L B M
Bioaccumulation and Effects of Polychlorinated Biphenyls (PCBs) in the Sea Star Asterias rubens L.
Bruno Danis March 2004
Thesis submitted in fulfilment of the degree of :
Doctor in Sciences
Supervisors:
Dr Michel Warnau
Dr Philippe Dubois
A ceux qui nous manquent et qui veillent sur nous…
A CKNOWLEDGEMENTS -R EMERCIEMENTS
Quand vient 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 affuté 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 relevions 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.
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
2Stanislas 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 Ideal, 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, « Gomez da Costa » Sergio, «DivX le Breton » Vinz, « le Dubbe » Phil,
« Sand » Cristi, « le Dim-Dim », « le Loron », « Chipito» GillesD, « Snore » Geoff, « Mac Guy-ver » Guy, « 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 » GillesR, « Aussie » Raph, « Cotten » Guillemette, « 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…).
Je remercie les GMs de mon Comité d’Accompagnement, les Drs Christiane Lancelot et Guy 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 Gwenaëlle Leclercq et le presque-Dr Grégory 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. Gilles 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.
Zoé, mon bébé-crabe, et Liam, mon bébé-grogne, papa vous dédie ce travail.
T ABLE OF C ONTENTS
ACKNOWLEDGEMENTS-REMERCIEMENTS... 5
TABLE OF CONTENTS... 9
SUMMARY ... 11
I. GENERAL INTRODUCTION ... 15
I.1. POLYCHLORINATED BIPHENYLS... 15
I.1.1. General information ... 15
I.1.2. Analysis ... 16
I.1.3. International recommendations ... 18
I.2. PCBS IN THE MARINE ENVIRONMENT... 19
I.2.1. Caracterization of PCB contamination ... 20
I.2.2. Biological effects of PCBs... 22
I.2.3. Biomarkers of PCB exposure ... 27
I.3. CONTAMINATION OF THE NORTH SEA BY PCBS... 31
I.3.1. The North Sea ... 31
I.3.2. Origin and fluxes of PCB contamination in the North Sea... 32
I.3.3. PCBs in benthic ecosystems of the North Sea ... 33
II. OBJECTIVES ... 35
III. EXPERIMENTAL CONDITIONS ... 37
III.1 NON-COPLANAR VS. COPLANAR CONGENER-SPECIFICITY OF PCB BIOACCUMULATION AND IMMUNOTOXICITY IN SEA STARS... 39
III.2 DELINEATION OF PCB UPTAKE PATHWAYS IN A BENTHIC SEA STAR USING A RADIOLABELLED CONGENER ... 59
III.3 COPLANAR PCB UPTAKE KINETICS IN THE COMMON SEA STAR ASTERIAS RUBENS AND SUBSEQUENT EFFECTS ON ROS PRODUCTION AND CYP1A INDUCTION... 73
III.4 CONTRASTING EFFECTS OF COPLANAR VS NON-COPLANAR PCB CONGENERS ON IMMUNOMODULATION AND CYP1A LEVELS (DETERMINED USING AN ADAPTED ELISA METHOD) IN THE COMMON SEA STAR ASTERIAS RUBENS L. ... 95
IV. FIELD CONDITIONS...113
IV.1 CONTAMINANT LEVELS IN SEDIMENTS AND ASTEROIDS (ASTERIAS RUBENS L., ECHINODERMATA) FROM THE BELGIAN COAST AND SCHELDT ESTUARY: POLYCHLORINATED BIPHENYLS AND HEAVY METALS...115
IV.2 BIOACCUMULATION AND EFFECTS OF PCBS AND HEAVY METALS IN SEA STARS (ASTERIAS RUBENS, L.) FROM THE NORTH SEA: A SMALL SCALE PERSPECTIVE...141
IV.3 ECHINODERMS AS BIOINDICATORS, BIOASSAYS AND IMPACT ASSESSMENT TOOLS OF SEDIMENT- ASSOCIATED METALS AND PCBS IN THE NORTH SEA...159
IV.4 LEVELS AND EFFECTS OF PCDD/FS AND C-PCBS IN SEDIMENTS, MUSSELS AND SEA STARS OF THE INTERTIDAL ZONE IN THE SOUTHERN NORTH SEA AND THE CHANNEL...185
V. GENERAL DISCUSSION ...205
THE BIOACCUMULATION OF PCBS IN SEA STARS...205
THE EFFECTS OF PCBS IN SEA STARS...208
CONCLUSIONS-RECOMMENDATIONS...211
VI. REFERENCES...213
VII. ANNEX STUDIES ...249
VII.1 BIOACCUMULATION OF PCBS IN THE SEA URCHIN PARACENTROTUS LIVIDUS: SEAWATER AND FOOD EXPOSURES TO A 14C-RADIOLABELLED CONGENER (PCB 153). ...251
VII.2 BIOACCUMULATION OF PCBS IN THE CUTTLEFISH SEPIA OFFICINALIS FROM SEAWATER, SEDIMENT AND FOOD PATHWAYS. ...263
VII.3 MEASUREMENT OF EROD ACTIVITY!: CAUTION ON THE SPECTRAL PROPERTIES OF THE STANDARDS USED
...279
VII.4 EFFECTS OF PCBS ON REACTIVE OXYGEN SPECIES (ROS) PRODUCTION BY THE IMMUNE CELLS OF PARACENTROTUS LIVIDUS (ECHINODERMATA)...287
APPENDIX I!: CAPTIONS TO FIGURES ...299
APPENDIX II!: CAPTIONS TO TABLES ...304
APPENDIX III!: CAPTIONS TO EQUATIONS...308
S UMMARY
PCBs are among the most problematic marine contaminants. Converging towards the oceans via the rivers and the atmosphere, they concentrate in sediments where they become a permanent threat to organisms living at their contact. PCBs are extremely resistant, 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 region.
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 present study focused on the characterization of PCB bioaccumulation in A. rubens exposed through different routes (seawater, food, sediments) and on subsequent 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 (CYP1A IPP) which, in vertebrates, is involved in PCB detoxification.
Experimental exposures carried out have shown that A. rubens efficiently accumulates 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 displayed the highest bioaccumulative potency and can therefore be considered as particularly interesting for field biomonitoring applications. Rectal caeca, which play a central role in digestion and excretion processes in sea stars, have also rised particular interest as results suggest these organs could be involved in the elimination of PCB 77 degradation products.
The field work carried out during the present 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 sediments 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 present 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 ELISA was used to measure CYP1A 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 CYP1A IPP according to a dose-dependent relationship. These results have
highlighted many similarities between the dioxin-like responsiveness of CYP1A 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, CYP1A IPP induction
was found to be significantly related to PCB levels determined in bottom sediments. It can
thus be considered as a valuable biomarker. Further research is however needed to better
characterize the influence of physico-chemical and physiological parameters on CYP1A induction to refine the interpretation of the information gathered via this biomarker.
Results obtained in our study have lead to questionning international regulations applying to PCB biomonitoring in the marine environment. For instance, we strongly suggest that the selection 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, determination of c-PCB concentrations was extremely difficult due to analytical limitations!; however, nowadays, these problems have been overcome and do no more justify their exclusion from monitoring studies.
Although A. rubens appeared to be quite resistant 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 region undergo similarly low risks, or that sea star-
structured ecosystems may not become affected in the long-term.
I. G ENERAL I NTRODUCTION
I.1. Polychlorinated biphenyls
I.1.1. General information
Polychlorinated biphenyls (PCBs) have 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 thermic insulation) led to their proliferation 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 molecule.
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 (IUPAC).
Figure 1. Numbering system for sites of chlorine on a biphenyl molecule (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 have been detected in the
environment. The relative abundance of the different congeners in environmental samples depends 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 considerable distances, as a consequence 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 have been detected in the most remote regions, such as the Arctic, the North Atlantic and even the Antarctic and deep-sea (e.g. AMAP 1998), where there is no anthropogenic emission sources.
Data on the global production and use of PCBs has been collected for decades, but more work is needed for the interpretation of past, present 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 atmosphere (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.
I.1.2. Analysis
Most analyses of PCB levels in the environment have been reported as Aroclor equivalents.
This was once due to necessity because traditional packed-column gas chromatography (GC) were not able to resolve 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
dechlorination by anaerobic bacterial action when present in sediment 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 aerobic degradation in particular conditions (Furukawa 1982). Consequently, original PCB mixtures become depleted in the most degradable congeners and enriched in metabolites of the latter ones and in “resistant”
congeners. In biota, PCB composition can also be altered, depending on the uptake, metabolization and depuration 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 electron capture detection (HRGC-ECD) or high resolution gas chromatography with low resolution electron 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 dependent 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-p-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 equivalents 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).
Figure 2. Molecular configuration of 2,3,7,8 TCDD and PCB 169 (Metcalfe 1994)
I.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 regional and global scales in the last decades (Wania & Mackay 1996, UNECE 1998, UNEP 2001).
Due to their beyond-boundaries transport, political problems have also arisen. International
agreements have thus come 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, emission 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 regional 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 regimes 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
disposal of POPs wastes and provisions on the reduction of emissions 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 have 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 Conference 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 sediments 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 all non-ortho substituted congeners should be included in analysis programmes (Metcalfe 1994).
I.2. PCBs in the marine environment
The ultimate sink for many contaminants is the marine environment, following either direct
discharges or hydrologic and atmospheric processes (Stegeman & Hahn 1994). Since the late
1960s, PCBs are known to be present in substantial quantities in marine sediments, as well as
in marine biota (Jensen et al. 1969). PCBs accumulate in the organic phase, such as biota and
the organic fraction of sediments, transfering between these compartments according to the
model presented in Fig. 3. PCBs persist in the marine environment for several decades: most
PCBs only exist in trace concentrations, but all have extensive half-lives (degradation 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).
Figure 3. Contaminants transfers between compartments in a coastal model (Moore et al. 2002)
I.2.1. Caracterization of PCB contamination
a. Seawater
PCBs are hydrophobic compounds, i.e. they have extremely low water solubilities.
Concentrations in ocean water are generally very low, making reliable quantification technically difficult. PCB concentrations in filtered ocean water are usually reported to be in the low pg l
-1range. In contrast, PCBs are highly lipophilic 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 µm 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 have 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
considerable 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 have a high productivity and attract a wide range of
animals, which thus receive a PCB-enriched diet. Since the upper millimetre of the sea is also
enriched in microorganisms and zooneuston (including larvae), great concern has been
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 animals 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 animals, entering food chains by this route.
b. Sediments
Sediments are repositories for physical and biological debris and are considered as sinks for a wide variety of chemicals (Clark 1997). The concern associated with PCBs sorption to sediments is that many organisms spend a considerable portion of their life-cycle on or in marine sediments. This provides a path for PCBs to reach higher trophic levels. Direct transfer of contaminants from sediments or interstitial water to organisms is considered to be a major route of exposure (Walker & Peterson 1994). PCBs are present in much higher concentrations in sediments than in overlying water. Sorption to sediments is the predominant removing mechanism for PCBs from the water column. The analysis of PCBs in sediments has the advantage of integrating time variations. Once contaminated, sediments 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 consequence of their hydrophobic 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 sediments.
Once they have entered the organism, PCBs are stored within the fatty tissues, or in other
lipophilic 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 have the capacity to bioaccumulate and
to biomagnify PCBs, which results in body concentrations several orders of magnitude higher
than in seawater or in the food (OSPAR 2000). In marine animals, contaminants tend to concentrate in specific organs (Walker et al. 1996).
The fact that PCBs 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 various 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 age, 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 all chemical toxicity: without some degree of accumulation, even if slight, toxic action in organism target site(s) cannot take place.
I.2.2. Biological effects of PCBs
Experimental studies have 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 concern 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 dependent 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 lipophilic xenobiotics in organisms (Hodgson & Levi 1993)
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 consequent 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 specific 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).
Figure 5. Pathways for activation and detoxification of organic chemicals (Walker et al. 1996)
Subcellular effects of pollutants can be out of two types: those which serve to protect the organism against the harmful effects of the chemical (viz. detoxification 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 function 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 function is
to increase the rate of production of water-soluble metabolites and conjugates of low toxicity,
which can be excreted. However, in some cases, metabolism leads to the production of highly
reactive metabolites, that can cause more damage than the parent compound.
Table 1. Protective and non-protective responses to chemicals (Walker et al. 1996).
Type of effects Example Consequences
Protective Induction of monooxygenases Increase in rate of metabolism of pollutant to more water- soluble metabolite and thus increase in rate of excretion Induction of metallothionein Increase the rate of binding sites with metals to decrease
bioavailability
Non-protective Inhibition of AChE Toxic effects seen above 50% inhibition Formation of DNA adducts May cause harmful effects if leading to mutation
These chemical surveillance systems have evolved as mechanisms for recognizing a broad range of chemical structures and initiating appropriate responses, such as the biotransformation and elimination 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).
Figure 6. Hypothesized induction mechanism of CYP1A (Bucheli & Fent 1995)
The adaptive function of the AhR has been studied for more than 30 years, leading to the
prediction and then discovery of the AhR as an ‘induction receptor’ that controls the
induction of adaptive enzymes, especially CYP1A (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 function, 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 functions 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 irreversible 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 interferences 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 determine 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 effects
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 specific- immunity and innate -or nonspecific- immunity (Roitt et al. 1993). Acquired immunity provides rapid, specific, and selective reaction against a given infectious agent, but requires a previous exposition to the same agent. Innate immunity is less specific, but protects the organism whithout previous contact with the infectious agent. Innate immunity is present in all metazoan animals, whereas acquired immunity would be present only in vertebrates.
Owing to the complexity of the immune system, several authors have 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 immunocompetence for invertebrates can therefore be approached at different
levels of organization (Pipe et al. 1995)!: (1) the apparatus for immunity (total/differential
blood cell counts, hemopoietic tissues), (2) the mechanisms of immunity (phagocytosis, blood cell proliferation, release of antimicrobial molecules), and (3) the efficiency of the immune response (susceptibility to infection by model agents). Most studies of environmental modulation of the immune function in marine invertebrates have focused on heavy metals, but some organic compounds have been adressed, both in laboratory and field studies (Fischer 1988, Anderson 1993, Pipe & Coles 1995, Coteur et al. 2001, 2003a).
c. Individual effects
Chemically-induced disorders have been attributed to PCBs at the level of the individual. The existence of repairing and detoxification 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 integrate 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 specific of a given contaminant. Individual responses include for instance direct increase in mortality rates or interference with processes of resource acquisition. These effects may result in slower population growth or population decline. 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, Warnau et al. 1996a).
d. Population and ecological effects
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 tolerant 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 tolerant
species in other trophic levels. The direct effects of contaminants on sensitive species may
also alter competitive interactions within the resistant 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 decomposition rates of the organic matter, oxygen dynamics
and nutrient cycling (Walker & Livingstone 1992). It has also been suggested that localized toxicant-induced mortality may alter metapopulation dynamics, and have significant impacts on non-exposed groups (Spromberg et al. 1998). Mechanisms associated to population and community dynamics can vary in potentially complex fashion following PCB exposure. The presence of PCBs in marine ecosystems 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 discrete parts and potentially intense indirect influences of realistic PCB exposures should be incorporated into an integrated ecotoxicological approach (Walker et al. 1996) . I.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 various organizational levels (Huggett 1992, Livingstone 1991, Livingstone et al. 2000). Several definitions have 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 detection and quantitation of an exogenous substance or its metabolites or the product of an interaction between this xenobiotic and some target molecules or cells (e.g. DNA or protein adducts, formation of specific metabolites,…),
•!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 inherent or acquired ability of an organism to
respond to the challenge of an exposure to a specific 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 detoxification of a specific chemical or
DNA repair capacity for specific types of DNA damage).
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. Molecular markers of PCB exposure
Molecular markers have 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 demonstration of causality through mechanistic understanding
• the identification of different routes of exposure and their relative importance
• the detection 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 (CYP1,
CYP2 and CYP3) can also catalyze the oxidation of xenobiotics (Nelson et al. 1996). CYP1
may be induced in organisms exposed to specific aromatic and chlorinated hydrocarbons,
such as dioxins, furans, polyaromatic hydrocarbons (PAHs), or PCBs (Stegeman & Hahn
1994). An elevation of CYP1A
*levels may therefore indicate exposure to these inducers. The use of CYP1A 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 have been studied extensively in vertebrates, less is known in invertebrates (Livingstone et al. 2000). Different invertebrates have been screened for the occurrence of the CYP1A 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 detoxification system were present in the blue mussel Mytilus edulis, but with limited metabolizing capacity for organic xenobiotics. The use of CYP1A as a biomarker was assessed with several molluscs (Stegeman 1985, Livingstone et al. 1989). Yawetz et al. (1992) observed an induction of CYP1A content by PCBs in molluscs. Enhanced CYP1A 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 CYP1A activities and lacked or had undevelopped gametes (Fries & Lee 1984). Studies on crustaceans provided controversial results, but in any case, crustacean CYP1A is less sensitive to induction than fish (James 1989). The CYP1A system is also present in several echinoderms species (den Besten et al. 1991). Evidence for the presence of P450 enzymes belonging to the CYP1, CYP2, and CYP3 subfamilies have 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, immunocompetence assays have 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 determine which
* According to Stegeman et al. (1992), the hydrocarbon-inducible isoenzyme cytochrome P4501A is referred to as CYP1A. Hitherto, only the respective isoenzyme of rainbow trout (Oncorhyncus mykiss) can conclusively be termed CYP1A1 (Heilmann et al. 1988), whereas, for all other species, CYP1A is more appropriate.
chemical has caused the observed effect as none of the changes in immune function can be attributed to a specific 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 specific 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 fundamental information on invertebrate immune responses and disease susceptibility is not available.
Among the immune functions, oxidative stress is widely investigated. The interest of
oxidative stress in ecotoxicological applications is based on the oxygen paradox: this
molecule is fundamental for many biochemical pathways in aerobic 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
generation 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). All these possibilities (and their combinations) have 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.
I.3. Contamination of the North Sea by PCBs
I.3.1. The North Sea
The depth of the North Sea is not uniform!: shallowest region is located near Dover (30 m), getting deeper towards the west (up to 100 m) and the north (up to 700 m). Water masses result from the mixing of NE Atlantic waters, precipitations, and river inputs. Seasonal variations of salinity are relatively low (salinity remains around 35‰ all year long), except in coastal regions, where the influence of large estuaries can bring it down to 32‰. 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 region, 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 region 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
bight (seawater, sediments, biota) which is not altered anthropogenically in a way or another (Rygg 1985, Kersten et al. 1994).
I.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 regions, and according to contaminants considered (OSPAR 2000). Direct inputs of contaminants arise mainly as a consequence 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 atmosphere.
Ocean 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 depend (Olsen et al. 1982, Balls 1988). The residence 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 sedimentation areas such as Wadden Sea, Skagerrak, Norwegian Trench and estuaries (Eisma 1973, Eisma & Kalf 1987, Eisma & Irion 1988). Whereas the Atlantic Ocean 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).
Sediments 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 sediments. Although it is not possible to derive reliable estimates of inputs because
most PCB concentrations are below the limit of detection, estimated fluxes derived for the
North Sea are in the range of 0.13 – 2.4 t yr
-1for the 1990 to 1995 period (OSPAR 2000).
I.3.3. PCBs in benthic ecosystems of the North Sea
PCB contamination levels in North Sea biota have 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 geographic 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 present a major flaw as they are known to migrate during spawning periods over relatively long distances, which makes difficult to determine 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 concentrates PCBs to levels well above those present 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 series of biochemical (sublethal) responses to pollutants (see Livingstone 1991 for a review) which may be used as early warning signals 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 sediments (the main reservoir 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.
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, Warnau 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 have shown that these organisms are affected (den Besten 1998, den Besten et al. 1990a, 1993). Existing studies have 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).
II. O BJECTIVES
Oceans are the ultimate receptacle for many anthropogenic contaminants that converge towards them via the rivers and the atmosphere. Often displaying low solubility in seawater, these contaminants concentrate in sediments, where they become a persistant threat for benthic communities, particularly in coastal areas. PCBs are among the marine contaminants of highest concern. Their unique physico-chemical characteristics, at first exploited by the industry, have rapidly become a cause of concern!: PCBs are resistant to degradation, readily accumulated by marine organisms and highly toxic. Among the 209 possible congeners, coplanar PCBs display a dioxin-like conformation and have the highest toxic potential.
Information available about the impact of PCBs on marine benthic species in the North Sea is scarce. In addition most of the existing studies addressed species that are poorly representative of North Sea benthic ecosystems, which leads to uncertain characterization of the principal ecological impact of these contaminants.
The common NE Atlantic sea star Asterias rubens (L.) can be considered as an ecosystem- structuring species in the North Sea. In addition it is ubiquist, abundant and generally recognized as a good bioindicator species. Therefore, A. rubens was chosen as an experimental model to study PCB bioaccumulation and effects. Sea stars have already been used in the field to characterize the contamination status of a region, but no data regarding bioaccumulation efficiencies, body distribution or relative importance of the different uptake routes are available in the literature.
The main objectives of the present study were to examine -both in experimental and natural conditions- the bioaccumulation and body distribution of PCBs in A. rubens, and the biological consequences attributable to PCB exposure.
For this purpose, the accumulation biokinetics of structurally-contrasting PCB congeners were studied by exposing sea stars via different exposure routes (seawater, sediments or food), using either a mixture of PCB congeners (Chap. III.1) or single-congeners (Chap. III.2
& III.3). The effects of PCB exposure on the immune response (ROS production) and on the induction of detoxification mechanisms (CYP1A induction) have been studied in parallel (Chap. III.4).
The existing relationships between PCB levels measured in the environment and those
measured in sea stars were also examined (Section IV) in order to «!calibrate!» the
bioindicating value of A. rubens. The health status of sea stars was studied alongside (same
Section) by measuring different biological responses (ROS production and CYP1A induction)
in field-collected individuals. Finally, the last objective of this fourth section was to assess the
contamination status of the southern North Sea by measuring concentration levels of PCBs
and of other contaminants of concern, viz. heavy metals and dioxins, in sediments and biota.
III.1 Non-coplanar vs. coplanar congener-specificity of PCB bioaccumulation and immunotoxicity in sea stars
Danis B
1, Cattini Ch
2, Cotret O
2, Teyssié JL
2, Coteur G
1, Villeneuve JP
2, Fowler SW
2& Warnau M
21
!Laboratoire de Biologie marine (CP 160/15), Université Libre de Bruxelles, Av. F.D.
Roosevelt 50, B-1050 Brussels, Belgium
2
