Judit Kalman

Inmaculada Riba

Julián Blasco

T Ángel Delvalls

Is δ-aminolevulinic acid dehydratase activity in bivalves from south-west Iberian Peninsula a good biomarker of lead exposure?

Is

Abstract

The coast of Huelva is considered as a heavily contaminated area where the Tinto and Odiel rivers discharge after running through a metalliferous mining area in the Iberian Pyrite Belt and end in common estuary called Ría of Huelva. Lead is a highly toxic and widely distributed element in the aquatic environment; therefore there is a great interest in assessing the impact of this contaminant on aquatic organisms. To study the bioavailability and sources of lead, the bivalve species Chamelea gallina was collected from six sampling sites along the coast of Huelva. Besides the effect of this metal on the enzyme activity, -aminolevulinic acid dehydratase (ALA-D) was determined in the whole tissue. Results show the inverse relationship between the lead concentration and the activity of ALA-D measured in whole tissues (r 2 =0.7). Individuals collected from the Ría of Huelva estuary had the highest levels of total lead concentrations and the lowest activity of ALA-D. Lead isotope analyses demonstrated the different sources of lead contamination and also confirmed the influence of Ría of Huelva on the metal input to the marine environment.

Keywords: ALA-D activity; Bivalve; Lead; Coast of Huelva *Corresponding autor: Tel.: +34-956-016040; fax: +34-956-016040 E-mail address: angel.valls@uca.es Lead is a highly toxic contaminant due to its ability to accumulate in living organisms and inhibit the activities of essential enzymes. The main sources are mining and industrial activities, sewage waste and the remains of leaded gasoline used in the past. The study area, the coast of Huelva, is situated on the Iberian Pyrite Belt, which is one of the most famous sulfide mining regions in the world. The coast receives considerable fluvial contaminant input from Ría of Huelva formed from the Odiel and Tinto Rivers [START_REF] Machado | [END_REF]. Pb contamination of the study area is related to sulfide minerals outflow which, besides wastes from the industrial complex located at the junction of the two rivers, contribute to considerable lead input to the marine environment [START_REF] Martins | VI Congresso Ibérico de Geoquímica[END_REF]. The epibenthic suspension feeder, Chamelea gallina is abundant along the whole Atlantic coast of southern Spain and highly appreciated for human consumption (Usero et al., 2005). The activity of -aminolevulinic acid dehydratase (ALA-D) is considered as a specific and useful biomarker of lead pollution (Aisemberg et al., 2005). This enzyme catalyzes the condensation of two molecules of -aminolevulinic acid to porphobilinogen (PBG), which is an important step in the heme biosynthesis. In the presence of this pollutant the activity of ALA-D is inhibited.

To investigate the influence of lead contamination and its adverse biological effect, bivalves were collected along the coast of Huelva and total lead concentration and activity of ALA-D were determined in the whole tissues. This survey was carried out as a part of a multidisciplinary study (INTERREG IIIA, UTPIA) to investigate the distribution and sources of lead. In the framework of this project lead isotopes (204Pb,206Pb,207Pb,208Pb) were determined to identify the origins of lead contaminant.

Figure 1 shows a map indicating the six sampling sites. At each sampling site approximately twenty individuals were collected. Three of them were frozen individually in liquid nitrogen immediately after dissection, and the rest (about seventeen organisms) were placed in plastic containers filled with seawater and transferred to the laboratory. Samples for analysis of ALA-D inhibition were stored at -80˚C, while samples for lead determination were placed in clean aerated seawater for 48 h to depurate the digestive tract. After dissection pooled samples were freeze-dried and prepared for total Pb and lead isotope analyses. In our laboratory total lead concentration was determined in duplicate by ICP-OES (Optima 2000 DV, Perkin Elmer), bicinchoninic acid assay (Smith et al., 1985) was used for the measurement of total protein concentration and the method described by Nakagawa et al. (1994) was followed for determination of the inhibition of ALA-D. The identification of distinctive isotope compositions of lead was carried out by GEOMEDIC Laboratory using multi collector ICP-MS.

Total lead concentrations were highest in samples collected near the estuarine mouth (1.58 g g -1 d.w. for stations 1), while organisms found further from the river outflows accumulated less lead (0.60-0.84 g g -1 d.w.). Although C. gallina is commonly found not only the Atlantic coast, but also in the Mediterranean Sea, limited surveys were conducted on the bioaccumulation of lead. Castro et al. (1999) recorded much lower lead concentrations in C. gallina (0.07-0.30 g g -1 d.w.) collected from fishing areas located in the coast of Almería (SE Spain), suggesting that coast of Huelva is a heavily lead-contaminated area. It is well established that higher concentration of Pb causes higher inhibition of ALA-D [START_REF] Hodson | Contaminant Effects on Fisheries[END_REF], which is in accordance with our findings, as the lowest activities of ALA-D were found in tissues collected from the estuarine mouth (Fig. 2). For several reasons it is difficult to compare our results to other reports; most studies were performed in the blood or particular tissues of vertebrates and not in the whole tissues; moreover the experiment-related variables (acclimation period, temperature) differed as well. ALA-D inhibition was determined by Campana et al. (2003) in the blood, liver and kidneys of Halobatrachus didactylus, with considerably higher levels obtained than those of the present study (7.54; 37.7; 75.4 ngPBG/min/mg protein). The strong inverse relationship (r 2 =0.7) between ALA-D and Pb concentrations found in this survey also confirms that ALA-D inhibition increased with the increase of Pb concentrations. The activity of this enzyme is sensitive to a variety of metals due to its sulfhydryl groups; therefore to propose the activity of ALA-D as a specific biomarker for lead contamination, further investigation is required.

Alves Costa et al. (2007) found that the activity of ALA-D in blood of the fish Hoplias malabaricus was affected not only by lead but by methylmercury as well.

The isotope studies indicated that the anthropogenic lead input is related to mining and industrial activities. On the other hand the strong increase in 206Pb/204Pb ratio with decreasing lead contamination also confirms that Ría of Huelva contributes to the lead input into the marine ecosystem on a large scale [START_REF] Martins | VI Congresso Ibérico de Geoquímica[END_REF].

Figure Captions

a

  Instituto de Ciencias Marinas de Andalucía. CSIC Polígono Río San Pedro s/n. 11510 Puerto Real (Cádiz). Spain. b Cátedra UNESCO/UNITWIN/WiCop. Facultad de Ciencias del Mar y Ambientales. Universidad de Cádiz. Polígono Río San Pedro s/n. 11510 Puerto Real Spain.

Fig. 1 .

 1 Fig. 1. Map of the studied area in southwest Spain showing the locations of six

Fig. 2 .

 2 Fig. 2. Relationship between the activity of ALA-D (ng PBG min -1 mg prot -1 ) and Pb concentration ( g g -1 d.w.) measured in the whole tissues of Chamelea gallina. Vertical

  Figure 1.

Acknowledgement

This work has been funded by the joint contract between the University of Huelva and the Instituto de Ciencias Marinas de Andalucía, CSIC under INTERREG IIIA, UTPIA, framework.


F Akcha
email: fakcha@ifremer.fr

G Arzul

S Rousseau

M Bardouil

Comet assay in phytoplankton as biomarker of genotoxic effects of environmental pollution

Keywords: Phytoplankton, Pesticides, Genotoxicity, Comet assay

The alkaline comet assay was tested on different microalgae: the dinoflagellates, Karenia mikimotoi and Alexandrium minutum, and the diatom, Chaetoceros gracilis. The microalgae were exposed during their exponential growth to the model direct genotoxicant, hydrogen peroxide (1h, 5 and 100 µM H 2 O 2 ). Following H 2 O 2 exposure, the comet assay was validated only for K. mikimotoi for which genotoxicity was observed from the lowest tested concentration of 5 µM with a concentration-dependent effect. C. gracilis was too small in size (4 µm) to be correctly analysed. For A. minutum, our lysis buffer was not strong enough to digest the cellulosic thecal plates. For K. mikimotoi, the comet assay was thus applied for the study of the genotoxic effects of different pesticides: epoxiconazole (as Opus® formulation), chlorpyriphos-ethyl (as Dursban® formulation) and endosulfan at 1, 10 and 100 µg of active substance /L for 24h. Exposure to epoxiconazole in formulation resulted in an increase in the extent of DNA strand breaks at the highest tested concentration of 100 µg/L. Genotoxicity

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was also observed for chlorpyriphos-ethyl in formulation from 1 µg/L (p<0.05), with a significant increase at 10 µg/L. Endosulfan exposure resulted in DNA damage for K. mikimotoi nuclei. Genotoxicity was observed from 1 µg/L of endosulfan and was not concentration dependent.

The phytoplankton is a model of choice for the study of the long term effects of pollutant exposure at population level. As primary producers, phytoplankton constitutes the first level of marine trophic chains. Due to its microscopic size, it is possible to get sample at population and community levels. Some species can be cultivated in photobioreactors under controlled conditions. Due to a high growth rate, phytoplankton offers the possibility to study the trans-generational effects of pollutant exposure. With a view to obtaining a better insight into the long term consequences of genotoxicity at population level, it appeared so valuable in a first step to test our alkaline comet assay on different microalgae. Up to now, most of the genotoxicity studies conducted in microalgae focussed on the formation of cyclobutane pyrimidine dimers, a DNA damage typical from UV irradiation (Fafandel et al., 2001, Häder et al., 2005).

For the exposure experiment, each phytoplankton strain [Karenia mikimotoi (GATIN95) isolated in 1995 in the bay of Brest, France; Alexandrium minutum (AM89BM) isolated in 1989 during a bloom in the Morlaix river, Brittany, France; Chaetoceros gracilis was obtained from SATMAR, Saint-Vaast-La Hougue, France] was cultivated in f/2 medium (Guillard and Ryther, 1962) at a temperature of 18°C and under controlled light (60µmol quanta/h/cm 2 ).

The microalgae were exposed during their exponential growth (concentration of about 50 x 10 6 cells/L) to hydrogen peroxide (1h, 5 and 100 µM), a model direct genotoxicant, and to different pesticides: epoxiconazole (as Opus® formulation form BASF), chlorpyriphos-ethyl (as Dursban® formulation from Dow AgroSciences) and endosulfan (in 0.01% DMSO) at 1, 10 and 100 µg of active substance/L for 24h exposure. The toxicity of these compounds was previously demonstrated for several strains of phytoplankton (unpublished data): for epoxiconazole and chlorpyriphos-ethyl, the effective concentration (EC50) obtained was lower when the active substance was tested in formulation.

Following exposure, the cell suspension was concentrated by centrifugation at 300g for 10 min and pellet was resuspended in 50 µL of filtrated seawater. Cell viability was assessed under the microscope by a neutral red viability test. The comet assay was then carried out as described previously in Akcha et al. (2003) by using 30 µL of the concentrated cell suspension. For each sample, two slides were prepared and the comet parameter Olive Tail Moment (OTM) of 150 cells was calculated using an image analysis system (Komet 4, Kinetic Imaging Ltd).

Following exposure, cell viability was at least 75% allowing the conduction of the comet assay. Comet data were normalised (Ln x) and analysed by an ANOVA (Statistica 6.0 Soft.), taking exposure concentration as a random factor. Following H 2 O 2 exposure, a significant dose-dependent increase in comet parameters was observed only for K. mikimotoi nuclei (p < 0.005) (Fig. 1), validating the use of the comet assay for genotoxicity study in this microalgal specie. C. gracilis nuclei was too small in size (4 µm) to be correctly scored by our image analysis system. It is noteworthy that an application of the comet assay was already reported for another genus of this specie, C. tenuissimus (3-5 µm), following cadmium exposure but damage was not assessed with the image software but by counting the number of comet cells against total normal cells (Desai et al., 2006). For A. minutum, our lysis buffer was not strong enough to digest the cellulosic thecal plates.

Epoxiconazole in formulation was genotoxic for K. mikimotoi only at the highest tested concentration of 100 µg/L (Fig. 2A). DNA damage was also induced by chlorpyriphosethyl in formulation from the lowest tested concentration of 1 µg/L (p<0.05), with a significant increase at 10 µg/L (Fig. 2B). Endosulfan exposure resulted in DNA strand breaks from the concentration of 1 µg/L (p < 0.05), but effect did not appear to be concentrationdependent (Fig. 2C). Effective concentrations for chlorpyrifos-ethyl (Auby, personal communication) and endosulfan (Nwankwoala and Osibanjo, 1992;Leonard et al., 2001, ) correspond to the higher concentrations recorded in the aquatic environment.

These results demonstrate the ability of K. mikimotoi to bioactivate the selected pesticides to reactive species for the DNA. Bioactivation of organic contaminants into genotoxic metabolites was already demonstrated by the SOS chromotest and by the comet assay for freshwater and marine green algae, Selenastrum capricornutum (Harwood et al., 1989) and Euglena gracilis, respectively (Aoyama et al., 2003). Despite the ability of microalgae to repair DNA strand breaks (Desai et al., 2006), genotoxic insult can have long term consequences on algal growth and diversity.

The applicability of the comet assay on phytoplankton offers valuable perspectives to study 1) the genotoxic effects of pollutants at population level, 2) the trans-generational effects of genotoxicant exposure and 3) the effects of genotoxicity on population genetics.

Fig. 1 .

 1 Figure captions:

Acknowledgements

This project was supported by IFREMER (French Institute for Exploitation of the Sea) and by The French Ministère de l'Ecologie et du Développement Durable.


J Couteau

J-M Flaman

C Minier

J Cachot

Detection of environmental mutagens using the FACIM assay

Keywords: Yeast, In vitro mutagenesis assay, UV, NQO, Sediment organic extract

A genetically engineered diploid yeast strain named yJC 2 , was specifically developed for environmental mutagen detection and characterization of induced mutations. This strain contains one copy of the human TP53 tumour suppressor gene coding sequence which is used as a molecular target for mutagens and two copies of the ADE2 reporter gene allowing accurate measurement of the TP53 transcriptional activity. The strain sensitivity to mutagens was evaluated by exposing cells to UVC, 4-nitroquinoline (NQO) or to an organic extract of sediment from the Seine estuary. For all studied mutagens, a significant and dose-dependent increase of mutant frequency was observed. The present assay named FACIM II (Functional Analysis of Chemical-Induced TP53 Mutations) is more convenient than the FACIM I and more inducible than the SOS Chromotest to detect direct-acting mutagens in the environment.

Email address: jerome.couteau@univ-lehavre.fr (J. Couteau)

In the aquatic environment, living organisms are exposed to complex mixtures of pollutants, some are potent mutagen including polycyclic aromatic hydrocarbons (PAHs), organic solvents and aromatic amine, etc (Watenabe et al., 2005;Cachot et al., 2006). These compounds can induce DNA damage which may lead to deleterious or heritable mutations, cancer or embryo lethality. It has been shown that co-exposure to UV can potentiate toxicity of certain mutagens such as PAH to aquatic larvae (Peachey, 2005). Detection of environmental mutagens is thus a major issue for preservation of natural ecosystems.

Accordingly a new assay named FACIM II was specifically designed to assess mutagenicity of chemicals or radiations and to allow identification of induced mutations at the TP53 locus.

An in vitro version of the FACIM assay was previously developed (Cachot et al., 2004).

In the present assay the human TP53 tumour suppressor gene cDNA as well as two functional copies of the reporter ADE2 gene under the transcriptional control of the TP53 protein were integrated into the genome of a diploid yeast strain by gap repair. The TP53 cDNA serves as a molecular target for mutagens while the reporter system allows accurate evaluation of the TP53 functional status. Yeasts expressing a mutant form of the TP53 protein are unable to activate expression of the ADE2 reporter gene. After plating onto an adenine-depleted medium, mutant ADE2 -mutant colonies turn red or pink due (no activation or impaired activation at the reporter gene) to accumulation of an intermediate in adenine biosynthesis while wild-type ADE2 + colonies are white.

For UV exposure, yJC 2 cells were grown to log phase (A 600nm of 0.4) and then plated onto selective solid medium. Plates were then exposed to 254 nm UV C radiations generated from a germicidal lamp (Bioblock) for 0, 106, 213, 320 and 426 J/m 2 . For exposure to 4-nitroquinoline-1-oxide (NQO) or organic extract, yJC 2 cells were grown up to log phase in an appropriate selective medium and were then exposed for 2 hours to 0-520 nM of NQO or to 1% (v/v) DMSO. Three replicates were performed for each treatment condition. After four days of culture at 30°C and one more day at 4°C (for colour development), red, pink and white colonies were counted in each plate. Mutant frequencies were calculated by dividing the number of red and pink colonies by the total number of colonies. Yeast cell survival was determined by calculating the ratio of the total number of colonies in each assay to the total number of colonies in control plates. Results were compared to the SOS developed by Quillardet and Hofnung (1996).

Following exposure to UVC a significant and dose-dependent increase in mutant frequency and decrease of cell survival were observed (Fig. 1A). The mutant frequency was always significantly (p<0.005) above the background level, except with the lowest UV dose (106 J/m 2 ). At the highest tested dose (426 J/m 2 ), a 79-fold increase in mutant frequency over the background level was observed. In the meantime cell survival decreased 6-fold. Using the haploid yeast indicator yeast strain yIG397 and an in vitro exposure protocol, Moshinsky and Wogan (1997) and Inga et al. (1998) reported a 17 to 18-fold increase of mutant frequencies following exposure to 300 J/m 2 of UVC.

Following exposure to NQO a significant and dose-dependent decrease in cell survival and increase in mutant frequency were observed (Fig. 1B). All treatments except the lowest NQO concentration (0.005 µM) induced a significant (p<0.005) increase of mutant frequencies in comparison to DMSO-exposed cells. At the higher test concentration (0.5 µM) a 78-fold increase of mutant frequency and a 18-fold decrease of cell survival were observed.

At the same NQO concentration and using a micro-plate version of the SOS Chromotest (White et al., 1996), a 8-fold increase of the SOS induction factor (IF) was observed (data not

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shown). In the meantime, limits of detection for both tests were very close, 0.016 µM and 0.025 µM for the SOS Chromotest and the FACIM II, respectively.

The genotoxic potency of a crude organic extract of sediments collected in Oissel (Seine estuary, France) in June 2005 (Cachot et al., 2006) was evaluated using both the SOS Chromotest (Quillardet and Hofnung, 1985) and the FACIM II. Concordant results were obtained with both tests. A clear dose-dependent increase of the SOS IF (Fig. 2A) and the mutant frequency (Fig. 2B) were observed. The maximum SOS IF achieved 1.7 while the mutant frequency increased 3-fold in comparison to the control solvent. These positive responses were obtained without addition of a S9 fraction indicating that this sediment sample contained organic pollutants with direct genotoxic activity. A previous study has already shown that sediments from the upper Seine estuary were genotoxic (Cachot et al., 2006) but this the first evidence of the presence of direct-acting mutagens.These preliminary results suggest that the FACIM II represent a sensitive bioassay to evaluate mutagenicity of directacting mutagens either pure or in complex mixture. This assay could also allow identification of mutations and characterization of mutation spectra induced by different classes of mutagens. Finally it was demonstrated that sediments from the upper Seine estuary contain direct-acting mutagens, which may represent a threat to aquatic species living in this area.

Figure captions

Organic extract per assay (%) Mutant frequency (%)

Linear regression polynomial regression

Fig. 1 .

 1 Fig. 1. Mutant frequencies and cell survival for diploid yJC 2 yeast cells exposed to several

Fig. 2 .

 2 Fig. 2. Genotoxic activity of a sediment organic extract from the upper Seine estuary

  Figure 1

  Figure 2

Acknowledgements

This study was supported by the French National Program of Ecotoxicology (PNETOX III) and the Seine-Normandy Water Agency.


Till Luckenbach
email: till.luckenbach@ufz.de

Rolf Altenburger

David Epel

Teasing apart activities of different types of ABC efflux pumps in bivalve gills using the concepts of independent action and concentration addition

Keywords: Concentration addition, Independent action, Mixtures, MRP, Multixenobiotic resistance, Mytilus californianus, P-gp

Fluorescent dyes and inhibitor compounds are commonly used to detect activity of multixenobiotic resistance (MXR) efflux pumps in marine invertebrates. We here address the question whether compounds acting as specific inhibitors of certain mammalian transporters can be used in dye efflux assays to distinguish different transporter activities in gill tissue from a marine mussel. We quantified effects of PSC833, a specific inhibitor of mammalian P-gp (P-glyoprotein, ABCB1), and MK571, which blocks MRP (Multidrug resistance associated protein, ABCC) type transporters, on calcein-am efflux in gill tissue of Mytilus californianus. Calcein-am acts as a substrate of both P-gp and MRP. Effects of single compounds and mixtures were determined and combined effect models predicting independent action (IA) and concentration addition (CA) of the chemicals were applied.

Effect values predicted by IA showed better correspondence with the experimentally obtained data. This indicates that the inhibitor compounds target different mechanisms of calcein-am efflux and points to P-gp and MRP activities in mussel gills. Our approach could be a simple way for identifying the efflux transporter types targeted by chemosensitizers, including environmentally relevant compounds, in native tissues from marine invertebrates.

Multixenobiotic resistance (MXR) is conferred by membrane-based proteins from the ABC (ATP binding cassette) transporter family. These proteins act as efflux pumps against multiple, chemically unrelated compounds and provide a cellular defense against toxicants. High activity of these transporters has been found in a range of species and this activity can protect against xenobiotics.

In earlier studies, transporter activity in aquatic organisms was mainly associated with P-glycoprotein (P-gp) (Kurelec, 1992). However, more recent papers show that the MXR system of aquatic organisms could arise from many different transporter types. For instance, the genome of the sea urchin Strongylocentrotus purpuratus contains 65 ABC transporters, with many being homologous to known mammalian xenobiotic transporters (Sodergren et al., 2006). Transcripts and activities of different xenobiotic transporter types were also identified in tissues from bivalves (Luedeking and Koehler, 2002, Luckenbach and Epel, submitted), including full-length P-gp and MRP type transcripts with 40-49% amino acid identity with human P-gp and 33-45% identity with human MRP transporters (Luckenbach and Epel, submitted).

We here ask whether one can distinguish the activities of these different transporters from the mussel Mytilus californianus with specific inhibitors of mammalian ABC transporters. Studies on mammalian transporters indicate specific inhibition of P-gp by PSC833 (Mayer et al., 1997) and of MRP by MK571 (Gekeler et al., 1995). However, functional properties of homologous transporters may vary among species and it is therefore not clear whether those compounds are also useful to determine specific activities of transporters in other taxa.

Transporter activity was assessed using calcein-am (c-am) (Molecular Probes; Eugene, OR, USA), which is a substrate of both the P-gp and MRP transporters (Essodaigui et al., 1998). The compound is non-fluorescent, but once it enters the cell it is hydrolyzed by endogenous esterases and fluorescent calcein is formed. This fluorescent product is trapped in the cell since it is a poor substrate for these transporters.

Accumulation of calcein in gill tissue was determined in the presence of different concentrations of PSC833 (kind gift of Novartis; Basel, Switzerland) (range 0.1 -20 µM) and MK571 (Cayman Chemicals; Ann Arbor, MI, USA) (range 0.1 -50 µM).

Additionally, combinations of PSC833 and MK571 (range 0.2 -20 µM) were applied as equimolar mixtures (e.g., the mixture with a total of 10 µM of inhibitor contained 5 µM of each, PSC833 and MK571). The dye efflux assay was performed according to Luckenbach and Epel (2005). Briefly, equally sized discs (5 mm diameter) prepared from gill tissue of a freshly dissected mussel were incubated in filtered seawater (FSW) for 90 min at 15°C with 0.25 µM c-am and varying concentrations of test compounds and their mixtures along with a control with c-am only. The discs were then rinsed, homogenized with a lysis buffer (10 mM KCl, 1.5 mM MgCl 2 , 10 mM Tris HCl, pH 7.4) and centrifuged to remove tissue debris. Fluorescence of the tissue extracts with calcein was then measured with a microplate reader (emission: 530 nm, excitation: 485 nm). Each concentration was tested in at least 3 independent experiments with gill tissue from different animals.

The effects of the PSC833 and MK571 mixtures were then analyzed using models that predict actions of chemical mixtures by concentration addition (CA) and by independent action (IA) based on the dose-response functions of the individual compounds (Fig. 1). The CA model is based on the pharmacodynamic assumption that the mixture components act through a similar mechanism, whereas IA describes effects by components acting on different targets (Altenburger et al., 2003). Their usefulness as a pharmacological probe for similar or dissimilar action has recently been demonstrated (Altenburger et al., 2005).

As seen in Fig. 1, there was a concentration dependent increase in calcein accumulation in the gill tissue with both inhibitors. This shows presence of transporter activity in the gills that is sensitive to both inhibitors since inhibiting the activities resulted in increased levels of the substrate. Low levels of either inhibitor caused a slow linear increase in calcein accumulation between zero and 2 µM. Above 2 µM however, the PSC833 inhibition continued to increase linearly, whereas the MK571 inhibition had more of an exponential increase. In the lower concentration range inhibition by the PSC833 and MK571 mixture did not largely differ from inhibition by the single compounds, but at concentrations above 2 µM inhibitory effects by the mixture were greater than by any of the single compounds.

Applying the CA and IA modelling it is observed that the inhibition of transporter activity by the PSC833 and MK571 mixture corresponds better to the independent action mode of inhibition (IA) than to the concentration addition mode (CA). This result is consistent with the assumption that PSC833 and MK571 affect two different mechanisms of calcein-am efflux in mussel gill. Although specificities of PSC833 and MK571 for mussel P-gp and MRP need to be corroborated in more specific test systems, such as transfected cells, our finding corresponds to the presence of P-gp and MRP transcripts in gill tissue (Luedeking and Koehler, 2002;Luckenbach and Epel, submitted) and therefore is an indication of both types of activities in the gills.

Our data show that analyzing effects of chemical mixtures based on the concepts of IA and CA enables to distinguish different kinds of efflux in mussel gills. This approach can be useful to identify transporters targeted to different chemicals and also assess the contribution these transporters make to MXR. The approach will also be useful in risk assessment to determine which components of the MXR defense are involved in

Figure caption

Fig. 1 .

 1 Fig. 1. Effects of pharmacologic ABC transporter inhibitors PSC833 and MK571 on

  Figure 1

Acknowledgements

This publication was supported in part by the Deutsche Forschungsgemeinschaft (DFG), the National Sea Grant College Program of the U.S. Department of Commerce's National Oceanic and Atmospheric Administration under grant R/CZ-182 through the California Sea Grant College Program, and the California State Resources Agency.


Xavier Denier

Jérome Couteau

Magalie Baudrimont

Elisabeth M Hill

Jeanette Rotchell

Christophe Minier

In vitro study of the effects of cadmium on the activation of the estrogen response element using the YES screen

Keywords: Cadmium, Yeast, Estradiol, Estrogen-receptor, Endocrine disruption

Estrogenic potential of environmental samples is frequently assessed using receptorbased functional assays. Using the Yeast Estrogen Screen (YES) developed by Routledge and Sumpter, we assessed the ability of cadmium to activate the estrogen receptor-mediated response. No induced transcriptional activity was observed with a range of CdCl2 concentrations (1nm-1mM). But, when combining cadmium with the model compound 17βestradiol, cadmium was able to significantly potentiate the induced estrogenic response for concentrations ranging from 15 nM to 1 µM. A maximal effect was observed at 0.5 µM with a ten fold reduction of the 17β-estradiol EC50.

Corresponding author: Email address: xavier.denier@univ-lehavre.fr (X. Denier) Important efforts have been made for the development of screening strategies to evaluate and understand endocrine disruption by estrogenic compounds in human and wildlife. Various in vitro tests have been developed and are used to assess the estrogenicity of isolated compounds or complex environmental mixtures. One of the most widely used in vitro tests is the yeast estrogen screen (YES; Routledge and Sumpter, 1996).

Metals, and particularly cadmium, have already been shown to act in vivo as endocrine disruptors. But action through a direct estrogenic pathway is still unclear. Several studies have found an in vitro estrogenic activity of cadmium. Cadmium induced growth in the human breast cancer cell line MCF-7 (Garcia-Morales et al., 1994, Choe et al, 2003). Stoica et al. (2000) showed that cadmium is able to activate the estrogen receptor by binding to the hormone binding domain of the receptor. Choe et al. (2003) and Wilson et al. (2004) also observed induction of an estrogen receptor-dependent transcriptional expression assay (MCF-7-ERE or TD47 cell lines).

But contradictory results have been reported. Le Guevel et al. (2000) used yeast cells transfected with rainbow trout estradiol receptor and observed that cadmium alone is unable to activate the receptor. Silva et al. (2006) using the YES and E-screen assays (MCF-7 proliferation) concluded a lack of activity of cadmium in both tests. In view of these contradictory results, this study aimed to assess the estrogenicity of a wide range of cadmium concentrations and to evaluate its potential to modulate the 17--estradiol activity in the YES assay.

CdCl 2 was obtained from Sigma-Aldrich. From 10mM CdCl 2 stock solution, 21 different concentrations ranging from 1 mM to 1 nM were obtained by serial dilution in water reference for the test. 10 µl of each dilution were added to wells. "Metal alone" and blank wells received 10 µl ethanol instead. 20 µl of cadmium solutions were then added to wells, while "estradiol alone" and blank wells received 20 µl water. Finally, 180 µl of assay medium and yeast were added to the wells. Plates were incubated 48 hours at 30°C before absorbance reading. In the yeast estrogen screen, the activation of estrogen receptor leads to its binding to estrogen response element then subsequent synthesis of -galactosidase. The galactosidase hydrolyses the chlorophenol red-galactopyranoside present in the medium. The chlorophenolred produced was measured by absorbance at 540nm. Absorbance at 620 nm allows correction for turbidity: corrected absorbance = Abs 540nm -(Abs 620nm -Abs 620nm blank).

We first tested the ability of CdCl 2 to directly activate the ER-mediated response in the YES assay. Fig. 1 indicates that none of the tested cadmium concentrations resulted in any significant galactosidase activity as previously reported in other studies using yeast cells (Le Guevel et al., 2000;Silva et al., 2006). Although the human estrogen receptor and the estrogen response element have been inserted in the yeast strain, it appears that the YES assay can generate results that are different from those obtained with mammalian cell lines. Indeed several studies have reported that cadmium can activate the human reporter by interacting with the ligand-binding domain and exerts effects characteristic of the endogenous hormone in human cells (Stoica et al., 2000;Choe et al., 2003;Wilson et al., 2004).

In a second experiment, we assessed the putative joint action of cadmium and the model compound 17β-estradiol (E2). 21 different concentrations (from 1nM to 1 mM) of cadmium were tested in combination with 11 E2 concentrations (10pM to 10nM). Results are shown in Fig. 2. Different effects were observed according to the cadmium concentration in the test medium. Cadmium concentrations from 1 to 500 nM potentiated the estrogenic activity in a dose-dependant manner (although significant differences were only obtained for concentrations of 15 nM CdCl 2 and higher). For each curve, the maximum induction was not affected but the EC 50 decreased. Maximum effect (EC 50 reduced to one tenth of the control value) was observed with 0.5 µM CdCl 2 .

Cadmium concentrations from 1 to 20 µM reduced the enzymatic activity in a dose dependent manner. The decreased activity appeared to be, at least, partly due to toxicity as indicated by the reduced absorbance at 620nm (not shown). The curve obtained with 2 µM CdCl 2 did not differ significantly to the standard curved obtained with E2 alone but a rapid decrease in galactosidase activity was observed for higher concentrations. Finally, at cadmium concentrations from 30 µM to 1 mM CdCl 2 , yeast growth was inhibited.

These results show that cadmium can either potentiate the estrogenic response or have toxic effects on yeast metabolism. Reports from Silva et al. (2006) and Le Guevel et al. (2000) did not show potentiating effect of cadmium on their yeast-based assay but significant decrease in the response possibly due to toxicity. It should be stressed that only a few cadmium concentrations together with E2 were tested in these studies. Le Guevel and colleagues incubated the cells in 10 µM cadmium only, a concentration which was already toxic in this study. Cadmium toxicity at 10 µM or higher concentrations of cadmium has recently been reported either in the YES or in the E-screen assays.

The reason why cadmium cannot directly activate the response mediated by the human estrogen receptor in the YES assay is unclear. But results obtained in this study indicate that cadmium can potentiate the estrogenic response induced by the interaction of E2 with the ER. This effect was not due to an enhanced metabolic activity as cadmium did not significantly increase yeast growth at effective concentrations (data not shown). Predki and Sarkar (1994) showed that cadmium, copper and chromium can interact with the zinc finger DNA-binding domain of the ER and activate gene transcription. These effects may have also occurred in our

Fig. 1 .

 1 Fig. 1. Effect of 17 -estradiol (E2) and cadmium in the recombinant yeast estrogen screen

Fig. 2 .

 2 Fig. 2. Variation of yeast estrogen screen response to 17 -estradiol with increasing

  ). Added cadmium concentrations are indicated in the table for each numbered curves. Plain lines indicate a decreased E2 EC50, dotted lines indicate an increased E2 EC50 with increasing cadmium concentrations. Curves 2, 3, 4, 6 and 7 are not displayed to gain clarity. Curves with E2 + cadmium were derived from two independent experiments.

	Figure 2
	Figure 1
	3 3.5											curve	cadmium
												number	concentration (M)
	3											1 2	9.54E-10 1.91E-09
												3	3.81E-09
	1.5 2 2.5 1 Corrected absorbance 1.5 2 Corrected absorbance 2.5			Estradiol 10 9	11	8	5	12 1	13	14	15	14 15 16 17 4 5 6 7 8 9 10 11 12 13	7.81E-06 1.56E-05 3.13E-05 6.25E-05 7.63E-09 1.53E-08 3.05E-08 6.10E-08 1.22E-07 2.44E-07 4.88E-07 9.77E-07 1.95E-06 3.91E-06
												16-21	18 19	1.25E-04 2.50E-04
	1 0.5 1.E-11			1.E-10	Estradiol concentration (M) 1.E-09	Cadmium	1.E-08	20 21	5.00E-04 1.00E-03
	0.5
	1.E-12	1.E-11	1.E-10	1.E-09	1.E-08 Concentration (M) 1.E-07	1.E-06	1.E-05	1.E-04	1.E-03	1.E-02
	.

Acknowledgments

This work is part of the European Interreg III program "Risk Analysis associated with Endocrine Disruption in the Manche Regions (RAED)".


Alex T Ford
email: alex.ford@thurso.uhi.ac.uk

Christine Sambles

Peter Kille

Intersexuality in Crustaceans: Genetic, Individual and Population level effects

Keywords: Endocrine disruption, Invertebrates, Biomarkers, Microarray, Population modelling, Intersexuality

Scientists investigating toxicants such as endocrine disrupting chemicals (EDCs) at the cellular at the sub-cellular level are often faced with criticisms as to how these effects can be extrapolated to the level of individuals and their populations. This report aims to provide an overview of the studies undertaken on crustacean model, Echinogammarus marinus LEACH (AMPHIPODA), and intersex phenotypes, at the individual and population levels, and provide additional emergent data at the genomic level. These, normal and intersex, males and females have been investigated by cross-hybridisation microarray analysis and specific sexually dimorphic genes and corresponding properties identified between each sexual phenotype. The morphology, physiology and histology of these intersexes have been investigated in detail and a number of reproductive costs have been identified including reduced fecundity and fertility. These costs have been incorporated into a population model and simulated over a ten-year period to ascertain how different levels of intersexuality affect the stability of populations. Based on the information

ACCEPTED MANUSCRIPT

2 gained through study of intersex models (with known endocrine dysfunction) together with the substantial quantity of historical data relating to effects of chemicals on amphipod fecundity, growth and mortality, the development of appropriate biomarkers is nearer to being assessed from the level of genes to that of the population.

Many of the studies of endocrine disruption (ED) have arisen through the discovery of sexual abnormalities (including intersexuality) in wildlife (Colburn et al., 1996). One of the problems faced by toxicologists investigating ED in invertebrates is that the genetics and endocrinology are not as fully understood as those in vertebrate models [START_REF] Defur | Endocrine Disruption in Invertebrates: Endocrinology, Testing and Assessment[END_REF].

Consequently, the search for and pace of development of suitable biomarkers for laboratory and field monitoring has been relatively slow. Invertebrates do however provide some advantages over vertebrates due to their simple and inexpensive methods of culture, as well as fast generation times making them more amenable to population studies.

Male and female amphipods can be distinguished from each other by the presence of genital papillae (males) and brood plates/oostegites (females). Intersex specimens can be observed displaying both genital papillae and brood plates. Intersexuality in the amphipod E. marinus exists as two phenotypes, namely, intersex male and intersex female (for descriptions see Ford et al., 2004a). These abnormal specimens provide a unique opportunity to study varying types of endocrine dysfunction and develop endocrine specific biomarkers at several levels of biological organisation. Ford et al. (2004a) observed higher proportions of intersex E. marinus at sites associated with pollution when compared to reference sites. These intersexes have subsequently been reported to be caused by feminising microsporidian parasites possibly enhanced through polluted conditions (Ford et al., 2006). Recently, intersex has been demonstrated to be induced chemically in Crustacea by juvenile hormone mimics (Olmstead and LeBlanc, 2007).

Intersexuality can also be induced in amphipods by the ablation or implantation of androgenic glands, from/into males and females, respectively (Charniaux-Cotton, 1958).

A number of costs associated with intersexuality at the individual level have been reported in E. marinus including lower fecundity and fertility; and delayed maturation (Ford et al., 2004b; Table 1). Morphologically, histologically and behaviourally intersex specimens separate into two phenotypes, intersex male and intersex female. When comparing the morphology of normal male E. marinus from clean and polluted sites with the standard morphology of an intersex male, discriminant analysis revealed normal males from an industrially polluted site in Scotland had a morphology most similar to that of intersex males (Ford et al., 2004a). Considering feminising parasites are infrequently found in normal male E. marinus, the authors suggested that subtle endocrine disruption maybe occurring, possibly relating to effects associated with the androgenic gland. Martins et al. (2002) modelled populations of E. marinus in response to environmental parameters (salinity, temperature) based upon a population with no intersexes and equal numbers of males and females (1:1). In view of the costs associated with intersexuality, this model was subsequently adapted by to include proportions of intersex females and the reproductive costs associated with intersexuality (Ford et al., 2007). Simulations were undertaken with varying sex ratios, fecundity, mortality and intersex proportions and revealed that the inclusion of intersexes can have dramatic effects on the population. (for details please refer to Ford et al., 2007; summarised in Table 1).

To date, genomics resources available for Crustacea species encompass a complete genome sequence for the water flea Daphnia pulex (http://wfleabase.org/) together with limited EST (Expressed Sequence Tag) studies for a small number of other crustaceans. Gammarus pulex cDNA libraries were generated representing a range of developmental stages and phases of the moult-cycle for both sexes together with sex-specific subtractive suppression hybridisation libraries. These libraries have been used to generate 9600 full-length directional ESTs, together with 3,840 EST fragments from a male enriched SSH library.

We used these ESTs to generate a glass-slide microarray that demonstrated consistent hybridisation (83% significant signal above background) to total RNA from E. marinus. Normal and intersex male and female E. marinus were sampled from Inverkeithing Bay (Scotland). Whole organisms (n = 24 ) were used to generate total RNA which was subsequently used as a template for cDNA synthesis in the presence of aminoallyl dUTP and subsequently coupled to monofunctional NHS-ester of Cy3. Samples were then hybridised against a common reference oligonucleotide labelled with Cy5. Hybridisations were imaged and spot intensities quantified using image analysis software Spot (http://spot.cmis.csiro.au/spot/index.php). Results were subsequently imported into GeneSpring GX (http://www.chem.agilent.com/) for normalisation and statistical analysis (see Figure 1 for details). Analysis of variance of all samples revealed 582 genes which differed significantly between all groups (normal and intersex males and females).

Subsequent PCA analysis of the transcript profile of these differentially expressed genes within the individual specimens showed a clear separation between normal males and females and intersex groups separated from their normal counterparts (Fig. 1). Interestingly, the clustering of intersex males and females together runs contrary to their phenotypic variation, suggesting a relatively few genes may influence large changes in secondary sexually dimorphic characteristics.

Preliminary analysis of microarray data has revealed a vast array of potential genetic biomarkers for assessing (de)-feminisation and/or (de)-masculinisation in Crustacea. The role of individual genes and their potential for biomarkers of EDCs and toxicants is currently under further investigation. E. marinus provides a unique species for studying intersexuality and currently remains one of the only intersex models to have been studied at the genome level up to that of the population.

Figure caption

Fig. 1: Principal component analysis of individual transcript profiles of normal and intersex males and female E. marinus ( male; i intersex male; female; i intersex female). Array data was normalised by dividing by the common reference channel followed by normalising each gene expression by its average over all samples together with normalising slide distribution to the 50 th percentile. Genes that showed >2 change and statistically significant differences between 'Sex' groups (normal and intersex males and females) were analyzed using a non-parametric test (Kruskal-Wallis test) p-value cut-off 0.05 yielding 582 genes. The expression of these genes was then used in a PCA analysis of the individual specimen transcript profiles. [* outlier]. • Reduced androgenic gland activity in intersex (male and female) versus normal males quantified by MALDI-TOFF spectrometry (Ford et al., 2005). Individual

• External morphometric differences between normal and intersex specimens (Ford et al., 2004a). Intersex males externally show the development of 'female' immature brood plates, whilst intersex females externally display 'male' genital papillae.

• Internal differences in gonodal structure between normal and intersex specimens (Ford et al. 2005). Intersex male E. marinus internally display oviducts and vas defera have been observed in intersex females (Ford et al., 2005).Histological analysis has revealed both the inclusion of ovarian and testicular tissues in some intersex amphipods.

• Behaviourally intersex females and intersex males act as do normal females and males, respectively (Ford et al., 2007).

• Normal male E. marinus at polluted sites display close morphology to intersex specimens at polluted sites. Suspected de-masculinisation and anti-androgenic activity (Ford et al., 2004a).

• Reduced fecundity and fertility in female intersexes. Intersex female E. marinus produce ~20% fewer eggs than normal females. 10% fewer eggs from intersex females develop fully as embryos compared to normal females (Ford et al., 2004b).

• Increased size and delayed maturation in intersexes. Intersexes grow ~10% larger than normal specimens (Ford et al., 2004b).

• Reduced pairing success in intersex females. Possible pheromone dysfunction (Ford et al., 2004b;and papers within).

• Reduced fertility in intersex male Corophium volutator (Barbeau and Grecian, 2003). Population

• Simulated population extinction within approximately 6 years if 5% of normal females are replaced by intersex females (Ford et al., 2007).

• Rapid population growth if normal males are replaced by intersex females. This population growth however can be negated by ~10% reductions in normal female fecundity (Ford et al., 2007).

  Figure 1

Olmstead, A.W., and Leblanc, G. A. (2007). International Journal of Biological Sciences, 3, 77-84.


Letizia Marsili

Silvia Casini

Daniela Bucalossi

Serena Porcelloni

Silvia Maltese

Maria Cristina

Maria Cristina Fossi

Use of immunofluorescence technique in cultured fibroblasts from Mediterranean cetaceans as new "in vitro" tool to investigate effects of environmental contaminants

Keywords: Mediterranean cetaceans, Fibroblast culture, Immunofluorescence technique

de niveau recherche, publiés ou non, émanant des établissements d'enseignement et de recherche français ou étrangers, des laboratoires publics ou privés.

Use of immunofluorescence technique in cultured fibroblasts from Mediterra nean cetaceans as new "in vitro" tool to investigate effects of environmental contaminants Letizia Marsili, Silvia Casini, Daniela Bucalossi, Serena 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. Immunofluorescence was quantified with a specially designed Olympus macro, DetectIntZ.

A major result was the possibility of using this "in vitro" assay to quantify induction of endogenous proteins.

Xenobiotic contaminants *Corresponding author: Email address: marsilil@unisi.it (L. Marsili)

Cetaceans, especially odontocetes, are particularly subject to chemical stress from xenobiotic compounds, as they accumulate large quantities of these contaminants (Marsili and Focardi, 1996;Marsili, 2000;Fossi and Marsili, 2003). Among the various techniques applied to integument biopsies, developed at the University of Siena (Fossi et al., 2003;Fossi et al., 2006), here we present the immunofluorescence technique in cultured fibroblasts as a new "in vitro" tool to explore the susceptibility of these marine mammals to different xenobiotics.

The development of a non-lethal sampling method to obtain viable tissue samples for fibroblast cell cultures from integument biopsies of free-ranging cetaceans was described in Marsili et al. (2000). Using the indirect immunofluorescence assay, we evaluated qualitative and quantitative presence of CYP1A1 and CYP2B induced by some POPs (DDTs and PCBs)

and emerging contaminants (PBDEs) (Nims et al., 1998;Sierra-Santoyo et al., 2000;Meredith et al., 2003, Stoker et al., 2005) A major result was the possibility of using this "in vitro" assay to quantify induction of endogenous proteins such as CYP1A1 and CYP2B: the measurement of immunofluorescence intensities was used to determine protein levels and different induction phenomena correlated with the contaminant treatment.

In this paper we present the preliminary results of two odontocete species representative of the Mediterranean Sea: striped dolphin (Stenella coeruleoalba) and bottlenose dolphin (Tursiops truncatus). The fibroblasts were treated with the OC mixture (Fig. 1 and Table 1) and PBDE (Table 1). The results confirm the capability of this methodology to detect CYP1A1 and CYP2B induction in the striped dolphin, particularly at the highest treatment dose; only CYP2B induction with OC mixture at the intermediate dose and only CYP 1A1 with PBDE at the highest treatment dose was revealed in the bottlenose dolphin.

In conclusion, this methodology, applied to cultured fibroblasts of cetaceans, appears as a powerful "in vitro" technique to assess susceptibility of these marine mammals to different xenobiotic compounds.

Sierra-Santoyo, A., Hernandez, M., Albores, A., and Cebrian, M. E. (2000). Toxicological

Sciences, 54, 81-87.

Stoker, T.E., Cooper, R.L., Lambright, C.S., Wilson, V.S., Furr, J., Gray, L.E. ( 2005).

Toxicology and Applied Pharmacology, 207, 78 -88.

Figure caption

  Please cite this article as: Marsili, L., Casini, S., Bucalossi, D., Porcelloni, S., Maltese, S., Fossi, M.C., Use of immunofluorescence technique in cultured fibroblasts from Mediterranean cetaceans as new "in vitro" tool to investigate effects of environmental contaminants, Marine Environmental Research (2008), doi: 10.1016/ j.marenvres.2008.02.068

  12 h). Using the indirect immunofluorescence assay, we detected endogenous proteins induced by different contaminants. Here we present the method used for qualitative and quantitative evaluation of cytochromes P450 (CYP1A1 and CYP2B) induced by some POPs (DDTs and PCBs) and emerging contaminants (PBDEs) in fibroblast cell cultures of striped dolphin (Stenella coeruleoalba) and bottlenose dolphin (Tursiops truncatus).

  . Cells were treated for 48 h with contaminants in sterile culture plates. Treatments were an OC mixture of Arochlor 1260, pp'DDT and pp'DDE in DMSO (0.05%) at three doses: 1 µg/ml, 5 µg/ml and 25 µg/ml, plus a DMSO (0.05%) chemical control and a mixture (BDE-MXE, Wellington, Canada) containing 27 PBDEs, from mono-to deca-brominated, in nonane (0.01 µg/ml) at three doses: 0.1 µg/ml, 0.05 µg/ml and 0.01 µg/ml, plus a nonane (0.01 µg/ml) chemical control. After fixing and extraction with methanol (4 min) and acetone (4 min) at -20°C, they were washed three times in PBS for 10 min. Then we conducted a first reaction with primary polyclonal antibodies: goat anti rabbit cytochrome P450 1A1 and goat anti rabbit cytochrome P450 2B (Oxford Biochemical Research, Oxford, USA), diluted 1:500 for 1A1 and 1:100 for 2B, for 2 h. The cells were washed three times in PBS for 10 min and treated with the secondary antibody (Alexa Fluor 594 goat anti rabbit IgG (H+L)) labelled with red-fluorescent Alexa Fluor 594 dye, diluted 1:400, in the dark. The coverslips were again washed three times in PBS for 10 min. Fluorochrome was detected using 15µl per coverslip of a solution containing 40% CITIFLUOR, which is anti-fading solution that stabilizes fluorescence, and 60% PBS, whereas 3µl DAPI diluted 1:500 (ex. 345 nm -em. 455 nm; AT selective) was used as marker of chromatin for cell count. The reaction was read using an Olympus mod. BX41 fluorescence microscope with BX-URA2 HBO100W fluorescence device and U-MSWG2 dichroic cube excitation filter for green and U-MWU2 dichroic cube filter for UV. Images were obtained with an Olympus mod. DP-70 digital telecamera. Immunofluorescence was quantified with a specially designed Olympus Soft Imaging Systems macro, DetectIntZ, which works with the image acquisition, processing and analysis system, analySIS^B (Olympus). The image analysis procedure has the objective of quantifying, with an adimensional index generated for this purpose, the amount of Alexa Fluor 594 localized in the membrane of cytoplasmatic area of sample cells. The sample cells are imaged using DAPI and this image is presented to the operator for threshold selection of cytoplasmatic and nuclei Region of Interests (ROIs) across the field. The procedure then utilizes these ROIs to measure fluorescence intensity of Alexa Fluor 594 sample cell and summarizes the results in a worksheet. The system generates index values which are unitless until compared with other units, such as number of cells to obtain mean fluorescence per cell or the area in which it is calculated to obtain mean fluorescence per mm 2 . Images are all obtained with a magnification of 20X, a calibration of 0.65 µm/pixel and a resolution of 1360 x 1024 x 8 pixel. Exposure times were maintained fixed while reading the CYP1A1 and CYP2B for each species and for each treatment. A series of images of each slide was acquired so that a minimum of 250 cells/slide could be counted. The total fluorescence revealed by the program is divided by number of cells to obtain Arbitrary Unity of Fluorescence (AUF) per cell.Nine slides for CYP1A1 and CYP2B were made for each culture, making a total of 18 slides for each toxicological treatment: of the nine, one was a blank (only primary and secondary antibodies), one was a secondary blank (only with secondary antibody), one was a chemical blank (treated with contaminant vehicle), two were for the first dose, two for the second dose and two for the third dose of contaminants. The blank enabled the natural presence of cytochromes in cultured fibroblasts to be checked. The secondary blank enabled validation of the dose of secondary antibody without cross reaction as the primary antibody was absent.

Fig. 1 .

 1 Fig. 1. Immunofluorescence of CYP1A1 and CYP2B in cultured fibroblasts of striped dolphin

  Fig.1

Table 1 .

 1 Mean values of immunofluorescence of CYP1A1 and CYP2B revealed in cultured fibroblasts of striped dolphin and bottlenose dolphin treated with the OC and PBDE mixture. The immunofluorescence is expressed in index numbers respect to solvent control. Explication about treatment doses in the text.

			CYP 1A1 -Organochlorine mixture
		Solvent control Dose 1	Dose 2	Dose 3
	Striped dolphin	100	93.16	90.42	166.92
	Bottlenose dolphin	100	63.00	96.48	81.43
			CYP 2B -Organochlorine mixture
		Solvent control Dose 1	Dose 2	Dose 3
	Striped dolphin	100	94.07	107.86	219.28
	Bottlenose dolphin	100	88.56	136.23	85.72
			CYP 1A1 -PBDE
		Solvent control Dose 1	Dose 2	Dose 3
	Striped dolphin	100	316.68	399.04	1304.62
	Bottlenose dolphin	100	71.14	102.50	131.92
			CYP 2B -PBDE
		Solvent control Dose 1	Dose 2	Dose 3
	Striped dolphin	100	186.09	212.86	1776.41
	Bottlenose dolphin	100	71.46	69.24	94.15

Acknowledgements

Thanks to Laura Carletti for laboratory technical support, Prof. Tommaso Renieri, Fabrizio Cancelli and Centro Studi Cetacei onlus for sampling stranded animals. This project was supported by the Italian Ministry for Environment and Territory (Pelagos Project, ACCOBAMS Project).


Mathijs G D Schipper

Nicholas H B M Smit

A Dick Kaag

Vethaak

Cor A Schipper
email: cor.schipper@deltares.nl

Mathijs G D Smit

Nicholas H B M Kaag

A Dick

A weight-of-evidence approach to assessing the ecological impact of organotin pollution in Dutch marine and brackish waters; combining risk prognosis and field monitoring using common periwinkles (Littorina littorea)

Keywords: Gastropods, harbours, tributyltin, intersex, risk assessment, sediments, SPM

A weight-of-evidence approach to assessing the ecological impact of organotin pollution in Dutch marine and brackish waters; combining risk prognosis and field monitoring using common periwinkles ()

Introduction

Tributyltin (TBT) was first used as an antifouling agent in the late 1960s. It proved to be very effective, but also extremely toxic to aquatic life (Champ and Seligman, 1996). TBT effects were first recognized in the marine environment in the 1970s and attracted particular attention in the 1980s, when the effects of TBT started to become apparent in the Netherlands and elsewhere [START_REF] Marquenie | Effecten van verontreinigingen in zoutwater ecosystemen (Effects of pollutants in marine ecosystems)[END_REF]Champ and Seligman, 1996). In the 1980s, there was a sharp decline in the population of dog whelk (Nucella lapillus) in the Delta area of the south-western Netherlands and in the 1990s the occurrence of imposex was demonstrated among whelks (Buccinum undatum) in the same area [START_REF] Mensink | Assessment of imposex in the common whelk, Buccinum undatum (L.) from the Eastern Scheldt, the Netherlands[END_REF] as well as in the open North Sea [START_REF] Hallers-Tjabbes | Imposex in whelks (Buccinum undatum) from the open North Sea: Relation to shipping traffic intensities[END_REF].

Given the concern about the environmental effects of TBT in the Dutch marine environment a lot of data has been collected in the past. The available data relate to TBT concentrations in water, sediments and suspended matter and field measurements of intersex in, and the distribution and abundance of gastropods. The Chemical Toxicity Test (CTT) has shown TBT to be the most problematic of all the chemicals tested in toxicological terms in assessing whether dredged material may be relocated in the marine environment [START_REF] Schipper | Towards implementation of a chemistry-toxicity testing tool[END_REF]. Up to now, in assessing the impact of TBT on the ecosystem, the available information has been evaluated separately. This paper describes the results of an integrated assessment including all the information collected in recent years. Comparable to the sediment quality Triad concept, which uses information from the toxicity line, the biological field line and the chemistry line, this assessment is also based on results of three lines of evidence (LOEs) (Chapman, 1996;[START_REF] Rutgers | Site-specific ecological risks: a basic approach to the function-specific assessment of s soil pollution[END_REF]. However, instead of using measurements from multiple species in the biological field line, as normally done in the Triad approach, the assessment in the present study uses information from one indicator species only. In the present study the LOEs are primarily represented by measurements of: (i) chemical concentrations (exposure levels), (ii) (eco)toxicological indicators and results from ecotoxicological laboratory experiments, and (iii) abundance of indicator species Littorina littorea in the field. The Potentially Affected Fraction of species (PAF) was used to represent the ecological risk and quantify the degree of TBT stress on the ecosystem (Aldenberg et al., 2002). All three LOEs provide indications of the ecological health status related to TBT pollution, though all approaches based on this level include uncertainties: i. TBT pollution is directly chemically measured in samples collected in the field. However, the relevance of chemical analysis of sediments to the quality of water bodies is not always clear.

Processes such as bioavailability determine the actual relation between chemical concentrations and possible ecological effects. Ecological risk prognosis, using accepted generic procedures, predicts the health status related to the stressor under consideration based on the measured or estimated chemical concentrations assuming equilibrium partitioning and constant bioavailability. The prognosis will provide the likelihood of effects to occur. Whether or not these effects will actually occur depends on location specific conditions.

ii. Biological effect parameters can be used to assess the potential and actual impacts of TBT on marine coastal life. However, it can be difficult to link biological effects to exposure to a single contaminant. In this study, the degree of intersex index (ISI) in the common periwinkle, L. littorea, was selected as an indicator of TBT pollution, as this species is the most common prosobranch species on hard substrates in Dutch coastal waters. Intersex in this species has proven to be sensitive to the degree of environmental organotin pollution in coastal waters [START_REF] Oehlmann | Imposex in Nucella lapillus and intersex in Littorina littorea: Interspecific comparison of two TBT-induced effects and their geographical uniformity[END_REF]. Although periwinkles are not the most sensitive gastropod and ISI is not a rapid response, measuring ISI levels in L. littorea in the field is still a useful way of assessing effective levels of TBT pollution. Several studies have shown that ISI can be used as biological tool to determine the degree of TBT pollution, which can be expressed as approximate TBT concentrations in seawater [START_REF] Stroben | Hinia reticulata and Nucella lapillus. Comparison of two gastropod tributyltin bioindicators[END_REF][START_REF] Oehlmann | Imposex in Nucella lapillus and intersex in Littorina littorea: Interspecific comparison of two TBT-induced effects and their geographical uniformity[END_REF]. However, the ISI indicator in periwinkles cannot easily be translated to a generic health indicator for the water body under consideration.

iii. Observations from biological effect monitoring at gastropod level [START_REF] Hallers-Tjabbes | Imposex in whelks (Buccinum undatum) from the open North Sea: Relation to shipping traffic intensities[END_REF][START_REF] Mensink | Assessment of imposex in the common whelk, Buccinum undatum (L.) from the Eastern Scheldt, the Netherlands[END_REF]De Wolf, 2001) are used to assess effects of the distribution and contamination levels in the Dutch marine environment. The abundance of L. littorea and the far more sensitive dog whelk, N. lapillus, can be used as indicators. TBT contamination has made the dog whelk a rare species in Dutch coastal waters and it is almost entirely absent from harbours and marinas [START_REF] Harding | TBT imposex survey of the North Sea[END_REF]Van Moorsel, 1996). TBT potentially accumulates in sediments. However, there is no direct link between sediment concentrations and effects in gastropods, as the preferred habitats of some of the most common gastropods (e.g. N. lapillus, L. littorea) are hard substrates, including natural substrates such as rock and artificial substrates such as harbour walls, breakwaters, etc. [START_REF] Oehlmann | Baseline study on intersex in Littorina littorea with recommendations for biological TBT assessment criteria[END_REF].

In this study the results of the LOEs are compared in order to establish their similarity and reduce the uncertainty of the overall assessment. The methodological links between different LOEs must be evaluated so that the results of monitoring effects on L. littorea and risk assessment studies focusing on the impacts of TBT can be compared. This paper describes a WOE approach [START_REF] Smith | Weight of Evidence (WOE): Quantitative Estimation of Probability of Impact[END_REF] focusing on TBT in Dutch harbours and open waters. Focusing on the integration of indicator monitoring and risk prognosis, the present study applies a more holistic approach to impact assessment [START_REF] Suter | Applicability of indicator monitoring to ecological risk assessment[END_REF].

Materials and Methods

Study area

An integrated ecological risk assessment approach including evaluation of three LOEs is applied to the collected data to assess ecological health status related to TBT pollution in Dutch marine and estuarine waters. Figure 1 shows the harbour locations 1 to 30 and the seven open water areas (A to G) from which chemical, (eco)toxicological and/or ecological data have been included in this study.

Chemistry

The top 5 cm of sediment was sampled using a box core at 120 locations along the Dutch coast between 1998 and 2003. Over this period, 106 locations were sampled twice with an interval of three years. The locations covered the main open North Sea waters, the Wadden Sea, the Eastern Scheldt and estuarine areas such as the Western Scheldt and Ems-Dollard (see Figure 1). TBT was also analysed in sediment from a number of harbour locations from 1999 to 2002 (Table 1). The sediments were freeze-dried and homogenized using a Retsch planetary ball mill. 1.5 ml acetic acid, 15 ml methanol and 7 ml hexane were added to 1 g of sediment. This mixture was stirred for 30 min, after which 3 ml of 4 M sodium acetate solution (pH=5) was added to obtain optimal ethylation. Then, 4 ml of 5% sodium tetraethyl borate solution was slowly added using a peristaltic pump over 20 min, during which time the mixture was continuously stirred. The mixture was shaken vigorously and 5 ml 10 M sodium hydroxide solution added to remove boroxin formed at ethylation. About 5 ml of water was added, and after shaking and subsequent centrifugation the hexane layer with the ethylated organotin compounds was collected. This hexane extract was cleaned over a column of 4-gramme aluminium oxide (Alumina B-super deactivated with 10% water) using hexane as an eluent. 1 ml of iso-octane was added to the hexane eluate and this extract was then concentrated to a final volume of 0.5 ml using Kuderna-Danish solvent evaporation. The final extracts were analyzed for ethylated organotin compounds by electron impact (EI) GC-MS using selective ion monitoring. The method uses ethylated Quasimeme standards and tetrapropyltin as internal standard. Quality assurance recovery control with carried out with monopropyltin and tripropyltin in each sample and blank control in every analysis batch. Yearly participation in Quasimeme laboratory proficiency exercises for organotins resulting in satisfactory zscores.

Probabilistic risk assessment methods

The measured TBT-Sn concentrations were used as a basis for a probabilistic risk assessment. In those samples where the fraction organic carbon (f oc ) was reported, the sediment and suspended matter concentrations from the field surveys were normalised to water concentrations, applying the formulas in the European Technical Guidance Document for Risk Assessment [START_REF] Ec | Technical Guidance Document on Risk Assessment in Support of Commission Directive 93/67/EEC on Risk Assessment for new notified substances[END_REF]. The ecological risk expressed as the Potentially Affected Fraction of species (PAF) was calculated in accordance with Aldenberg et al. (2002). This method can be summarised as follows. The calculated water concentrations were compiled in a log-normal distribution representing the spatial variation of TBT concentrations in a specific area. The concentration distribution was then compared to the species sensitivity distribution (SSD) for TBT. The SSD describes the average sensitivity of species and the variation among species [START_REF] Posthuma | General introduction to Species Sensitivity Distributions[END_REF]. The SSD for TBT reported by [START_REF] Lepper | Towards the derivation of quality standards for priority substances in the context of the Water Framework Directive[END_REF] is based on toxicity data for 29 marine and freshwater species representing seven taxonomic groups (algae, annelida, crustacea, echinodermata, insecta, mollusca, pisces). The SSD for TBT-Sn used in this study is provided in the supplementary data for this paper.

The PAF indicates the likelihood that adverse ecological effects will occur as a result of TBT exposure in a particular area. This value can be interpreted as the probability that a randomly selected exposure concentration will exceed species sensitivity. The probabilistic risk assessment was performed on the (1995). ISI is expressed as the average intersex stage within the population.

Overview of chemical and biological datasets

Results

Observed and predicted exposure levels

Predicted ecological risk

The two distributions presented in Figure 2 were used as input for a probabilistic risk assessment as described by Aldenberg et al. (2002). This resulted in an expected total ecological risk for open waters of 6.7% and 19.1% for harbours. This implies that, on average, for open waters it is likely that 6.7 % of species will be negatively affected as a result of exposure to TBT. For the harbours this fraction is 19.

Comparison of the LOE results

To check the correspondence between the maximum ISI levels that can be expected in a harbour and the ecological risk, PAF values have been plotted against the calculated ISI value corresponding to the 95 th percentile of the exposure concentration distribution (Figure 3). The 95 th percentile was chosen because this value corresponds to the most polluted part of the harbour or water body corresponding to a maximum ISI value that might be encountered during field monitoring. In ecological risk prognosis a Potentially Affected Fraction of 5% of species is generally considered acceptable (e.g. Van Straalen and Denneman, 1989;Aldenberg and Slob, 1993;[START_REF] Newman | Applying species-sensitivity distributions in ecological risk assessment: Assumptions of distribution type and sufficient numbers of species[END_REF]Van der Hoeven, 2001;[START_REF] Ec | Technical Guidance Document on Risk Assessment in Support of Commission Directive 93/67/EEC on Risk Assessment for new notified substances[END_REF].

This study shows that at this 5% level intersex in L. littorea is unlikely to occur in the field. But even at higher risk levels, up to approximately 10%, ISI levels in the field are not expected to exceed the ISI signalling level of 0.3 [START_REF] Oehlmann | Baseline study on intersex in Littorina littorea with recommendations for biological TBT assessment criteria[END_REF]. This study showed that high ISI levels up to a maximum of 3.35 are expected in the field only when the predicted risk level exceeds 20%. These results indicate that either the ISI measurements in L. littorea are not sensitive enough to assess ecosystem health, or the risk prognosis is too conservative (predicting risks while there are no actual effects).

In order to check whether the conservatism in the risk assessment is a result of the estimated concentrations (chemistry LOE), the calculated ranges of ISI values based on the 5 th and 95 th percentile of TBT concentrations are compared by the observed ISI values in the field (Ecotoxicity LOE). If there is a good correspondence between calculated and observed ISI, it can be concluded that the concentrations are a proper basis for the risk calculations as well. Figure 4 intersex levels are predicted for three harbours (locations 8, 14 and 18) where no periwinkles were found at some stations. In IJmuiden (location 15) and Delfzijl (location 1), where periwinkles were in low abundance, the maximum calculated ISI level is around 1.5.

Discussion

In the present study a weight-of-evidence approach derived on the basis of three LOEs has been evaluated in order to better assess the risk from TBT in Dutch harbours and open waters. Based on chemical data, risk values were determined according to Aldenberg et al. (2002) and ISI values were calculated based on the dose-response curve reported by [START_REF] Oehlmann | Baseline study on intersex in Littorina littorea with recommendations for biological TBT assessment criteria[END_REF]. Actual ISI observations were made in the field and the presence of gastropods was recorded. In the overall approach evidence from the different LOEs was combined.

The partition coefficient (K p ) was used for the conversion of TBT concentrations in sediment and SPM to TBT concentrations in water. This K P for TBT was calculated by multiplying the organic carbon partition coefficient (K oc ) with the measured fraction organic carbon (f oc ). Consequently, the K oc value has a strong impact on the final results of the risk prognosis. Generally, in risk predictions the lowest K oc value is applied to calculate concentrations in water [START_REF] Ec | Technical Guidance Document on Risk Assessment in Support of Commission Directive 93/67/EEC on Risk Assessment for new notified substances[END_REF]. This results in a worst-case water concentration, in accordance with the precautionary principle. However, with literature values for the K oc of TBT ranging from 3.0 -6.2 [START_REF] Lepper | Towards the derivation of quality standards for priority substances in the context of the Water Framework Directive[END_REF], it is more appropriate to base an assessment on local measured values. In this study the K oc value was estimated from measured aqueous and sediment concentrations in 30 Dutch harbours evenly distributed over the Netherlands. This resulted in an average K oc value of 5.34 (see also the supplementary data). This outcome is within the range of reported K oc values from the literature [START_REF] Lepper | Towards the derivation of quality standards for priority substances in the context of the Water Framework Directive[END_REF].

The highest ISI value, measured in the Western Scheldt (G), still fell within the predicted ISI range.

Higher ISI values for samples from the Western Scheldt (G) have, however, been reported by De Wolf et al. (2001 and2004), showing a decline from 1.3 in 1998 and 2001 to 0.5 in 2002. This decline was confirmed by the surveys reported here, where an ISI of 0.15 was found at the same location in 2002, using only young adults. This is comparable to the ISI value reported by De Wolf et al. ( 2004), as these were adults that had been exposed some years before the young adults were exposed. In contrast to the trend in the ISI, no downward trend in the sediment-derived water concentrations from 1998 to 2003 was observed in the Western Scheldt (G). We can conclude from the data that observed ISI levels in open waters are all in the lower range of predicted ISI levels.

Combination of LOE results and WOE

The results of the risk predictions confirm that L. littorea is not the most sensitive species regarding to TBT pollution [START_REF] Oehlmann | Imposex in Nucella lapillus and intersex in Littorina littorea: Interspecific comparison of two TBT-induced effects and their geographical uniformity[END_REF][START_REF] Ospar | Harmonisation of Criteria for the Assessment of TBT-specific Biological Effects, presented by the Netherlands[END_REF]. In the field, significant effects on intersex levels are expected for this species only when the risk level exceeds 10%, whereas a risk level of 5% is generally considered acceptable for the protection of ecosystem health (e.g. Van Straalen and Denneman, 1989).

High ISI levels, up to a maximum of 3.35, are expected only when the risk level exceeds 20%. At these high ISI levels, up to a maximum of 3.35, the population of L. littorea is expected to be completely sterile [START_REF] Oehlmann | Baseline study on intersex in Littorina littorea with recommendations for biological TBT assessment criteria[END_REF]. It may even be argued that a risk level of 5% is far too high for specifically acting toxicants such as TBT. Even though L. littorea is not the most sensitive species to TBT pollution, the only species that are more sensitive are other species of prosobranch molluscs. Protecting 95% of species in general would still exterminate the carnivorous neogastropods and also a number of deposit-feeders such as Scrobicularia plana, which can be considered as key species for ecosystem function. Therefore for specifically acting toxicants even stricter criteria should be argued for. In absence of other indicator species L. littorea can serve as a species for monitoring, however, when observed ISI levels in L. littorea do not exceed natural environmental background levels it is still likely that 10% of other species might be affected by TBT contamination.

Both calculated ISI and PAF values are based on the spatial distribution of concentrations. This distribution is derived from TBT concentrations in sediment and SPM. One complicating factor lies in the fact that, in most water systems, periwinkles do not live near the sampled sediment (which is also used to derive the water concentrations), but considerably higher up, on the hard substrates where the macro algae they feed on grow. There might therefore be no direct relationship between TBT in sediment in harbours and ISI values in periwinkles. Instead, the periwinkles are probably predominantly exposed to TBT dissolved in the water and adsorbed to SPM circulating in the water. The fair correspondence between observed and calculated ISI values for L. littorea as presented in this study shows that sediment and SPM concentrations, combined with a realistic partition coefficient (K oc ), can be usefully applied to estimate effective concentrations in the water column.

With this relationship between predicted ISI in periwinkles and the PAF value, the ISI indicator might potentially not only serve as an indicator of TBT pollution but also to represent generic ecological health status. However, this relationship is useful only if many ISI measurements are available for an area, There is good correspondence between observed ISI levels and the predicted ISI ranges for each harbour. Since almost all observed ISI levels fell within the ranges of predicted ISI levels in open waters, it can be concluded that the spatial distribution of TBT concentrations is a good basis for determining predicted ISI levels, and therefore ecological risk.

Periwinkles were found in low abundance at several harbour locations: Delfzijl (location 1); Harlingen (location 8); Den Helder (location 14); IJmuiden (location 15); and Rijnmond (location 18). It is unclear from the observations alone whether this was caused by high levels of TBT or other stress factors (e.g. salinity, excessive sand transport, etc.). The WOE approach can help to indicate the likelihood of TBT causing this absence. High risk values (PAF> 20%) and maximum ISI levels have been predicted for Harlingen (location 8), Den Helder (location 14) and Rijnmond (location 18), so it is not unlikely that high TBT pollution caused the absence of periwinkles from these harbours.

Based on the evaluation of TBT effects on intersex development in the periwinkle L. littorea, and considering ecological health status as a criterion for decision-making, With reference to Table 3, LOEs were derived for L. littorea, representing the species used in the OSPAR monitoring guideline. The relationships between LOEs obtained from comparison of observed effects in the field aid the environmental assessment process. Table 3 brings together the objectives of the ecological health strategy for TBT substances and existing action by water managers.

As gastropods are the most sensitive species to TBT, few of these species are likely to be present in a TBT-contaminated harbour environment. The consequences of the absence of this group for the functioning of the system should be evaluated. Probabilistic risk assessment based on chemical data might provide a generic ecosystem health indicator related to TBT exposure to improve our understanding of the impact of organotin pollution. The results of such an assessment should be validated in the field, in order to reveal the actual field status. Given the relationships presented, observations with ISI as a specific indicator could serve to validate the generic assessment. Inspection of location-specific conditions alongside generic risk prognosis improves the assessment of impacts related to TBT exposure. This study provides evidence of the poor ecological health status of some Dutch harbours as a result of TBT pollution and promotes the use of integrated approaches like the WOE to help practitioners improve ecological impact assessment.

Conclusion

In this study an integrated ecological risk assessment including three LOEs has been applied to assess the risk from TBT pollution in coastal and harbour environments. Only two measured ISI values for coastal environments fell within the predicted ISI ranges (North Sea and Western Scheldt), while the rest of the observations were below the calculated ranges. In harbours there is good correspondence between observed ISI levels and the predicted ISI ranges for each harbour. This study provides evidence of the poor health status of some Dutch harbours as a result of TBT pollution and promotes the use of integrated approaches like the WOE to help practitioners improve ecological impact assessment.

Figure 1

Map showing sampled harbours and open waters (A-G) along the Dutch coast. 1. Delfzijl;2. Emshaven,3. Lauwersoog;4. Schiermonnikoog;5. Holwerd;6. Ameland;7. Terschelling;8. Harlingen;9. Kornwederzand;10. Breezand;11. Den Oever;12. Vlieland;13. Texel;14. Den Helder;15. IJmuiden;16. Noordzeekanaal;17. Scheveningen;18. Rijnmond;19. Herkingen;20. Bruinisse;21. Zierikzee;22. Yerseke;23. Colijnsplaat;24. Hansweert;25. Vlissingen;26. Walsoorden;27. Perkpolder;28. Terneuzen;29. Breskens;30. Appelzak -4 -

  TBT in suspended particulate matter (SPM) was measured four times a year at eight locations in four different open water areas ( A. Ems-Dollard; B. Wadden Sea; C. North Sea; E. Eastern Scheldt; G. Western Scheldt, Figure 1) along the Dutch coast between 1998 and 2003, giving information on fluctuations between and within years. A further 17 locations (data not shown) from the French-Belgian border to the Ems-Dollard estuary on the Dutch-German border were sampled in 2003. Suspended matter was sampled by pumping surface water through a flow-through centrifuge lined with Teflon plates to collect the suspended matter. At harbour locations 23 and 29 (see Figure 1) the seasonal variation of water-dissolved TBT was sampled from 1999 to 2002.

  clustered data for harbours and for open waters, as well as for each harbour and open water separately. Only water systems for which exposure concentrations were measured at two or more locations were selected (all open waters and 19 out of 30 harbours). concentrations. The range of possible ISI values from the 5 th and 95 th percentile of the exposure concentration distributions were calculated applying Formula 1.

  ISI level[-] PEC seawater =concentration of TBT-Sn in seawater [ng.l-1 ] 2.4. Biological effects and abundance surveysPeriwinkles were sampled from 2002 to 2005 at 42 locations along the Dutch coast, representing the main Dutch open waters areas and harbour locations (Figure 1). Grevelingen (D) and Veersemeer (F) lakes were not included. At the sampling locations the presence of L. littorea was recorded at every location randomly selected in 5 to 7 areas of 1 m 2 . If present, at least 40 individuals of L. littorea with a shell height between 15 and 25 mm (approx. 1.5-year-old new adults) were collected. The occurrence of intersex in the periwinkles was analysed within seven days, in accordance with the methods described by Bauer et al.

  Table 1 lists the chemical and biological datasets compiled for this study. The number of 'areas' refers to the open waters presented in Figure 1. The number of 'locations' refers to different sampled harbours within the areas.

  The observed concentration ranges of TBT in sediment or SPM in Dutch open waters and 19 Dutch harbours are shown in Table 2. The highest observed TBT-Sn concentration in Dutch harbours was a factor 30 higher than the highest TBT-Sn concentrations in open waters. The lowest TBT concentrations were more or less in the same range. In order to normalise the sediment and suspended matter concentrations to water concentrations a value for the K oc (organic carbon partition coefficient) needs to be estimated. From the distribution of measured sediment and water concentrations in Dutch harbours a value for the K oc representative of the Dutch situation was estimated. A value of 5.34 gave the best correspondence between observed and predicted water concentrations in Dutch waters (More information on how this value was obtained can be found in the supplementary data). As no temporal trends were observed in the concentrations of TBT-Sn in either the open waters or the harbours, the average concentration of TBT-Sn was taken to represent the exposure at a sampling location. This average concentration was used for further analysis in the risk predictions. Since the concentration distributions of the normalised sediment samples and SPM samples did not differ significantly (Student's t test, equal variances, p=0.42), the normalised data of the two distributions were pooled. Figure 2 presents the overall exposure concentration distributions for TBT-Sn for 138 open water and 270 harbour locations respectively. Dedicated exposure concentration distributions were derived for all areas and locations (open waters and harbours), representing the spatial distribution of water concentrations for each area. In Table 2 the 5 th and 95 th percentile of the distribution represent the range of water concentrations for each area.

  1%. The risk values for the individual open waters and harbours are given in Table 2. Relatively high-risk values were obtained for open waters in the Western Scheldt (G) and Veersemeer (F), while the risk values for the Wadden Sea (B) and the Eastern Scheldt (E) were below 5%. For harbours the highest risk value was predicted at Harlingen in the Wadden Sea (location 8). Four harbours, all small harbours in the Wadden Sea, yielded a risk value equal to or lower than the lowest risk value for the individual open water bodies (see Table 2). The table presents the observed and predicted concentrations and ISI levels in L. littorea. The ecotoxicological risk presented as PAF can be interpreted as the characterization of the potential adverse health effects of exposure to environmental hazards. The 5 th and 95 th percentiles of the exposure concentration distribution in the dedicated open water or harbour were used to calculate the range of ISI values likely to occur in the area. The upper and lower limits of the calculated ISI values are reported in Table 2.3.3 Observed ISI values and species abundance ISI observations are available for the open waters, except for Grevelingen (D) and Veersemeer (F) lakes, where no ISI sampling was carried out. The observed ISI values for open waters were relatively low. The highest ISI value was observed in the Western Scheldt. Ten harbours were selected for ISI sampling. However, since no periwinkles were found in Harlingen (location 8), ISI observations are available for only 9 harbours. Periwinkles with low abundance were found at the harbours in Delfzijl (location 1), Den Helder (location 14), IJmuiden (location 15) and Rijnmond (location 18). Much higher ISI levels were observed in harbours than in open waters. The highest value (3.0) was observed in Den Helder (location 14). The wide range of ISI levels observed in Vlissingen harbour (location 25) indicates that a high variation in ISI levels can exist within a single harbour.

  presents the comparison between calculated and observed ISI for open waters (A) and Dutch harbours (B). The observed ISI values were relatively low for open waters, often below the 5 th percentile of the predicted ISI range based on the TBT concentrations. Given the TBT concentrations in the Western Scheldt (G) and Veersemeer (F), predicted ISI values around 1.0 might be expected. Although intersex was observed at a low level (ISI = 0.15) at one location in the Western Scheldt, the observations never exceeded the signalling level of 0.3 (Oehlmann, 2002). Nearly all observed ISI levels in harbours fall within the ISI ranges corresponding to the median and 95 th percentile of the exposure concentrations. The ISI levels observed in IJmuiden (location 15), Den Helder (location 14), Vlissingen (location 25) and Lauwersoog (location 3) corresponded to the upper 25% of the exposure concentrations. At 4 other harbours -Delfzijl (location 1), Yerseke (location 22), Scheveningen (location 17) and Rijnmond (location 18) -the observed ISI levels corresponded to the upper 50% of the exposure concentrations. Maximum

  covering the whole distribution of possible ISI values. The wide range of ISI levels observed in Vlissingen harbour (location 25) indicates that a high variation in ISI levels can exist within a single harbour. For open waters, only two measured ISI values fell within the predicted ISI ranges (North Sea (C)) and Western Scheldt (G), while the rest of the observations were below the calculated ranges. This is presumably a consequence of the asymptotic minimum of 0.07 in Formula 1. Differences in local K oc values may have caused the predicted ISI levels in open waters to exceed the observed ISI values. In this study the K oc is based on measured concentrations in sediment and water from harbours. This is in line with the fact that the correspondence in harbours is much better. It might be possible to derive a dedicated K oc for open waters from the comparison of water and sediment concentrations in open waters, potentially improving the correspondence between predicted and measured ISI levels for these water systems. However, such data are not available. The possible over-estimation of water concentrations in open waters would also imply that the ecological risk values presented in this study have been overestimated for open waters.

in

  Figure 2Exposure distributions for all sample locations in Dutch open/estuarine/marine waters and harbours.

Figure 4 .

 4 Figure 4. Comparison of the observed ISI values and the predicted ISI ranges based on the concentration distribution in the corresponding open waters (A) and harbours (B). The calculated ISIvalues corresponding to the median and the 5th and 95th percentile of the exposure concentration distribution are also shown. At some sampling locations in harbours periwinkles were rare, while at other locations( †) no periwinkles were observed.

Table 1

 1 Chemical and biological datasets used in this study

	Dataset	Area type Medium	Time range	# areas	# locations
	TBT concentrations	Open water Suspended matter 1998-2003	5	25
	TBT concentrations	Open water Sediment	1998-2003	7	123
	TBT concentrations	Harbour	Suspended matter 1999-2002	4	8
	TBT concentrations	Harbour	Water	1999-2002	2	6
	TBT concentrations	Harbour	Sediment	1999-2002	31	270
	ISI values and abundance Open water Periwinkle	2002-2005	5	24
	ISI values and abundance Harbour	Periwinkle	2002-2005	9	18

Table 2

 2 Observed and predicted concentrations, ecological risk values and observed and predicted ISI levels in different Dutch open waters.

Table 3

 3 Combinations of LOE results and likely ecological health status in relation to TBT pollution. Integrated ecological risk assessment is based on multiple Lines Of Evidence

	LOE -	LOE -	LOE -	Likely	Action
				ecological
	Chemistry	Toxicology	Ecology	health status in
				relation to TBT
	Predicted	Observed ISI	Presence of
	risk;		gastropods at
	Predicted ISI		all locations
	PAF < 10%;	ISI < 0.3	+	Good	No action
	ISI < 0.3
	PAF < 10%;	ISI < 0.3	-	Good	Other stressors
	ISI < 0.3				present?
	PAF > 10%;	ISI < 0.3	+ or -	Probably poor	Extend
	ISI > 0.3				monitoring
	PAF > 10%;	ISI > 0.3	+	Poor	Mitigating
	ISI > 0.3				measures and
					monitoring of
					recovery
	PAF > 10%;	ISI > 0.3	-	Poor	Mitigating
	ISI > 0.3				measures and
					monitoring of
					recovery
	LOE =

The likely range of the intersex indicator (ISI) for L. littorea was also calculated based on the doseresponse relationship for L. littorea as published by[START_REF] Oehlmann | Baseline study on intersex in Littorina littorea with recommendations for biological TBT assessment criteria[END_REF] and the spatial distribution of water

Acknowledgements

We would like to acknowledge the valuable contribution of the laboratory personnel at the Directorate-General of Public Works and Water Management's (RWS) National Institute for Coastal and Marine Management (RIKZ) for the TBT analyses. The authors wish to express their gratitude to Johan Jol (IMARES) for his invaluable support in conducting the fieldwork. Ivonne Rietjens and Tinka Murk (WUR) are thanked for their critical review of a first version of the manuscript.

Appendix A. Supplementary data

Supplementary data associated with this article can be found, in the online version, at…..

Aldenberg, T., Slob, W., 1993. Confidence limits for hazardous concentrations based on logistically distributed NOEC toxicity data. Ecotoxicology and Environmental Safety,25,[48][49][50][51][52][53][54][55][56][57][58][59][60][61][62][63] Aldenberg, T., Jaworska, J.S., Traas, T.P., 2002. Normal species sensitivity distributions and probabilistic ecological risk assessment. In Posthuma, L., Suter II, G.W., and Traas, T.P., Species Sensitivity Distributions in Ecotoxicology Boca Raton: CRC Press, USA. Bauer, B., Fioroni, P., Ide, I., Liebe, S., Oehlmann, J., Stroben, E., Watermann, B., 1995. TBT


Andreia Cruz

Tânia Caetano

Satoru Suzuki

Sónia Mendo

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    Corrigendum to "Aeromonas veronii, a tributyltin (TBT)-degrading bacterium isolated from an estuarine environment, Ria de Aveiro in Portugal" [

Corrigendum

  Corrigendum to "Aeromonas veronii, a tributyltin (TBT)degrading bacterium isolated from an estuarine environment, Ria de Aveiro in Portugal" [Marine Environmental Research 64 (2007) 639-650] Andreia Cruz a,b, Taˆnia Caetano a, Satoru Suzuki b, So´nia Mendo a,* a CESAM and Departamento de Biologia, Universidade de Aveiro, Campus Universita´ rio de Santiago, 3810-193 Aveiro, Portugal b Centre for Marine Environmental Studies (CMES), Ehime University, Bunkyo-cho 3, Matsuyama 790-8577, Japan original article: doi:10.1016/j.marenvres.2007.06.006

Corrigendum to "Aeromonas veronii, a tributyltin (TBT)-degrading bacterium isolated from an estuarine environment, Ria de Aveiro in Portugal" [Marine Environmental Research 64 (2007) 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.