Environmental genotoxicity and cytotoxicity in flounder (Platichthys flesus), herring (Clupea harengus) and Atlantic cod (Gadus morhua) from chemical munitions dumping zones in the southern Baltic Sea

Accepted Manuscript Environmental genotoxicity and cytotoxicity in flounder (Platichthys flesus), herring (Clupea harengus) and Atlantic cod (Gadus morhua) from chemical munitions dumping zones in the southern Baltic Sea Janina Baršienė, Laura Butrimavičienė, Wlodzimierz Grygiel, Thomas Lang, Aleksandras Michailovas, Tomas Jackūnas PII:

S0141-1136(13)00146-3

DOI:

10.1016/j.marenvres.2013.08.012

Reference:

MERE 3784

To appear in:

Marine Environmental Research

Received Date: 26 June 2013 Revised Date:

16 August 2013

Accepted Date: 21 August 2013

Please cite this article as: Baršienė, J., Butrimavičienė, L., Grygiel, W., Lang, T., Michailovas, A., Jackūnas, T., Environmental genotoxicity and cytotoxicity in flounder (Platichthys flesus), herring (Clupea harengus) and Atlantic cod (Gadus morhua) from chemical munitions dumping zones in the southern Baltic Sea, Marine Environmental Research (2013), doi: 10.1016/j.marenvres.2013.08.012. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

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Highlights:

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Data on genotoxicity and cytotoxicity in fish from chemical munitions dumpsites Increased genotoxicity risk in southern Baltic Sea The reference, background and threshold levels of genotoxicity in fish described

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Environmental genotoxicity and cytotoxicity in flounder (Platichthys flesus), herring (Clupea harengus) and Atlantic cod (Gadus morhua) from chemical munitions dumping zones in the southern Baltic Sea

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Janina Baršien÷a, Laura Butrimavičien÷a, Wlodzimierz Grygielb, Thomas Langc, Aleksandras Michailovasa, Tomas Jackūnasa a

Nature Research Centre, Akademijos 2, 08412 Vilnius, Lithuania National Marine Fisheries Research Institute in Gdynia, 1 Kollataja Street, 81-332 Gdynia, Poland c Thünen Institute of Fisheries Ecology, Deichstraße 12, 27472 Cuxhaven, Germany

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Corresponding author – Janina Baršien÷ Nature Research Centre Institute of Ecology Akademijos str. 2 08412, Vilnius, Lithuania Tel: +370 6 8260979 Fax: +370 5 2729257 E-mail: [email protected]

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Short Running Head – Genotoxicity in fish from chemical munitions dumpsites

Abstract The data on environmental genotoxicity and cytotoxicity levels as well as on genotoxicity risk in flounder (Platichthys flesus), herring (Clupea harengus) and cod (Gadus morhua) collected in 2010-2012 at 42 stations located in chemical munitions dumping areas of the southern Baltic Sea are presented. The frequency of micronuclei, nuclear buds and nucleoplasmic bridges in erythrocytes was used as genotoxicity endpoint and the induction of fragmented-apoptotic, binucleated and 8-shaped erythrocytes as cytotoxicity endpoint. The most significantly increased geno-cytotoxicity levels were determined in fish collected near known chemical munitions dumpsites. Extremely high genotoxicity risk for flounder were identified at 21 out of 24 stations, for herring at 29 out of 31 and for cod at 5 out of 10 stations studied. The reference level of genotoxicity was not recorded at any of the stations revealing that in the sampling area fish were affected generally.

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Keywords: Genotoxicity, Cytotoxicity, Flounder, Herring, Cod, Chemical munitions, Baltic Sea

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1. Introduction

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The marine environment is increasingly contaminated with xenobiotic compounds, including hydrocarbons such as petroleum and its derivates, heavy metals, chlorinated pesticides, radionuclides and many other substances discharged from industries, agriculture or municipal activities (Korpinen et al., 2012). International attention over the last few decades has been attracted by the problem of sea-dumped chemical warfare agents (CWA), which after World Wars I and II were accumulated and dumped into the seas in hundreds of thousands tons (Kaffka, 1995). For a long time, the problem was concealed and politically sensitive, as the munitions dumping operations were secret with undefined responsibility. The full extent of munitions dumping operations has remained uncleared to date and the real picture of dumping locations is still incomplete. Consequently, there is considerable uncertainty as to ecological risks associated with CWA in the marine environment due to the lack of information on the location, amount and subsequent distribution of chemical warfare agents (Missiaen et al., 2010) as well as their biological effects on marine organisms. About thirteen thousand of tons of chemical warfare agents have been dumped at different sites of the Baltic Sea, mainly in the Bornholm and Gotland basins. The main CWAs were sulphur mustard (63%), arsenic-containing compounds such as Clark and arsine oils, Adamsite (31%) and α-chloroacetophenone (5%) (HELCOM, 1994). Due to unvalued, imperfect and indecorous dumping operations and later relocation by bottom fishing gears, warfare agents appeared in over a much wider territory than the one registered in operation documents (HELCOM, 1993; Astot et al., 2007). Chemical munitions east of the Bornholm Island were dumped in a considerably large area. Warfare agents were mainly dumped in munitions, mostly in bombs, shells and containers. Munitions were also packed into wooden boxes, which could drift before sinking. Later these wooden boxes with warfare agents were reported to have drifted to Bornholm or the Swedish coastline (HELCOM, 1996). Munitions that have been retrieved since 1992 proved to be heavily corroded or empty (Sanderson and Fauser, 2008). Studies of the Bornholm chemical munitions dumpsite showed that contamination of bottom sediments with CWA was patchy and was recorded up to about 100 meters away from the shipwrecks (Missiaen et al., 2010). Several authors reported significant concentrations of As in sediments from dumping sites, suggesting that leakage of CWAs contributed to arsenic contamination (Munro et al., 1999; Tørnes et al., 2002; Garnaga and Stankevičius, 2005). Mustard and organoarsenic warfare agents were also detected in sediments from CWA dumping areas located in the North Sea (Granbom, 1996; Tørnes et al., 2006) and the Southern Adriatic Sea (Amato et al., 2006). Attributed to the leakage of yperite from rusted bomb shells dumped in the Southern Adriatic Sea, 3-4 times higher levels of As and Hg were recorded in fillets of two benthic fish species and genotoxicity effects of CWA were described. The suitability of fish for the environmental quality assessment was emphasised (Della Torre et al., 2010, 2013). The main objective of the present study was to evaluate the level of environmental genotoxicity and cytotoxicity in three native fish species inhabiting the chemical munitions dumping zones located in the Bornholm Basin of the Baltic Sea. Furthermore, environmental genotoxicity risk was assessed at each of the study stations. This basin is regarded as the most important spawning area for commercially important fish species in the Baltic Sea (Köster et al., 2003). A number of studies have indicated that fish respond to low concentrations of genotoxic substances that may be present in chronically polluted areas, and, therefore fishes are regarded as appropriate environmental genotoxicity bioindicators (Baršien÷ et al., 2006; Schiedek et al., 2006; Kopecka et al., 2006; Yadav and Trivedi, 2009; Nahrgang et al., 2010). Furthermore,

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2. MATERIALS AND METHODS 2.1. Sampling of fish

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considering a continuous intake of contaminants through water (respiration) and food, fish may be particularly sensitive to arsenic compounds released to the environment (Bears et al., 2006). In the present study, the induction of micronuclei (MN), nuclear buds (NB), nuclear buds on filament (NBf) and bi-nucleated cells with nucleoplasmic bridges (BNb) were used as genotoxicity endpoints and the induction of fragmented-apoptotic (FA), bi-nucleated (BN) and 8shaped nucleus cells were assessed as cytotoxicity endpoints in erythrocytes of three fish species. The micronucleus (MN) test is a sensitive and fast approach to detect structural and numerical chromosomal alterations induced by clastogenic and aneugenic agents (Heddle et al., 1991). The formation of nuclear buds may reflect an unequal capacity of organisms to expel damaged, amplified, failed replication or condensed improperly DNA, chromosome fragments without telomeres and centromeres from the nucleus (Lindberg et al., 2007). Induction of bi-nucleated cells with nucleoplasmic bridges has been used as a marker of dicentric chromosomes, and thereby served as an index of cytogenetic damage in fish (Summak et al., 2010). The investigated cytotoxicity endpoint, fragmented-apoptotic erythrocytes, is a form of regulated cell death. A higher rate of cytokinesis failure can result in an increase of bi-nucleated and 8-shaped nucleus cells.

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For the environmental genotoxicity and cytotoxicity assessment, flounder (Platichthys flesus), herring (Clupea harengus) and cod (Gadus morhua) specimens were collected at 42 study stations located in the Bornholm Basin from November 2010 to February 2012 (Fig. 1). Samples were obtained from research catches carried out by the RVs “Baltica” and “Walther Herwig III” using standard bottom or pelagic trawls. The list of study stations, geographic coordinates of trawling stations, depth of trawling (m) and hydrological parameters such as water temperature (°C), salinity and O2 concentration (mg/l) are presented in Table 1. The dates of sampling surveys and numbers of collected fish specimens are presented in Table 2.

Table 1 Table 2

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Fig. 1.

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Blood samples were collected from 188 flounder (23 stations), 307 herring (31 stations) and 127 cod specimens (10 stations). Most study stations for flounder and herring were located along the chemical munitions transportation (to the Bornholm and Gotland dumpsites) route in the Polish EEZ. Study stations 17a, 19a, 21a, 23, 16b and 45d were located the closest to the known CW dumpsite southeast of Bornholm, while stations 28, 28a and 26d the closest to the warfare dumping sites east of Bornholm. Geno-cytotoxicity responses in these fish species were analysed by comparing them with those in fishes from stations B03 and B05 visited in 2003, 2004 and 2009 (Fig. 1). The main Bornholm CW dumping area studied (station B13/06) was situated in the primary chemical munitions dumping zone. The other stations (B13/09, B13/11 and Z24) were located in the secondary zone outside the primary dumping zone. Stations B13/07, B13/08 and B13/12 were 3

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situated close to the primary dumping zone, stations B13/10 and Z12 close to the secondary dumping zone. Cod samples were collected from reference stations B09/17, B09/18 and B09/19, which were located in the Gdansk Deep (Fig. 1). 2.2. Sample preparation and analysis

Fig. 2.

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A drop of blood taken from the caudal vein of fish was directly smeared on glass slides and air-dried. Smears were fixed in methanol for 10 min and were stained with 5% Giemsa solution in phosphate buffer pH=6.8 for 8 min. The stained slides were analyzed under light microscopes Olympus BX51 or Nikon Eclipse 50i at final magnification of 1,000×. Blind scoring of micronuclei and other nuclear abnormalities was performed on coded slides. Micronuclei (MN) were identified according to the following criteria: (1) round and ovoid-shaped non-refractory particles in the cytoplasm, (2) colour and structure similar to chromatin, (3) diameter of 1/3-1/20 of the main nucleus, (4) particles completely separated from the main nucleus (Heddle et al., 1991). Nuclear buds, bi-nucleated and fragmented-apoptotic cells were identified using criteria described by Fenech et al. (2003). Nuclear buds (NB) and nuclear buds on filament (NBf) were characterized as extruded nuclear material conjugated to the main nucleus by nucleoplasmic connection. Nucleoplasmic bridges in bi-nucleated cells (BNb) are described as a continuous nucleoplasmic link between two nuclei with the same staining and focusing pattern as the nuclei and the width, which may vary but does not exceed one-fourth of the diameter of the nuclei. Fragmented-apoptotic cells (FA) in early stages were identified by the presence of chromatin condensation within the nucleus and intact cytoplasmic and nuclear boundaries, as late apoptotic cells exhibit nuclear fragmentation into smaller nuclear bodies within an intact cytoplasm/cytoplasmic membrane. Nuclei of bi-nucleated cells (BN) should not overlap, be approximately equal in size, staining pattern and staining intensity, have intact nuclear membranes and be located within the same cytoplasm membrane. Eight-shaped erythrocytes were evaluated also as cytotoxicity markers. The morphological features of the studied nuclear abnormalities are shown in Fig. 2. For each studied specimen of fish, 4000 intact erythrocytes were analyzed. Final results were expressed as the mean value (‰) of sums of the analyzed individual lesions scored in 1,000 cells per fish collected from every study station (Baršien÷ et al., 2004).

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2.3. Environmental genotoxicity risk assessment in fish Assessment of genotoxicity risk at each of the 42 studied stations was done on the basis of the calculated background response of the total genotoxicity level in flounder (<0.40 MN+NB+BNb/1000 erythrocytes), in herring (<0.85 MN+NB+BNb/1000 erythrocytes) and in cod (<0.55 MN+NB+BNb/1000 erythrocytes) specimens. The background level of total genotoxicity frequencies was calculated in the same way as that for micronuclei (Baršien÷ et al., 2012a), following the methodology presented in the ICES background document for genotoxicity assessment in marine organisms (Baršien÷ et al., 2012b). The empirical 90% percentile (P90) was determined in flounder, herring and cod specimens collected in the 2001-2010 period from the reference sites Kvädöfjärden, Palanga, Leba, Pärnu, 1a-1, 2a-1 and 2b-1 that are characterized by no known local sources of contamination and no impact from urban or industrial activity. The 4

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P90 value separates the upper 10% of all the genotoxicity values in the group of data from the lower 90%. In general, an elevated genotoxicity frequency lies above the P90 percentile, whereas the majority of values below the P90 value belong to individuals that are unexposed, weaklyexposed, non-responding or adapted to stressful conditions. Additionally, the 99% percentile was calculated as a threshold level of genotoxicity in studied fish species. At each of the study stations, reference, background and threshold levels in studied fish species were mapped in the GIS system. The study stations were grouped as risk groups according to genotoxicity levels, using a 4-grade scale, i.e. low (reference), increased (≥ P90), high (≥P99) and extremely high risk (>P99) levels. 2.4. Statistical analysis

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The statistical analysis was carried out using the GraphPad PRISM 5.0 statistical package. The mean and standard errors were calculated for each studied group of fish. Spearman's correlation and regression analysis (linear model Y=a+Bx) was performed to illustrate possible relations between geno-cytotoxicity biomarker responses in fish, and environmental variables, or biometrical measurements in fish from different study stations of the Bornholm region. 3. Results

3.1. Environmental genotoxicity and cytotoxicity levels in flounder

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In peripheral blood erythrocytes of flounder, the genotoxicity levels assessed as a sum of the frequency of micronuclei (MN), nuclear buds (NB+NBf) and bi-nucleated erythrocytes with nucleoplasmic bridges (BNb) varied from 0.40‰ in fish from station Z10 to 3.13‰ in fish from station 18a (Fig. 3). Genotoxicity response values above 2‰ were found in flounder from stations 24 (2.98‰), 28a (2.91‰), 22 (2.57‰), 21a (2.49‰), 19 (2.46‰), 23 (2.29‰), 25 (2.22‰) and Z10 (3.13‰). In flounder from these stations there was a markedly high frequency of nuclear buds. In general, genotoxicity levels were significantly higher in flounder specimens caught in November 2010 (stations 19-28) and in February 2011 (stations 15a-28a) than in fish collected later in June and November 2011 as well as in February 2012 (Fig. 3). Environmental cytotoxicity levels were assessed as the sum of the frequency of fragmentedapoptotic (FA), bi-nucleated (BN) and 8-shaped erythrocytes. Cytotoxicity responses in flounder specimens varied between 0.25‰ (stations 18b and 35d) and 1.01‰ (station 23). The latter station was situated close to the warfare dumping area. At most of the stations there cytotoxicity levels were found to be lower compared to genotoxicity levels. Overall, the levels of genotoxicity and cytotoxicity in flounder specimens collected in 2010-2012 in most cases were higher (especially in fish collected in November 2010 and February 2011) than in those sampled from closely located sites B03 and B05 in December 2003 (Fig. 3). Comparatively higher genotoxicity and cytotoxicity levels were detected in flounder specimens from stations 28 and 28a located at a comparatively close distance to warfare dumping sites in the Polish waters of the Baltic Sea. The lowest levels of the studied cellular alterations were recorded in fish from stations 18b and Z10. Both stations are located outside Bornholm CW dumping sites (Fig. 3).

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3.2. Environmental genotoxicity and cytotoxicity levels in herring

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The measurement of environmental genotoxicity in herring blood erythrocytes revealed exceptionally high levels in specimens collected at stations 16b (4.95‰), 15a (4.40‰), 33d (4.25‰), 45d (3.18‰), 17a (3.10‰) and Z24 (2.88‰) (Fig. 4). Single specimens from stations 16b (30.0 MN/1000 erythrocytes), 33d (20.0 MN/1000 erythrocytes) and Z10 (9.76‰) showed a strongly deviating frequency of micronuclei. Station 16b is located close to the known CW dumping site in the Polish EEZ. At each of the studied stations 15a, 17a, 45d and Z24, there were 2-3 specimens of herring caught with a very high level of nuclear buds formation (up to 10‰) and a high percentage (up to 80%) of erythrocyte nuclei with multiple morphological abnormalities (Fig. 5). Station Z24 is located in the secondary CW dumping area, stations 17a and 45d are located close to the known warfare dumpsite in the Polish EEZ. In general, the highest genotoxicity levels in herring were detected at stations close to the CW dumping site (stations 23, 17a, 19a, 21a, 16b and 45d). In contrast, low genotoxicity responses were found in herring from station 19 (0.75‰) located outside the CW dumpsite. Environmental cytotoxicity levels in herring specimens sampled at 31 study stations in the Bornholm region varied from 0.19‰ (station 19) to 0.89‰ (station 15a). Cytotoxicity levels were 2-12 times lower than those of genotoxicity. Levels of genotoxicity and cytotoxicity in herring specimens collected in 2010-2012 were much higher than those recorded in December 2009 at the closely located site B03/46 (Fig. 4). Fig. 4. Fig. 5.

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3.3. Environmental genotoxicity and cytotoxicity levels in cod

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The total genotoxicity (MN+NB+NBf+BNb) levels in blood erythrocytes of cod specimens varied from 0.41‰ (station B13/09) to 0.98‰ (station B13/06 located at the primary dumping site of chemical munitions). A low induction of genotoxicity responses (0.52‰, 0.50‰ and 0.44‰) was recorded in fish from the suspected reference stations B09/17, B09/18 and B09/19. The highest genotoxicity responses were recorded in cod specimens collected from stations B13/06, B13/12, B13/11, B13/07 and B13/10. Comparatively high frequencies of cytotoxicity parameters (FA+BN+8-shaped erythrocytes) were found in fish from stations B13/11, B13/06 and B13/07. The response detected in fish from the reference station B09/18 was three times lower. In general, the levels of genotoxicity in cod collected in December 2011 were 1.5-2 times higher than those recorded at the closely located site B03 in September 2004 (Fig. 6). Fig. 6.

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3.4. Relationships between environmental geno-cytotoxicity levels and environmental and biological variables in fish The Spearman correlation analysis of genotoxicity levels and six biometrical variables (fish age, total length, total weight, and liver weight, liver-somatic index, (LSI) and condition factor (CF)) in flounder showed a positive correlation at two out of the 23 stations investigated. In herring, correlations were found at 8 out of 31 study stations. The most significant relationships 6

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were detected in flounder from station 19a, where a very strong positive correlation was found between genotoxicity and liver weight (rs = 0.890), fish weight (rs = 0.816); strong correlation was between genotoxicity and length (rs = 0.787), age (rs = 0.736) and CF (rs = 693). In herring from station 17a, there was a very strong positive correlation recorded between genotoxicity response and liver weight (rs = 0.890), CF (rs = 0.889), a strong relationships with fish age (rs = 0.790), weight (rs = 0.716) and length (rs = 0.686). In herring caught at station 18b, a positive correlation between fish age (rs = 0.830), length (rs = 0.732) and weight (rs = 0.718) was found. Analysis in cod showed mainly positive correlation between genotoxicity and fish length (rs = 0.747), weight (rs = 0.743) and liver weight (rs = 0.614) in samples from station B13/06. Genotoxicity response was also correlated to fish weight (rs = 0.631) and spleen weight (rs = 0.607) of cod sampled from station B13/07. Very strong correlation between cytotoxicity responses and biological variables was found only in those specimens of flounder that were collected from stations 16b (fish age, rs = 0.898; weight, rs = 0.883; length, rs = 0.836), 17b (fish length and weight, rs = 0.999) and 14b (CF, rs = 0.949). Strong relationships were estimated in flounder specimens from stations 19a (CF, rs = 0.790) and 23b (age, rs = 0.676). In herring, very strong positive correlation with fish biological parameters was found mainly in specimens collected from stations 17a (CF, rs = 0.900; length, rs = 0.835; weight, rs = 0.830), 15a (fish length, rs = 0.855; liver weight rs = 0.848, fish weight rs = 0.838, age, rs = 0.836), 36d (fish length, rs = 0.858) and 23a (CF rs = 0.874). Strong correlation was recorded between cytotoxicity and herring age (rs = 0.747 – station 18b; rs = 0.765 – station 36d) and fish weight (rs = 0.753 – station 23a; rs = 0.793 – 36d). The performed regression analysis of genotoxicity, as well as cytotoxicity responses and environmental variables (water temperature, salinity, oxygen concentration and depth of sampling) revealed only a few correlations in flounder and herring samples collected in November 2010 and 2011. Seawater temperature was found to be the determining factor in the formation of genotoxicity in herring (November 2011; R2 = 0.8881) and flounder (November 2010; R2 = 0.6967). Oxygen content in water showed a statistically significant correlation with genotoxicity in flounder (November 2010; R2 = 0.7543); bottom depth had the greatest influence on herring (November 2010; R2 = 0.6316), and salinity was relevant both to flounder (November 2011; R2 = 0.7498) and herring (November 2011; R2 = 0.5152). Cytotoxicity levels in flounder were significantly correlated only with salinity (R2 = 0.6757) and bottom depth in November 2010; a lower correlation was found with salinity (R2 = 0.5724) and depth (R2 = 0.4728) in November 2011. There was no correlation between geno-and cytotoxicity levels and environmental variables in fish collected in February 2011 and 2012.

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3.5. Environmental genotoxicity risk assessment in fish collected from the Bornholm Basin The results of the analysis revealed a background level of genotoxicity at all stations sampled in 2003, 2004 and 2009 and also at stations B09/17, B09/18, B09/19 (2011), as well as in herring caught at stations 19 and 26d, in cod from stations B13/08 and B13/09 and in flounder from station Z10. A high genotoxicity risk level was detected only in flounder collected from stations 14b and 18b. All other responses in the three fish species studied indicated an extremely high genotoxicity risk level. At none of the stations studied, the reference level of genotoxicity was reached (Fig. 7). Fig. 7.

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4. Discussion

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The present paper, to our knowledge, represents the first attempt to evaluate genotoxicity and cytotoxicity responses in flounder, herring and cod from study stations located at or in the vicinity of dumpsites of chemical munitions in the southern Baltic Sea, as well as the first endeavour to detect the background and threshold level of environmental genotoxicity. The assessments of genotoxicity risk at each of the 42 stations studied revealed effects in all the studied fish species. Zones of extremely high genotoxicity risk for flounder were found at 21 out of 24 stations, for herring at 29 out of 31 and for cod at 5 out of 10 stations. A high genotoxicity risk was determined only in flounder from two stations. It should be pointed out that the background level of genotoxicity responses was detected in fish collected from all reference stations, as well as in flounder from one study station, and in herring and cod caught at two study stations. The reference level reflecting a low genotoxicity risk was exceeded in all the fish species investigated and at all the stations studied. It can be concluded from these findings that the study area is characterised by a high genotoxicity risk, most likely associated with the presence and effects of genotoxic agents. Since there is an indication that the highest risk is at or close to the dumpsites of chemical munitions, it cannot be excluded that the major causative agents affecting not only the primary dumpsites, but also neighbouring areas are chemical warfare agents. According to HELCOM (1994), the Bornholm Basin chemical munition dumpsite covers a large area, where about 11,000 tons of toxic chemical warfare agents were dumped east of the island of Bornholm. The main types of warfare agents dumped in the Bornholm Basin include sulfur mustard, phosgene, α-chloroacetophenone as well as arsenic-containing compounds such as Adamsite (diphenylaminechloroarsine, DM), Clark I (diphenylarsine chloride, DA), Clark II (diphenylarsine cyanide, DC) and arsine oil (HELCOM, 1993; Politz, 1994). Arsine oil is a technical mixture of Clark I (35%), phenyldichloroarsine (PDCA; 50%), trichloroarsine (TCA; 5%) and triphenylarsine (TPA; 5%). In water, most of the released CWAs can be degraded during hydrolytic processes, while sulfur mustard and arsenic-containing chemicals are normally degraded through oxidation. Sulfur mustard degrades into thiodiglycol (TDG) (Munro et al., 1999), which can oxidize into thiodiglycol sulfoxide (TDG[ox]) (Black et al. 1992). The main degradation products of Clark I, Clark II are diphenylarsinous acid (DPA), phenylarsinic oxide, triphenylarsine (TPA) and triphenylarsine oxide (Hanaoka et al., 2005). Adamsite degrades in a similar manner as Clark I and Clark II, producing both hydrolyzed and oxidized products (Haas et al., 1998). Many of the dumped chemical warfare agents also contain additives, such as chlorobenzene (CB). CB is an environmentally hazardous compound (HELCOM, 1994) which was added to sulfur mustard and tabun (Franke, 1977). Likewise, it is a precursor of Clark I (Lohs, 1967). This compound was found in sediments of the Bornholm dumpsite (Missiaen et al., 2010). Chlorobenzene and TPA are stable components, and are, thus, highly resistant to hydrolysis and oxidation in the environment (Munro et al., 1999). Hydrolysis products of Clark I and Clark II are reported to have the same toxicity as their parent chemicals, posing a long-lasting threat to the marine environment (Francken and Hafez, 2009). It is of particular interest that the analysis of sediments and near-bottom water samples, collected within the Bornholm dumpsite and from the surrounding area in 2007 and 2008, has shown that CWA degradation products were detected in 56 of the 65 analyzed sediment core samples. Moreover, the highest it’s concentrations in pore water were found in the same samples in which the highest sediment concentrations were detected (Missiaen et al., 2010). The

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analytical study revealed that concentrations of different CWAs and their degradation products were distributed over a larger area than that of the main Bornholm CWAs dumpsite. Degradation products of Adamsite (DM) and phenyldichloroarsine (PDCA), i.e. oxidized forms of DM[ox] and PDCA[ox], were found at increasing concentrations outside the primary Bornholm dumpsite. In general, the highest concentrations of Clark (I and II) degradation products, propyl diphenylarsinothioite (DPA[SPr]) and it’s oxidation product diphenylarsinic acid (DPA[ox]), were detected within the primary dumpsite as well as the concentrations of chlorobenzene (CB) and triphenylarsine (TPA). Nevertheless, an evident spread of these toxic agents within the secondary dumpsite was observed (Missiaen et al., 2010). Active parental CWA (Clark I, Adamsite) and arsenic oil consisting of TPA and TCA were detected in sediments and porewater at CWA monitoring stations along the Nord Stream pipeline route in the Bornholm Deep in 2008 and 2010. It should be stressed that there were elevated geno-cytotoxicity levels in herring specimens collected from stations Z13 and 34d, i.e. the zone where the highest concentrations of CWAs (Clark I, Adamsite and arsenic oil constituent compounds) were detected in sediments (Sanderson et al., 2012). Arsenic compounds were identified in samples collected close to shipwrecks with dumped munitions and in samples collected from sites situated at a longer distance from wrecks in the Skagerrak. These compounds are very stable and can be found a long time after they are released from the ammunition (Tørnes et al., 2002). Marine organisms can accumulate arsenic compounds, and further metabolic transformation to complex methylated compounds has been observed (Sanders et al., 1989; Francesconi and Edmonds, 1994). Complex methylated arsenic compounds are able to accumulate in aquatic organisms via the food web, for example, in algae up to 1000-10,000 fold higher compared to their concentrations in sea water (Borak and Hosgood, 2007). Arsenobetaine, a tri-methylated pentovalent (As[V]) compound, is the predominant arsenic compound in marine organisms (Edmonds et al. 1997) and can have great toxicity (Styblo et al., 2000). Amato et al. (2006) found a total of 29.7mg of As/kg fish (dw) in specimens from a CWA dumpsite in the southern Adriatic Sea, whereas total As levels detected in fish under reference conditions were more than 10-fold lower (1.9 mg/kg (dw)). Among fishes from the Danish waters, including the Bornholm Basin, cod showed the highest total As levels with a maximum of 11.5 mg As/kg (ww). The majority of total As in fish was in the form of arsenobetaine (35–100%) (Sanderson et al., 2009). Maher et al. (1999) showed that most of the organs and blood in sea mullet (Mugil cephalus) contained a large percentage (35–100%) of arsenobetaine (AsB), with the concentration of total As ranging from 0.54 µg/g dry mass in gills to 19.2 µg/g dry mass in liver. Ventura-Lima et al. (2009) observed higher accumulation of As in the gills than in the liver of the common carp (ranging from 0.30 to 10.8 and 0.19 to 0.73 µg/g, respectively). Fattorini and Regoli (2004) reported that polychaete Sabella spallanzanii accumulated mostly dimethylated forms of As, while estuarine polychaete Laeonereis acuta accumulated not only mono and/or dimethylated forms, but also arsenobetaine (AsB) and arsenocholine (AsC) (Ventura-Lima et al., 2007). It was observed that although arsenic accumulates primarily in retina, it may accumulate in liver and kidney of fish as well (Ghosh et al., 2007). The research results obtained by Della Torre et al. (2010) confirmed the data indicating arsenic release from munitions and accumulation in fish tissues. Literature data on degradation products of chemical warfare agents and their genotoxicity in aquatic organisms are still limited. However, it is important to emphasize that dumped chemical warfare substances are leaking and harmful effects on aquatic organisms are suspected. Della Torre et al. (2010) performed an ecotoxicological study using two benthic species, blackbelly

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rosefish (Helicolenus dactylopterus) and European conger (Conger conger), as sentinel species for the CW risk assessment at a dumpsite in the Adriatic Sea. Significantly higher levels of As and Hg were detected in muscles of both species from the dumping site than in muscles of those from the reference site. DNA damage (quantified as 28.77% and 28.51% DNA migrated in the tail) was found only in gills of conger fished at CW dumping site (Della Torre et al., 2010, 2013). Experimental treatment with different concentrations of NaAsO2 evoked elevated micronuclei induction in blood erythrocytes of fish Oreochromis mossambicus after 96 and 192 h. The fourday exposure of fish to 3 ppm of NaAsO2 increased the MN frequency up to 2.2%, exposure to 28 ppm – to 4.4% and the treatment of fish with 56 ppm of NaAsO2 increased the MN frequency to 5.8% (Ahmed et al., 2011). Arsenicals increased the frequency of MN in gill cells of zebra fish (Danio rerio) even 56-fold compared to control levels (Oliveria and Francisco, 2005). Yadov and Trivedi (2009) reported an increased MN induction in fish after 96h treatment with As2O3 and decreased MN levels after 192 hours. We also detected increased levels of cytotoxicity in fish collected from the study stations located close to the known CWA dumpsites in the Bornholm Basin of the Baltic Sea. Diphenylarsinous acid, a degradation product of Clark II or Clark I, has been investigated in terms of its capacity to induce cytotoxic effects, numerical and structural changes of chromosomes, and abnormalities of centrosome integrity and spindle organizations in conjunction with the effects of glutathione (GSH) depletion in mammalians (Ochi et al., 2004). Increases in cytotoxicity in fish cell lines after exposure to sodium arsenide was described by Wang et al. 2004 and Seok et al. 2007. Akter et al. (2009) reported the death of fish cells involving fragmentation of chromosomal DNA due to arsenic toxic effects. The study performed by Ahmed et al. (2008) also showed the apoptotic cell death of Channa punctatus due to arsenic toxicity. On the other hand, chronic effects of other hazardous compounds on geno-cytotoxicity in fish cannot be ruled out. Elevated concentrations of PAH and PCB were detected in the Bornholm Deep after an extensive water inflow from the Odra River (Kowalewska et al., 2003). In the area where our study stations 18, 19 and 21 are located, elevated metal concentrations were measured in mussels and sediments in summer 2006 (Hendožko et al., 2010). Investigations of PCB concentrations and their profiles in Baltic fish (during the period 1997-2006) inhabiting the Kołobrzeg–Darłowo area have shown higher concentrations of PCBs (congeners 101, 118, 153, 138 and 180), which were determined in sprat and in cod (Szlinder-Richert et al., 2009). It must be emphasized that most of our sampling stations were located in the same area. According to Sundqvist et al. (2009), the concentration of polychlorinated dibenzofurans (PCDFs) in surface sediments from the south of Sweden coast (collected during 2003-2007) was in the range 5002500 pg g-1 (dw). Increases in environmental genotoxicity in relation to pollution levels were described in fish collected from the Wismar Bay (Schiedek et al., 2006) and Lithuanian coast of the Baltic Sea (Baršien÷ et al., 2006). Higher levels of geno-cytotoxicity in the Ekofisk oil extraction field were described in flounder and cod by Rybakovas et al. (2009). The linear correlation between MN induction and tissue concentrations of 16 PAHs was detected in mussels caged in pollution gradient at Ekofisk oil platform in the North Sea (Sundt et al., 2011). Statistical data analyses carried out in the present study revealed only weak relationships between geno-cytotoxicity responses in fish and the environmental variables measured. There were only a few correlations between geno-cytotoxicity responses and environmental variables (water temperature, salinity, oxygen concentration and depth of sampling) recorded in flounder and herring samples collected in November 2010 and 2011. It should be noted that although the investigated fish at certain study stations were caught in deep Bornholm Basin layers

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characterized by hypoxic conditions, very strong relationships between genotoxicity responses and oxygen content in water were found only in flounder fished in November 2010. A relationship between genotoxicity and hypoxia-related conditions was pointed out by Kreisberg et al. (in press). Hypoxia is supposed to increase water flow through fish gills and, consequently, an uptake of chemicals that are present in it. It has also been shown that hypoxia affects gene expression in the model fish species Japanese medaka (Oryzias latipes) (Zhang et al., 2009). The Spearman correlation analysis revealed a strong positive correlation of six biometrical variables (fish age, total length, total weight, and liver weight, liver-somatic index (LSI) and condition factor (CF) with geno-cytotoxicity levels in flounder at two out of 23 stations, in herring at 8 out of 31 stations and in cod at 2 out of 10 study stations. Since a strong correlation was found between geno-cytotoxicity and liver weight and CF in fish collected from the study stations located the closest to the known CWAs dumping sites, an impact of released CWAs on fish health could be suspected. In summary of the present study results, it can be stated that in general, the highest genotoxicity and cytotoxicity levels and ultimate genotoxicity risk for the studied fish species was found at stations closely located to the known CWAs dumping sites. Extremely high genocytotoxicity levels, exceeding 3.5‰, were determined in flounder from stations 18a, 24 and 28a and in herring from stations 15a, 17a and 45d. All the three stations are situated the closest to the warfare dumping sites in Polish waters. Comparatively high genotoxicity levels were found in fish collected from station Z24, located in the secondary CW dumping area and almost at all stations (23, 17a, 19a, 21a, 16b and 45d) located close to the southeast Bornholm CWAs dumping site. From the results of the present study, it can be concluded that the application of pooled genotoxicity and cytotoxicity endpoints in native fish species can provide information regarding marine organisms’ exposure to genotoxic and cytototoxic compounds in situ and can, thus, be of great value for the assessment of the environmental status of chemical munitions dumping sites. The pooled geno-cytotoxicity incidences in flounder, herring and cod from the Baltic Sea as well as genotoxicity risk levels were assessed for the first time. Significantly increased genocytotoxicity levels in fish collected from primary and secondary chemical munitions dumpsites and from the other study stations located the closest to the known CWAs in the Polish EEZ confirmed the distribution of genotoxic and cytotoxic agents and the existence of extremely high genotoxicity risk in the Bornholm Basin environment. The comparatively low influence of environmental and biological factors estimated in the present study on geno-cytotoxicity responses in the studied fish species most likely indicates that the source of the elevated genocytotoxicity induction is environmental pollution. An increased distribution of contamination during the implementation of pan-Baltic industrial projects (the Nord Stream pipelines and cable St. Petersburg-Kaliningrad construction) and mine clearance program performed in 2008-2011 (Möller, 2011) can be suspected. Data of our long-term investigations have provided evidence that the emergence of genotoxicity effects was concurrent with the above mentioned pan-Baltic interventions. Therefore, there is an evident need to monitor biological impacts of CWAs on marine organisms and to highlight its ecological importance in dumped chemical munitions areas. Considering the danger from dumped CWAs, laboratory-controlled studies using environmentally realistic doses of genotoxic compounds should help to describe cause-effects relationships between concentrations of CWAs in fish tissues and genotoxicity and cytotoxicity parameters.

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5. Conclusions

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The comparison of the present study findings covering the period 2010-2012 with those for the period 2003-2009 reveals that environmental geno-cytotoxicity levels and the associated genotoxicity risk in the three fish species studied from almost all the stations has increased. The highest induction of nuclear abnormalities and, thus, an extremely high genotoxicity risk was detected in fish collected from the main CWA dumpsite east of Bornholm, as well as at stations located close to the known warfare dumping sites in Polish waters of the Bornholm Basin. Genotoxicity levels were found to be lower at stations located further away from dumping sites, but these stations were still characterised by a high genotoxicity risk. The comparatively weak influence of environmental and biological factors on the geno-cytotoxicity responses detected in the investigated fish species indicates that the most likely source of the observed elevated genocytotoxicity induction is environmental pollution. Furthermore, the study results implicitly indicate an increased leakage of chemicals from the corroded munitions, increasing the already considerable pollution of the Baltic Sea ecosystem with geno- and cytotoxic anthropogenic contaminants. Conflict of interests None Acknowledgements

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This study was funded by the Research Council of Lithuania through the GENOTOX-CG project MIP-33/2012. We are thankful to Aleksandras Rybakovas for the collection of cod samples and their partial analysis and to Nicolai Felix Fricke for the provided biological data on the cod. This publication has been produced with the assistance of the EU, BSR Programme, CHEMSEA project. The content of this publication is the sole responsibility of its authors and can in no way be taken to reflect the views of the European Union.

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eelpout (Zoarces viviparus) from the southwestern Baltic Sea. Mar. Poll. Bull. 53, 387– 405. 53. Seok, S.H., Baek, M.W., Lee, H.Y., Kim, D.J., Na, Y.R., Noh, K.J., Park, S.H., Lee, H.K., Lee, B.H., Ryu, D.Y., Park, J.H., 2007. Arsenite-induced apoptosis is prevented by antioxidants in zebrafish liver cell line. Toxicol. In Vitro 21, 870–877. 54. Styblo, M., del Razo, L.M., Vega, L., Germolec, D.R., LeCluyse, E.L., Hamilton, G.A., Reed, W., Wang, C., Cullen, W.R., Thomas, D.J., 2000. Comparative toxicity of trivalent and pentavalent inorganic and methylated arsenicals in rat and human cells. Arch. Toxicol. 74, 289–299. 55. Summak S., Aydemir N.C., Vatan O., Yilmaz D., Zorlu T., Bilaloglu R. 2010. Evaluation of genotoxicity from Nilufer Stream (Bursa/Turkey) water using piscine micronucleus test. Food Chem. Toxicol. 48: 2443-2447. 56. Sundqvist, K.L., Tysklin, M., Cato, I., Bignert, A., Wiberg, K., 2009. Levels and homologue profiles of PCDD/Fs in sediments along the Swedish coast of the Baltic Sea. Environ. Sci. Pollut. Res. 16, 396–409. 57. Sundt, R.C., Pampanin, D.M., Grung, M., Baršien÷, J., Ruus, A., 2011. PAH body burden and biomarker responses in mussels (Mytilus edulis) exposed to produced water from a North Sea oil field: Laboratory and field assessments. Mar. Pollut. Bull. 62, 1498–1505. 58. Szlinder-Richert, J., Barska, I., Mazerski, J., Usydus, Z., 2009. PCBs in fish from the southern Baltic Sea: Levels, bioaccumulation features, and temporal trends during the period from 1997 to 2006. Mar. Poll. Bull. 58, 85–92. 59. Tørnes, J.A., Voie, Ø.A., Ljønes, M., Opstad, A.M., Bjerkeseth, L.H., Hussain, F., 2002. Investigation and risk assessment of ships loaded with chemical ammunition scuttled in Skagerrak TA-1907/2002. 60. Tørnes, J.A., Opstad, A.M., Johnsen, B.A., 2006. Determination of organoarsenic warfare agents in sediment samples from Skagerrak by gas chromatography-mass spectrometry. Sci. Total Environ. 356, 235–246. 61. Ventura-Lima, J., Sandrini, J.Z., Ferreira-Cravo, M., Piedras, F.R., Moraes, T.B., Fattorini, D., Notti, A., Regoli, F., Geracitano, L.A., Marins, L.F., Monserrat, J.M., 2007. Toxicological responses in Laeonereis acuta (Annelida, Polychaeta) after arsenic exposure. Environ. Int. 33, 559–564. 62. Ventura-Lima, J., Fattorini, D., Regoli, F., Monserrat, J.M., 2009. Effects of different inorganic arsenic species in Cyprinus carpio (Cyprinidae) tissues after short-time exposure: bioaccumulation, biotransformation and biological responses. Environ. Poll. 157, 3479–3484. 63. Wang, Y.C., Chaung, R.T., Tung, L.C., 2004. Comparison of the cytotoxicity induced by different exposure to sodium arsenite in two fish cell lines. Aquat. Toxicol. 69, 67–69. 64. Yadav, K.K., Trivedi, S.P., 2009. Sublethal exposure of heavy metals induces micronuclei in fish, Channa punctata. Chemosphere 77, 1495–1500. 65. Zhang, Z., Ju, Z., Wells, M.C., Walter, R.B., 2009. Genomic approaches in the identification of hypoxia biomarkers in model fish species. J. Exp. Mar. Biol. Ecol. 38, S180-S187.

AC C

703 704 705 706 707 708 709 710 711 712 713 714 715 716 717 718 719 720 721 722 723 724 725 726 727 728 729 730 731 732 733 734 735 736 737 738 739 740 741 742 743 744 745 746 747 748 749

16

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Fig. 6. Genotoxicity (MN+NB+BNb) and cytotoxicity (FA+BN+8) levels in cod collected respectively from the Bornholm CW dumpsite (December 2011) and from the reference station B03 (September 2004).

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Fig. 7. Results of environmental genotoxicity (MN+NB+BNb, summed means measured in ‰) risk assessment in flounder, herring and cod specimens collected from the Bornholm Basin chemical munitions (CW and WG) dumping zones. Genotoxicity risk scale: green colour reflects low (reference), yellow – increased (background, ≥P90), orange – high (threshold, ≥P99) and red – extremely high (over threshold) responses.

AC C

772 773 774 775 776 777 778 779 780 781

18

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Table 1 The list of fish catch-stations and measurements of hydrological parameters in the southern Baltic Sea (data from the r/v “Walther Herwig III” and the r/v “Baltica).

AC C

EP

TE D

M AN U

SC

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Depth Salinity O2 Temperature Study of at the concentration at the Latitude Longitude stations trawling bottom at the bottom bottom (°C) (m) (PSU) (mg/l) B03 54°37.84'N 15°03.52'E 51.5 10.78 10.67 – (2003) B05 55°05.30'N 16°32.24'E 43 9.71 9.9 – (2003) B03 54°38.19'N 15°13.63'E 59 7.31 14.44 – (2004) B03/46 54°37.25'N 15°05.58'E 51 8.08 10.13 6.01 (2009) 54°26.20'N 15°09.50'E 19 32 6.90 8.13 6.44 54°26.40'N 15°12.00'E 54°27.50'N 15°37.40'E 51 5.41 9.92 4.81 21 54°27.20'N 15°40.00'E 54°23.00'N 15°46.30'E 22 31 7.73 7.26 7.63 54°22.90'N 15°48.80'E 54°31.70'N 15°47.30'E 54–56 5.71 9.03 5.48 23 54°30.70'N 15°49.20'E 54°37.60'N 16°03.50'E 48 7.67 7.35 7.64 24 54°36.20'N 16°02.00'E 54°32.00'N 16°00.00'E 25 44–47 7.80 7.28 7.77 54°33.60'N 16°00.20'E 54°52.60N 16°39.50E 28 20 7.46 7.28 7.85 54°52.80N 16°41.90E 54°39.40'N 15°09.80'E 15a 61 5.01 10.43 8.25 54°39.50'N 15°08.90'E 54°34.00'N 15°38.50'E 60 5.59 11.73 5.60 17a 54°34.80'N 15°36.80'E 54°33.30'N 15°24.60'E 18a 58 3.58 8.51 8.06 54°33.90'N 15°22.30'E 54°39.20'N 15°34.10'E 19a 68 5.91 11.97 4.27 54°38.90'N 15°31.70'E 54°38.00'N 15°53.70'E 21a 55 5.55 11.99 3.85 54°36.60'N 15°52.50'E 54°44.20'N 15°54.70'E 22a 54 6.03 12.60 3.76 54°43.10'N 15°54.70'E 54°48.70'N 16°01.00'E 52 4.94 11.10 5.99 23a 54°47.50'N 15°59.30'E 55°02.60'N 16°24.00'E 54 4.35 9.31 5.45 28a 55°02.30'N 16°21.60'E

Z24 14b 16b 17b 18b 26d 33d 34d 35d 36d 45d 47d B13/06 B13/07 B13/08 B13/09 B13/10 B13/11 B13/12 B09/17 B09/18 B09/19

12.34

4.57

43–46

3.22

7.12

6.85

57–74

3.43

11.85

4.66

35–61

3.22

9.42

50–52

2.70

8.42

66

4.66

12.76

2.81

51

5.68

9.86

3.66

54

5.99

48 34

54 60–61 69 76 76 74 51 92 91 92 91 85 93 87 77 67 59

RI PT

Z13

8.55

SC

Z12

46

TE D

Z11

14°16.50'E 14°16.50'E 14°18.60'E 14°20.20'E 15°13.70'E 15°11.50'E 15°17.10'E 15°17.00'E 15°15.30'E 15°15.60'E 15°41.90'E 15°45.00'E 16°01.70'E 16°00.20'E 15°54.40'E 15°52.60'E 16°04.10'E 16°04.00'E 16°18.40'E 16°20.30'E 16°24.60'E 15°09.10'E 15°19.30'E 15°38.70'E 15°41.00'E 15°42.70'E 16°03.00'E 15°35.53'E 15°37.10'E 15°34.03'E 15°31.48'E 15°50.91'E 15°44.33'E 15°48.09'E 18°20.16'E 18°14.49'E 18°10.85'E

EP

Z10

55°02.50'N 55°02.50'N 55°17.60'N 55°19.00'N 55°27.90'N 55°26.50'N 55°17.60'N 55°15.80'N 54°51.10'N 54°49.20'N 55°07.30'N 55°07.40'N 54°48.90'N 54°47.50'N 54°39.60'N 54°40.70'N 54°38.90'N 54°40.50'N 54°36.50'N 54°37.60'N 55°03.10'N 54°39.40'N 54°46.20'N 54°50.30'N 54°50.70'N 54°47.80'N 54°50.20'N 55°20.32'N 55°21.01'N 55°17.42'N 55°18.93'N 55°07.27'N 55°16.03'N 55°25.10'N 55°08.30'N 55°08.75'N 55°13.57'N

AC C

Z1

M AN U

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11.32

5.47 5.45

3.00

5.58

10.20

2.47

5.53

8.37

6.26

2.16 6.57 6.78 6.91 6.91 6.86 2.30 6.32 6.32 6.31 6.23 6.61 6.43 6.24 5.05 5.16 5.62

7.93 12.74 13.42 14.31 14.31 14.77 7.93 14.79 14.78 14.81 14.81 14.72 14.78 14.73 10.73 10.38 9.33

8.72 3.72 2.64 2.36 2.36 2.72 8.92 0.21 0.28 0.24 0.25 0.39 0.19 0.24 0.34 1.33 6.17

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Table 2 Materials for the analysis of genotoxicity and cytotoxicity in peripheral blood erythrocytes of flounder, herring and cod specimens collected from different stations of the Baltic Sea (data from the Polish r/v “Baltica” and the German r/v “Walther Herwig III” surveys) in the period of Nov. 2011 – Feb. 2012.

19 (10), 22 (10), 23 (8), 24 (5), 25 (9), 28 (6)

February 2011 „Baltica“

15a (10), 17a (10), 18a (10), 19a (10), 21a (10), 22a (10), 23a (10), 28a (10)

June 2011 „Baltica“

Z1 (10), Z10 (10)

November 2011 „Baltica“

14b (5), 16b (7), 17b (5), 18 b (3)

December 2011 “Walther Herwig III”

–

February 2012 „Baltica“

26d (10), 35d 6), 36d (7)

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–

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188 specimens from 23 stations Earlier investigated study stations December 2003 “Walther Herwig B03 (20), B05 (17) III” September 2004 “Walther Herwig – III” December 2009 “Walther Herwig – III”

Total

Cod sampling (investigated specimens)

RI PT

November 2010 „Baltica“

Herring sampling stations (investigated specimens) 19 (10), 21 (10), 22 (10), 23 (10), 24 (10), 25 (10), 28(10) 15a (10), 17a (10), 18a (10), 19a (10), 21a (10), 22a (10), 23a (10), 28a (10) Z10 (9), Z11 (10), Z12 (10), Z13 (10), Z24 (10) 14b (10), 16b (9), 17b (10), 18b (10)

–

SC

Flounder sampling stations (investigated specimens)

M AN U

Sampling date and vessel

26d (10), 33d (9), 34d (10), 35d (10), 36d (10), 45d (10), 47d (10) 307 specimens from 31 stations

–

–

B09/17 (15), B09/18 (22), B09/19 (8), B13/06 (12), B13/07 (13), B13/08 (15), B13/09 (11), B13/10 (11), B13/11 (10), B13/12 (10) – 127 specimens from 10 stations

–

–

–

B03 (15)

B03/46 (6)

–

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Figure caption: Fig. 1. Location of sampling stations in the Baltic Sea (CW and WG – known warfare dumpsites).

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Fig. 2. Nuclear abnormalities in peripheral blood of fish: a) erythrocyte with micronucleus (MN), b) erythrocyte with nuclear bud (NB), c) nuclear bud on filament (NBf), d) bi-nucleated erythrocyte with nucleoplasmic bridge (BNb), e) fragmented-apoptotic erythrocyte (FA), f) binucleated erythrocyte (BN), g) erythrocyte with 8-shaped nucleus (8-shaped).

b)

f)

d)

g)

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e)

c)

Fig. 3. Genotoxicity (MN+NB+BNb) and cytotoxicity (FA+BN+8) levels in flounder specimens collected from 23 stations in 2010-2012 and from stations B03 and B05 in December 2003. Fig. 4. Genotoxicity (MN+NB+BNb) and cytotoxicity (FA+BN+8) levels in herring specimens collected from 31 stations in 2010-2012 and from the station B03/46 in December 2009. Fig. 5. Multiple morphological abnormalities of herring blood erythrocytes.

AC C

761 762 763 764 765 766 767 768 769 770

a)

EP

760

M AN U

SC

750 751 752 753 754 755 756 757 758 759

771

17

AC C EP TE D

SC

M AN U

RI PT

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4

RI PT

3.5

0.12

3

0.09 0.19

0.12

0.10

2.5

0.35

0.19

2

0.12

SC

0.13

0.32

0.07

M AN U

0.14

0.61

0.21 0.42

AC C

EP

0.41

0.38

0.29

21a

0.36

19a

0.31

18a

24

TE

23

22

19

D

0.5

28

0.50

17a

1

0

0.41

0.25

15a

0.32

23a

0.23

22a

1.5

25

NA/1000 erythrocytes

0.16

RI PT

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5.5 5

0.19

SC

4.5

3 2.5 0.28

2

D

TE

AC C

0.12

1.59

0.35

1.41

0.25

0.35

0.69

22a

21a

19a

18a

17a

15a

28

25

24

19

22

0.13

EP

0

0.12

0.36

21

0.5

0.10

0.45 0.58

0.08

0.40

0.14

0.25

0.13

0.49

0.32

0.19 0.97

0.21

0.08

1.5 1

M AN U

3.5

23

NA/1000 erythrocytes

4

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2

RI PT

1.8

0.24

1.4

SC

1.2

0.4 0.2

0.11

AC C

EP

TE

D

B09/17

0

0.07

0.18

0.20

0.24 0.14 0.17

0.18

B09/19

0.07

0.6

0.13

B13/07

0.18

M AN U

0.8

B13/06

1

B09/18

NA/1000 erythrocytes

1.6

AC C EP TE D

SC

M AN U

RI PT