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Thèse de doctorat/ PhD Thesis Citation APA:

Danis, B. (2004). Bioaccumulation and effects of polychlorinated biphenyls (PCBs) in the sea star Asterias rubens L. (Unpublished doctoral dissertation).

Université libre de Bruxelles, Faculté des Sciences – Sciences biologiques, Bruxelles.

Disponible à / Available at permalink : https://dipot.ulb.ac.be/dspace/bitstream/2013/211171/4/61b21da4-ee03-482a-8ef2-495d4c088a3d.txt

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f ^

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I___________________________________ ^

UNIVERSITE L ibre de B ruxelles - F aculté des S ciences

L aboratoire de B iologie M arine

Bioaccumulation and EfTects of Polychlorinated Biphenyls (PCBs) in the Sea Star Asterias rubens L.

Bruno Danis March 2004

Thesis submitted in JiilJilment of the degree of:

Université Libre de Bruxelles 1

Supervisors:

Dr Michel Wamau Dr Philippe Dubois

X

U niv ' ersite L ibre de B ruxelles - F aculté des S ciences

L aboratoire de B iologie M arine

Bioaccumulation and Effects of Polychlorinated Biphenyls (PCBs) in the Sea Star Asterias rubens L.

Bruno Danis March 2004

Thesis submitted in Jûlfilment of the degree of :

Doctor in Sciences

Supervisors;

Dr Michel Wamau

Dr Philippe Dubois

Si les pétroliers transportaient de l'eau de mer, on s'en foutrait qu'ils fassent naufrage...

Philippe Geluck

A ceux qui nous manquent et qui veillent sur nous...

A

-R

A cknowledgements -R emerciements

Quand \âent le moment d’écrire les remerciements, des tonnes de souvenirs s’engouffrent dans votre tête, faisant apparaître les personnes sans qui une thèse ne serait pas ce qu’elle est, sans qui tous les éléments qui la composent ne se seraient pas ajustés comme ils le sont...

Beaucoup de mes phrases vont commencer par «je remercie le Dr... », mais bon c’est comme ça... [on se croirait un peu à une cérémonie de remise des Oscars...]

Je commence évidemment par remercier le Professeur Michel Jangoux, qui m’a ouvert les portes de son laboratoire il y a six ans déjà (!) et qui, j’en suis sûr, a gardé un œil bienveillant sur moi pendant les moments difficiles ou joyeux qui ont échelonné cette période.

Je remercie le Dr Philippe Dubois, mon copromoteur et voisin de palier. Ton sens critique affûté m’a plus d’une fois poussé à puiser au fond de mes ressources. Merci d’avoir été toujours présent dans les moments de doutes, et merci pour ta franchise.

Je remercie le Dr Michel Warnau, mon autre copromoteur, mon quasi-grand frère. Généreux et impartial, sans compter, tu m’as donné le goût de la recherche, celle que l’on mène en poussant toujours plus loin ses limites, celle pour laquelle on a tant besoin de ses proches.

Merci à Geneviève et aux trois petits warnouilles: Nathan, Luane et Max, pour la gentillesse sans fin que vous avez montré à mon égard.

Je remercie le Dr Scott Fowler, un autre protagoniste bienveillant de ma thèse, qui m’a accueillit au sein de son laboratoire à plusieurs reprises, et m’a permis de concrétiser mes fantasmes expérimentaux pratiquement sans limite.

Je remercie Jean-Louis Teyssié, technicien de l’extrême dopé à la salade qui, tel un compteur Geiger, détecte la moindre trace de radioactivité. Tu m’as appris tous les rouages de la manipulation « à chaud », et tu n’as jamais reculé devant les défis que je souhaitais que nous relerions ensemble.

Je remercie Olivier Cotret, autre technicien de l’extrême, qui a été ma main droite pendant

toute la durée des expérimentations monégasques. Merci pour tes Hénaurmes coups de main !

Je remercie le Dr Jean-Paco Bustamante, mon binôme Rochelais, avec qui nous avons bataillé

à mort, à coup de cerises, dans le jardin de M. Verola, ex-champion de boules de son petit

état.

ACKNOW’LEDGEMEXTS-REMERCIEMENTS

Je remercie le Dr Jean-Pierre Villeneuve pour son aide précieuse dans l’analyse des PCB

« froids », et pour sa grande patience.

Je remercie le Dr Chantal Cattini pour sa gentillesse, son aide, ainsi que pour le temps qu’elle a passé pour moi devant sa colonne, à voir couler de l’extrait d’étoile de mer.

Je remercie le Dr Véronique Flamand pour ses précieux conseils en matière d’ELISA et son ouverture d’esprit.

Je remercie le Dr Virginie De Backer pour son aide et ses conseils concernant la famille des dioxines.

Je remercie le Dr Patrick Flammang pour ses conseils pour la réalisation des Western Blots.

Je remercie le Dr Pascale Wantier pour sa gentillesse et pour m’avoir initié aux joies de l’analyse des PCBs.

Je remercie le Prof Robert Flammang pour son intérêt pour notre projet, et l’implication de son laboratoire pendant plusieurs années.

Je remercie le Dr^ Stanislas Goriely, qui m’a initié aux joies de l’ELISA, mais qui est surtout mon ami depuis quinze ans. Merci aussi à Nanou et au petit Kolya pour les inoubliables brunchs du dimanche. Je suis sûr que nos bambins joueront encore longtemps ensemble.

Je remercie le Dr Drossos, dextre chirurgien de la main, à qui je dois le sauvetage de mon annulaire droit après une rencontre indésirée, un soir de noël 2003, entre les tendons de mon cher doigt et une rogneuse mal léchée de la marque Idéal, qu’au passage je ne remercie pas, car elle ne place pas de garde sur de tels engins.

Je remercie évidemment l’ensemble du Biomar, dans l’ordre du début à la fin du couloir

« Hippie» Chantal, « Cornez da Costa » Sergio, «DivX le Breton » Vinz, « le Dubbe » Phil,

« Sand » Cristi, « le Dim-Dim », « le Loron », « Chipito» CillesD, « Snore » Ceoff, « Mac Cuy-ver » Cuy, « Never-short-of-a-joke » Herwig, « Oui-oui » Richard, « Vivi » Viviane,

« Psycho » Marcelle, « Happy hour » Edwin, « Drine-drine » Sandrine, « Radar» Phil,

« Buffalo » Beber, « Coup’-coup’ » Dev, « MasterMind » Didje, « Isaac » CillesR, « Aussie » Raph, « Cotten » Cuillemette, « Never-steack never again» den Daaav-id, « Tam-tam » Tamar, « Civette » Yves, « Jedi » Eugene, « Chewbacca » Jean-Marc, « Batman » Walter, et

« Dites-Edith » Edith. Merci à vous tous de m’avoir supporté pendant toutes ces (courtes) années (six ans, je n’en reviens pas !).

Je remercie chaleureusement l’équipage du Belgica, composé d’hommes dévoués, qui me

laisseront des souvenirs de moments hors du commun, parfois surréalistes (surtout au cours de

nos premières sorties...).

acknovvxedgements

-R

Je remercie les GMs de mon Comité d’Accompagnement, les Drs Christiane Lancelot et Cuy Josens, pour leur constante attention tout au long de ma thèse.

Je remercie le FRIA, qui m’a accordé une bourse de thèse pendant les premières années de celle-ci.

Je remercie la fondationVan Buuren qui m’a apporté un ballon d’oxygène non négligeable pendant les mois de disette.

Je remercie l’ONEM pour son soutien financier (pendant les nombreux mois de disette encore et quand le ballon d’oxygène s’est envolé...).

Je remercie la Communauté Française de Belgique pour la bourse de voyage qu’elle m’a accordée, et qui m’a permis de -presque- joindre les deux bouts au cours de mes séjours monégasques.

Je remercie le Dr Cwenaëlle Leclercq et le presque-Dr Crégory Sempo pour leur amitié si simple qu’elle m’a fait oublier bien des prises de têtes. Spéciale dédicace au petit Loup, dont les piles ne sont jamais à plat pour jouer avec ma petite Zoé.

Je remercie mon cousin David pour le soin avec lequel il a court-circuité mon ordinateur tout neuf au champagne -s’il vous plaît- à quelques semaines du dépôt de ma thèse (je t’avais dit que je te louperais pas !!).

Je remercie Chanda & Maxime, de chez Macline, et qui ont réussi à sauver in extremis toutes mes données après l’incident cité ci-dessus.

Je remercie ma sœur, Muriel, pour sa présence rayonnante dans les moments très difficiles qui ont émaillés la période de thèse. Cilles t’es quelques lignes au dessus...

Je remercie ma marraine Janine, ainsi que Jacques pour nous avoir accueillis mille fois autour d’un incroyable festin qui a toujours eu l’art de nous remonter le moral à l’infini.

Je remercie ma belle-famille, Monique -notre ange gardien-, Aude et Yvan, toujours prêts à nous changer les idées.

Je remercie évidemment tout le reste de la famille pour l’affection et l’attention qu’elle m’a toujours prodigué.

Je remercie mes parents-adorés qui m’ont soutenu sans faillir depuis de nombreuses années (je n’ose même pas les compter) et m’ont montré l’importance de l’ouverture d’esprit, de la curiosité, de la ténacité et du bon vin.

Enfin, bien sûr, je remercie Céline, avec qui j’ai traversé en très peu de temps les plus dures

tempêtes, mais aussi les plus grandes joies de ma vie. Je crois qu’après ce que nous avons vécu,

aucun ouragan ne pourra nous éloigner.

acknow

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Zoé, mon bébé-crabe, et Liam, mon bébé-grogne, papa vous dédie ce travail.

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T able of C ontents

ACKNOWLEDGEMENTS-REMERCIEMENTS...5

TABLE OF CONTENTS...9

SUMMARY... 11

I. GENERAL INTRODUCTION...15

1.1. POLYCHLORINATED BIPHENYLS...15

1.1.1. General information...15

1.1.2. Analysis...16

1.1.3. International recommendations...18

1.2. PCBS IN THE MARINE ENVIRONMENT...19

1.2.1. Caracterization ofPCB contamination...20

1.2.2. Biological effects ofPCBs... 22

1.2.3. Biomarkers ofPCB exposure... 27

1.3. Contaminationofthe North Seaby PCBs... 31

1.3.1. The North Sea...31

1.3.2. Origin and fluxes ofPCB contamination in the North Sea...32

1.3.3. PCBs in benthic ecosystems ofthe North Sea...33

II. OBJECTIVES...35

III. EXPERIMENTAL CONDITIONS... 37

III. 1 NON-COPLANAR VS. COPLANAR CONGENER-SPECIFICITY OF PCB BlOACCUMULATION AND IMMUNOTOXICITY IN SEA STARS...39

111.2 DELINEATION OF PCB UPTAKE PATHWAYS IN A BENTHIC SEA STAR USING A RADIOLABELLED CONGENER 59 111.3 COPLANAR PCB UPTAKE KINETICS IN THE COMMON SEA STAR ASTERIAS RUBENS AND SUBSEQUENT EFFECTS ON ROS PRODUCTION AND CYPl A INDUCTION... 73

111.4 CONTRASTING EFFECTS OF COPLANAR VS NON-COPLANAR PCB CONGENERS ON IMMUNOMODULATION AND CYP IA LEVEES (DETERMINED USING AN ADAPTED ELIS A METHOD) IN THE COMMON SEA STAR ASTERIAS RUBENS L... 95

IV. FIELD CONDITIONS... 113

IV. 1 C^{ontaminantlevelsinsédimentsandasteroids} (A^{steriasrubens} L„ Echinodermata) ^from THE BELGIAN COAST AND SCHELDT ESTUARY: POLYCHLORINATED BIPHENYLS AND HEAVY METALS... 115

IV.2 Bioaccumulationandeffectsof PCBsandheavymetalsinseastars (Asteriasrubens, L.) FROM THE North Sea: asmallscaleperspective...141

IV.3 ECHINODERMS ^as BIOINDICATORS, BIOASSAYS ^andimpact ASSESSMENT TOOLS OF SEDIMENT- ASSOCIATED METALS AND PCBS IN THE NORTH SEA... 159

IV.4 LevelsandeffectsofPCDD/Fsandc-PCBsinsédiments, musselsandseastarsofthe INTERTIDAL ZONE IN THE SOUTHERN NORTH SEA AND THE CHANNEL... 185

V. GENERAL DISCUSSION... 205

The BIOACCUMULATION ofPCBsinseastars... 205

Theeffectsof PCBsinseastars... 208

Conclusions-recommendations...211

VI. REFERENCES... 213

VII. ANNEX STUDIES...249

VII. 1 Bioaccumulationof PCBsintheseaurchin Paracentrotuslividus: seawaterandfood EXPOSURES TO a '“C-RADIOLABELLED CONGENER (PCB 153)... 251

VII.2 Bioaccumulationof PCBsinthecuttlefish Sepiaofficinalisfromseawater, sédimentand FOOD PATHWAYS... 263

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VII.3 M

EROD

: C

... 279

VII.4 EFFECTS

PCBS

OXYGEN SPECIES (ROS)

BY THE IMMUNE CELLS OF P

(E

)... 287

APPENDIXI : CAPTIONS TO HGURES... 299

APPENDIX II : CAPTIONS TO TABLES... 304

APPENDIX in : CAPTIONS TO EQUATIONS... 308

SUMMARY

SUMMARY

PCBs are among the most problematic marine contaminants. Converging towards the océans via the rivers and the atmosphère, they concentrate in sédiments where they become a permanent threat to organisms living at their contact. PCBs are extremely résistant, bioaccumulated and some congeners are considered as highly toxic. The North Sea is considered as a highly contaminated area ; however little information is available regarding the impact of PCBs on key benthic organisms of this région.

Ubiquist, abundant and generally recognized as a good bioindicator species, the common NE Atlantic sea star Asterias rubens (L.) is an ecosystem-structuring species in the North Sea and was chosen as an experimental model. The présent study focused on the characterization of PCB bioaccumulation in A. rubens exposed through different routes (seawater, food, sédiments) and on subséquent biological responses, at immune and sucellular levels. The considered responses were respectively (i) the production of reactive oxyggen species (ROS) by sea stars amoebocytes, which constitutes the main line of defence of echinoderms against pathogenic challenges and (ii) the induction of a cytochrome P450 immunopositive protein (CYPl A IPP) which, in vertebrates, is involved in PCB détoxification.

Experimental exposures carried out hâve shown that A. rubens efficiently accumulâtes PCBs.

Exposure concentrations were always adjusted to match those encountered in the field. PCB

concentrations reached in sea stars during the experiments matched the values reported in

field studies ; therefore our experimental protocol was found to accurately simulate actual

field situations. Uptake kinetics were related to the planar conformation of the considered

congeners : non-coplanar PCB uptake was described using saturation models, whereas

coplanar PCBs (c-PCBs) were bioaccumulated according to bell-shaped kinetics. Non-

coplanar congeners generally reached saturation concentrations whithin a few days or a few

weeks, which means that sea stars can be used to pinpoint PCB contamination shortly after

occurrence. On the other hand, c-PCB concentrations reached a peak followed by a sudden

drop, indicating the probable occurrence of c-PCB-targeted metabolization processes in sea

stars. Our experimental studies also demonstrated that seawater was by far the most efficient

route for PCB uptake in sea stars and that even if PCB levels in seawater are extremely low

compared to sediment-associated concentrations, seawater constitutes a non-negligible route

for PCB uptake in marine invertebrates. Among the different body compartments, bodywall

SUMMARY

displayed the highest bioaccumulative potency and can therefore be considered as particularly interesting for field biomonitoring applications. Rectal caeca, which play a central rôle in digestion and excrétion processes in sea stars, bave also rised particular interest as results suggest these organs could be involved in the élimination of PCB 77 dégradation products.

The field work carried out during the présent study showed that PCB concentrations measured in A. rubens tissues reflect environmental levels of certain congeners. As it was the case in experimental conditions, A. rubens differentially accumulated PCB congeners according to their planarity. Strong relationships were found between concentrations measured in sédiments and those determined in sea stars body wall for certain non-coplanar congeners (e.g. 118 and 138), thus allowing to consider A. rubens as a suitable bioindicator species for medium-chlorinated PCB congeners. On the other hand, sea stars appeared to be able to regulate -to a certain extent- their content in coplanar PCBs. This implies that (i) A. rubens cannot be strictly considered as an indicator organism for c-PCBs and (ii) c-PCBs probably affect essential aspects of sea star biology, potentially leading to deleterious effects.

The présent study addressed effects of PCB exposure on A. rubens biology, in both experimental and field conditions. In experimental conditions, PCBs were found to significantly alter ROS production by sea stars amoebocytes. This alteration also occurred in a congener-specific way : c-PCBs were found to significantly affect, and probably impair sea stars immune System, whereas non-coplanar congeners had no effect. In the field, the PCB contribution to immunotoxicity could not be determined because none of our studies considered ROS production along with c-PCB concentration measurements. However, the levels of ROS production by sea stars amoebocytes measured in field and experimental conditions were found to potentially lead to altered immunity, and therefore to impair sea stars defence against pathogenic agents.

A specially designed ELIS A was used to measure CYPIA IPP in experimental and field

conditions. Experimental work has shown that the induction of this protein was related to

PCB exposure in a congener-specific fashion : c-PCBs alone were found to strongly induce

the production of CYPIA IPP according to a dose-dependent relationship. These results hâve

highlighted many similarities between the dioxin-like responsiveness of CYPIA IPP

induction in sea stars and that occurring in vertebrates. This strongly suggests similarities in

the toxicity-triggering mechanism of dioxins and c-PCBs. In the field, CYPIA IPP induction

was found to be significantly related to PCB levels determined in bottom sédiments. It can

thus be considered as a valuable biomarker. Further research is however needed to better

SUMMARY

characterize the influence of physico-chemical and physiological parameters on CYPIA induction to refîne the interprétation of the information gathered via this biomarker.

Results obtained in our study hâve lead to questionning international régulations applying to PCB biomonitoring in the marine environment. For instance, we strongly suggest that the sélection of congeners to be systematically considered should be revised to include c-PCBs.

Indeed, in our experiments PCB toxicity was almost always attributable to the sole c- congeners. Historically, détermination of c-PCB concentrations was extremely difficult due to analytical limitations ; however, nowadays, these problems hâve been overcome and do no more justify their exclusion from monitoring studies.

Although A. rubens appeared to be quite résistant to PCB contamination, levels measured in

sea stars from the Southern North Sea can possibly affect their immune and endocrine Systems

in a subtle way, but with relatively low risk for this species at the short-term. However, this

does not mean that other species in this région undergo similarly low risks, or that sea star-

structured ecosystems may not become affected in the long-term.

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I. G eneral I ntroduction

1.1. Polychlorinated biphenyls

1.1.1. General information

Polychlorinated biphenyls (PCBs) hâve been in use in the industry since the 1930s, in electrical equipment and in the manufacture of paints, plastics, adhesives, coating compounds and pressure-sensitive copying paper (Clark 1997). Their remarkable physico-chemical properties (very high stability, excellent electric and thermie insulation) led to their prolifération in the industry to which they were marketed under various trade names (Metcalfe 1994) : Aroclor (Monsanto, United States), Clophen (Bayer, Germany), Phenoclor (Caffaro, Italy), Pyralene (Prodelec, France), Kanechlor (Kanegafushi, Japan), Sovol (Russia).

These commercial PCBs vary in texture from clear oils to powders, according to the needs of industrial applications. These mixtures also vary in composition, containing between 20 and 60 % per weight of chlorine, with a varying number of chlorine atome per molécule.

Theoretically, PCBs include 209 possible compounds (congeners) with varying degree and pattern of chlorine substitution (Fig. 1). Ballschmiter & Zell (1980) set up a nomenclature System for PCBs that assigns each congener a number comprised between 1 and 209. This System was adopted by the International Union of Pure and Applied Chemistry (lUPAC).

3’ 2' 23

5- 6' 65

nwta orttio

Figure 1. Numbering System for sites of chlorine on a biphenyl molécule (Metcalfe 1994)

PCBs were first reported in environmental samples in the 1960s (Jensen, 1966). Although

there are 209 possible PCB congeners, only around 90 of them hâve been detected in the

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environment. The relative abundance of the different congeners in environmental samples dépends on factors such as the congener abundance in the initial commercial products, the relative sales and uses of the products, and the relative persistence and transport in the environment of the compounds.

Due to high persistence and relative volatility, on the long term PCBs may be transported over considérable distances, as a conséquence of mass movements of air or water. These movements can also occur by diffusion, which may be very localized, but can take place over large distances, especially in air. PCBs hâve been detected in the most remote régions, such as the Arctic, the North Atlantic and even the Antarctic and deep-sea (e.g. AMAP 1998), where there is no anthropogenic émission sources.

Data on the global production and use of PCBs has been collected for décades, but more work is needed for the interprétation of past, présent and future contamination levels around the World: it is likely that PCB compounds will remain in the environment for a very long time (Cummins 1988, Tanabe 1988, Voldner & Li 1995, Wania & Mackay 1996, Vallack et al.

1998). In the late 1980s, estimations indicated that there were still 374,000 tons of PCBs in the environment, of which 232,400 tons dissolved in seawater, 3,500 tons dissolved in freshwater, and 1580 tons circulating in the atmosphère (Tanabe 1988). According to the same author, 783,000 additional tons of PCBs were remaining in storages or in landfills, of which an undetermined part could be released in the environment.

1.1.2. Analysis

Most analyses of PCB levels in the environment hâve been reported as Aroclor équivalents.

This was once due to necessity because traditional packed-column gas chromatography (GC) were not able to résolve individual PCB congeners. This lack of resolution limited the capacity of analyses to accurately describe environmental PCB levels and patterns. Moreover, these analyses were mostly based on Aroclor peaks from the packed-column chromatogram, assuming that ratios among PCB congeners in the environment were the same as those found in commercial mixtures (Metcalfe 1994).

It is now well-known that, once released in the environment, the composition of a commercial PCB mixture changes over time, since different congeners display very different physico- chemical properties (e.g. water solubility, vapor pressure, tendency to sorb to organic matter).

Individual congener partition behaviour differs among water, air and solid phases (Dickhut et

al. 1986, Lara & Ernst 1989, Brunner et al. 1990). Moreover, some congeners undergo

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dechlorination by anaérobie bacterial action when présent in sédiment at threshold concentrations, while others do not (Brown et al. 1987, Quensen et al. 1990, Mohn & Tiedje 1992). Also, “light” congeners can be subject to aérobic dégradation in particular conditions (Furukawa 1982). Consequently, original PCB mixtures become depleted in the most degradable congeners and enriched in métabolites of the latter ones and in “résistant”

congeners. In biota, PCB composition can also be altered, depending on the uptake, metabolization and dépuration rates of individual congeners. PCBs in the environment take on a congener composition that becomes dissimilar to the original Aroclor mixture. Thus, since the early 1990s, congener-specific analysis of PCBs has progressively replaced traditional Aroclor-equivalent based methods (Duinker et al. 1991, Eganhouse & Gossett 1991).

For routine analyses of PCB congeners in marine biota samples, the commonly used methods are high resolution gas chromatography with capillary columns and électron capture détection (HRGC-ECD) or high resolution gas chromatography with low resolution électron impact mass spectrometry in selected ion mode (HRGC-LRMS-SIM) (Metcalfe 1994). Most individual PCB congeners can be resolved using these methods at low parts-per-billion concentrations (Schultz et al. 1989). An advantage of the ECD over LRMS for PCB analysis is that it is halogen sensitive (Cairns et al. 1989): many coextractive compounds (e.g.

polynuclear aromatic hydrocarbons, phthalates) are not detected. A disadvantage of the method over HRGC-LRMS is that the response is highly dépendent on the degree and pattern of chlorination, reducing sensitivity and accuracy of the method for lesser chlorinated congeners.

Another approach used to address the toxic potential of PCBs is the use of toxic equivalency (TEQ). In this approach, the biological or toxic potencies of individual congeners are expressed related to a benchmark contaminants, usually 2,3,7,8 tetrachloro-dibenzo-/i-dioxin (TCDD), an extremely potent toxicant (Fig. 2). Using a variety of endpoints or responses, a relative biological potency or toxic equivalency factor (TEF) can be determined for each congener. The TEQ approach is an attempt to provide integrated assessment of the toxic potential of environmental mixtures. It relies on a number of assumption, including the absence of non-additive interactions (i.e. possible synergism or antagonism is not taken into account) among the components of the mixture (Safe 1990, Ahlborg et al. 1992, 1994).

TCDD équivalents are being used increasingly in risk assessments as a replacement for

exposure measures based only on TCDD or total PCBs (Barron et al. 1994, Van den Berg et

al. 1998).

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Figure 2. Molecular configuration of 2,3,7,8 TCDD and PCB 169 (Metcalfe 1994)

1.1.3. International recommendations

Persistent Organic Pollutants (POPs) is the common name refering to a group of organic contaminants that comprises PCBs. POPs are semi-volatile, bioaccumulative, persistent and toxic (Vallack et al. 1998). Although the occurrence of POPs at elevated levels is of great concern in “hot spots”, the POPs issue has received increasing attention at régional and global scales in the last décades (Wania & Mackay 1996, UNECE 1998, UNEP 2001).

Due to their beyond-boundaries transport, political problems bave also arisen. International

agreements hâve thus corne into effect, such as the 1998 Aarhus Protocol on POPs (UNECE,

1998). The overall and long-term objective of the Aarhus Protocol on POPs is to eliminate

any discharge, émission and loss of POPs to the environment. The international community

has called for action to reduce and eliminate production, use and releases of these substances

through: (i) the Protocol to the régional UNECE Convention on Long-Transboundary Air

Pollution (CLRTAP) on POPs, opened for signatures in June 1998 and (ii) the global

Stockholm Convention on POPs, opened for signatures in 2001. These instruments establish

strict international régimes for initial lists of POPs (16 in the UNECE Protocol and 12 in the

Stockholm Convention). Both instruments also contain provisions for including additional

Chemicals into their list. They lay down the following control measures: prohibition or severe

restriction of the intentional production of POPs and their use, restrictions on export and

import of the intentionally produced POPs (Stockholm Convention) , provisions on the safe

handling of stockpiles (Stockholm Convention), provisions on the environmentally sound

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disposai of POPs wastes and provisions on the réduction of émissions of unintentionally produced POPs (e.g. dioxins and furans).

Regarding PCBs, the International Council for the Exploration of the Sea (ICES) has recommended that congeners 28, 52, 101, 153, 138 and 180 should be selected for routine analysis (Duinker et al. 1988). Several European Union (EU) countries hâve adopted these congeners, with the addition of congener 118, for defining maximal levels of PCBs in edible marine resources. These environmental quality standards and other international commitments also arise from the 1984 International Conférence on the Protection of the North Sea, the 1995 Barcelona Convention for the Protection of the Mediterranean Seas against Pollution, the Baltic States HELCOM, etc.

One of the most frequent objectives of monitoring is to assess seafood quality using estuarine and marine water and sédiments as a check for sources of possible pollution. The recent emphasis on the monitoring of non-ortho and mono-ortho PCB congeners has necessitated an expansion of the list of congeners to be considered in routine analysis. Because of their high toxic potential (Safe 1990), it is most probable that ail non-ortho substituted congeners should be included in analysis programmes (Metcalfe 1994).

1.2. PCBs in the marine environment

The ultimate sink for many contaminants is the marine environment, following either direct

discharges or hydrologie and atmospheric processes (Stegeman & Hahn 1994). Since the late

1960s, PCBs are known to be présent in substantial quantities in marine sédiments, as well as

in marine biota (Jensen et al. 1969). PCBs accumulate in the organic phase, such as biota and

the organic fraction of sédiments, transfering between these compartments according to the

model presented in Fig. 3. PCBs persist in the marine environment for several décades: most

PCBs only exist in trace concentrations, but ail hâve extensive half-lives (dégradation half-

lives ranging up to 200 years) in the environment (Howard et al. 1991, Haynes et al. 2000, Oh

2000, Moore et al. 2002, Wania & Daly 2002).

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Bioaccumuîatloni Ç Transfer

How?

Where?r.; M Spatial .distribution

^When?^"’ ^ Temporal distribution

Figure 3. Contaminants transfers between compartments in a Coastal model (Moore et al. 2002)

1.2.1. Caracterization of PCB contamination

a. Seawater

PCBs are hydrophobie compounds, i.e. they hâve extremely low water solubilities.

Concentrations in océan water are generally very low, making reliable quantification technically difficult. PCB concentrations in filtered océan water are usually reported to be in the low pg 1' range. In contrast, PCBs are highly lipophilie and adsorb readily onto particles.

Their distribution in sea is thus far from being uniform.

The sea surface microlayer (SSM) is a film varying from a few pim to 1 mm in thickness. It is extremely difficult to study, but is known to contain high levels of particulate organic carbon and lipids compared to bulk water, thus allowing PCBs to accumulate (Daumas et al. 1976, Hardy et al. 1988, Xhoffer et al. 1992, Garabetian et al. 1993). Elevated levels of dissolved organic contaminants in the SSM hâve been reported with enrichment factors reaching one to three orders of magnitude for PCBs (Duce et al. 1972, Bidleman 1973, Napolitano 1995).

While the total quantity may not be great, the PCB enrichment of the SSM may be of

considérable importance to surface-living organisms. Where water masses with variable

physico-chemical characteristics meet, they form a front where floating material gets

accumulated including surface oil. Fronts hâve a high productivity and attract a wide range of

animais, which thus receive a PCB-enriched diet. Since the upper millimétré of the sea is also

enriched in microorganisms and zooneuston (including larvae), great concern has been

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expressed on the toxic effects of the high contaminant levels in the SSM (Hardy et al. 1990, Hardy & Cleary 1992, Stebbing et al. 1992). The PCB enrichment in SSM microorganisms also poses analytical difficulties in distinguishing the portion that is incorporated and the one that is adsorbed onto it. The former may affect them, but the latter is bioavailable to animais feeding on the contaminated organisms. PCBs adsorbed onto inorganic particles may ultimately be carried to the seabed, which acts as a sink for these compounds. Moreover, suspended or re-suspended particles are commonly ingested by filter-feeding animais, entering food chains by this route.

b. Sédiments

Sédiments are repositories for physical and biological débris and are considered as sinks for a wide variety of Chemicals (Clark 1997). The concern associated with PCBs sorption to sédiments is that many organisms spend a considérable portion of their life-cycle on or in marine sédiments. This provides a path for PCBs to reach higher trophic levels. Direct transfer of contaminants from sédiments or interstitial water to organisms is considered to be a major route of exposure (Walker & Peterson 1994). PCBs are présent in much higher concentrations in sédiments than in overlying water. Sorption to sédiments is the prédominant removing mechanism for PCBs from the water column. The analysis of PCBs in sédiments has the advantage of integrating time variations. Once contaminated, sédiments can act themselves as a slowly releasing source of PCBs, which causes chronic exposure of biota long after the primary source of contamination has discontinued (Moore et al. 2002).

c. Organisms

As a conséquence of their hydrophobie and persistent characteristics PCBs are bioaccumulated and high concentrations are found in biota (Stebbing et al. 1992, Clark 1997, OSPAR 2000). PCBs are efficiently accumulated by marine organisms by absorption across outer surfaces (e.g. gills, skin), or by ingestion of contaminated food, seawater or sédiments.

Once they hâve entered the organism, PCBs are stored within the fatty tissues, or in other

lipophilie sites, such as cell membranes or lipoproteins. In the long term, release from storage

may occur (e.g. in times of low food availability) during which organisms mobilize and use

their fat reserves, so increasing the concentration of PCBs in their body up to possibly

harmful levels (Walker et al. 1996). Delayed toxicity may therefore be observed some time

after initial exposure to the contaminant. Organisms hâve the capacity to bioaccumulate and

to biomagnify PCBs, which results in body concentrations several orders of magnitude higher

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than in seawater or in the food (OSPAR 2000). In marine animais, contaminants tend to concentrate in spécifie organs (Walker et al. 1996).

The fact that PCB s accumulate preferentially in fatty tissues implies that caution must be taken in comparing levels of contamination in different organisms. Different amounts of PCBs can be accumulate in the varions organs, having quite different implication for a fat animal than for an emaciated one. Accumulation rates vary among species, but also within a species according to factors such as âge, sex, stage in the breeding cycle, as well as exposure concentrations or feeding habits (Van der Oost et al. 2003). Bioaccumulation is a precursor to ail Chemical toxicity: without some degree of accumulation, even if slight, toxic action in organism target site(s) cannot take place.

1.2.2. Biological effects of PCBs

Experimental studies hâve shown that PCBs are capable of producing a wide variety of toxic effects in exposed organisms, some of the most common include neurotoxicity, immune dysfunction, reproductive and developmental effects, and cancer (Harding & Addison 1986, Zabel et al. 1995, Chapman 1996, Krogenaes 1998, Coteur et al. 2001). PCBs are of concem primarily because of their potential for causing chronic effects following long-term, low-level exposure (Walker et al. 1996, OSPAR 2000). The effects of substances on biota are dépendent on a number of factors and processes including bioavailability, bioaccumulation, toxic potency and the capacity of the organism to metabolize the substance (Fig. 4). Marine contamination by PCBs poses a relatively well-documented risk to the health of marine organisms, which can occur at levels ranging from subcellular effects to ecosystem effects (Tanabe & Tatsukawa 1992, Elkus et al. 1992, Norstrom & Muir 1994, Bello et al. 2001).

Figure 4. Model describing the fate of lipophilie xenobiotics in organisms (Hodgson & Levi 1993)

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a. Subcellular and cellular effects

To gain a full understanding of the toxic effects of a Chemical, it is necessary to link initial molecular interactions to conséquent effects at higher levels of organization. The extent to which such a molecular interaction occurs is, in general, related to the dose received, although the relationship is rarely a simple one (Walker et al. 1996). Molecular interactions between the xenobiotic and sites of action, which lead to toxic manifestations, may be highly spécifie for certain types of xenobiotics and organisms or non-specific, because of the variety of sites of action, which can occur in one species and not in other ones (Fig. 5).

Détoxication

n

Monooxygenase System

Initial Chemical

Repair ofDNA

Original State

Figure 5. Pathways for activation and détoxification of organic Chemicals (Walker et al. 1996)

Activation

w'jthlsNA'--- ^ Mutation--- Carcinogenicity

Subcellular effects of pollutants can be out of two types: those which serve to protect the organism against the harmfui effects of the Chemical (viz. détoxification via e.g. induction of monooxygenases or induction of metallothioneins), and those which do not (e.g. inhibition of AchE, formation of DNA adducts) (Table 1). Protective mechanisms fonction by reducing the contaminant concentration in the cell (e.g. some PCB congeners induce enzymes that metabolize them) or by reducing the bioreactive fraction of the contaminant concentration.

One of these mechanisms is achieved through the monooxygenase System, whose fonction is

to increase the rate of production of water-soluble métabolites and conjugates of low toxicity,

which can be excreted. However, in some cases, metabolism leads to the production of highly

reactive métabolites, that can cause more damage than the parent compound.

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Table 1. Protective and non-protective responses to Chemicals (Walker et al. 1996).

Type of effects Example Conséquences

Protective Induction of monooxygenases Induction of metallothionein

Increase in rate of metabolism of pollutant to more water- soluble métabolite and thus increase in rate of excrétion Increase the rate of binding sites with metals to decrease bioavailability

Non-protective Inhibition of AChE Formation of DNA adducts

Toxic effects seen above 50% inhibition May cause harmful effects if leading to mutation

These Chemical surveillance Systems hâve evolved as mechanisms for recognizing a broad range of Chemical structures and initiating appropriate responses, such as the biotransformation and élimination of toxic compounds (Brattsen 1979, Nebert & Gonzalez 1987, Gonzalez & Nebert 1990). The enzymatic components of this inducible biotransformation System are now well-known and include monooxygenases in the cytochrome P450 (CYP) superfamily as well as conjugating enzymes such as the glutathionetransferases and glucuronosyltransferases. The sensory component of this System consists of soluble receptors that regulate the expression of the biotransformation and transporter genes in response to environmental Chemicals. These receptors include several members of the steroid/nuclear receptor superfamily (Kliewer et al. 1999a,b, Savas et al.

1999, Waxman 1999, Honkakoski & Negishi 2000) as well as the aryl hydrocarbon receptor (AhR, Fig.6).

BTCDCa BtcddB

AhR --- > AhR --- > AhR

■arÎ^

cvtoplasma

;. ; ccl] nucieus' ;

^---

>1

CYPlAmRNAl CYPIA protcin

Figure 6. Hypothesized induction mechanism of CYPIA (Bucheli & Fent 1995)

The adaptive fonction of the AhR has been studied for more than 30 years, leading to the

prédiction and then discovery of the AhR as an ‘induction receptor’ that Controls the

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induction of adaptive enzymes, especially CYPIA (Poland et al. 1976, Whitlock 1999). The AhR is now known to recognize an impressive range of Chemical structures, including non- aromatic and non-halogenated compounds (Denison et al. 1998). In regulating biotransformation enzymes, the AhR serves an important adaptive fonction, but the function of this protein is much more complex: studies dealing with the toxicities associated with exposure to TCDD and related compounds showed that these Chemicals are interfering with important physiological fonctions in addition to inducing biotransformation enzymes (Poland

& Knutson 1982, Pohjanvirta & Tuomisto 1994).

Physiological and morphological parameters are higher-level responses that follow Chemical and cellular interactions. They are generally indicative of irréversible damages (Hinton et al.

1992). When a pollutant enters a cell, it may trigger certain biochemical responses, or it may be stored within a compartment, preventing interférences with essential biochemical components of the cells.

Many alterations may persist even after the exposure to a toxicant has ceased so that host responses to prior toxicity can also be used to détermine effects. Responses are relatively easily recognized, provided that proper reference and control data are available. Nowadays, sufficient information is at hand to assemble cellular or histopathological biomarker approaches and to apply them in integrated field studies (Hinton 1994).

b. Immunological ejfects

The immune System of an organism maintains a close and efficient surveillance of the body in order to react against infection and infestation or eliminate dysregulated protein expression.

This System can be divided into two forms of immunity : acquired -or spécifie- immunity and innate -or nonspecific- immunity (Roitt et al. 1993). Acquired immunity provides rapid, spécifie, and sélective reaction against a given infections agent, but requires a previous exposition to the same agent. Innate immunity is less spécifie, but protects the organism whithout previous contact with the infections agent. Innate immunity is présent in ail metazoan animais, whereas acquired immunity would be présent only in vertebrates.

Owing to the complexity of the immune System, several authors hâve suggested using a tiered approach for examining immunotoxicity in mammals and lower vertebrates (Vos 1980, Miller 1985, Luster et al. 1988, Weeks et al. 1992). Although invertebrate immunity relies on the innate System for host defence there are wide ranging strategies for eliminating microbes.

Assays of immunocompétence for invertebrates can therefore be approached at different

levels of organization (Pipe et al. 1995) : (1) the apparatus for immunity (total/differential

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blood cell counts, hemopoietic tissues), (2) the mechanisms of immunity (phagocytosis, blood cell prolifération, release of antimicrobial molécules), and (3) the efficiency of the immune response (susceptibility to infection by model agents). Most studies of environmental modulation of the immune fonction in marine invertebrates hâve focused on heavy metals, but some organic compounds hâve been adressed, both in laboratory and field studies (Fischer

1988, Anderson 1993, Pipe & Coles 1995, Coteur et al. 2001, 2003a).

c. Individual ejfects

Chemically-induced disorders hâve been attributed to PCBs at the level of the individual. The existence of repairing and détoxification mechanisms (e.g. mixed-function oxidases) involves that a biological response measured at a given level of biological organization might not be detected at a higher organization level (Luoma & Carter 1991, George & Olsson 1994, Goldstein 1995). Effects on individuals intégrale these latter mechanisms and are therefore highly relevant of the actual deleterious effects of a contaminant from a biological viewpoint, but these responses are generally slow and not spécifie of a given contaminant. Individual responses include for instance direct increase in mortality rates or interférence with processes of resource acquisition. These effects may resuit in slower population growth or population décliné. At the individual level, fertilization rate or embryonic development are commonly used as markers, and represent a good compromise: these responses are quite fast (from a few hours to a few days) and ecologically relevant because the reproductive success and the maintaining of populations rely directly on these processes (Dinnel et al. 1988, Gray 1989, Langston 1990, Weis & Weis 1991, Wamau et al. 1996a).

d. Population and ecological ejfects

The presence of PCBs into the marine environment is known to provoke toxic effects in biota,

which vary with the intensity and duration of exposure (Long et al. 1995). PCBs may exert

dramatic effects on relatively tolérant species (as determined by laboratory testing) by a

number of ecological mechanisms (Walker et al. 1996). Indeed, the direct influence of

contaminants on predators and grazers can lead to cascading indirect effects on more tolérant

species in other trophic levels. The direct effects of contaminants on sensitive species may

also alter compétitive interactions within the résistant populations of producers and

consumers of a given community. Similarly, disturbance rates or resource availability may be

influenced by the presence of contaminants such as PCBs, leading to important modifications

in ecosystem processes, such as décomposition rates of the organic matter, oxygen dynamics

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and nutrient cycling (Walker & Livingstone 1992). It bas also been suggested that localized toxicant-induced moitality may alter metapopulation dynamics, and bave significant impacts on non-exposed groups (Spromberg et al. 1998). Mecbanisms associated to population and community dynamics can vary in potentially complex fasbion following PCB exposure. The presence of PCB s in marine ecosy stems can clearly cause a wide range of indirect ecological effects that can be as or more significant than the direct (toxic) effects triggered by the contaminant (Feldman et al. 2000).

It is now widely adopted that communities and ecosystems are much more than the sum of their discrète parts and potentially intense indirect influences of realistic PCB exposures should be incorporated into an integrated ecotoxicological approach (Walker et al. 1996).

1.2.3. Biomarkers of PCB exposure

The need to detect and assess the impact of contaminations in the marine environment has led to the development of markers of biological effect (biomarkers) at varions organizational levels (Huggett 1992, Livingstone 1991, Livingstone et al. 2000). Several définitions hâve been given for the term ‘biomarker’, which is generally used in a broad sense to include almost any measurement reflecting an interaction between a biological System and a potential hazard, which may be Chemical, physical or biological (WHO 1993). A biomarker is defined as a change in a biological response (ranging from molecular through cellular and physiological responses to behavioural changes) which can be related to exposure to environmental Chemicals or to their toxic effects (Peakall 1994). According to NRC (1987) and WHO (1993), biomarkers can be subdivided into three classes:

• Biomarkers of exposure: allow détection and quantitation of an exogenous substance or its métabolites or the product of an interaction between this xenobiotic and some target molécules or cells (e.g. DNA or protein adducts, formation of spécifie métabolites,...),

• Biomarkers of effect: indicate measurable biochemical, physiological or other alterations within tissues or body fluids of an organism that can be recognized as associated to an established or possible health impairment or disease (e.g. reproductive, developmental, endocrine or genetic toxicity),

• Biomarkers of susceptibility : indicate the inhérent or acquired ability of an organism to

respond to the challenge of an exposure to a spécifie xenobiotic substance, including genetic

factors and changes in receptors which alter the susceptibility of an organism to that exposure

(e.g. activity of enzymes implied in activation or détoxification of a spécifie Chemical or

DNA repair capacity for spécifie types of DNA damage).

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Table 2. Biomarkers at different organizational levels (Walker et al. 1996)

Organizational level Example of biomarker Binding to a receptor TCDD binding to Ah receptor

Nonphenyls binding to oestrogenic receptor Biochemical response Induction of monooxygenases

Vitellogenin formation Physiological alterations Eggshell thinning

Feminization of embryos Effects on individuals Behavioural changes

Scope for growth

a. Motecular markers of PCB exposure

Molecular markers hâve been used extensively in environmental monitoring as part of integrated programmes (Bayne et al. 1988, Hylland et al. 1996, Schlenk et al. 1996). The main advantages provided by this level of organization are ;

• an integrated measure of the bioavailable fraction of contaminants

• the démonstration of causality through mechanistic understanding

• the identification of different routes of exposure and their relative importance

• the détection of exposure to readily metabolized contaminants.

The most widely and best studied biomarker of PCB exposure is the induction of cytochrome P450 (CYP)-dependent monooxygenase. Payne & Penrose (1975) and Payne (1976) were among the first to make use of this enzymatic complex as a biomarker, reporting elevated cytochrome P450 activity in fish from petroleum-contaminated sites. The multiple forms of CYP catalyze a wide variety of monooxygenation reactions that contribute to cellular oxidative metabolism in both prokaryotes and eukaryotes (Gibson & Skett 1994, Nelson et al.

1996). The products of the CYP super gene family undertake the oxidation of endogenous

substrates, e.g. fatty acids and steroid hydroxylation but some CYP gene families (CYPl,

CYP2 and CYP3) can also catalyze the oxidation of xenobiotics (Nelson et al. 1996). CYPl

may be induced in organisms exposed to spécifie aromatic and chlorinated hydrocarbons,

such as dioxins, furans, polyaromatic hydrocarbons (PAHs), or PCBs (Stegeman & Hahn

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1994). An élévation of CYPIA* levels may therefore indicate exposure to these inducers. The use of CYPIA induction as a biomarker for the pollution of aquatic ecosystems by organic contaminants has mainly been based on fish.

Whereas the enzyme System and its inductibility hâve been studied extensively in vertebrates, less is known in invertebrates (Livingstone et al. 2000). Different invertebrates hâve been screened for the occurrence of the CYPIA System. It has been reported in four phyla : Annelida. Arthropoda, Echinodermata and Mollusca (Lee 1981). Moore et al. (1980) found that some components of a xenobiotic détoxification System were présent in the blue mussel Mytilus edulis, but with limited metabolizing capacity for organic xenobiotics. The use of CYPIA as a biomarker was assessed with several molluscs (Stegeman 1985, Livingstone et al. 1989). Yawetz et al. (1992) observed an induction of CYPIA content by PCBs in molluscs. Enhanced CYPIA activity was found in the marine polychaete Nereis virens after exposure to benzo[a]pyrene or PCBs (Lee et al. 1981). Organisms from oil-contaminated sites showed several times higher CYPIA activities and lacked or had undevelopped gametes (Fries & Lee 1984). Studies on crustaceans provided controversial results, but in any case, crustacean CYPIA is less sensitive to induction than fish (James 1989). The CYPIA System is also présent in several echinoderms species (den Besten et al. 1991). Evidence for the presence of P450 enzymes belonging to the CYPl, CYP2, and CYP3 subfamilies hâve been obtained in sea stars (den Besten et al. 1993). Recently, the first echinoderm CYP genes were identified by Snyder (1998) in digestive tissues of an echinoid {Lytechinus anamesis).

b. Immunological markers

Pollution-induced suppressive effect on the immune System was found to lead to enhanced disease in organisms (Pipe & Coles 1995). Therefore, immunocompétence assays hâve been increasingly used as biomarkers of environmental contamination in the last years. Monitoring the immune System as a target for toxicity is difficult, given the complexity and self- regulatory nature of the immune network, so that conventional dose-response relationships may not always be observed. As with other biomarker responses, immune responses provide an integrated measure of exposure over time and may reflect the combined results of simultaneous exposure to several Chemicals. It is, however, not possible to détermine which

“ According lo Slcgcman cl al. (1992), the hydrocarbon-induciblc isoenzyme cyloebrome P4501A is referred to

as CYPIA. Hitberto, only tbe respective isoenzyme of rainbow trout {Oncorhyncus mykiss) can conclusively bc

lcrmed CYPl Al (Hcilmann et al. 1988), wbercas, for ail otber species, CYPIA is more appropriatc.

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Chemical has caused the observed effect as none of the changes in immune function can be attributed to a spécifie compound or class of Chemicals (Wester et al. 1994).

The use of invertebrate immunotoxicology, although of increasing interest whithin environmental monitoring studies, is still very much in its infancy (Livingstone et al. 2000).

Risk assessment of spécifie compounds in terms of immunomodulation leading to enhanced disease susceptibility for individuals or populations whithin a particular ecosystem has not yet been attempted. Indeed, much of the fondamental information on invertebrate immune responses and disease susceptibility is not available.

Among the immune fonctions, oxidative stress is widely investigated. The interest of

oxidative stress in ecotoxicological applications is based on the oxygen paradox; this

molécule is fondamental for many biochemical pathways in aérobic organisms, but its

consumption generates the intracellular formation of potentially toxic reactive oxygen species

(ROS). Despite the fact that basal oxyradical production is normally counteracted by a

complex antioxidant System, several pollutants are known to enhance the intracellular

génération of ROS through different mechanisms including the redox cycle, the cytochrome

P450-dependent oxidative metabolism of aromatic hydrocarbons and the Fenton reaction in

the presence of some transitional metals (Livingstone 1998). From the standpoint of

biomarkers it is useful to understand how antioxidants react to xenobiotic-mediated

enhancement of oxyradical production but the complexity of interactions between pro-oxidant

factors and cellular targets often precludes this possibility. Variations of individuals

antioxidants are difficult to predict and they often vary according to the class of Chemicals

tested, species sensitivity and several environmental and biological factors (Winston & Di

Giulio 1991). Induction of antioxidant defences is referred to as a counteracting response of

exposed organisms but the same antioxidants can be depleted when overwhelmed. Depending

on the duration and intensity of the pro-oxidant stressor, antioxidant defences might only be

induced during the first phase of the response, while in other conditions organisms can exhibit

no variations or transitory responses before adaptive mechanisms occur (Regoli & Principato

1995). Ail these possibilities (and their combinations) hâve been reported (Winston & Di

Giulio 1991) and the complexity of antioxidant responses to pollutant exposure often leads to

a controversy about the use of oxidative stress in ecotoxicological applications.

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1.3. Contamination of the North Sea by PCBs

1.3.1. The North Sea

The depth of the North Sea is not uniform : shallowest région is located near Dover (30 m), getting deeper towards the west (up to 100 m) and the north (up to 700 m). Water masses resuit from the mixing of NE Atlantic waters, précipitations, and river inputs. Seasonal variations of salinity are relatively low (salinity remains around 35%o ail year long), except in Coastal régions, where the influence of large estuaries can bring it down to 32%o. Water masses circulation has been modelized using radionuclide data ; grossly, water masses circulate according to an anti-clockwise direction in the North Sea (Fig. 7).

Figure 7. Diagram of the general water circulation in the North Sea (NSTF 1993)

Approximately 164 million people live in the North Sea catchment area. Numerous large rivers (e.g. Rhine, Scheldt, Elbe, Thames) flow through this heavily urbanized and industrialized région, providing significant inputs of several different pollutants into the North Sea. Hence, the North Sea constitutes the ultimate repository for a large range of domestic and industrial contaminants.

The Southern région of the North Sea (the Southern bight) is considered as one of the most

contaminated area. Certain researchers consider that there is no compartment of the Southern

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bight (seawater, sédiments, biota) which is not altered anthropogenically in a way or another (Rygg 1985, Kersten et al. 1994).

1.3.2. Origin and fluxes of PCB contamination in the North Sea

Inputs of contaminants in the North Sea occur via three main routes: direct, riverine and atmospheric inputs. The relative importance of each input route differs among the régions, and according to contaminants considered (OSPAR 2000). Direct inputs of contaminants arise mainly as a conséquence of municipal and industrial discharges in Coastal waters and from offshore activities and dumping. Riverine inputs extend along the coasts, and constitute another important contribution to contamination. Atmospheric inputs are an important source to the marine environment for several substances including heavy metals (e.g. mercury and lead), PCBs and some nitrogen compounds. The sources of atmospheric inputs may be located within or outside the North Sea area as PCBs can be transported on a global scale through the atmosphère.

Océan currents are also important in the transport and distribution of PCBs in the North Sea.

Although PCB concentrations in seawater are extremely low, the largeness of the water volumes transported implies that fluxes are large. PCBs, as many other contaminants, get adsorbed onto particulate matter upon which the transport path and fate of substances largely dépend (Olsen et al. 1982, Balls 1988). The résidence time of dissolved substances in the North Sea is 1 to 3 years (Otto et al. 1990). However, over 70% of the substances associated with the suspended matter remain in the North Sea or in associated sédimentation areas such as Wadden Sea, Skagerrak, Norwegian Trench and estuaries (Eisma 1973, Eisma & Kalf 1987, Eisma & Irion 1988). Whereas the Atlantic Océan is the major source of suspended matter in the Southern North Sea (McManus & Prandle 1997), in the Dutch Coastal zone the dumping of dredged material and the riverine input from the Scheldt and Rhine are also relatively important sources of suspended matter and associated substances (Eisma 1973, Eisma & Kalf 1987, Van Alphen 1990, Lourens 1996).

Sédiments are subject to resuspension and bioturbation, which can lead to the remobilization

of PCBs (which become available again to organisms) or to their burial in deeper layers of

bottom sédiments. Although it is not possible to dérivé reliable estimâtes of inputs because

most PCB concentrations are below the limit of détection, estimated fluxes derived for the

North Sea are in the range of 0.13 - 2.4 t yr ' for the 1990 to 1995 period (OSPAR 2000).

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1.3.3. PCBs in benthic ecosystems of the North Sea

PCB contamination levels in North Sea biota hâve been mainly characterized in dab {Limanda limanda), blue mussels (Mytilus edulis), and common sea stars (Asterias rubens) (e.g.

Stebbing et al. 1992, den Besten et al. 2001, Coteur et al. 2003a, Stronkhorst et al. 2003).

Dab is the most abundant flatfish species in the North Sea, with an estimated biomass of about 2 million tons (Daan et al. 1990). Because it is a demersal fish with a large géographie distribution and abundance, and because it is sensitive to PCB exposure (Sleiderinck et al.

1995), it has been routinely used in pollution monitoring programmes in the North Sea (North Sea Task Force; Joint Monitoring Programme) (Stebbing et al. 1992, NSTF 1993a,b).

However, regarding indicating purposes, dabs présent a major flaw as they are known to migrate during spawning periods over relatively long distances, which makes difficult to détermine the origin of the contamination (Rijnsdorp et al. 1992).

The blue mussel {M. edulis) is a sedentary, filter-feeding bivalve of commercial importance, which has long been considered amongst the best suited sentinel organisms for monitoring marine pollutions (Goldberg et al. 1978). It has been therefore widely used as bioindicator in North Sea pollution studies. M. edulis efficiently takes up and concentrâtes PCBs to levels well above those présent in the surrounding seawater. It provides information on spatial and temporal pollution trends and enables the identification of contamination « hot spots » in Coastal areas (e.g. Phillips 1990). Moreover, the blue mussel exhibits a sériés of biochemical (sublethal) responses to pollutants (see Livingstone 1991 for a review) which may be used as early warning signais of exposure (McCarthy & Shugart 1990, Huggett 1992).

The common NE Atlantic sea star Asterias rubens (Fig. 8) is also an interesting test organism

because of its key position as top predator in the food chain “seston-mussels-sea stars”. In the

North Sea, this echinoderm is known to influence the structure and functioning of benthic

communities (Menge 1982, Hayward & Ryland 1990, Hostens & Hammerlink 1994). It lives

on or in proximity of sédiments (the main réservoir of contaminants in the marine

environment) and can be found in very diversified biotopes from the surface to depths

reaching 650 m. A. rubens is also able to colonize low salinity areas, such as estuaries, which

are under direct influence of contamination carried by large rivers.

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Figure 8. The common NE Atlantic sea star Asterias rubens L. (Hayward & Ryland 1996)

In semi-field studies, A. rubens has been demonstrated to efficiently accumulate PCBs, leading to deleterious effects on reproductive processes (den Besten et al. 1989, 1990a). This sea star has largely proved its value or potential value as a bioindicator species for a wide range of anthropogenic contaminants (e.g. PCBs, metals, organometals) in laboratory and/or in field studies (e.g. Bjerregaard 1988, Everaarts & Fischer 1989, Temara et al. 1997a, 1998a,b, Wamau et al. 1999, Coteur et al. 2003a, Stronkhorst et al. 2003). Although there is a wealth of studies showing the quality of the sea star as a bioindicator, no study has investigated bioaccumulation processes of PCBs in the sea star. However, such data is a prerequisite to assess the value of A. rubens as a bioindicator of PCB contamination.

Available data on effects of PCB exposure in sea stars are also scarce, but hâve shown that these organisms are affected (den Besten 1998, den Besten et al. 1990a, 1993). Existing studies hâve focused on the subcellular level using the microsomal activity of the CYP enzyme System (den Besten et al. 1991, 1993) and steroid metabolism (den Besten et al.

1991). In vitro or in vivo exposure of sea stars to PCBs has elsewhere been reported to

decrease DNA integrity (Sarkar & Everaarts 1995).