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Aquatic Toxicology 99 (2010) 473–478

Contents lists available at ScienceDirect

Aquatic Toxicology

journa l homepage: www.e lsev ier .com/ locate /aquatox

esponses of metabolic pathways to polycyclic aromatic compounds in flounderollowing oil spill in the Baltic Sea near the Estonian coast

andel Kreitsberga,b,∗, Irina Zemitb, Rene Freibergb, Meelis Tambetsc, Arvo Tuvikeneb

Department of Zoology, Institute of Ecology and Earth Sciences, University of Tartu, Vanemuise 46, 51020 Tartu, EstoniaCentre for Limnology, Estonian University of Life Sciences, Rannu, 61117 Tartu County, EstoniaDepartment of Hydrobiology, Institute of Ecology and Earth Sciences, University of Tartu, Vanemuise 46, 51020 Tartu, Estonia

r t i c l e i n f o

rticle history:eceived 24 January 2010eceived in revised form 16 June 2010ccepted 22 June 2010

eywords:AHlounderil spilliotransformationAC

a b s t r a c t

In January 2006 an oil spill that involved approximately 40 tons of heavy fuel oil affected more than30 km of the north-west coast of Estonia. The aquatic pollution of the coastal area of the Baltic Seawas monitored by measuring the content of selected polycyclic aromatic hydrocarbons (PAHs and PAHmetabolites) in flounder (Platichthys flesus trachurus Duncker). One hundred and thirty-one fish werecollected: muscle and liver tissues were analyzed by high-performance liquid chromatography (HPLC);bile and urine samples were analyzed using fixed wavelengths fluorescence. Fifteen different types ofPAHs were analyzed in liver and muscle, and four types of PAH metabolites were analyzed in bile andurine (2-, 3-, 4- and 5-ringed PAH metabolites represented by naphthalene, phenanthrene, pyrene andbenzo(a)pyrene). Fluorescence analyses were carried out using excitation/emission wavelength pairs:290/380, 256/380, 341/383 and 380/430 nm, respectively. There was a time-dependent decrease of PAH

he Baltic Sea concentrations in liver (83%), bile (82%) and urine (113%). HPLC analysis of muscle tissues demonstratedlow concentrations of single PAHs, but a decrease of concentrations during the study period was notobserved. During the analyses concentrations of PAH metabolites in bile and urine were compared. Livermetabolic transformation activity is believed to exceed that of the kidney but the analyses demonstratedhigh metabolite concentration in fish urine, particularly of 4- and 5-ring PAH metabolites. The results

er capc path

indicate remarkable buffof kidney-urine metaboli

. Introduction

It is well known that polycyclic aromatic hydrocarbons (PAHs)ay originate from natural as well as from anthropogenic

ources. Anthropogenic sources in the aquatic environmentnclude petroleum spills and discharges, industrial and municipal

aste water, refuse incineration and internal combustion enginesatmospheric deposition). The aromatic content of petroleum, par-icularly PAHs, has been generally assumed to be the principaleterminant of the toxicity of oil to aquatic organisms. Natu-al sources of PAHs have existed for millions of years and fishave developed different metabolic pathways to transform PAH

olecules into water-soluble metabolites (Hahn et al., 1988). Even-

ually these metabolites are excreted via the gall bladder (Varanasi,989) or urinary bladder (Pesonen et al., 1987).

∗ Corresponding author at: Department of Zoology, Institute of Ecology and Earthciences, University of Tartu, Vanemuise 46, 51020 Tartu, Estonia.el.: +372 5229383.

E-mail address: randel.kreitsberg@ut.ee (R. Kreitsberg).

166-445X/$ – see front matter © 2010 Elsevier B.V. All rights reserved.oi:10.1016/j.aquatox.2010.06.005

acity of hydrodynamically active sea as well as considerable importanceways in flounder physiology.

© 2010 Elsevier B.V. All rights reserved.

Organic pollutants, including polycyclic aromatic compounds,are absorbed easily via passive diffusion through the gills, intes-tine and skin, and accumulate in lipid-rich tissues (Varanasi, 1989).Recent advances in molecular and cell biology have led to anincreased understanding of the influence of toxic PAH metaboliteson various factors including DNA, RNA, fertility and sex hormones(Carlson et al., 2002; Chen and White, 2004; Schober et al., 2006).During metabolic transformation processes PAH metabolites formadducts with DNA molecules and cause several lesions by changingthe structure of coding sequences (Carlson et al., 2002; Chen andWhite, 2004). Although it is generally known that PAHs do not bio-magnify markedly via the food chain, their cellular effects still leadto changes in organisms and their ecosystems.

Hundreds of thousands of tons of crude oil have been releasedinto the aquatic environment through the world’s largest oil spills(Galt et al., 1991; Gonzalez et al., 2006). The effects of these acci-dents can still be detected after 10 years (Peterson et al., 2003;

Payne et al., 2008). The Baltic Sea is one of world’s most activelynavigated seas; hundreds of major and minor oil discharges affectits ecology every year. HELCOM’s annual report reveals that 224illicit oil spills were observed during 5637 h of surveillance flightsconducted by coastal countries over the Baltic Sea during 2005

474 R. Kreitsberg et al. / Aquatic Toxicology 99 (2010) 473–478

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(1:1600) (Aas et al., 2000) and measured with fixed wavelengthpairs (Lee and Anderson, 2005) to detect PAH metabolites (Table 1).

Fig. 1. R

HELCOM, 2006). Nevertheless, the number of illegal oil dischargesrom ships, observed by surveillance planes in the Baltic Sea area,ecreased by more than 50% between 1998 and 2006.

On the 28th of January 2006 a notable oil spill (40 tons) wasiscovered on the beach of Keibu Bay (59◦14′59′′; 23◦44′38′′), Esto-ia (Fig. 1). More than 35 km of coastline was covered with heavy

uel oil and ornithologists estimated that 10,000–30,000 birds died.ccording to the International Tanker Owners Pollution Federation

ITOPF) classification, this amount of hazardous substance is classeds a middle-sized pollution (Historical data, 2007). At present, theil has settled to the bottom of the sea but approximately 8 m3 ofil is still thought to be buried under a layer of sediments (Martin,006).

The aim of this study was to quantify residues of PAHs from theil spill and its change over time using sediment-exposed flounderPlatichthys flesus trachurus Duncker). We used two different meth-ds: muscle and liver tissues were analyzed by high-performanceiquid chromatography (HPLC); whereas bile and urine samples

ere analyzed by fixed wavelengths fluorescence (FF). Using flu-rescence at fixed wavelengths to detect metabolites in fish bile istill a rather new method which has been modified several timesn recent years (Aas et al., 2000; Vuorinen et al., 2006). Fluores-ent aromatic compounds (FAC) could be promising biomarkersor quantifying the exposure of fish to PAHs (Schiedek et al., 2006;arsiene et al., 2006). However, this method has apparently noteen used for detecting PAH metabolites in fish urine previously.

n addition to FF analyses of bile and urine, we used HPLC to detectAH in the liver and muscular tissues of fish. Use of these meth-ds provides a thorough overview concerning the concentrationsf PAHs in pollution-exposed flounder.

. Materials and methods

Oil samples were collected from the affected coastline at the

eginning of March 2006 from the beach at Keibu Bay. The floun-ers (Platichthys flesus trachurus Duncker) analysed were caught onth of June, 22nd of August and 23rd of October 2006, 14th of Augustnd 1st of December 2007, and 9th of August 2008, using gillnets,rom polluted areas at depths of 1.5–4 m. The control fish samples

ch area.

for HPLC were collected next to the Saarnaki islet (HELCOM CoastalFish Monitoring area), more than 80 km from the polluted area. Onehundred and thirty-one specimens were analyzed. The fish were5–9 years of age: average body length (L) 193 ± 43 mm in Saarnakiand 275 ± 34 mm in Nõva; average body mass (G) 114 ± 114 and231 ± 116 g, respectively (Mean ± Standard Deviation). All sampleswere collected in situ: bile and urine samples were collected with1 ml disposable syringes directly from the bladder, after openingthe abdominal cavity. The liver and muscle samples were collectedusing a surgical knife, and the middle part of the tail area was usedfor muscle analysis. All samples were kept at +4 ◦C during trans-portation to the laboratory and stored at −20 ◦C pending analysis.

2.1. Chemicals

Acetonitrile and n-hexane were purchased from Merck KGaA(Germany), ethanol from Rakvere Ethanol Factory (Estonia), andKOH and anhydrous Na2SO4 from RdH Laborchemikallen GmbH& Co. (Germany). PAH standards were obtained from Ultra Scien-tific Co. (USA). Ultrapure water was produced with the Elga WaterSystem (Great Britain). Solvents for HPLC analysis were gradientgrade.

2.2. Fixed wavelength fluorescence (FF) analyzes

A Hitachi F-2500 fluorescence spectrophotometer equippedwith excitation and emission monochromators (Hitachi, Japan) wasused for the FF-analyses. 4-sided quartz cuvette size was 1 cm;spectral band slit was 5 nm. Software Fl Solution 2.0 (Hitachi, Japan)was used to operate the fluorescence spectrophotometer and toacquire results. Bile and urine samples were diluted in 48% ethanol

The results are presented as fluorescence units (FU), which are pro-portional to the actual concentration of PAH metabolites in thedilution. Relying on the analogy of enzyme complexes in liver andkidney (Pesonen et al., 1987), the same wavelength pairs and dilu-tions were used for urine samples (Table 1).

R. Kreitsberg et al. / Aquatic Toxicology 99 (2010) 473–478 475

Table 1Fixed wavelength pairs used to detect molecularly different PAH equivalents usinga fluorescence spectrophotometer (Lee and Anderson, 2005).

PAH’s equivalent Wavelength

Excitation Emission

Naphthalene (2-ring PAHs) 290 nm 380 nm

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demonstrated low concentrations of single PAHs, and there was no

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Phenanthrene (3-ring PAHs) 256 nm 380 nmPyrene (4-ring PAHs) 341 nm 383 nmBenzo(a)pyrene (5-ring PAHs) 380 nm 430 nm

.3. High-performance liquid chromatography (HPLC) analyses

The method described by Tuvikene et al. (1997) was modifiedor measuring parent PAHs in muscle and liver tissues of fish byPLC.

For homogenization and fat hydrolysis, the liver and muscleamples (2–4 g wet weight) were ground and boiled for an hourn 20 ml of 3% solution of potassium hydroxide in 96% ethanol. Theomogeneous mass was cooled and diluted with deionised water.

PAHs from samples were extracted twofold with 10 ml of n-exane. The n-hexane fraction was isolated with separating funnel,hen dehydrated in glass column filled with anhydrous Na2SO4, andoncentrated to 1 ml with a Rotavapor R110 rotor-evaporator.

A CECIL 1100 high-performance liquid chromatography gra-ient system (CECIL Instruments, Cambridge, UK) with a UV-etection wavelength of 254 nm was used for chromatographicnalyses. An acetonitrile-water combination was used as eluenthe mobile phase parameters were: the initial gradient by volumeas held at 50:50 (AcN:H2O) for 5 min, then there was a linear

radient change to 100:0 for 15 min and held at 100:0 for anothermin. The flow rate was 2 ml/min and the injection volume 50 �l.Hypersil Green PAH (Thermo Electron Corporation, USA) columnith 5 �m particle size and dimensions of 100 mm × 4.6 mm IDas used. Chromatograms were obtained and analysed using the

omputer program Data Control (CECIL Instruments, Cambridge,K).

The average lengths of fish varied in different locations and atifferent time-points, therefore we included fish length and bodyass as co-variants in data analyses. Nonparametric Spearman cor-

elation was used to describe the relationship between fish bodyize and PAH concentrations. For time-scale dynamics of pollution,

ne-way co-variation method and post hoc Tukey test were used.rincipal components (PC), calculated from the concentrations ofour types (2-, 3-, 4- and 5-ring PAH metabolites) of fluorescentromatic compounds were used in ANOVA for comparing the con-

able 2istribution of selected PAHs measured in the oil sample (collected in March, on the beach

n June; n = 5, means presented) and flounder liver tissues from the control area (collected

PAH levels in oilsample (�g/g DW)

Pyrene 3.50Fluoranthene 3.46Benzo(b)fluoranthene 0.34Benzo(a)anthracene 0.77Chrysene 0.51Dibenzo(a,h)anthracene 0.29Phenanthrene 0.71Acenaphthylene 0.04Benzo(k)fluoranthene 0.16Benzo(g,h,i)perylene 0.14Acenaphthene 0.00Fluorene 0.48Anthracene 0.11Indeno(1,2,3-cd)perylene 0.09Benzo(a)pyrene 0.12

Fig. 2. PAHs in muscle. The concentration of total PAHs in flounder muscle tis-sues as dependent on body length (L): Spearman’s correlation: n = 16 (all musclemeasurements); Spearman R = −0.46; t = −1.98; p = 0.06.

centrations of PAH metabolites in bile and urine. In addition, PAHmetabolite concentrations were analysed on the basis of moleculesize (2-, 3-, 4- and 5-ring) and tissue type (bile and urine).

3. Results

The dominant PAHs were identical in oil and in liver samplesfrom the polluted area of Nõva, which indicates a common ori-gin of the pollutants. The dominant PAHs in Nõva liver sampleswere fluoranthene and pyrene and the highest concentrations inliver reached 25 �g/g dry weight (DW) (Table 2). HPLC analysesproduced detectable results in some muscle samples and revealedthat benzo(b)fluoranthene was one of the most abundant PAH infish muscle tissues (highest concentration 1.4 �g/g DW).

The total concentrations of PAHs in fish muscle tended to cor-relate negatively with the length of the fish (p = 0.06; Fig. 2).

Total PAH concentrations in liver from polluted areas were sig-nificantly different (p = 0.01) at different sampling times during thesummer of 2006 (Fig. 3A). A significant decrease in PAH contentof liver was identified during this period; from June to Octoberthe levels of PAH diminished 83% in liver tissues (means 30.2 and5.1 �g/g DW, respectively; Fig. 3A). HPLC analysis of muscle tissues

time-dependent decrease in the concentrations (p = 0.81).The principal component scores of four types of fluorescent

aromatic compounds (FACs) were different at various time-points(June 2006 to August 2008) in the polluted area. A remarkable

of the polluted area; n = 1), flounder liver tissues from the polluted area (collectedin August; n = 10, means presented).

PAH levels in liver sample(polluted area) (�g/g DW)

PAH levels in liver sample(control area) (�g/g DW)

8.37 21.2711.62 2.84

0.00 10.203.12 6.160.00 7.520.00 7.552.70 0.750.11 2.470.00 2.140.00 1.450.32 1.170.81 0.000.42 0.100.17 0.000.00 0.00

476 R. Kreitsberg et al. / Aquatic Toxicology 99 (2010) 473–478

Fig. 3. Decrease of pollutants following oil spill. Total PAH concentrations (LSmeans ± 95%) in flounder liver (A) and principal component scores of fluorescentaromatic compounds (LS means ± 95%) in bile (B) and urine (C) at different time-points in the polluted area (Nõva) from June 2006 to August 2008. Saarnaki (A)rcpf

decb

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Table 3ANOVA testing the effects of tissue (bile and urine) and PAH molecule size (2-,3-, 4- and 5-ring) on PAH metabolite concentrations. The dependent variable waslogarithmically transformed to allow interpretation of the interaction in relativeterms.

Type III SS DF F p

Tissue 8.3 1 44.6 <0.00001Molecule size 672.5 3 1193 <0.00001Tissue × molecule size 4.1 3 7.43 0.000067

Fig. 4. PAH metabolites in bile and urine. The concentrations of final PAH conjugates(logarithmically transformed fluorescence units, FU) are presented, measured by

epresents a control area for HPLC analyses. ANCOVA testing the effect of spe-ific time points on the concentrations of PAHs and PAH metabolites: p = 0.01 (A),< 0.001 (B) and p < 0.001 (C) respectively. Different letters denote significant dif-

erences between single points of time (p < 0.05).

ecrease in PAH metabolite concentrations in bile and urine wasvident (p < 0.001; Fig. 3B and p < 0.001; Fig. 3C, respectively). Mean

oncentrations for PAH metabolites decreased by 82% in bile andy 113% in urine samples.

PAHs with different molecular sizes had different metabo-ite concentrations in bile and urine (Table 3). The significantnteraction term indicates that PAH metabolites with different

using the FF method (LS means ±95% confidence intervals). ANOVA testing the effectof tissue (bile and urine) on concentrations of differently sized PAH metabolites;2-ring molecules: F1.16 = 34.6, p < 0.0001; 3-ring molecules: F1.16 = 15.3, p = 0.0001;4-ring molecules: F1.16 = 2.8, p = 0.09 and 5-ring molecules: F1.16 = 1.0, p = 0.3.

molecular sizes were disproportionally distributed between the tis-sues; there were relatively higher concentrations of low-molecularPAH metabolites in bile (Fig. 4). In contrast, the amounts of 4- and5-ring PAH metabolites remained approximately at the same levelin bile and urine samples.

4. Discussion

4.1. PAHs in liver

Relatively high PAH concentrations were measured in the liv-ers of fish even 5 months after the oil spill, indicating that flounderwere exposed to highly polluted sediments. Recovery from oil pol-lution ranges from a few months (Lee and Anderson, 2005) toseveral years (Hayakawa et al., 2006; Boehm et al., 2007). We con-clude that in the case of Nõva, the majority of the pollution wasbuffered by a hydrodynamically active sea (heavy waves, currentsand upwelling area) and this was evident in October when the con-centrations of PAHs were close to the reference level. The total levelof PAHs in fish livers from the polluted area decreased six timesbetween June and October, an effect similar to analogous studies inSpain, Scotland and Alaska (George et al., 1995; Lee and Anderson,2005; Gonzalez et al., 2006). No toxicological studies had been car-ried out in this area before the accident, therefore there is no directinformation about the natural level of PAH concentrations in thesewaters. Nevertheless, total PAH content in liver reached 5.1 �g/g

DW in October, which is comparable to the level in our controlarea, Saarnaki.

PAHs abundance in the polluted area during October was com-parable to the concentrations of Estonia’s inland waters in the NarvaRiver where total PAH concentrations in roach (Rutilus rutilus) liver

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aried between 1.9 and 2.9 �g/g DW (Tuvikene et al., 1997). Ithould be noted that the Narva River is moderately affected by oilhale mining.

No data are available for previous PAH concentrations in fishiver from Estonian coastal waters, but it is possible to comparehe results with PAH concentrations in the Mediterranean andlack Sea. Red mullets (Mullus barbatus) from coastal areas of theediterranean Sea near Spain and France had total PAH concen-

rations in different locations of between 0.0147 and 0.0496 �g/gW (Blaumard et al., 1998). During 1994–1996 the levels of PAHsere measured in the Mediterranean and Black Sea from the livers

f three mullet species (Lisa ramada, Oedolechilius labeo and Mugilaliens). In METU and Mersin harbours on the southern coast ofurkey the concentrations of PAH were 9.3 ± 1.2 and 37.1 ± 7.6 �g/gW, respectively, and in Yomra harbour on the northern coast ofurkey the concentrations were 55.7 ± 16.6 �g/g DW (Telli-Karakoct al., 2001). This study presents the PAH content following an oilpill but the Baltic concentrations are significantly lower than inarious harbour areas in the Mediterranean and the Black Sea.

.2. PAHs in muscle

An unexpected and remarkable result was the negative corre-ation between the body size of fish and the PAH concentrationsn muscle tissues (p = 0.06). The majority of toxicological studiesemonstrates that high levels of pollutants (heavy metals, PCBs)nd fish size have a positive correlation (DeVault et al., 1986;tafford et al., 2004), but the opposite is true for PAHs (Vuorinent al., 2006). However, there is general agreement that high con-entrations of toxic substances are correlated with the age of thepecimen (Stow and Carpenter, 1994) and not the size of the body;he time that fish have been exposed to the pollutant is the mostmportant factor. It is, however, important to remember that PAHsre chemically active when entering an organism and biotransfor-ation of PAHs starts immediately; no considerable cumulative

ffect via the food chain has been found (Varanasi, 1989).There have been few studies analyzing PAHs from fish mus-

le in Estonian coastal waters, leaving us with scarce referenceoints. Research on eelpout (Zoarces viviparus) from the Gulf of Rigaemonstrated total PAH concentrations in fish from three differ-nt locations of 0.33, 0.056, and 0.64 �g/g DW in muscle (Ojaveernd Tanner, 1996). The most commonly recorded PAH was phenan-hrene, followed by anthracene and benzo(b)fluoranthene.

.3. PAH metabolites in bile and urine

Two different methods were used during our study; a relativelyast and inexpensive fixed wavelength fluorescence analysis, and

time-consuming and more expensive HPLC method. This workemonstrates that using both methods gave more detailed resultsquantitative HPLC method) and allowed a quicker throughput ofamples (FF analysis).

Recent biomarker studies conducted in the Baltic Sea haveemonstrated the suitability of PAH metabolite measurements insh bile to investigate PAH pollution. In particular, elevated lev-ls of pyrene-type PAH metabolites in flounder were registeredfter an oil spill at the Butinge oil terminal in 2001 (Barsiene etl., 2006); FACs (fluorescent aromatic compounds) in the bile ofelpout caught from Wismar Bay were also higher at more pol-uted sites (Schiedek et al., 2006). Still, PAH metabolite values

easured using the FF method are semi-quantitative and therefore

omparisons to other studies are difficult (Schiedek et al., 2006),articularly when there are no data regarding urine samples andew with regard to bile samples after oil spillages (Barsiene etl., 2006). Barsiene et al. (2006) reported an approximately 66%ecrease in PAH metabolites in the bile of flounder 7 months after

ology 99 (2010) 473–478 477

the Butinge oil spill of November 27, 2001 in Lithuania (Baltic Sea).Compared to the latter measurements, the present data demon-strate a more gradual decrease in pollutants (82% in 2 years, Fig. 3B),which could be due to sedimentary effects in Nõva and Keibu Bay.Martin (2006) reported a large amount of oil buried under a slightlayer of sediments in underwater observations.

Previous studies using several extra-hepatic tissues of differentfish species have revealed that almost every tissue or organ exam-ined had detectable enzyme (monooxygenase) activity towardsPAH, with maximum activity being detected in liver (Varanasi,1989). Two metabolic pathways excrete biotransformed PAH com-pounds, via the liver and bile into faeces, or through the kidney andinto urine. Metabolites from bile could be further biotransformedby intestinal microbes before final excretion (Karasov and Martínesdel Rio, 2007).

We used a novel method to compare the activities of twometabolic pathways (liver-bile and kidney-urine) in oil-pollutedfish. FF measurements have not been used previously to detectPAH metabolites in fish urine. The results demonstrated that PAHmetabolite concentrations are different in bile and urine. Further-more, their distribution is based on the metabolite’s molecule size.We found that metabolic reactions in kidneys were notably high,particularly for 4- and 5-ring molecules, causing approximatelythe same transformation level as in liver. At the same time, for2- and 3-ring PAHs, there is higher metabolic activity in liver,as it has been described before (Pesonen et al., 1987; Nakayamaet al., 2008). Previous studies have demonstrated that the maintransformation enzymes for PAHs are components of cytochromeP-450 (Varanasi, 1989) and levels of P-450 are approximately 5-fold higher in liver than in kidney (Pesonen et al., 1987; Nakayamaet al., 2008). Therefore, higher concentrations of metabolites inbile would be expected. This is the case for 2- and 3-ring PAHs inthis study but not for metabolite levels of 4- and 5-ring PAHs. Weconclude that the activity of metabolic reactions in kidney is unex-pectedly high in the present study, particularly for 4- and 5-ringmolecules, which demonstrate approximately the same transfor-mation level as in liver. This suggests a much higher importance ofkidney in PAH’s metabolism than assumed up to the present.

In conclusion, 40 tons of heavy fuel oil caused moderate harm tofish and there was no evidence of large-scale deaths in the pollutedarea. Nevertheless the highest level of contaminants was present inearly spring, when the spawning and maturation of eggs and larvaewas present. Flounders in this part of Baltic have short migrations,and according to Ojaveer and Drevs (2003), the spawning areas inthe Gulf of Finland partly overlap with oil spill areas. Longstandingstudies have demonstrated that early stages of the reproductiveprocess in female fish (e.g. English sole (Parophrys vetulus)) can bedisrupted by contaminant exposure, and if exposed to substantiallevels of pollutants, female fish can be partially excluded from thespawning population (Collier et al., 1992). Although our data do notshow that, it is possible that there was an impact on the populationof coastal spawning flounder in north-west Estonian sea areas dueto the oil spill.

Following the spill the oil settled to the sea bed; approximately8 m3 of oil could be buried under a layer of sediments (Martin,2006). Although we found considerable decrease in polycyclic aro-matic hydrocarbons in fish liver, the contamination remains in thisarea of the Baltic. It is generally known that sediment hydrocarboncontamination can be harmful to ecosystems health. For example,biochemical effects such as changes in monooxygenase enzymelevels and the fat content of fish liver occur at very low hydro-

carbon levels (∼1.0 �g/g) (Payne et al., 1988). Kotta et al. (2007)demonstrated that the oil spill in Keibu Bay had a moderate effecton macrophytes and macrozoobenthos.

PAHs are readily absorbed by fish and other aquatic animals dur-ing exposure to contaminated food, water and sediments, reaching

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evels higher than those in the ambient medium. The relative con-entrations of PAHs in aquatic ecosystems are generally highestn the sediments, medium in the aquatic biota, and lowest in the

ater column. The current study partly confirmed this relation-hip. The average PAH content in the sediments collected from theolluted area 3 months after the oil spill was 87 �g/g DW, and theontent in water plants collected from the same area during sum-er 2006 was 34 ng/g DW (Kotta et al., unpublished data). In the

orth-western Estonian coastal area, accumulation of oil in sedi-ents enables harmful pollutants to affect the benthic flora and

auna.

cknowledgements

This research was supported by the Estonian Environmentalnvestment Centre, Estonian Ministry of Education and Sciencestarget funding projects No. SF0170011s08, and SF0180122s08),nd BONUS+ project BEAST. Our team is grateful to everyone whoupported this study, particularly Anu Kisand, Merit Kreitsberg,nnela Hendrikson, Toomas Tammaru, Anno Põld, EMI Baltic Seaonitoring team and two anonymous reviewers for their com-ents and advice.

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