioxins
PCDFs Polychlorinated dibenzo furan
6 PCP Pentachlorophenol
PFASs Perfluoroalkyl substances
PFDA Perfluorodecanoic acid
PFDoA Perfluorododecanoic acid
PFHxS Perfluorohexane sulfonate
PFNA Perfluorononanoic acid
PFOA Perfluorooctanoic acid
PFOS Perfluorooctane sulfonate
PFOSA Perfluorooctane sulfonamide
PFTrA Perfluorotridecanoic PFUnA Perfluoroundecanoic
pg Picogram
PhCs Phenolic compounds
SRM Standard reference material
TBAS Tetrabutylammonium silicate
TBBPA Tetrabromobisphenol-A TBT Tetrabutyltin
TCS Trichlosan
TEF Toxic equivalency factor
TEQ Toxic equivalency quantity
TPhT Tetraphenyltin US-EPA United state- environmental protection agency
w.w. Wet weight
Zn Zinc
7 TABLE OF CONTENTS
SUMMARY ... 3
LIST OF ABBREVIATIONS ... 4
TABLE OF CONTENTS ... 7
1. BACKGROUND ... 9
1.1THE BALCOFISH PROJECT ... 9
1.2CONTAMINANTS... 9
1.2.1 Dioxins (PCDD/Fs) and dioxin-like PCBs (dl-PCBs) ... 9
1.2.2 Polybrominated diphenyl ethers (PBDEs) ... 10
1.2.3 Organotin compounds (OTCs) ... 11
1.2.4 Metals (Me) ... 11
1.2.5 Perfluoroalkyl substances (PFASs) ... 11
1.2.6 Phenolic compounds (PhCs) ... 11
1.3AIM OF STUDY ... 12
2. METHODS ... 13
2.1SAMPLING AND SAMPLE PREPARATION ... 13
2.2CHEMICAL ANALYSES ... 15
2.2.1 Dioxins and dl-PCBs ... 15
2.2.2 Polybrominated diphenyl ethers ... 15
2.2.3 Organotin compounds ... 16
2.2.4 Metals ... 16
2.2.5 PFASs ... 16
2.2.6 Phenolic compounds ... 17
2.3DATA TREATMENT, STATISTICAL ANALYSES AND GRAPHICAL PRESENTATION ... 17
3. RESULTS AND DISCUSSION ... 18
3.1BIOLOGICAL VARIABLES ... 18
3.2CONTAMINANT CONCENTRATIONS ... 19
3.2.1 Dioxins and dl-PCBs ... 20
3.2.2 Polybrominated diphenyl ethers ... 22
3.2.3 Organotin compounds ... 23
3.2.4 Metals ... 24
3.2.5 Perfluroalkyl substances ... 26
3.2.6 Phenolic compounds ... 27
8
3.3SPATIAL PATTERN OF CONTAMINANTS IN EELPOUT FROM THE BALTIC SEA AREA ... 28
4. CONCLUSION ... 29
5. REFERENCES ... 30
APPENDIX A ... 35
9 1. BACKGROUND
1.1 The BALCOFISH project
BALCOFISH is an EU funded BONUS-169 project, with the aim to explore and establish whether the coastal fish, eelpout (Zoarces viviparus), can be used as an indicator species for assessing impacts of pollutants in the Baltic Sea areas (Förlin, 2009). The eelpout is a viviparous blenny, giving birth to fully developed larvae after ca. 5 months of gestation (Hedman et al., 2011). They are relatively stationary, inhabiting vegetated shallow sandy sediments, feeding on benthic invertebrates and small fish. Pregnant female eelpout from around the Baltic Sea have been collected within the project to link environmental stressors, such as contaminants and eutrophication, to gene responses and early warning biomarkers, reproduction, larvae development and population dynamics.
In this report, the results of chemical analyses of a wide range of contaminants in eelpout adults and larvae are presented.
1.2 Contaminants
A wide range of contaminants with different chemical properties and potential toxic effects were analyzed (table 1). Not all substances were analyzed in all matrices. Further details are in table 2.
Below is a short description of each contaminant group.
1.2.1 Dioxins (PCDD/Fs) and dioxin-like PCBs (dl-PCBs)
Polychlorinated dibenzodioxins/furans (PCDD/Fs) and polychlorinated biphenyls (PCBs) are organochlorine compounds with low water solubility and a high environmental persistency. They tend to accumulate in sediments and in the fatty tissue of organisms in the aquatic environment.
Dl-PCBs generally bio-magnify in the food chain. Dioxins were never created intentionally but are formed as by-products in many industrial processes, e.g. the production of chlorinated pesticides, and from most combustion processes, such as waste incineration and forest fires.
Formerly, pulp bleaching using chlorine gas was a source of dioxins. PCBs are synthetic compounds that do not occur naturally in the environment. They were produced and used in a variety of industrial products e.g., plasticizers, insulators, and hydraulic lubricants. Most countries banned the production of PCBs in the 1980s. Today, the main sources of PCBs to the aquatic environment include leaching from contaminated sites and improper waste disposal (Beyer and Biziuk, 2009).
The most relevant toxic effects of dioxins and dl-PCBs are developmental toxicity, carcinogenicity, immunotoxicity and endocrine disruption. They mainly exert their toxicity via binding to the aryl hydrocarbon (Ah) receptor. Of the 210 and 209 possible PCDD/F and PCB congeners, respectively, only 17 dioxin and 12 PCB congeners are considered to be of toxicological importance. 2,3,7,8-TCDD is the most toxic and well-studied congener and is used as a reference for the toxicity of the other congeners. A toxic equivalency factor (TEF) is assigned each relevant congener, where 2,3,7,8-TCDD equals 1 (Van den Berg et al., 1998; Van den Berg et al., 2006). Dioxin concentrations are commonly reported as toxic equivalents (TEQ), which is the sum of the individual congener concentration multiplied with its specific TEF.
Within the European Union, the maximum level of dioxins and dl-PCBs allowed in fish for human consumption is 8.0 pg WHO98-TEQ /g wet weight, (EC/1881/2006). For the protection of
10 wildlife (i.e. secondary poisoning of predators), an unofficial target level (Environmental Quality
Standard, EQS) in biota has been set to 0.23 pg PCDD/F g-1 wet weight.
Table 1. Contaminant groups, individual congeners and compounds analyzed in eelpout (Zoarces viviparus) adults (muscle and liver) and larvae.
Contaminant group Individual congener/compound
Dioxins and dl-PCBs
PCDD/Fs and dl-PCBs
PCDDs: 2,3,7,8-TCDD, 1,2,3,7,8-PeCDD, 1,2,3,4,7,8-HxCDD, 1,2,3,6,7,8-HxCDD, 1,2,3,7,8,9-HxCDD, 1,2,3,4,7,8,9-HpCDD, and OCDD
PCDFs: 2,3,7,8-TCDF, 1,2,3,7,8-PeCDF, 2,3,4,7,8-PeCDF, 1,2,3,4,7,8-HxCDF, 1,2,3,6,7,8-HxCDF, 2,3,4,6,7,8-HxCDF, 1,2,3,7,8,9-HxCDF, 1,2,3,4,6,7,8-HpCDF, 1,2,3,4,7,8,9-HpCDF, and OCDF
dl-PCBs: CB-77, CB-105, CB-118, CB-126, CB-156, CB-157, CB-167, CB-169, CB-189, CB-81, CB-114, CB-123
Polybrominated
diphenyl ethers PBDEs BDE-17, BDE-28, BDE-47, BDE-49, BDE-66, BDE-85, BDE-99, BDE-100, BDE-153, BDE-154, BDE-183, BDE-209
Organotin
compounds OTCs Tributyltin (TBT), dibutyltin (DBT), monobutyltin (MBT), and tetraphenyltin (TPhT)
Metals Me Mercury (Hg), cadmium (Cd), lead (Pb), zinc (Zn), nickel (Ni), copper (Cu), chromium (Cr), manganese (Mn), and cobalt (Co)
Perfluoroalkyl
substances PFASs PFOA, PFNA, PFDA, PFUnA, PFDoA, PFTrA, PFHxS, PFOS, and PFOSA.
Phenolic compounds PhCs
Bisphenol-A, 4-Nonyphenol, 4-t-Octylphenol, trichlosan, pentachlorophenol, 2,4-Dibromophenol, 2,4,6-Tribromophenol, pentabromophenol, and tetrabromo-bisphenol-A
1.2.2 Polybrominated diphenyl ethers (PBDEs)
Polybrominated diphenyl ethers (PBDEs) are lipophilic, persistent organic compounds used in large quantities since the 1970s as flame retardants in a wide variety of products, such as plastics, textiles and electronic products (de Wit, 2002). Penta- and octa- BDEs are partially banned in the EU since 2006. Sources to the environment include direct discharges at production and processing facilities, as well as diffuse leaching from treated products during the entire life cycle.
PBDEs bio-accumulate and bio-magnify in the aquatic environment and increasing levels are being detected in wildlife (de Wit, 2002; Norstrom et al., 2002; Sellström et al., 1993). The toxicity of PBDEs includes endocrine disruption (thyroid system), neurobehavioral effects and immunotoxicity. Generally, lower brominated compounds are more toxic. PBDEs also have the ability to interact with the Ah receptor (Chen and Bunce, 2003).
11 1.2.3 Organotin compounds (OTCs)
Tributyltin (TBT) and triphenyltin (TPhT) have been widely used as agricultural pesticides and antifouling agents in marine paints, while dibutyltin (DBT) and monobutyltin (MBT) are mainly used as heat and light stabilizers for PVC materials (Hoch, 2001). Leaching from marine paints used on ships, leisure boats and marine constructions is the main source of TBT and TPhT to the aquatic environment. OTC compounds are hydrophobic and strongly associate to particles in natural waters and are ultimately deposited in the sediments where they have a long persistency.
TBT and TPhT bio-accumulate in individual organisms, and potential to bio-magnify in the aquatic food chain (Strand and Jacobsen, 2005). TBT and TPhT are well-known to be highly toxic to aquatic organisms, particularly mollusks, even at low concentrations (Hoch, 2001). In fish, TBT and TPhT have been shown to have negative effects on reproduction, larvae development and the immune system. The use of TBT in anti-fouling paints has been effectively banned in the Baltic Sea on pleasure boats since 1992 and on larger commercial ships since 2008.
1.2.4 Metals (Me)
Several metals cause serious environmental problems due to their toxicity to wildlife and humans. Metals of particular concern include mercury, cadmium, lead, zinc, nickel, copper, chromium, manganese and cobalt. These metals have both natural and anthropogenic sources, e.g. from the production of batteries, papers, pesticides, and fertilizers (Fu and Wang, 2011).
Zinc, nickel, copper, chromium, manganese and cobalt are nutritionally essential elements for many plants and animals to a certain level, above which they can cause a wide range of toxic effects (Paulino et al., 2006; Richard and Bourg, 1991). Mercury, cadmium and lead have no known beneficial biological effects (non-essential) and are toxic even at low concentrations.
Mercury concentrates in muscles, while most other metals concentrate in the liver.
1.2.5 Perfluoroalkyl substances (PFASs)
Per/polyfluoroalkyl substances (PFASs) are a diverse group of chemicals widely used as industrial additives and surfactants in e.g. detergents, flame retardants and lubricants (Prevedouros et al., 2005). They consist of a hydrophilic functional group and a hydrophobic fluorinated chain. PFASs are extremely stable chemicals and persistent in the environment. The most investigated compounds are PFOS, PFOA and PFNA (table 1), which are bio-accumulative and have been found in wildlife and humans as a result (Giesy and Kannan, 2001; Prevedouros et al., 2005). PFASs are released to the environment at production and product use, and indirectly through biotic and/or abiotic degradation of precursor molecules used in surface treatment applications (Prevedouros et al., 2005). PFASs can cause toxic effects in the reproductive and developmental systems of wildlife and humans (Lau et al., 2004).
1.2.6 Phenolic compounds (PhCs)
Several phenolic compounds are acutely toxic to aquatic organisms and have estrogen mimicking effects. Simple brominated phenols (2,4-DBP, 2,4,6-TBP, PBP) are used as flame retardants and wood preservatives but are also known to be naturally produced by some marine organisms.
They are generally not readily biodegradable and will persist in the environment. The toxicity is expected to increase with increasing bromination of the phenol (Howe et al., 2005).
12 Pentachlorophenol (PCP) and trichlosan (TCS) are chlorinated PhCs used as antimicrobials and
disinfectants in e.g. agriculture (PCP) and consumer products (TCS). They are not readily dissolved in water and have a potential to bio-accumulate but bio-magnification in the food chain is probably not significant. Bisphenol A (BPA) is primarily used in the production of plastics and formerly as a fungicide. BPA is not classified as bio-accumulative, but have been shown to have endocrine effects in fish at low exposure levels, as well as affecting growth, reproduction and development in aquatic organisms (Colerangle and Roy, 1997; Fürhacker et al., 2000). BPA is a precursor to the flame retardant tetrabromobisphenol A (TBBPA). TBBPA is bio-accumulative and may cause developmental toxicity in fish at environmental concentrations. Alkylphenols, including 4-nonylphenol (4-NP) and 4-t-octylphenol (4-t-OP), is a major class of non-ionic surfactants widely used in industrial detergents (alkylphenol ethoxylates) and pesticide formulations (Colerangle and Roy, 1997; Fürhacker et al., 2000; Staples et al., 1999).
Alkylphenols have a hydrophilic and a hydrophobic unit, are persistent and may accumulate in aquatic organisms (Ekelund et al., 1990; Lewis and Lech, 1996; Mayer et al., 2002; Veldhoen et al., 2006; Zorrilla et al., 2009). Alkylphenolic compounds have estrogen mimicking properties and the potential to disrupt the endocrine system (Colerangle and Roy, 1997).
1.3 Aim of study
The aim of this study was to provide the BALCOFISH project with information on contaminant concentrations in adult eelpout and larvae from polluted and less polluted areas in the Baltic Sea (Task 3.1 in the BALCOFISH project). A variety of contaminants of current interest in aquatic ecotoxicology, with different chemical properties and sources, were chosen. Contaminant concentrations are compared with international target levels and the spatial pattern is analyzed.
Toxic effects and spatial differences measured in eelpout from the same areas (in most cases the same individuals as used for chemical analysis) will be further evaluated together with the contaminant concentrations within the BALCOFISH project to find potential correlations between pollution and effect.
13 2. METHODS
2.1 Sampling and sample preparation
Female eelpout were collected at 17 sites in the Baltic Sea, Kattegat and Skagerrak (fig. 1).
Twelve sites were sampled in November 2009 within the BALCOFISH project and five sites during autumn 2009 and 2010 within the BEAST project (table 3, BALCOFISH report, BEAST report). Sites were chosen to represent areas of low pollution (reference sites) and polluted (contaminated) sites. Biological effect parameters, e.g. biomarker responses and larval development, have been measured in female eelpout and larvae collected concurrently (results presented in separate BALCOFISH and BEAST reports, respectively). All fish were stored at -20 ºC until sample preparation.
Muscle samples were prepared by dissecting an equal amount of muscle tissue (not including skin and sub-cutaneous fat) from individual fish to a pooled weight of ca. 40 g w.w. from each site. The muscle tissue of each pool was homogenized and divided into clean glass jars (dioxins, BDEs, OTCs, PhCs) or plastic tubes (Me/Hg) and kept frozen at -20 ºC until chemical analysis.
Liver samples were only available from the Swedish sites and two Danish sites (table 2). At Fjällbacka (5SF) and Kvädöfjärden (6SK), not enough eelpout liver was collected in November 2009. It was therefore supplemented with liver from eelpout collected at the same sites in November 2008 and 2010. An equal amount of liver tissue was dissected from each fish, using plastic utensils, to a pooled weight of ca. 2.5 g w.w.. Homogenized samples were divided into two plastic jars (Me/Hg, PFASs) and kept frozen until chemical analysis.
5SF
4SS 3SG
6SK
7SGs 8SSm
13DF 12D R
14DA
1 1GE 2GW1GS
19PP
15LR1 17FF4
16LR3 18SGl
5SF
4SS 3SG
6SK
7SGs 8SSm
13DF 12D R
14DA
1 1GE 2GW1GS
19PP
15LR1 17FF4
16LR3 18SGl
Figure 1. The 17 sampling sites throughout the Baltic Sea, Kattegat and Skagerrak. See table 1 for further information about site names and ID.
14
Table 2. Number of individuals in pooled samples, and chemical analysis in different matrices.
Pool ID Nr of Ind.
in pool
Sampled matrix
Chemical analysis
Dioxins BDEs OTCs Me/Hg PFASs PhCs
1GS 20 Muscle x x x x - x
2GW 21 Muscle x x x x - x
11GE 20 Muscle x x x - - x
3SG 20 Muscle x x x x - x
4SS 20 Muscle x x x x - x
5SF 20 Muscle x x x x - x
6SK 20 Muscle x x x x - x
7SGs 20 Muscle x x x x - x
8SSm 20 Muscle x x x x - x
12DR 20 Muscle x x x x - -
13DF 20 Muscle x x x x - -
14DA 11 Muscle x x x x - -
15LR1 10 Muscle x x x x - x
16LR3 10 Muscle x x x x - x
17FF4 21 Muscle x x x x - x
18SGl 20 Muscle x x x x - x
19PP 9 Muscle x x x x - x
1GS - Liver - - - - - -
2GW - Liver - - - - - -
11GE - Liver - - - - - -
3SG 17 Liver - - - x x -
4SS 20 Liver - - - x x -
5SF 54 Liver - - - x x -
6SK 55 Liver - - - x x -
7SGs 20 Liver - - - x x -
8SSm 11 Liver - - - x x -
12DR 20 Liver - - - x x -
13DF 20 Liver - - - x x -
14DA 11 Liver - - - x x -
15LR1 - Liver - - - - - -
16LR3 - Liver - - - - - -
17FF4 - Liver - - - - - -
18SGl - Liver - - - - - -
19PP - Liver - - - - - -
1GS - Larvae - - x x x x
2GW - Larvae - x x x x x
11GE - Larvae - x x x x x
3SG - Larvae x x x x x x
4SS - Larvae x x x x x x
5SF - Larvae - x x x x x
6SK - Larvae - - x x x x
7SGs - Larvae - - x x x -
8SSm - Larvae - - - - - -
12DR - Larvae x x x x x -
13DF - Larvae x x x x x -
14DA - Larvae x x x x x x
15LR1 - Larvae - - - - - -
16LR3 - Larvae - - - - - -
17FF4 - Larvae - - - - - -
18SGl - Larvae - - - - - -
19PP - Larvae - - - - - -
15
Table 3. Description of sampling sites.
Country Site Pool ID Sampling Date Project Pollution
pressure1 Germany
Salzhaff 1GS 2009-11-01 BALCOFISH R Wismar 2GW 2009-11-04 BALCOFISH C Eggers Wiek 11GE 2009-11-11 BALCOFISH R
Sweden
Göteborg harbor 3SG 2009-11-12 BALCOFISH C Stenungsund 4SS 2009-11-10 BALCOFISH C Fjällbacka 5SF 2009-11-09* BALCOFISH R Kvädöfjärden 6SK 2009-11-03* BALCOFISH R Gåsö 7SGs 2009-11-05 BALCOFISH C Slakmöre 8SSm 2009-11-04 BALCOFISH C Denmark
Roskilde 12DR 2009-11-06 BALCOFISH C Fredriksvaerk 13DF 2009-11-03 BALCOFISH C Agersö 14DA 2009-11-11 BALCOFISH R
Latvia Gulf of Riga 1 15LR1 2009-12-14 BEAST C
Gulf of Riga 3 16LR3 2009-12-16 BEAST C
Finland Gulf of Finland 17FF4 2009-08-27 BEAST C
Sweden Gävle off-shore 18SGl 2010-12-09 BEAST R
Poland Coast 19PP 2009-09-01 BEAST R
1R = reference site/low pollution; C = contaminated site.
* Liver samples pooled from fish collected 2008, 2009 and 2010.
Larvae were only available from the BALCOFISH sites (table 2). The larvae were dissected on- site directly after collection and kept frozen. There is no record of the number of larvae from each individual female reserved for chemical analysis. When available, ca. 40 g w.w. of larvae was pooled from each site, homogenized, divided into clean containers and kept frozen until chemical analysis. In sites where this amount of larvae was not available, some chemical analyses had to be omitted (table 2). All samples were frozen or semi-frozen throughout the whole preparation process.
2.2 Chemical analyses 2.2.1 Dioxins and dl-PCBs
The samples were homogenized and an extraction spike (13C-labelled PCDD/Fs and PCBs) was added before extraction (0.5 ng for each congener). The sample was left to dry for 2 days in a fume cupboard. The extraction was performed with toluene in a Soxhlet apparatus for 20 hours.
The solvent was evaporated and 100 ml hexane was added. Before clean up, the extract was treated with acid silica. Clean up was performed following the method published by Aries et al.
(Aries et al., 2006). The method is based on combined multilayer silica and florisil clean up, so that a PCDD/F fraction and a WHO-PCB fraction are obtained separately. Each fraction was evaporated and reconstituted in dodecane. The analysis was performed by GC-HRMS at 10,000 resolution using a DFS from Thermo Scientific. Concentrations were calculated by the isotope dilution method as recommended by the European Standard EN 1948-3 (2006).
2.2.2 Polybrominated diphenyl ethers
Method description still missing from analyzing laboratory.
16 2.2.3 Organotin compounds
The organotin analysis was performed on sub-samples consisting of 2 g w.w. of the homogenised sample material. These were digested with HCl and 30 min sonication before being adjusted to pH5 with ethylacetate and NaOH. Further steps included 3 times in situ derivatisation with 1 ml of 10% sodium tetraethylborate in methanol, extraction with pentane, and centrifugation before the organic phase was gently evaporated to ~0.1 ml using a flow of argon. The organotin compounds tributyltin (TBT), dibutyltin (DBT), monobutyltin (MBT), and triphenyltin (TPhT) were determined with GC-PFPD i.e., gas chromatography (Varian 3500 with a 30 m 0.25 mm/0.25 mm ZB-5 column, Phenomex) and a pulsed flame photometer detector (Varian, USA).
Organotins were quantified from a tripropyltin internal standard.
Laboratory procedures and quality assurance followed accredited methods. Detection limits for organotins ranged from 0.1–1 mg Sn kg-1 w.w. The certified reference material CRM 477 was used for quality control material (Institute for Reference Materials and Measurements (IRMM), Geel, Belgium).
2.2.4 Metals
Metal analyses were performed on freeze-dried sub-samples of homogenised fish tissues, by taking 0.4 g for closed vessel microwave digestion in a Multiwave microwave oven (Anton Paar, Graz, Austria) with 20 ml of 7M nitric acid. A 12 rotor system with six temperatures and one pressure sensor was used. Samples were checked for residues after digestion, and poured directly into pre-weighed 100 ml HDPE containers. Exactly 80 ml of 18 Mohm milliQ water and 0.5 ml 1 ppm Ir internal standard were added before weighing the HDPE container.
Metal concentrations in the digested tissue were quantified using a 7500ce ICP-MS (Agilent) with a Babington nebuliser connected to a cooled spray chamber (5 °C). The mist was introduced into an Ar-plasma operating at 1500 W with 15 l argon per minute. Standard mass-overlap correction from US-EPA 6020 was used to correct the signal before the calibration was calculated, and drift was corrected by using Rh, Ir and In as internal standards. Zn was determined using acetylene-air flame atomic absorption (AA) on a Perkin Elmer 5100PC, Cd using a transversally heated graphite furnace AA on a Perkin Elmer 3300 AA, and finally Hg was determined using cold vapour AA spectroscopy on a Perkin Elmer FIMS 400 system.
Laboratory procedures and quality assurance followed accredited methods. Detection limits and quality assurance followed OSPAR guidelines. Detection limits for metals ranged from 0.05 to 10 mg kg-1 dry weight (d.w.). For every 10 samples, two quality control samples of freeze-dried mussel Reference Material (internal reference material ‘‘MARINA’’ or certified SRM 2976 (National Institute of Standards and Technology, United States)) were included.
2.2.5 PFASs
The extraction method for biota was based on ion pairing as described by Hansen et al. (Hansen et al., 2001) with some small modifications. The tissue was homogenized and 5 g was diluted with 25 ml deionized water. The homogenate (1 ml) was transferred to a polypropylene centrifuge tube and 100 ng each of the internal standards (13C-PFOA, 13C-PFDA and 13C-PFOS) were added. The sample was extracted with 1 ml of a 0.5 M TBAS solution (pH adjusted to 10), 2 ml of sodium carbonate/sodium bicarbonate buffer and 5 ml MTBE by shaking for 20 minutes.
17 The organic phase was isolated by centrifuging the sample for 25 minutes at 3500 rpm. The
MTBE supernatant (4 ml) was transferred to a polypropylene tube and the solvent was evaporated to dryness under nitrogen flow. The extract was reconstituted in 1000 l methanol/2 mM ammonium acetate (1:1, v/v), vortexed for 15 s. and filtered through a 0.2 µm nylon filter.
All sample extracts were analysed by liquid chromatography-tandem mass spectrometry (LC- MS-MS) with electrospray ionization (ESI) operated in negative mode. The extracts (20 l injection volume) were chromatographed on a C18 Betasil column (2.1 x 50 mm, Thermo Hypesil-Keystone, Bellafonte, PA) using an Agilent 1100 Series HPLC (Agilent Technologies, Palo Alto, CA). The mobile phase A was 2 mM ammonium acetate and the mobile phase B was methanol/2 mM ammonium acetate (90:10, v/v). The HPLC was interfaced to a triple quadrupole API 2000 (Sciex, Concorde, ON, Canada) equipped with a TurboIon Spray source operating in negative ion mode. The analyses were performed with a multiple reaction monitoring (MRM) method that monitored two mass transitions (parent ion/product ion) for each analyte. For PFNA only one mass transition was obtained. Positive identification of the analytes was based on retention time with the most abundant mass transition corresponding to an authentic standard.
Confirmation of analyte identity was based on the relative response of the secondary mass transition to the primary mass transition.
2.2.6 Phenolic compounds
Method description still missing from analyzing laboratory.
2.3 Data treatment, statistical analyses and graphical presentation
Values below the quantification limit were substituted using the reported LOQ divided by the square root of 2 (Helsel, 2005). PCDD/Fs were converted to TEQ pg g-1 by multiplying concentrations with their TEF values. Some contaminants e.g., PCDD/Fs, PCBs, and PhCs, were adjusted for fat content by converting from wet weight to lipid weight using the reported fat percentage for each pooled sample. Analytical results are displayed using standard and stacked bar charts, and in some instances, where detected concentrations were especially low, scatter plot points overlay the bar graphs so all substances can be viewed in the same picture.
The condition factor (CF) was calculated as:
CF = total weight* 100/length3 (Vibert and Lagler, 1961). Eq. 1 (Weight is reported in grams (g) and length in centimeters (cm)).
Principal Component Analysis (PCA) was used on the proportions of various groups of chemical concentrations to the sum of all groups to identify contaminant patterns. The percentage of each PFAS relative to the sum was calculated and log-transformed prior to the PCA analysis. Before the PCA-scores were plotted, they were centered and scaled to 100%. The eigenvector loadings were added to the PCA plot as vectors using the software PIA (Plot and Image Analysis, Bignert, 2007, available at www.amap.no). Geometric means and 95% confidence intervals were calculated for the eelpout biological data i.e., length, weight, and condition factor.
18 3. RESULTS AND DISCUSSION
3.1 Biological variables
Eelpout from all sites, except Poland (19PP, 39.3 cm, 279.5 g), had a mean length and weight ranging from 19.8 - 27.9 cm and 34.8 - 108.6 g, respectively (table 4). Eelpout from the northern Baltic Sea (Sweden, Finland, and Latvia) were generally smaller than fish from further south (Denmark, Poland and Germany). This may be related to differing temperature and/or salinity conditions; however, these parameters were not examined here. The condition factor (CF) was highest at the Danish sites (0.59 - 0.76), but otherwise was similar between sites (0.38 - 0.50) (table 4). Fat content varied across sites (range 0.35 - 1.43%), with eelpout from Latvia and Finland having the highest fat content (0.90 and 1.43%, respectively) i.e., a percentage more than two fold higher compared with other sites.
Table 4. Biological variables of eelpout. Geometric mean with 95% confidence interval (CI) in brackets.
Pool ID Length (cm) (CI)
Weight (g) (CI)
Condition factor (Eq. 1)
(CI)
Muscle fat content (%) 1GS 26.97
(26.43 – 27.51)
82.22 (76.52 – 88.35)
0.42
(0.40 – 0.43) 0.518 2GW 27.91
(27.17 – 28.67)
103.6 (95.43 – 112.4)
0.48
(0.46 – 0.50) 0.577
11GE 27.94
(26.80 – 29.13)
108.58 (94.88 – 124.26)
0.50
(0.48 – 0.51) 0.837 3SG 22.70
(21.58 – 23.88) 56.90
(47.44 – 68.26) 0.48
(0.46 – 0.51) 0.520 4SS 22.94
(21.78 – 24.16)
58.32 (49.42 – 68.81)
0.48
(0.46 – 0.50) 0.401 5SF 24.06
(22.30 – 25.96)
64.13 (49.88 – 82.46)
0.46
(0.43 – 0.49) 0.574 6SK 21.56
(20.73 – 22.42)
40.05 (35.43 – 45.27)
0.40
(0.39 – 0.42) 0.531
7SGs 20.84
(20.04 -21.68)
34.51 (30.49 – 39.07)
0.38
(0.37 – 0.39) 0.532
8SSm 19.76
(18.70 – 20.88) 30.59
(25.96 – 36.04) 0.39
(0.38 – 0.41) 0.542
12DR 20.84
(19.95 – 21.77) 68.38
(60.63 – 77.15) 0.76
(0.72 – 0.79) 0.570
13DF 21.53
(20.70 – 22.39)
62.00 (55.09 – 69.79)
0.62
(0.59 – 0.65) 0.350
14DA 25.07
(23.79 – 26.42)
92.56 (78.79 – 108.75)
0.59
(0.55 – 0.63) 0.350
15LR1 23.15
(22.08 – 24.28)
57.31 (49.10 – 66.90)
0.46
(0.42 – 0.52) 0.893
16LR3 25.46
(24.40 – 26.56) 82.56
(75.33 – 90.50) 0.50
(0.47 – 0.53) 0.900
17FF4 19.76
(18.90 – 20.67)
35.33 (30.99 – 40.23)
0.46
(0.43 – 0.49) 1.431
18SGl 22.59
(21.81 – 23.39)
55.10 (49.43 – 61.43)
0.48
(0.45 - 0.50) 0.668
19PP 39.28
(35.67 – 43.25)
279.5 (227.1 – 343.9)
0.46
(0.40 – 0.53) 0.581
19 3.2 Contaminant concentrations
Tables 5-7 present the measured contaminant concentrations in eelpout muscle, larvae and liver.
The concentrations are the sum of all congeners (PCDD/Fs, BDEs) or individual compounds in each contaminant group. There is one pooled sample from each site. The results for each contaminant group are discussed separately below.
Table 5. Concentrations of examined contaminants in eelpout muscle. Summary of all individual congeners (PCDD/Fs, BDEs) or compounds in each contaminant group. One pooled sample per site.
ID
∑PCDD/Fs dl-PCBs BDEs OTCs Hg PhCs
pg g-1
l.w. ng g-1 l.w. ng g-1 d.w. µg g-1 w.w. ng g-1
l.w.
1GS 160.5 6.37E+04 21.41 4.27 4.59E-02 1.66E+03 2GW 180.7 2.53E+05 33.66 11.76 5.17E-02 4.24E+03
11GE 167.1 6.75E+04 10.18 4.49 - 1.06E+03
3SG 160.5 3.90E+05 43.03 10.74 9.37E-02 1.87E+03 4SS 205.8 1.59E+05 29.20 7.12 4.28E-02 2.41E+03 5SF 146.4 9.13E+04 35.90 3.35 1.04E-02 5.90E+02*
6SK 152.9 6.10E+04 64.12 4.07 1.47E-02 1.90E+03
7SGs 163.4 1.03E+05 25.98 4.66 3.21E-02 2.16E+03
8SSm 171.40 8.13E+04 28.29 3.52 2.08E-02 2.54E+03
12DR 400.4 1.33E+05 34.16 29.20 2.19E-02
13DF 396.7 1.14E+06 63.99 17.50 1.40E-02 No sample
14DA 314.6 4.26E+04 34.35 5.01 4.38E-02
15LR1 1603. 3.26E+05 33.00 4.68 6.18E-02 1.07E+03
16LR3 817.5 2.68E+05 61.45 4.58 6.41E-02 1.22E+03
17FF4 190.0 5.92E+04 8.97 5.42 4.35E-02 7.93E+02
18SGI 407.2 1.29E+05 26.70 5.33 5.15E-02 1.43E+03
19PP 217.1 2.44E+04 15.38 7.02 9.93E-02 1.55E+03
*Nonylphenol was removed from this value due to an analytical error.
- indicates no values for Hg at this site
Table 6. Concentrations of examined contaminants in eelpout liver. Summary of all individual congeners or compounds in each group. One pooled sample per site.
ID OTCs Cd Pb PhCs
ng g-1 d.w. µg g-1 d.w. µg g-1 d.w. ng g-1l.w.
3SG 230.46 0.06 0.07
No sample
4SS 280.29 0.110 0.04
5SF 323.40 0.092 0.03
6SK 351.49 0.517 0.04
7SGs 297.22 0.396 0.04
8SSm 324.83 0.28 0.08
12DR 73.63 0.141 0.04 6114.63
13DF 20.50 0.092 0.03 3908.09
14DA 67.60 0.211 0.08 2376.66
20
Table 7. Concentrations of examined contaminants in eelpout larvae. Summary of all individual congeners or compounds in each group. One pooled sample per site.
ID
∑PCDD/
Fs dl-PCBs BDEs OTCs Hg Cd Pb PFASs PhCs
pg g-1 l.w. ng g-1
l.w.
ng g-1 d.w.
µg g-1 w.w.
µg g-1 d.w.
µg g-1
d.w. ng g-1 l.w. ng g-1 l.w.
1GS
No sample
2.01 0.18 0.05 0.21 2.13+03 0.48E+03 2GW 10.45 9.24 0.14 0.05 0.20 1.74E+03 0.59E+03
11GE 6.53 3.19 0.09 0.05 0.54 1.75E+03 1.18E+03
3SG 47.22 3.71E+05 24.93 9.30 0.04 0.05 0.22 2.24E+03 1.65E+03 4SS 53.38 2.09E+05 14.09 1.98 0.05 0.05 0.20 1.83E+03 0.55E+03 5SF
No sample
7.26 2.23 0.04 0.05 0.47 1.59E+03 1.02E+03
6SK No
sample
1.77 0.03 0.05 0.20 2.15E+03 0.45E+03
7SGs 3.5 0.04 0.05 0.20 2.05E+03
No sample
8SSm No sample
12DR 150.49 8.91E+04 11.12 17.30 0.02 2.70 0.05 2.12E+03
13DF 240.03 7.23E+05 29.01 8.60 0.02 2.60 0.20 9.24E+03
14DA 102.48 2.69E+04 19.98 8.11 0.18 0.05 0.29 1.78E+04 0.47E+03
3.2.1 Dioxins and dl-PCBs
Different PCDD/F concentrations were detected in the eelpout sampled at the 17 sites. In general, dioxin concentrations in eelpout were lower than the detection limit. The highest concentrations in both eelpout and larvae are seen at site 15LR1, the Gulf of Riga 1, Latvia (fig.
2), followed by Fredriksvaek, Denmark. These high values are not surprising as those sites are contaminated and a high fat content was observed from eelpout in these areas. Higher concentrations of the lower chlorinated congeners e.g., TCDD/Fs and PeCDD/Fs, are observed.
In particular, 2,3,4,7,8-PeCDF had high concentrations compared to other congeners at the two sites in Latvia. In all samples, concentrations of PCDFs are high compared to PCDDs. Levels of PCDD/Fs in all selected sites are above 0.23 pg g-1 w.w., which is the suggested target value in fish for the protection of predators against secondary poisoning (Bignert et al., 2011). However, the observed concentration is still below the European limit for dioxin concentrations in the muscle of fish for human consumption i.e., 4 pg g-1 w.w. WHO98-TEQ (∑PCDD/Fs)(2006/199/EC, 2006).
PCDD/Fs were also found in larvae with quite similar congener patterns to eelpout. Larvae are fed by the transfer of maternal nutrients through ovarian fluid containing nutritive compounds e.g., amino acids and low molecular organic substances. Chemicals from the mother transfer to the larvae via an active exchange mechanism (Hedman et al., 2011). Therefore, larvae are likely to be affected by maternal exposure to contaminants in the environment if those substances bio- accumulate in the eelpout adult and are able to be transferred via ovarian fluid. The maternal – fetal trophic relationship has been reported in the literature (Hedman et al., 2011; Napierska and Podolska, 2006). Here, larvae had higher dioxin concentrations compared to adult females on a wet weight basis (fig. 2, left). However, if the concentration was normalized for fat content, the opposite trend was observed (fig. 2, left). This can be explained by the differences between fat content of larvae (1.94 - 2.57%) and eelpout (0.35 - 1.43%). The eelpout:larvae concentration
21 ratios, which were calculated based on lipid weight, ranged from 1.65 - 3.85. Similar congener
patterns between sites and between eelpout and larvae, but varying concentrations, were observed. Because eelpout are a stationary species, differences in these ratios between sites can probably be explained by local environmental parameters e.g., temperature, salinity, local discharges, community structure, diet.
Figure 2. PCDD/Fs in eelpout muscle (M) and larvae (L) in the Baltic Sea.
Left: pg WHO-98 TEQ g-1 w.w. Right: pg g-1 l.w. R = reference site; C = contaminated site. See table 2 for sampling site ID.
dl-PCBs were detected in eelpout at most sites, except for congener CB-169 (not found in Eggers Wiek and coastal Poland). The result was plotted based on wet (fig. 3, left) and lipid weight (fig.
3, right). There was no obvious relationship between dl-PCB concentrations and fish length. By contrast, a negative relationship was observed between dl-PCB concentrations and fat content i.e., as fat content increased, dl-PCB concentrations decreased. Here, the contaminated Danish site had the highest dl-PCB concentrations (Fredriksvaek: 1.14x106 pg g-1 l.w.), followed by a Swedish site (Göteborg Harbor: 3.90x105 pg g-1 l.w.) and Latvia (Gulf of Riga 1: 3.26x105 pg g-1 l.w.). CB-118, CB-105 and CB-167 were dominant congeners at most sites. The concentration in muscle was markedly higher at contaminated sites in comparison to reference sites.
Concentrations of dl-PCBs were lower in larvae compared to adult females (fig. 3, right), and the eelpout:larvae concentration ratios ranged from 0.8 - 1.6. No sites exceeded the permissible level for total PCDD/Fs and dl-PCBs in muscle meat of fish and fishery products (TEQ 8 pg g-1 w.w.)(2006/199/EC, 2006). However, CB-118 exceeds the Background Assessment Criteria (BAC) developed by OSPAR ( 0.1 ng g-1 w.w.) (Bignert et al., 2011) in most sites, except at Fredriksvaerk (0.05 ng g-1 w.w.) and coastal Poland (0.06 ng g-1 w.w.). CB-105 also exceeded the OSPAR BAC value (0.08 ng g-1 w.w.) (Bignert et al., 2011) in several contaminated sites.
Interestingly, CB-105 exceeds the OSPAR BAC value at both Fjällbacka and Gälve off-shore, which are reference sites. As it is unusual for a reference site to contain such high concentrations
22 of these substances, and as eelpout are largely stationary, the question is raised as to whether
there is an unknown point source in these areas.
Figure 3. Dl-PCBs in eelpout muscle (M) and larvae (L) in the Baltic Sea. Left: pg g-1 w.w. Right: pg g-1 l.w.
R = reference site; C = contaminated site. See table 2 for sampling site ID.
3.2.2 Polybrominated diphenyl ethers
Twelve BDEs were analyzed in muscle tissue of eelpout and larvae (table 1). In general, BDE- 17, 28, 49, 66, 85, 153, 154 and 183 were not detected at more than half of the total examined sites (lower than laboratory LOQ). BDE-77, 99, 100 and 209 are found in most sites, where BDE-47 and BDE-209 are the dominant congeners. This has also been observed in other species e.g., herring, sprat, and salmon (Szlinder-Richert et al., 2010). According to the Environmental Criteria for food consumption by humans via fishery products, the permissible level for PBDEs (congeners 28, 47, 99, 100, 153, and 154) is 0.27 µg g-1 w.w. (Lilja et al., 2010). Three contaminated sites exceeded this recommended limit i.e., Göteborg harbor (0.46 ng g-1 w.w.), the Gulf of Riga 3 (0.46 ng g-1 w.w.), and Fredriksvaek (0.49 ng g-1 w.w.). A negative relationship between fat content and PBDE concentrations were also observed, but this was not the case for length and weight. PBDE concentrations were higher in eelpout from contaminated than reference sites (fig. 4). PBDE concentrations in muscle are generally < 0.3 µg g-1 w.w., while concentrations in larvae in some contaminated sites were as high as 0.7 µg g-1 w.w. (fig.4, left).
Eelpout:larvae concentration ratios based on lipid weight ranged from 1.6 - 4.9, meaning adult females have higher concentrations than larvae (fig. 4, right); however, the opposite trend is observed when the ratios are calculated based on wet weight. This can be explained by the higher fat content found in larvae compared to adult females.
23
Figure 4. PBDEs in eelpout muscle (M) and larvae (L) in the Baltic Sea. Left: ng g-1 w.w. Right: ng g-1 l.w. R = reference site; C = contaminated site. See table 2 for sampling site ID
3.2.3 Organotin compounds
Shipyards and ports contain high amounts of TBT and TPhT from antifouling paint on ships.
These chemicals end up in sediment, and can remain unchanged for years. Therefore, benthic organisms in contaminated harbors can be exposed to high organotin concentrations. As benthic invertebrates form part of the eelpout diet, bioaccumulation of organotins is expected. Here, considerable variation is seen in organotin concentrations at different sites (fig. 5). High concentrations are observed at contaminated sites, particularly in Denmark where all four selected organotins were detected. Total concentrations varied from 2-30 ng g-1 d.w.. TBT, the most toxic organotin substance here, has higher concentrations compared to DBT and MBT. This same trend was observed in Polish mussels (Mytilus edulis) and European flounder (Platichthys flesus), indicating the dominance of TBT uptake from seawater (Albalat et al., 2002). Roskilde and Fredriksvaek have the highest TPhT concentrations, while at the other sites it is below LOD.
TPhT is observed at very low concentrations in muscle; however, higher concentrations are observed in liver (table 15). This agrees with Albalat et al. (2002). Concentrations of organotin in eelpout (0.9-2.2 ng g-1 w.w.) are low in comparison to fish from the German North Sea (eelpout 102 ng g-1 w.w.; bream 202 ng g-1 w.w. (Shawky and Emons, 1998)), or with mussels from the Polish coast (27-67.7 ng g-1 w.w.) (Albalat et al., 2002). Variable concentrations have been reported because organotin concentration depends on species type, habitat, and tissues examined.
As organotins bio-accumulate, the observed variation in concentrations in eelpout here could be related to diet, which would be influenced by ecosystem and community structure at each site.
These, in turn, are related to the physical conditions at each site, which were not investigated here. Maternal transfer of organotins was also observed, with adult females showing higher concentrations than larvae at all sites (fig. 5). Eelpout:larvae concentration ratios ranged from 0.6 - 3.5.
24
Figure 5. Concentrations (ng g-1 d.w.) of organotin compounds in eelpout muscle (M) and larvae (L) in the Baltic Sea. R = reference site; C = contaminated site. See table 3 for sampling site ID.
3.2.4 Metals
Mercury was analyzed in muscle tissue and larvae (fig. 6). Mercury bioaccumulates and biomagnifies in aquatic organisms, attaining its highest concentration in tissues of long-lived predatory fish (Essink, 1988; Schuhmacher et al., 1994). Eelpout feed on benthic invertebrates and small fish, and are therefore exposed to mercury at a medium trophic level. High mercury concentrations were observed in eelpout from contaminated sites i.e., Göteborg Harbour (Sweden), and the Gulf of Riga 1 and 3 (Latvia) (fig. 6). Polish eelpout have the highest observed mercury concentrations (100 ng g-1 w.w.), probably because they are the oldest fish sampled. Their concentration is five times higher than the European Community Environmental Quality Standard (EQS) limit for mercury in biota i.e., 20 ng g-1 w.w. (2008/105/EC). At other sites, mercury concentrations ranged from 10 - 52 ng g-1 w.w.. These fell inside the proposed range of ‘present background concentrations in pristine areas within the OSPAR Convention Area’ (10 - 50 ng g-1 w.w. in round fish, ICES, 1997). The same pattern is seen between adult female eelpout and larvae as seen for the other contaminants (fig. 6) i.e., a higher concentration in the adult female than in the larvae, with the eelpout:larvae concentration ratio varying from 1.9 - 14.8.
Figure 6. Concentration (µgg-1 w.w.) of mercury in eelpout muscle and larvae in the Baltic Sea.
C = contaminated sites; R = reference sites. See table 3 for sampling site ID.
25 The other metals (Cd, Pb, Zn, Mn, Cu, Cr, Co, Ni) were analyzed in liver and larvae (fig. 7, 8).
Nickel, chromium, manganese and copper were not analysed at the Danish sites. A relatively high abundance of zinc, manganese, nickel, chromium and copper were observed. Copper concentrations in eelpout liver ranged from 23.7 - 62.4 µg g-1 d.w., which is high compared to herring liver (10.3 – 16.4 µg g-1 d.w.) (Bignert et al., 2011). Copper is an essential metal and its concentration is regulated by homeostatic mechanisms, thus differences between species are expected. Cadmium concentrations here ranged from 0.23 – 1.84 µg g-1 d.w.. The highest cadmium concentration was observed at Kvädöfjärden (1.84 µg. g-1 d.w.), which concurs with the study about cadmium concentration in some marine biota (Bignert et al., 2011).
Concentrations are lower in samples from the Swedish west coast compared to other Baltic Sea sites. This can be explained by geographical differences between the two regions. The Baltic Sea has a high bioavailability of cadmium due to its low salinity compared to the Swedish west coast.
Cadmium concentrations in eelpout exceed the background levels according to OSPAR’s Background Assessment Criteria (ca 0.1 µg g-1 d.w.) (Hedman et al., 2011) and are also higher than the maximum allowable concentrations in food stuffs suggested by the European Comission (0.16 µg g-1 w.w.) (2000/60/EC, 2005). Lead concentrations are relatively similar between sites.
Lead in eelpout liver ranged from 0.12 - 0.28 µg g-1 d.w. (ca 30 - 80 ng g-1 w.w.), which is lower or close to the recommended limit for children’s food set by the Swedish National Food Administration (50 ng g-1 w.w.)(SLVFS, 1993).
Figure 7. Concentrations (µgg-1 d.w.) of selected metals in eelpout liver in the Baltic Sea.
C = contaminated sites; R = reference sites. Cadmium and lead are presented on the secondary y axis (µg g-1 d.w.) due to their lower concentrations meaning they were not visible otherwise. See table 3 for sampling site ID.
In larvae, zinc and manganese show considerably higher concentrations than the other metals.
All zinc concentrations were >120 µg g-1 d.w., while manganese concentrations at most sites were above 20 µg g-1 d.w. (fig. 8). Cadmium and lead concentrations are mostly below LOQ.
26
Figure 8. Concentrations (µgg-1 w.w.) of selected heavy metals in eelpout larvae in the Baltic Sea.
C = contaminated sites; R = reference sites. Mercury, cadmium and lead are presented on the secondary y axis (µg g-
1 d.w.) due to their lower concentrations meaning they were not visible otherwise. See table 3 for sampling site ID.
3.2.5 Perfluroalkyl substances
Eleven perfluoroalkyl substances (PFASs) were analyzed in larvae and liver (fig. 9). Higher concentrations are observed in larvae than liver. Fat content of liver was not measured; however, the differences in fat content may be an explanation for the differences in PFAS concentrations observed. PFOA, PFHxS and PFOSA were not detected in larvae (<1.2 ng g-1 w.w., <0.8 ng g-1 w.w., <0.5 ng g-1 w.w., respectively). PFNA and PFOS are observed in larvae and liver in high concentrations (fig. 9). A low PFOSA concentration and a high PFOS concentration were expected because PFOSA is a precursor compound of PFOS. Rüdel et al. (2011) researched patterns, levels and trends of perfluorinated compounds in aquatic organisms and bird’s eggs from the North Sea. Eelpout data here is in agreement with those results. However, while PFHxS was relatively high in herring liver (1.3 ng g-1 w.w.) (Bignert et al., 2011) , it is not detectable in eelpout liver (< 0.8 ng g-1 w.w.). PFOS found in herring liver (18.7 ng g-1 w.w.) (Bignert et al., 2011) was higher than that found in eelpout liver (12.4 ng g-1 w.w.). Differences between species are expected. Herring and eelpout feed at slightly different levels in the trophic system, therefore accumulation and bio-magnification may differ because of diet.
PFHxS concentration was lower than PFOS, because PFHxS is a by-product of technical PFOS.
Roskilde fjord has the highest total PFAS concentration in liver (ca. 30 ng g-1 w.w.), which was expected because this site is known as a point source of PFASs (Hedman et al., 2011). PFOS (29.2 ng g-1 w.w.) holds about 84% of the total PFAS concentration in this site and this was also the highest detected PFOS level in all sites. However, these values are still lower than a study on Danish eelpout by Strand et al. (2009), where PFASs are very high in liver (ca. 180 ng g-1 w.w.) and larvae (ca. 40 ng g-1 w.w.). Total concentrations of PFASs in the reference sites were similar to those of the contaminated sites.
27
Figure 9. Concentrations (ngg-1 w.w.) of selected perfluorinted substances in eelpout larvae (La) and liver (Li) in the Baltic Sea. R = reference site; C = contaminated site. See table 3 for sampling site ID.
3.2.6 Phenolic compounds
High concentrations of alkylphenolic compounds are seen at contaminated sites (fig. 10), in particular at Swedish sites. Concentrations at others sites are <10 ngg-1 w.w.. Bisphenol A and 4- nonylphenol are dominant in all samples. Bisphenol A is only detected in Swedish sites; at other sites, concentrations are <1.9 ng g-1 w.w.. While 4-nonylphenol was detected only at Wismar (Germany), concentrations at other sites are less than 10 ng g-1 w.w.; however there was an analytical error for this substance at Fjällbacka and it was therefore removed from this site in all figures and tables. Wismar and Göteborg harbor are known as areas contaminated by effluent from municipal treatment plants and industries (HELCOM, 2010). Nonylphenol concentration in eelpout is very low in comparison with mysid and shrimp from a contaminated site in the Netherlands (72-230 ng g-1 w.w.) (Verslycke et al., 2005). Trichlosan and 2,4-dibromophenol were not detected at any selected sites (<0.1 and <0.04 ng g-1w.w., respectively). 2,4,6- tribromophenol is seen at most sites, although concentrations are low. Adult female eelpout have lower concentrations of most analyzed phenolic compounds compared to larvae on a wet weight basis (fig. 10, left). This, however was not the case at Wismar in Germany, where adults had a higher concentration of phenolic compounds compared to larvae on a wet weight basis.
Eelpout:larvae concentration ratios based on lipid weight revealed two sites had high concentrations in larvae compared to adult females i.e., Eggers Wiek (0.9) and Fjällbacka (0.6).
PhCs in female eelpout adults are low at these sites.
