Dr Paco Bustamante

Pierre Miramand

Evaluation of the variegated scallop Chlamys varia as a biomonitor of temporal trends of Cd, Cu, and Zn in the field

Keywords: Heavy metals, Trace elements, Bioaccumulation, Biomonitoring, Seasonal variation, Pectinid

This study assess the potential of the scallop Chlamys varia as a biomonitor of metal contamination in the field. Cd, Cu, and Zn concentrations were determined in the soft tissues and organs of individuals from the French Atlantic coast sampled over a one year period and belonging to a large range of size. All metals were selectively distributed among the different considered body compartments, and their concentrations were influenced by the size of the specimens or the sampling-season. The present work shows the importance to consider the body compartment, the sampling period and the size as important factors in studies aiming to use this scallop as a biomonitor species. Among tissues, the digestive gland and kidneys exhibited the highest metal concentrations whatever the season or the size. The digestive gland contained 65 and 48 % of the whole Cd and Cu body burdens, respectively, and kidney accounted for 85 % of the Zn load. Those tissues are therefore particularly recommended for use in biomonitoring programs.

INTRODUCTION

The variegated scallop Chlamys varia is a common seashell on the rocky shore of the French Atlantic coast. Important shoals of this scallop are encountered in Brittany and in Charente-Maritime where this scallop is targeted by commercial fisheries as well as for leisure activities. This species has been previously reported to bioaccumulate high concentrations of several trace elements including metals, rare earth elements and radionuclides [START_REF] Martoja | Effets de la contamination expérimentale par l'argent chez Chlamys varia L. (Bivalve Pectinidé). Données quantitatives, histologiques et microanalytiques[END_REF][START_REF] Bustamante | Etude des processus de bioaccumulation et de détoxication d'éléments traces (métaux lourds et terres rares) chez les mollusques céphalopodes et bivalves pectinidés. Implication de leur biodisponibilité pour le transfert vers les prédateurs[END_REF][START_REF] Bustamante | Concentration and distribution of 210 Po in the tissues of the scallop Chlamys varia and the mussel Mytilus edulis from the coasts of Charente-Maritime (France)[END_REF], Bustamante & Miramand 2004, 2005). Therefore, it has been suggested that the variegated scallop could be useful biomonitor to survey marine pollution. However, metal concentrations in bivalves may vary with biological and environmental factors such as age (size) and seasons (e.g. [START_REF] Bryan | The occurrence and seasonal variation of trace metals in the scallops Pecten maximus (L.) and Chlamys opercularis (L.)[END_REF][START_REF] Boyden | Effect of size upon metal content of shellfish[END_REF]. Therefore, there is a need to assess the variations of elemental concentrations according to the size and to the different periods of the year before using a species as a biomonitor. On the other hand, whole bivalves are often employed to monitor metals, faking the actual variations of the concentrations due to natural dilution by non-storage organs. In this context, it appears particularly important to make the link between studies aiming to describe the detoxification processes leading to the bioaccumulation of metals in the storage organs and those aiming to determinate valuable biomonitor species. Indeed, the strong capacity of scallops to bioaccumulate numerous trace elements at high concentrations in their tissues has been related to their very efficient detoxification systems in the digestive gland and kidneys (e.g. George et al. 1980[START_REF] Ballan-Dufrançais | Physiological features of heavy metal storage in bivalve digestive cells and amoebocytes : EPMA and factor analysis of correspondences[END_REF][START_REF] Stone | Cadmium binding components of scallop (Pecten maximus) digestive gland. Partial purification and characterisation[END_REF], Fowler & Gould 1988[START_REF] Mauri | Heavy metals in the Antarctic scallop Adamussium colbecki[END_REF][START_REF] Viarengo | Cu, Zn and Cd content in different tissues of the Antarctic scallop Adamussium colbecki: role of the metallothionein in heavy metal homeostasis and detoxification[END_REF]. Therefore, these tissues have to be considered separately in biomonitoring studies. Furthermore, given the high filtration rates of bivalves, gills would constitute a significant pathway of incorporation of metals via seawater (eg. [START_REF] Lukyanova | Binding of zinc and cadmium with cytoplasmic proteins in the Yezo scallop at the increased concentrations of zinc in water[END_REF][START_REF] Roméo | Metal distribution in different tissues and in subcellular fractions of the Mediterranean clam Ruditapes decussatus treated with cadmium, copper, or zinc[END_REF][START_REF] Widdows | Scope for growth and contaminant levels in North Sea mussels Mytilus edulis[END_REF]. Gonads are likely to display important variations of their weight according to the reproductive cycle, which might affect its metal concentrations.

The aim of this study was thus to investigate the variations of Cd, Cu, and Zn concentrations in the tissues and organs of the variegated scallop Chlamys varia over a one year period in order to assess the potential of this species as a biomonitor.

Concentrations and distribution of metals in the different body compartments were determined at different seasons and at different sizes. Statistical analyses were then performed to discriminate the influence of the different factors (i.e. body compartment, sampling period and size) in the variability of the metal concentrations.

MATERIAL AND METHODS

Sampling and sample preparation

Individuals of the variegated scallop Chlamys varia (Linné, 1758) were collected at low tide in the infralittoral zone in the Bay of La Rochelle, Atlantic coast of France. This site harbours a stable scallop population. Moreover, forbidden seashore fishing in this area generally allowed good sampling of the different size classes.

Scallop samplings were performed in November 1995 (n=139), February (n=265), June (n=228) and September 1996 (n=152). These periods correspond to the main stages of the reproductive cycle of the variegated scallop in this area [START_REF] Letaconnoux | Contribution à l'étude du pétoncle (Chlamys varia L.)[END_REF].

After collection, the scallops were depurated for 48 hours in filtrated seawater (close circuit, constant aeration, 34 p.s.u., 12h/12h light/dark cycle) to eliminate faecal and pseudo-faecal material. Whenever possible, they were then pooled by size classes of 10 individuals (up to 30 for the smallest ones) according to the height of the shell (called size in the rest of this work), and dissected. Scallop tissues were pooled to avoid, as much as possible, individual variations of the metallic concentrations. The characteristics of the pooled samples are given in Table 1.

For each individual, the digestive gland, kidneys, gonad, gills, and adductor muscle were separated from the rest of the soft parts (mantle, foot, heart and intestine), and pooled for each defined size class. However, for scallop below 20 mm, only the digestive gland and the adductor muscle allowed an accurate dissection. Thus, in the case of these small scallops, the remaining tissues also included gills, kidneys and gonad. In both case, remaining tissues were analysed in order to calculate the whole trace element contains of the soft parts.

Trace element analyses

Tissue samples were dried for several days at 80°C to constant weight. Whenever it was possible, two aliquots (100 to 300 mg) of each homogenised dry sample were digested with 4 ml of 65% ultrapure HNO 3 and 0.3 ml of ultrapure 70 % HClO 4 at 80°C until the solution was clear. Then acids were evaporated and residues were dissolved in 10 ml 0.3N ultrapure nitric acid.

Cadmium, copper and zinc were determined by flame and graphite furnace atomic absorption spectrophotometry with a Varian spectrophotometer Vectra 250 Plus with Deuterium background correction. Reference standard materials, dogfish liver Canada) and dogfish muscle DORM-2 (NRCC) were treated and analysed in the same way as the samples. The results for standard reference materials displayed recoveries of the elements ranging from 98% to 102% and the detection limits calculated for 100 mg of dry material were (µg.g -1 dry wt): 0.004 (Cd), 0.5 (Cu), 1.5 (Zn).

Metal concentrations in tissues and proportion in each body compartment are given relatively to the dry weight (µg.g -1 dry wt and % of the metal loads, respectively).

Data analyses

Statistical analyses of the data were performed by 1-, 2-, or 3-way analysis of variance (ANOVA) followed by the Tukey's multiple comparison test [START_REF] Zar | Biostatistical analysis[END_REF]. The variability explained by each factor was derived from the sum of squares. The significance for statistical analyses was always set at α = 0.05.

RESULTS

Relative contribution of the investigated factors to metal concentration variability

The mean Cd, Cu and Zn concentrations measured in the body compartments of the variegated scallop are presented in Tables 2,3 & 4, respectively. Among tissues, metals greatly varied from 2 (i.e. Cd and Cu) to 3 orders of magnitude (i.e. Zn). Statistical analyses were therefore used to discriminate the relative contribution of the different factors in the variability of the metal concentrations. With the exception of the smallest sizes for which pools are lacking, the whole set of data was analysed using 3-way ANOVA (Table 5). These analyses showed that among the three factors of variation considered (body compartment, period of sampling and size), the body compartment was the predominant factor for the three metals, accounting for 79.3 to 93.7% of the variation (Table 5). Other factors appeared to be less important except the interaction between body compartment and sampling-period factors, which indicated that seasonal variations of the metal concentrations were not synchronous in the different body compartment.

Therefore, 2-ways ANOVA were used to discriminate the influence of the samplingperiod and the size of the scallops in each of the body compartment (Table 6). These analyses revealed that the variability of essential Cu and Zn concentrations in the different tissues was explained by the sampling period, except for the Cu in the digestive gland which variability was mainly related to the size of the scallops. The sampling period accounted for 57.2 to 84.8 % of the variation of Cu and for 33.9 to 72.3 % of the variation of Zn (Table 6). In the case of Cd, size was the predominant factor for digestive gland, gills, and whole soft parts (43.6, 44.9 and 42.8 %, respectively). In contrast, most of the variability of Cd concentrations in kidney, gonads, adductor muscles and remaining tissues was related to the sampling period (i.e. from 34.4 to 49.6 % of the variation).

Metal distribution

Among the investigated tissues, the digestive gland clearly appeared to be the main storage compartment of Cd and Cu (Table 7), while this organ only accounted for 10 to 20% of the soft part weight (Table 1). Zn occurred in larger proportion in the kidneys, which contained up to 85% of the total body burden of the metal despite its very low contribution in the soft part weight (less than 5 %). Body burdens of all elements in gills and gonads were generally low (< 5%) because these organs also represented a relatively small fraction of the soft part weight (Table 1). Among tissues, the adductor muscle accounting to up to 50% of the soft part weight, always displayed the lowest metal concentrations and only contained a small fraction of the total metal body burdens, i.e.

3.3, 5.9 and 4.4 %, of the total Cd, Cu and Zn, respectively.

Seasonal variations of metal concentrations

The mean concentrations of Cd, Cu and Zn in the body compartment varied significantly according to the considered sampling period except for Cu in the digestive gland (p = 0.124) (Tables 6 &8). Multiple comparison tests on the mean concentrations for each sampling period (i.e. all compartments considered together) indicated that the sampling period showing the highest or lowest concentrations differed from one metal to another.

Moreover, when considering the different compartments separately, the metal concentrations of several sampling periods were frequently not significantly different from each other (Table 8). However, the lowest Cd concentrations were always measured in September in all the compartments except in the gonads, for which the lowest concentrations corresponds to the sexual maturity period (i.e. June). For essential elements, scallops sampled in September had the highest Cu but the lowest Zn concentrations in most tissues. Highest Zn concentrations in most of the tissues were recorded in November.

Size variations of metal concentrations

Metals were measured in scallops belonging to the whole size range of the population.

The size of the scallops significantly affected the concentrations of Cd, Cu and Zn in most of the scallop body compartments but size was only the main factor explaining most of the variability for a few tissues, i.e. the digestive gland for Cd and Cu, and the gills for Cd (Table 6). However, the contribution of this factor to the variability of the concentrations was not significant (P>0.05) for Cd and Cu in the gonads and the muscle, and for Zn in the muscle and the remaining tissues (Table 6, Figures 1,2 & 3). When significant, the correlations between Cd concentration and size were always positive. The only positive relationship for essential elements was found for Cu in digestive gland and Zn in kidneys. In contrast, Cu in kidneys and gills and Zn in digestive gland, gills and gonads decreased significantly with size (Figures 2 &3).

DISCUSSION

Metal concentrations in the variegated scallop Chlamys varia appeared to be significantly affected by the three factors considered (body compartment, sampling period and size) in this study. Less than 3% of the variability in Cd, Cu and Zn concentrations is not explained by these three factors. The main factor was clearly the body compartment, which accounted for 79 to 94% of the variability (Table 5). Sampling period and size have a far lower weight in the variability but significantly influence the metal concentrations according to the tissue. Therefore, the influence of these two factors on the different compartments has been considered specifically (Table 6).

Metals were selectively distributed between the body compartments. This observation is consistent with other studies on different scallop species (e.g. [START_REF] Bryan | The occurrence and seasonal variation of trace metals in the scallops Pecten maximus (L.) and Chlamys opercularis (L.)[END_REF][START_REF] Mauri | Heavy metals in the Antarctic scallop Adamussium colbecki[END_REF][START_REF] Viarengo | Cu, Zn and Cd content in different tissues of the Antarctic scallop Adamussium colbecki: role of the metallothionein in heavy metal homeostasis and detoxification[END_REF], Bustamante & Miramand 2005), which indicated that the digestive gland and kidneys efficiently concentrated heavy metals. Such storage capacity is to be related with their role in the metabolism and to the detoxification processes occurring in both organs. The food is digested and absorbed in the digestive gland, where Cd, Cu and Zn are generally reported to be bound to cytosolic proteins such as metallothionein-like proteins [START_REF] Mauri | Heavy metals in the Antarctic scallop Adamussium colbecki[END_REF][START_REF] Viarengo | Cu, Zn and Cd content in different tissues of the Antarctic scallop Adamussium colbecki: role of the metallothionein in heavy metal homeostasis and detoxification[END_REF][START_REF] Bustamante | Etude des processus de bioaccumulation et de détoxication d'éléments traces (métaux lourds et terres rares) chez les mollusques céphalopodes et bivalves pectinidés. Implication de leur biodisponibilité pour le transfert vers les prédateurs[END_REF].

Once in the organism, heavy metals might be translocated to other organs, particularly to the kidneys because of their excretory function (e.g. [START_REF] Rainbow | Heavy metals levels in marine invertebrates[END_REF][START_REF] Rainbow | The significance of trace metal concentrations in marine invertebrates[END_REF]. In this organ, a high proportion of heavy metals are trapped in granules mainly composed of calcium and phosphate (Carmichael et al. 1979, George et al. 1980[START_REF] Stone | Cadmium binding components of scallop (Pecten maximus) digestive gland. Partial purification and characterisation[END_REF], Fowler & Gould 1988). However, to our best knowledge, the residence time and transfer processes of metals in and between the different tissues are still to be determined.

Despite clear excretion of residual bodies containing granules or residues of the digestion (George et al. 1980), trace element bioaccumulation in the digestive gland and kidneys lead to very high concentrations of heavy metals. This is particularly obvious for Zn in the kidney where the metal varied from 20 to more than 40 mg.g -1 dwt whatever the size or the sampling period (Table 4). Consequently to very high metal concentrations, the highest Cd, Cu and Zn loads are contained in the digestive gland and kidneys, which are therefore considered as the main accumulator body compartments. Very high concentrations and loads in the digestive gland and relatively low metal concentrations in the gills, suggest that food would be a major pathway of metal uptake in this species.

However, further investigation using metal radiotracers should be carried out to help in confirming this hypothesis and to characterise the uptake biokinetics of metals following both food and seawater pathways.

Most of the compartments displayed seasonal variations, which were however not synchronous among the different tissues. If we assumed that transfer between body compartments occurs relatively rapidly as reported for other bivalve species using radioisotopes (e.g. Fisher et al. 1996), it is likely that the asynchronous variations of metal concentrations are due to physiological factors rather than to environmental ones.

Among physiological factors, sexual maturity related to gametogenesis and gamete emission, is one process which leads to the loss of a fraction of the metal load. The lowest concentrations of all metals were measured in the gonads in the scallops sampled in June, which corresponds to the beginning of the spawning period of Chlamys varia in the Bay of Biscay [START_REF] Letaconnoux | Contribution à l'étude du pétoncle (Chlamys varia L.)[END_REF]. In this species, spawning could therefore constitute a way of metal elimination from the organism. Similar observations have been reported for different invertebrates or fish species such as the echinoid Paracentrotus lividus, the oyster Crassostrea iridescens or the dab Limanda limanda for metals and polychlorinated biphenyls [START_REF] Loizeau | Distribution of polychlorinated biphenyls in the dab (Limanda limanda) from the Baie de Seine (Eastern Channel)[END_REF][START_REF] Páez-Osuna | Trace metal concentrations in relation to season and gonadal maturation in the oyster Crassostrea iridescens[END_REF][START_REF] Warnau | Distribution of heavy metals in the echinoid Paracentrotus lividus from the Mediterranean Posidonia oceanica ecosystem: seasonal and geographical variations[END_REF]. Seasonal variations of metal concentrations might be explained by several factors such as the fluctuations in food supply [START_REF] Bryan | The occurrence and seasonal variation of trace metals in the scallops Pecten maximus (L.) and Chlamys opercularis (L.)[END_REF], the changes in run-off of particulate matter to the sea due to high precipitation (Fowler & Oregioni 1976), the variations related to the reproductive cycle [START_REF] Latouche | Seasonal variation in soft tissue weights and trace metal burdens in the bay mussel Mytilus edulis[END_REF] or the seasonal changes during the year (Páez-Osuna & Marmolejo-Rivas 1990). In other bivalves such as oysters or mussels, seasonal variations of metal concentrations have been related to changes in their tissue weight [START_REF] Boyden | Seasonal variation and inherent variability of trace elements in oysters and their implications for indicator studies[END_REF][START_REF] Latouche | Seasonal variation in soft tissue weights and trace metal burdens in the bay mussel Mytilus edulis[END_REF], Khristoforova & Chernova 1990). The variegated scallop described a winter-spring maximum and summer pre-spawning minimum of Cd and Zn concentrations, as previously reported for oysters and mussels. However, depletion of Cu at the end of the spawning period could indicate a drastic loss of this metal through gamete emission.

Because they poorly regulate metal concentrations in their tissues, bivalves might reflect the contamination of their environment. However, metal concentrations within these molluscs could vary with age (i.e. length and/or weight) (e.g. [START_REF] Boyden | Effect of size upon metal content of shellfish[END_REF][START_REF] Phillips | Quantitative Aquatic Biological Indicators[END_REF]. When considering the scallop body compartment, metal concentrations varied significantly with size. These size differences were dependent of the considered body compartment for all metals. However, trends of size variations might be related to the physiology of the metal in the scallop.

First, Cd is primarily accumulated through the digestive pathway as it is reflected by the large contribution of size in the variability of Cd concentrations in the digestive gland.

However, the contribution of the dissolved pathway could not be considered as negligible as the accumulation of Cd is positively correlated with the size for the gills (Figure 1).

Once incorporated, Cd accumulates linearly with size in most tissues except in gonads and muscles. This suggests a slow turn-over of the metal in most of the scallop tissues, which has however not been determined. Further studies should therefore aim to determine the residence time of this element in the variegated scallop.

Second, Cd concentrations in gonads and adductor muscle were not affected by size suggesting a metabolic control of this metal in those tissues. As Cd has no known physiological function, such constant levels might result from the detoxification processes occurring in the scallop tissues rather from a real homeostatic process.

However, the seasonal variations of Cd concentrations in gonads and adductor muscle show that the capacity of such detoxification processes would dependent of exogenous factors influencing the uptake of Cd and its subsequent redistribution (Bustamante & Miramand 2005). Similarly to Cd, Cu concentrations in gonads and adductor muscle were not affected by size. This also suggests a metabolic control of the metal in those tissues aiming to maintain the Cu concentrations in both tissues relatively constant. Thus, scallops appear to use both exogenous and endogenous Cu, the later use resulting in a decrease of Cu concentrations in kidneys, gills and remaining tissues due to a remobilization of this metal. Cu accumulation with size in the digestive gland might reflect a low bioavailability of the dissolved Cu, which is therefore mainly incorporated through the food pathway. Moreover, such an accumulation is related to efficient detoxification processes, resulting to the immobilisation of a fraction of the metal. As for Cu, the concentrations of Zn in adductor muscle and in the remaining tissues were not affected by size, also suggesting a metabolic control of Zn in those tissues. Consequently to these homeostatic needs, the concentrations of Zn tended to decrease in most of the tissues except in the kidney. In this organ, long term storage may be due to the precipitation of Zn on phosphate granules, which are slowly excreted (George et al.

1980).

Despite the importance of the sampling season and the size in metal concentration variability, a high degree of unexplained variability occurs in some organs, especially in the adductor muscle (Table 6). This phenomenon has been previously reported in bivalves (e.g. [START_REF] Boyden | Seasonal variation and inherent variability of trace elements in oysters and their implications for indicator studies[END_REF][START_REF] Lobel | Metal accumulation in four molluscs[END_REF]. Later on, [START_REF] Lobel | A major factor contributing to the high degree of unexplained variability of some elements concentrations in biological tissue: 27 elements in 5 organs of the mussel Mytilus as a model[END_REF] suggested that elements which are stored largely in an insoluble form (e.g. granules)

show a much higher degree of variability than those stored in soluble form (e.g. bound to cytosolic proteins). It would therefore be of particular interest to investigate in which extent the partitioning of metals between insoluble and soluble fractions could influence the residual variability in the tissues of the variegated scallop.

CONCLUSION

There is a great interest to use Pectinidae in biomonitoring studies because of their high bioaccumulation potential compared to other bivalves. However, the present work underlines the need to consider both sampling period and size in separated body compartments when aiming to use the variegated scallop as a biomonitor species. Indeed, the variations in metal concentrations could be higher between different body compartments than between seasons. The same also appears true for the variations in concentrations between sizes for some elements. Therefore, particular attention should be taken to compare metal concentrations only between the same compartments from scallop collected at the same period of the year. Among these compartment, the digestive gland, and the kidneys are proposed for consideration in biomonitoring programs, these compartments showing the highest metal concentrations. Fisher, N.S., Teyssié, J-L., Fowler, S.W., Wang, W-X., 1996. Accumulation and retention of metals in mussels from food and water: a comparison under field and laboratory conditions. Environmental Science and Technology 30, 3232-3242. Fowler, B.A., Gould, E., 1988. Ultrastructural and biochemical studies of intracellular metal-binding patterns in kidney tubule cells of the scallop Placopecten magellanicus following prolonged exposure to cadmium or copper. Marine Biology 97, 207-216. Fowler, S.W., Oregioni, B., 1976. Trace

  ACKNOWLEDGEMENTSBustamante, P., Miramand, P., 2004. Interspecific and geographical variations of trace element concentrations in Pectinidae from European waters. Chemosphere, 57 : 1355-1362. Bustamante, P., Miramand, P., 2005. Subcellular and body distributions of 17 trace elements in the variegated scallop Chlamys varia from the French coast of the Bay of Biscay. The Science of the Total Environment, 337, 59-73. Carmichael, N.G., Squibb, K.S., Fowler, B.A., 1979. Metals in the molluscan kidney : a comparison of two closely relative bivalves species (Argopecten), using X-Ray microanalysis and atomic absorption spectrophotometry. Journal of the Fishery Research Board Canada 36, 1149-1155.

Fig. 2 .

 2 Fig. 2. Variations of Cu concentrations with size in the tissues of the variegated scallop Chlamys varia from the French Atlantic coast. ○: November 1995; •: February 1996; ∆: June 1996; ×: September 1996. Significant correlation coefficients are shown in bold.

Table 1 .

 1 metals in mussels from the NW Mediterranean. Sea of Japan. Soviet Journal of Marine Biology 15, 340-346. Characteristics of the samples analysed : number of pools analysed (n), number of scallop per pool (N), scallop height (mm), fraction of each tissue relatively to the soft part weight (% of the soft part weight); values in brackets are the water fraction (%) of the wet body compartments (all data are means ± SD).

	Marine Pollution Bulletin 7, 26-29.
	George, S.G., Pirie, B.J.S., Coombs, T.L., 1980. Isolation and elemental analysis of
	metal-rich granules from the kidney of the scallop, Pecten maximus (L.). Journal of
	Experimental Marine Biology and Ecology 42, 43-156.
	Khristoforova, H.K., Chernova, E.H., 1990. Trace element composition of giant oyster
	from Poyset Bay,

Table 2 .

 2 Cd concentrations (mean ± SD, µg.g -1 dwt) in the tissues of the variegated scallop Chlamys varia of different size classes.

	Tissue	Period	< 11 mm	11 -17 mm 17 -24 mm	24 -30 mm	30 -36 mm	36 -41 mm	42 -47 mm	47 -53 mm
	Digestive gland	November 95	-	-	18.8	19.6 ± 1.9	22.3 ± 0.6	32.7 ± 5.0	39.6 ± 2.5	50.7 ± 5.4
		February 96	34.0 ± 2.2	25.6	23.7 ± 1.1	30.5 ± 3.9	28.8 ± 2.9	35.4 ± 4.8	37.9 ± 4.7	43.3 ± 0.5
		June 96	-	16.4 ± 1.0	-	24.6 ± 2.2	23.7 ± 2.6	31.1 ± 5.8	47.5 ± 11.5	41.7 ± 8.1
		September 96	13.2	-	10.8	9.3 ± 0.3	15.4 ± 1.9	19.4 ± 1.1	23.5 ± 3.6	26.9 ± 2.7
	Kidneys	November 95	-	-	-	35.3 ± 5.8	46.1 ± 1.7	47.0 ± 14.2	52.4 ± 10.0	60.9 ± 3.6
		February 96	-	-	-	31.8 ± 1.2	31.7 ± 4.1	34.3 ± 0.7	34.6 ± 5.8	47.5 ± 15.1
		June 96	-	-	-	30.3 ± 3.0	32.7 ± 1.4	29.8 ± 11.9	45.3 ± 5.8	48.8 ± 3.9
		September 96	-	-	-	32.5 ± 1.6	26.4 ± 5.3	30.9 ± 1.8	35.6 ± 4.7	34.6 ± 5.4
	Gill	November 95	-	-	-	3.29 ± 2.65	2.91 ± 0.06	3.65 ± 1.37	4.43 ± 0.84	4.91 ± 0.61
		February 96	-	-	-	3.25 ± 0.29	3.14 ± 0.26	3.60 ± 0.37	3.98 ± 0.26	4.33 ± 0.57
		June 96	-	-	-	3.25 ± 0.24	3.03 ± 0.35	3.53 ± 0.83	4.36 ± 0.36	4.46 ± 0.14
		September 96	-	-	-	1.63 ± 0.17	2.72 ± 0.07	2.96 ± 0.27	3.49 ± 0.44	3.67 ± 0.41
	Gonad	November 95	-	-	-	2.44 ± 0.30	1.89 ± 0.08	2.29 ± 0.22	2.38 ± 0.20	2.30 ± 0.18
		February 96	-	-	-	3.62 ± 0.21	2.42 ± 0.49	1.67 ± 0.35	2.11 ± 0.35	2.10 ± 0.21
		June 96	-	-	-	1.12 ± 0.21	1.16 ± 0.37	1.46 ± 0.35	1.54 ± 0.28	0.64 ± 0.21
		September 96	-	-	-	1.31 ± 0.43	4.67 ± 1.55	4.17 ± 1.36	3.94 ± 0.18	3.43 ± 0.75

Table 5 .

 5 Variability (%) in metal concentrations in the tissues of the variegated scallop Chlamys varia explained by the different factors

	considered (period of sampling and size of the scallops) and their
	interactions. Highest proportions of variation are shown in bold. ns:
	not significant (P ANOVA <0.05)
	Factors	Cd	Cu	Zn
	Compartment	86.3	79.3	93.7
	Period	1.2	3.6	0.5
	Size	1.7	0.2	0.1
	Compartment × Period	3.6	9.4	2.5
	Compartment × Size	3.1	2.2	0.6
	Period × Size	0.4	0.8	Ns
	Compartment × Period × Size	1.0	3.1	0.7
	Residual	2.6	1.4	1.8

Table 6 .

 6 Variability (%) in Cd, Cu and Zn concentrations in the tissues of the variegated scallop Chlamys varia explained by the different factors considered (period of sampling and size of the scallops) and their interactions. Highest proportions of variation are shown in bold. ns: not significant

	(P ANOVA >0.05)
	Factors	Digestive	Kidneys	Gills	Gonads	Adductor	Remaining	Whole soft
		gland				muscle	tissues	parts
	Cd Period	36.7	34.4	22.3	49.6	44.0	46.3	33.1
	Size	43.6	26.0	44.9	ns	ns	12.2	42.8
	Period × Size	8.4	ns	ns	35.2	ns	17.8	ns
	Residual	11.3	27.4	25.1	13.4	42.3	23.7	15.3
	Cu Period	ns	69.4	64.3	80.7	45.6	84.8	57.2
	Size	67.2	4.4	7.6	ns	ns	2.7	22.0
	Period × Size	14.4	19.9	21.0	ns	31.0	7.0	11.2
	Residual	7.8	6.3	7.0	10.3	20.9	5.5	9.7
	Zn Period	72.3	48.3	33.9	57.6	66.7	62.3	69.0
	Size	12.4	10.4	29.0	19.9	ns	ns	8.0
	Period × Size	10.5	ns	ns	13.8	ns	ns	11.3
	Residual	4.8	6.3	26.2	8.7	27.3	24.0	11.7

Table 7 .

 7 Proportions (%, mean ± SD) of metal loads in the body compartments of the variegated scallop Chlamys varia (all locations and periods together).

	Metal	Digestive	Kidney	Gonads	Gills	Adductor	Remaining
		gland				muscle	tissues
	Cd	65.3 ± 6.7	18.5 ± 5.0	1.4 ± 0.3	4.2 ± 1.7	3.3 ± 1.0	7.3 ± 2.3
	Cu	47.6 ± 10.9 28.8 ± 9.5	4.7 ± 3.6	3.8 ± 2.1	5.9 ± 3.1	9.1 ± 3.6
	Zn	3.0 ± 1.9	84.7 ± 9.4	1.0 ± 0.7	2.5 ± 2.0	4.4 ± 2.6	4.5 ± 3.8

Table 8 .

 8 Seasonal variations of metal concentrations in the tissues of the variegated scallop Chlamys varia: results of the multiple comparison tests of Tukey performed after ANOVA. Mean metal concentrations are ranking from the left to the right by increasing order. Concentrations in underlined seasons are not significantly different (α Tukey = 0.05). N: November 95, F: February 96, J : June 96, S : September 96.

	Tissues		Cd				Cu					Zn
	Digestive gland	S	N	J	F	F	J	N	S	S	J	N	F
	Kidneys	S	F	J	N	F	J	N	S	S	F	N	J
	Gills	S	F	J	N	F	J	N	S	S	F	J	N
	Gonads	J	N	F	S	F	J	N	S	J	S	F	N
	Adductor muscle	S	F	J	N	F	J	N	S	S	F	J	N
	Remaining tissues	S	F	N	J	F	J	N	S	S	J	F	N
	Whole soft parts	S	F	N	J	F	J	N	S	S	N	F	J

Shell height (cm)Shell height (cm)

We are grateful to Rhodia, the "Conseil Régional Poitou-Charentes" and the "Conseil Général de la Charente-Maritime" for supporting financially this study.


Thomas Merckx

Caroline Souffreau

Aurélien Kaiser

Lisa F Baardsen

Thierry Backeljau

Kristien I Brans

Marie Cours

Maxime Dahirel

Nicolas Debortoli

Katrien De Wolf

Jessie M T Engelen

Diego Fontaneto

Andros T Gianuca

Lynn Govaert

Frederik Hendrickx

Janet Higuti

Luc Lens

Koen Martens

Hans Matheve

Erik Matthysen

Elena Piano

Rose Sablon

Isa Schön

Karine Van Doninck

Luc De Meester

Hans Van Dyck

Hans Van Dyck

Body-size shifts in aquatic and terrestrial urban communities

Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B

Body size is intrinsically linked to metabolic rate and life-history traits, and is a crucial determinant of food webs and community dynamics 1,2 . The increased temperatures associated with the urban-heat-island effect result in increased metabolic costs and are expected to drive shifts to smaller body sizes 3 . Urban environments are, however, also characterized by substantial habitat fragmentation 4 , which favours mobile species. Here, using a replicated, spatially nested sampling design across ten animal taxonomic groups, we show that urban communities generally consist of smaller species. In addition, although we show urban warming for three habitat types and associated reduced community-weighted mean body sizes for four taxa, three taxa display a shift to larger species along the urbanization gradients. Our results show that the general trend towards smaller-sized species is overruled by filtering for larger species when there is positive covariation between size and dispersal, a process that can mitigate the low connectivity of ecological resources in urban settings 5 . We thus demonstrate that the urban-heat-island effect and urban habitat fragmentation are associated with contrasting community-level shifts in body size that critically depend on the association between body size and dispersal. Because body size determines the structure and dynamics of ecological networks 1 , such shifts may affect urban ecosystem function.

Body size is a fundamental species trait relating to space use and key life-history features such as longevity and fecundity 6 . It also drives interspecific relationships, thus affecting ecological network dynamics 1 . Size-biased species loss has profound effects on ecosystem function 7,8 . Ectotherms rely on ambient conditions to achieve operational body temperatures 9 . Because higher ambient temperature increases metabolic rates and the associated costs for a given body size 2 , global climatic warming is expected to drive shifts to communities consisting of smaller species 3 .

Our planet is urbanizing quickly 10 , which is a primary example of human-induced rapid environmental change. Cities are urban heat islands characterized by increased temperatures that are decades ahead of global averages 11 . Not only are cities warmer than surrounding areas, but they also experience extensive fragmentation of (semi-)natural habitats, and both of these effects increase with percentage built-up cover (BUC; a proxy for urbanization) 12,13 . This provides an opportunity to study the opposing effects of sizedependent thermal tolerance and dispersal capacity, as larger body size favours dispersal in some, but not all, taxa.

Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B

Here we test the hypothesis that urbanization causes shifts in community-level body size, and that these shifts are dictated by the community-specific association between body size and dispersal. We generally expect the urban-heat-island effect to drive shifts to species with smaller body sizes in communities of ectothermic species, in line with Atkinson's temperature-size rule 14 . For taxa characterized by a positive association between body size and dispersal, however, we also expect a filtering in favour of larger-bodied species associated with habitat fragmentation 5,15 . Filtering for increased mobility has been demonstrated for urban ground beetle and plant communities 16,17 . Hence, for taxa characterized by a positive body-size-dispersal link, we predict that the general communitylevel pattern of smaller species with increasing urbanization may be neutralized or even reversed.

To test our hypothesis, we engaged in an analysis of community-level shifts in body size across a broad range of both terrestrial and aquatic taxa along the same systematically sampled urbanization gradients. We studied the direction of change of community-level body size in ten taxa using a replicated, highly standardized and nested sampling design that covers urbanization gradients at seven spatial scales (50-3,200 m radii; Fig. 1). We sampled each taxon at up to 81 sites, sampling 95,001 individuals from 702 species, with species-specific body size varying by a factor of 400 (0.2-80 mm; Extended Data Table 1). Three of the ten groups are characterized by a positive association between body size and dispersal capacity (see Extended Data Table 1).

We show that the local temperature of pond, grassland and woodland habitats significantly increases with urbanization (linear mixed regression models, P < 0.002;

Extended Data Table 2). The intensity of these urban-heat-island effects is consistently larger during night and summer, in accordance with slower night-time city cooling and higher irradiation levels in summer 18 (Fig. 2, Extended Data Fig. 1, Extended Data Table 2). We also show that increased urbanization is linked to significant declines in habitat amount and the patch size of terrestrial habitats, and significant increases in distances among patches for both terrestrial and aquatic habitats (Pearson's r correlations, P ≤ 0.020; Extended Data Fig. 2).

Confirming our metabolism-based prediction that interspecific mean body size decreases with increasing temperature, urban communities for four out of the seven taxa (ground spiders, ground beetles, weevils and cladocerans) that did not have a positive sizedispersal link display reduced community-weighted mean body size (CWMBS). For ostracods, bdelloid rotifers and web spiders, no relationship with urbanization is found. By Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B contrast, all three taxa with positive size-dispersal links display increased CWMBS in response to urbanization (Figs. 3, 4, Extended Data Table 3). The positive shifts in size observed for these taxa are in line with our prediction that increased urbanization-mediated habitat fragmentation selects for larger species in taxa with positive size-dispersal links.

The Benjamini-Hochberg procedure 19 , which controls for false positives, confirms that all seven responses are significant at the study-wide level. Comparing the percentage changes in body size over a percentage BUC gradient of 0-25% shows a marked difference between taxa with a positive size-dispersal link (13.6% ± 8.3% (mean ± s.e.m.) body size increase) versus the other taxa (15.6% ± 5.3% body size decrease) (weighted two-sided analysis of variance (ANOVA): F 1,8 = 12.38; P = 0.0079). These community-level shifts in body size occur independently of shifts in species abundance and diversity along the urbanization gradients. For example, reduced diversity is apparent for taxa that display positive and negative size shifts, as well as for web spiders that lack a size shift. By contrast, cladocerans show size reduction without diversity change (Extended Data Table 4). For butterflies, macro-moths and orthopterans (that is, taxa with a positive size-dispersal link), the increase in the CWMBS ranges from 7% to 21% depending on the taxon, whereas size reductions of ground beetles, weevils and ground spiders (that is, terrestrial taxa with nonpositive size-dispersal links) range from -18% to -21% over an urbanization gradient of 0-25% BUC (Fig. 3). The cladocerans display the largest size reduction (-44%), in accordance with the temperature-size response generally being stronger in aquatic species than in terrestrial species as a result of the greater oxygen limitation in water 20 . However, the size reduction for the ostracods is much smaller (-13%) and non-significant (linear mixed regression model, P = 0.10), and for the rotifers no size shift is found. The absence of a size shift for the microscopic rotifers might indicate that their small size allows for sufficient oxygen exchange between warm, low-oxygen environments and body tissues, so that no community shift to smaller body sizes is induced by increased temperature. The absence of a size shift for web spiders may be explained by behavioural flexibility in their extended phenotype, as modified web designs help web-spider communities to adapt to urbanizationinduced lower average body size of aerial dipteran prey 21 .

Our multi-scale approach allows the pinpointing of the spatial scales at which urbanization best explains the observed effects. During winter, the urban-heat-island effect fades with increasing spatial scale during the day but not at night, whereas during summer both diurnal and nocturnal urban-heat-island effects are more pronounced at small scales (Fig. 2). The spatial scale at which most of the variation in CWMBS is explained varied considerably among taxa, with effects for smallersized taxa prevailing at small spatial scales (Figs. 3, 4, Extended Data Table 3).

Urbanization induces biodiversity loss and biotic homogenization 10 (see also Extended Data Table 4). Here, we demonstrate that urbanization also leads to community-wide shifts in body size for the majority of studied species groups. The size reductions within aquatic and terrestrial taxa follow metabolic rules in line with the urban-heat-island effect, especially as our data on various pollutants suggest no correlation with percentage BUC (data not shown).

By contrast, the increased fragmentation that is a result of urbanization appears to cause size increases for taxa with positive size-dispersal links. Hence, our multi-taxa study provides evidence of bi-directional shifts in community body size. In addition to the interspecific patterns reported here, shifts in body size can also occur at the intraspecific level, through phenotypic plasticity and genotypic change [22][23][24] . Our results should enable mechanistic studies that elucidate the cause of the variation in the observed shifts in body size along urban gradients and quantify their functional effects in urban ecosystems. A better insight into the mechanisms behind shifts in body size will allow prediction of the intertwined effects of climate change and urbanization on the body-size distribution of communities.

The size-biased species loss reported here is expected to strongly affect ecosystem function 7,8 . If taxa in urban areas are represented by smaller or larger species, ecosystem structure and function will be affected in several ways. Metabolic theory and a recent artificial-selection experiment predict that shifted size distributions affect whole-ecosystem properties such as primary productivity, carbon cycling and decomposition 25,26 . Shifts in body size also translate into altered life histories, demographic rates and interspecific relationships 1,2 . For example, consumer-resource dynamics have recently been modelled for warming-related intraspecific size shifts mediated by phenotypic plasticity 27 . A clear-cut effect of shifts in body size on ecosystem function can be predicted for cladoceran zooplankton. Smaller-sized cladoceran communities are typified by reduced densities of large Daphnia species (highly efficient filter feeders that consume phytoplankton), and are thus less able to maintain top-down control on algal blooms than larger-sized communities 28 . Also, the observed shifts in macro-moth body-size distributions may be functionally linked to flowering plant diversity through pollination 29,30 .

The shifts in body size that we observe across a range of animal taxa will be directly relevant to future efforts to understand, predict and mediate population resilience, trophic Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B

interactions, and ecosystem function in urban ecosystems 31,32 . Such insights will be essential to design the biodiverse towns and cities of the future. For example, urban planners could mitigate the micro-climatic effects and habitat fragmentation that result from urbanization with measures implemented at multiple spatial scales. Such interventions could involve the creation and/or modification of urban ponds and urban green infrastructure to increase the amount and quality of habitats 33 . Doing so would reduce the urban-heat-island effect and favour dispersal, and hence gene flow, in urban animal populations. Our results indicate that such impacts would maintain variation in the body-size distributions of urban communities and potentially mitigate the effect that shifts in body size may have on ecosystem function.

METHODS

Sampling design

Sampling was performed according to a nested design in which a local urbanization gradient (three classes: non-urban, semi-urban and urban) was repeatedly sampled within landscapes distributed along a landscape-scale urbanization gradient (three classes: non-urban, semiurban and urban). For each of ten taxa a total of up to 81 local-scale subplots (200 × 200 m) were sampled within 27 landscape-scale plots (3 × 3 km) situated in an 8,140 km 2 study area in northern Belgium (Fig. 1, Extended Data Table 1). The average human population density of this highly urbanized area amounts to 693 individuals per km 2 , with cities and urban sprawl embedded within an agricultural and semi-natural matrix 34 . As a proxy for urbanization we used percentage BUC, which was assessed in a geographic information system (GIS) using an object-oriented reference map of Flanders with the precise contours of all buildings, excluding roads and parking infrastructures, as a vectorial layer. Given that only buildings are considered, 15% BUC can be considered highly urbanized. Within each of the nine urban (BUC > 15%), nine semi-urban (5% < BUC < 10%) and nine non-urban (BUC < 3%) plots, one urban, one semi-urban and one non-urban subplot were chosen using identical BUC cutoff values, for a total of 81 subplots. Within each subplot, and for each of the ten taxa, a single grassland, woodland or pond habitat patch was targeted for sampling during the most appropriate season for each taxon (Extended Data Table 1). As each taxon was sampled in only one of three habitat types (that is, grassland, woodland or ponds), it was often impossible to sample all taxa within the same 200 × 200 m subplots. As such, independent subplots containing the corresponding habitats were sometimes selected among taxa, but these subplots were always of the same urbanization level and located within the same 3 × 3 km plot.

The classification of urban, semi-urban and non-urban (sub)plots on the basis of BUC cut-off values was used to establish the nested sampling design, which allowed samples to display a wide range of urbanization values at both local (subplot) and landscape (plot) scales.

To increase precision in the data analysis, however, we moved away from having BUC as a class variable with three levels, and instead quantified BUC as a continuous variable, at seven spatial scales around the sampling site (50, 100, 200, 400, 800, 1,600 and 3,200 m radii).

Owing to our nested design, BUC values at small scales were not correlated with values at large scales, hence allowing the pinpointing of the scales at which the effects of urbanization are most pronounced.

Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B

Using this highly replicated, nested sampling design, our sampling effort involved counting and assigning 95,001 individuals to 702 species in ten taxa: (i) aquatic: cladocerans and ostracods sampled in pond habitats; (ii) limno-terrestrial: aquatic bdelloid rotifers sampled within the water layers of terrestrial Xanthoria lichens; and (iii) terrestrial:

butterflies, orthopterans (that is, grasshoppers and bush crickets), macro-moths, ground beetles, weevils, web spiders and ground spiders sampled in grassland and woodland habitats (Extended Data Table 1).

Urban-heat-island effect

The urban-heat-island effect was quantified using hourly temperature readings that were collected automatically across 104 sampling sites for the three habitat types in which the ten taxa were sampled: ponds, grasslands and woodlands. Aquatic probes (HOBO, TidbiT v2

Habitat fragmentation

Correlations between urbanization (BUC) and three habitat fragmentation variables (that is, habitat coverage, mean size of habitat patches, and mean nearest-neighbour distance among habitat patches) were quantified using Pearson's r coefficients (Extended Data Fig. 2). This was done at a 3 × 3 km plot scale, on the basis of detailed land-use data from all 27 sampling plots (Fig. 1), and separately for terrestrial (that is, all types of (semi-)natural habitat) and aquatic habitat (that is, all pond types) 35,36 . Eutrophied, mono-specific intensive grasslands as well as orchards, plantations and conifer woodlands were not retained for analyses.

Statistical analyses

Temperature averages were analysed in relation to site-specific urbanization (BUC) values and habitat type (grassland, woodland and pond) using linear mixed regression models (R Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B

Extended Data Table 4 | Model output of abundance and diversity measures in relation to urbanization

Output of linear mixed regression models for ten taxa (n = 76, 12, 75, 80, 62, 60, 81, 81, 68 and 80 biologically independent communities, top to bottom), testing the relationship between abundance (N) and two diversity measures (species richness (S) and Shannon index (H)) and percentage BUC at the spatial scale (radius in metres) providing the best-fitting models. t indicates that log(x + 1) transformations improved residual fits. Modelled (back-transformed) percentage change across a 0-25% BUC gradient is also given. Credits for animal outlines are listed in the legend of Fig. 3.

  Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B 2, Extended Data Fig. 1, Extended Data Table

Fig. 1 |Fig. 2 |Fig. 3 |Fig. 4 |

 1234 Fig.1| Map of the study area. The configuration of 27 landscape-scale sampling plots (nine urban, magenta; nine semi-urban, yellow; nine non-urban, green) on an urbanization background (light and dark corresponds to non-urban and urban gradient, respectively). Solid lines refer to administrative province borders. Three plots are enlarged, showing the distribution of local subplot types within each plot. Subplots allowed sampling using a nested design that covers urbanization gradients at both the landscape and local scale. Different sets of subplots were selected among taxa, so that subplots always contained the corresponding habitats. Urbanization was quantified as the percentage BUC (assessed using a reference map with building contours; LRD, https://www.agiv.be/international/en/products/grb-en) for each sample site at seven spatial scales (50-3,200 m radii, depicted as black circles around one of the three sample sites of one non-urban plot). Photographs depict sites in a non-urban and urban subplot, for both aquatic (top row) and terrestrial (bottom row) systems. Corine Land Cover map for northern Belgium, copyright NGI (2017); map of western Europe, Ocean Basemap, copyright Esri.

  Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B Extended Data Fig. 1 | Micro-climatic urban-heat-island effect strengths. a, b, Slopes of the urban-heat-island effects (measured as the increase in temperature (°C) per 1% increase in percentage BUC) as a function of spatial scale (the at which urbanization was quantified) with 95% confidence intervals (CI). Separate measurements are shown for summer (red) and winter (blue) using merged readings for pond, grassland and woodland habitats (n = 104 sites). a, Diurnal measurements. b, Nocturnal measurements. Data points are offset from one another horizontally to improve clarity. Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B Extended Data Table 2 | Model output of average temperature in relation to urbanization and habitat type Output of linear mixed regression models that test the relationship between average local ambient temperatures and the interaction between percentage BUC (%BU) and habitat type (pond, grassland and woodland) (n = 104 sites). Only the output for the confidence set of models (∆AICc ≤ 2) is given, with scale referring to the associated radius scale (in metres) of percentage BUC. Model estimates (±s.e.m.) for percentage BUC regression coefficients are provided. Model output consistently shows clear temperature differences among habitats and a clear positive effect of urbanization on temperature, irrespective of habitat type (as shown by the non-significant interactions).

  Online content Any Methods, including any statements of data availability and Nature Research reporting summaries, along with any additional references and Source Data files, are available in the online version of the

		Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter
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Acknowledgements We thank M. De Cock, J. Dierick, P. Limbourg, E. van den Berg, M. Van Kerckvoorde and P. Vantieghem for help with sampling and identification of species. This is publication BRC419 of the Biodiversity Research Centre (UCL/ELI). This research is part of the SPEEDY-project, funded by the Interuniversity Attraction Poles program of the Belgian Science Policy Office BELSPO (IAP-grant P7/04).

Reviewer information

Nature thanks M. McDonnell and the other anonymous reviewer(s) for their contribution to the peer review of this work.

Data availability

The datasets generated and analysed during the current study are available from the corresponding author upon reasonable request.

<eref>34.

IBZ. Bevolkingscijfers per Provincie en per Gemeente op 1 Januari 2017 http://www.ibz.rrn.fgov.be/fileadmin/user_upload/fr/pop/statistiques/population-

L.G., F.H., J.H., L.L., K.M., E.M., E.P., R.S. and I.S. contributed to sampling and data collection. T.M. and A.K. performed the analyses. H.M. selected study plots, calculated fragmentation variables and designed the study area map. T.M. wrote the first draft of the manuscript with all authors contributing substantially to revisions.

Competing interests:

The authors declare no competing interests.

Additional information

Extended data is available for this paper at

Reprints and permissions information is available at www.nature.com/reprints.

Correspondence and requests for materials should be addressed to T.M.

Publisher's note: Springer Nature remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.

Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B package lme4). We ran separate models for both seasons (summer and winter) and for both day and night conditions (diurnal and nocturnal). Site ID and date (nested within year) were included as random factors. We used a multi-scale approach, running separate models with BUC values quantified at seven spatial scales (50-3,200 m radii). P values for the fixed effects were obtained using likelihood-ratio tests of nested models that were fitted with maximum-likelihood and parameter estimates from restricted maximum-likelihood models.

Residual plots were always visually inspected to evaluate the fit of models, and we compared maximum-likelihood-based AICc values (R package AICcmodavg) to select a confidence set of models whose AICc values did not differ substantially from the value of the best-fitting model, using ∆AICc ≤ 2 as a criterion 37 .

CWMBS was calculated for a given site as the average of the species-specific body sizes (mm) for all locally sampled species, weighted by species abundance. The raw data for calculating this metric are species-level count data for all taxa in all sites (based on taxonspecific sampling and identification protocols) and mean species-specific body-size values extracted from the literature or, in the case of web spiders and cladocerans, from our own measurements (Extended Data Table 1). An increase in CWMBS with increasing urbanization implies that the species assemblage of the site is increasingly composed of individuals belonging to larger species along the gradient from communities in more rural sites to communities in more urban sites. Our CWMBS index hence reflects the relative composition of large versus small species in local communities, and we use it here to quantify community response to urbanization. Although every sampling method introduces some bias in relative species abundances, the extent of the bias should be similar for non-urban and urban sampling sites. Therefore, using the relative species abundances that we obtained via sampling to calculate the CWMBS is appropriate to look into the relative effects of urbanization.

CWMBS was analysed for each taxon in relation to site-specific urbanization (BUC) values using linear mixed regression models with restricted maximum-likelihood estimation (R package lme4). Plot ID was used as a random variable to account for potential spatial autocorrelation of variables among sites belonging to the same landscape-scale plot. CWMBS values were log 10 -transformed for cladocerans and ostracods. For ostracods, we also transformed BUC values by taking the arcsine of square-rooted BUC values, which resulted in residual plots with a more homogeneous distribution. Analyses for the other taxa were run with untransformed data as residual plots proved to be homogeneous. The residual plots for orthopterans, ostracods and ground beetles each displayed one outlying data point, and the Publisher: NPG; Journal: Nature: Nature; Article Type: Biology letter ms: 2017-07-09177B

residual plot for weevils displayed two such points. Because these five data points are legitimate (that is, they are not due to measurement, data or sampling errors) we assessed their effect on the consistency of the regressions in the model output. Filtering these data points out of the regressions showed (i) that the best-fitting models remained linked to the identical spatial scales; (ii) that the positive slope for orthopterans remained positive and the negative slopes for the other taxa remained negative, and (iii) that the significance levels stayed equal for ground beetles and ostracods, got stronger for weevils, and decreased but remained significant for orthopterans. Because those five data points are legitimate and do not have a qualitative effect on the output, we opted to retain them in the analyses. We used a multi-scale approach, running separate models with BUC values quantified at seven spatial scales (50-3,200 m radii). For each taxon, we then selected the model (and hence the spatial scale) that fitted the data best using maximum-likelihood-based AICc values (R package AICcmodavg).

We also retained a confidence set of the models whose AICc values did not differ substantially from the value of the best model using ∆AICc ≤ 2 as a criterion 37 .

For each taxon, and at the spatial scale of the best-fitting model, we calculated the percentage change (with 95% confidence interval) in CWMBS over a 0-25% BUC gradient, on the basis of the modelled intercept and slope, or of back-transformed values for ostracods and cladocerans (Fig. 3). These values were then contrasted for taxa with a positive sizedispersal link against all other taxa using two-sided ANOVA, with the inverse of the taxonspecific error bars as weights to account for the differences among taxa in the variance of the estimated percentage change. This weighted ANOVA allows testing of the percentage change values for taxa with a positive size-dispersal link to determine whether they are significantly different from those of all other taxa.

All statistical analyses were performed using R v.3.2.3 38 .

Reporting summary

Further information on experimental design is available in the Nature Research Reporting Summary linked to this paper.


Paco Bustamante

Jean-Louis Teyssié

Scott W Fowler

J-L Teyssié

M Warnau

Assessment of the exposure pathway in the uptake and distribution of americium and cesium in cuttlefish (Sepia officinalis) at different stages of its life cycle

Keywords: Accumulation, Biokinetics, Cephalopods, Radionuclides, Retention

published or not. The documents may come

Assessment of the exposure pathway in the uptake and distribution of americium and cesium in cuttlefish (Sepia officinalis) at different stages of its life cycle

INTRODUCTION

Contamination of marine waters by radionuclides is a major concern in coastal areas which receive radioactive inputs from industries, accidents, and fallout from nuclear weapon testing.

Surveys estimating concentrations of such radionuclides in water or sediments are often complemented by biomonitoring programs, and marine mussels have been used as biological monitors for radionuclides and heavy metals [START_REF] Goldberg | The mussel watch -A first step in global marine monitoring[END_REF][START_REF] Goldberg | The Mussel Watch[END_REF][START_REF] Goldberg | Mussel Watch:1977-1978 results on trace metals and radionuclides[END_REF][START_REF] Goldberg | Beyond the Mussel Watch-New direction for monitoring marine pollution[END_REF]. However, previous studies on the trophic transfer of trace elements and radionuclides have shown that herbivores such as mussels do not assimilate to any extent transuranic elements ingested with their food (e.g., [START_REF] Fowler | Biological transfer and transport processes[END_REF][START_REF] Fisher | Interaction of marine plankton with transuranic elements. 3. Biokinetics of neptunium, plutonium, americium, and californium in phytoplankton[END_REF][START_REF] Warnau | Biokinetics of selected heavy metals and radionuclides in the common Mediterranean echinoid Paracentrotus lividus: sea water and food exposures[END_REF]. Nevertheless this aspect has been little studied in higher trophic levels which are also used in many contaminant surveys. Therefore, there is a need to determine the bioaccumulation potential of marine carnivorous species for such elements.

Previous investigations with cephalopods have shown that these carnivorous species do bioaccumulate radionuclides in their tissues that can at times reach high levels [START_REF] Suzuki | Accumulation of cesium-137 by useful Mollusca[END_REF], Guary et al. 1981, Yamada et al. 1999); however, little information is available on the pathways and rates of accumulation and retention of these radionuclides [START_REF] Suzuki | Accumulation of cesium-137 by useful Mollusca[END_REF][START_REF] Guary | Experimental studies on the biokinetics of plutonium and americium in the cephalopod Octopus vulgaris[END_REF]. Two long-lived radionuclides, 241 Am and 137 Cs (and 134 Cs), are present in fallout and also commonly found in nuclear wastes. The objective of our study was to examine the biokinetics of uptake and loss in cephalopods of these two contrasting radionuclides (particle-reactive americium and soluble cesium in sea water) in order to establish their bioaccumulation rates, tissue distribution and retention times as a function of (1) the uptake pathway, and (2) the life stage of the organism. The common cuttlefish Sepia officinalis was selected as a model species, and experimental exposures to these two radionuclides via seawater, food and sediment were studied in both juvenile and adult individuals.

MATERIAL AND METHODS

Experimental organisms

Eggs of the common cuttlefish (Sepia officinalis L.) were obtained from cultured adults and were maintained in an aquarium with flowing seawater until hatching. Newly hatched juveniles (n = 25; 0.387 ± 0.071 g wet wt) were selected and used in the experiments. Adult cuttlefish (n = 18; 138 ± 40 g wet wt) were either reared in the Monaco Oceanographic Museum from hatching to one year old individuals, or collected by net fishing off Monaco (n = 5; 253 ± 97 g wet wt). All organisms were subsequently maintained in filtered seawater in constantly aerated open circuit aquaria (salinity: 36 p.s.u.; temperature: 16.5 ± 0.5°C; 12/12 h dark/light cycle) until used in the radiotracer experiments.

Prior to experimentation, adults were anaesthetised in 2% ethanol in seawater for making biometric measurements, sex determination and for the insertion of a numbered plastic tag into the mantle fin to identify each animal during the experiments. In this way the same whole individual could be periodically radioanalyzed live in a small volume of sea water on the gamma well counter in order to reduce individual variability.

Radionuclides

241 Am [t ½ = 433 yr] and 134 Cs [t ½ = 2 yr] purchased from Amersham, UK, as nitrate and chloride salts, respectively, were used to trace americium and cesium biokinetics. The use of 134 Cs also served as an analogue for tracing the longer-lived 137 Cs with a t ½ of 33 yr. Stock solutions were prepared in their respective solutions (0.1 N) to obtain radioactivities which would allow using spikes of only a few microliters (typically 10 to 20 µL).

241 Am and 134 Cs uptake via sediments Sediments (2.5 kg dry wt) from the North Sea (Audresselles, Pas-de-Calais, France) were spiked for 4 d with 241 Am and 134 Cs tracer using the rolling jar method [START_REF] Murdoch | Spiking sediment with organochlorines for toxicity testing[END_REF].

Both 241 Am and 134 Cs are rapidly adsorbed onto sediments typically reaching near equilibrium after approximately 1 day [START_REF] Carroll | Radionuclide adsorption to sediments from nuclear waste dumping sites in the Kara Sea[END_REF]. Before initiating the experiment, radiolabelled sediments were held in flowing seawater overnight in order to leach weakly bound radiotracer. Sediments (50 g wet wt) were sampled at fixed intervals during the experiment to check for possible variations in radionuclide concentration. Juvenile cuttlefish (n = 9) were exposed for 29 d in a 20-L plastic aquarium containing 3 L of natural seawater running over a 4 cm layer of spiked sediment. The level of sea water was maintained low in order to minimise the movements required for feeding and to maximise the contact time with sediments. During the experiment, all juvenile cuttlefish were fed twice daily with brine shrimp Artemia salina and were periodically γ-counted live in a well counter to follow the radionuclide uptake kinetics over the 29 d. At the end of the uptake experiment, 3 individuals were dissected to determine the distribution of the radionuclides among digestive gland, cuttlebone and remaining tissues (including other organs).

241 Am and 134 Cs uptake via seawater and subsequent loss Newborn (n = 8) and adult (n = 5) cuttlefish were placed for 36 h and 8 h, respectively, in 70-L glass aquaria containing seawater spiked with 241 Am and 134 Cs (nominal activity: 6 kBq L -1 each). Cuttlefish were then radioanalyzed and transferred to another 70-L aquarium supplied with natural flowing seawater. Juvenile cuttlefish were fed A. salina twice daily and were periodically γ-counted to follow radionuclide loss kinetics over 29 d. At the end of the loss period, 4 juveniles were dissected to determine the radionuclide distribution among digestive gland, cuttlebone and remaining tissues.

During the loss phase, adults were fed daily with soft parts of the mussel Mytilus galloprovincialis. Three adults were dissected after 8 h and the remaining two were dissected after 6 d of depuration. For each individual, the branchial heart appendages, branchial hearts, gills, digestive tract (after removal of the gut contents), genital tract, ovary or testes, ink sack, digestive gland, kidneys, mantle skin, mantle muscle, head and cuttlebone were separated, weighed, and their radionuclide content measured.

Am and 134 Cs accumulation from food

To prepare radiolabelled food, mussels (M. galloprovincialis) and brine shrimp (A. salina) were exposed for 7 d in plastic aquaria containing 4 L of natural seawater spiked with 241 Am and 134 Cs (nominal activity: 6 kBq L -1 each). Radiolabelled seawater was renewed daily and the organisms were subsequently used as food for newborn (brine shrimp) and adult (mussels) cuttlefish.

For identification purposes, each individual juvenile cuttlefish (n = 8) was enclosed in a separate compartment allowing free circulation of seawater in a 70-L aquarium. After 1 h of ingesting radiolabelled brine shrimp, each individual was immediately γ-counted and replaced in the aquarium. From that time on cuttlefish were fed twice daily with non-contaminated A. salina and periodically γ-counted to determine radiotracer loss kinetics and assimilation efficiency. Throughout the depuration period (29 d), feces were removed 3 times per day to reduce possible indirect contamination by radiotracer recycling through leaching from the feces. At the end of the depuration period, 5 juveniles were dissected to determine the radiotracer distribution in their tissues.

Adult cuttlefish (n = 18) were held in a 3000 L aquarium and fed soft parts of the previously labelled mussels for 2 h. Immediately after ingestion, each individual was γ-counted and the same procedure was followed as for the juveniles. In addition, 3 adult cuttlefish were dissected at each counting time to determine the radiotracer distribution among their organs and tissues.

Radioanalyses.

Radioactivity was measured using a high-resolution γ-spectrometry system consisting of three coaxial Ge (N-or P-type) detectors Intertechnique) connected to a multichannel analyser and a computer with spectra analysis software (Interwinner, Intertechnique). The detectors were calibrated with appropriate standards for each of the counting geometries used, and measurements were corrected for background and physical decay of the radionuclides. Counting times were adapted to obtain relative propagated errors less than 5%. However, in a few cases, this counting precision could not be obtained even after 48 h of counting due to the very low activity in the extremely small dissected organs.

Counting times ranged from 10 min to 1 h for whole cuttlefish, mussels and brine shrimp, and from 10 min to 48 h for the dissected organs and tissues. Following the relatively short counting periods in the containers of sea water, none of the organisms displayed any obvious abnormal or stressed behaviour when returned to their aquaria.

Data and statistical analyses.

Uptake of 241 Am and 134 Cs from sediments and seawater was expressed, respectively, as whole-body transfer factors (TF) and concentration factors (CF) over time (Bq g -1 wet wt organism divided by the time-integrated Bq g -1 in sediments (TF) or seawater (CF).

Radionuclide loss was expressed in terms of percentage of remaining radioactivity over time, i.e. radioactivity at time t divided by initial radioactivity measured in the organisms at the beginning of the depuration period. Loss kinetics were described either by a single-component exponential model:

A t = A 0 e -k t ,
where A t and A 0 are remaining activities (%) at time t (d) and 0, respectively, or by a 2component exponential model:

A t = A 0s e -k s t + A 0l e -k l t ,
where the 's' subscript refers to a short-lived component (s component) and the 'l' subscript refers to a long-lived component (l component) [START_REF] Whicker | Radioecology: nuclear energy and the environment[END_REF][START_REF] Warnau | Biokinetics of selected heavy metals and radionuclides in the common Mediterranean echinoid Paracentrotus lividus: sea water and food exposures[END_REF]. The exponential model showing the best fit (based on calculation of the determination coefficients, R², and examination of the residuals) was selected.

The parameter k allows calculating of the radionuclide biological half-life (d) using the following equation:

T b½ = ln 2/k.
Constants of the models and their statistics were estimated by iterative adjustment of the model and Hessian matrix computation, respectively, using the non-linear curve-fitting routines in the Systat 5.2.1 Software [START_REF] Wilkinson | 239+240 Pu and 137 Cs concentrations in fish, cephalopods, crustaceans, shellfish, and algae collected around the Japanese coast in the early 1990s[END_REF]. Changes in radionuclide distribution among cuttlefish tissues and organs were tested for significance by the G procedure (adapted from the log-likelihood ratio test) for 2xk contingency tables [START_REF] Zar | Biostatistical analysis, 3 rd edn[END_REF]. Changes in % of radioactivity in a single tissue during the depuration period were tested by one-way ANOVA (after arcsin transformation of data) followed by the HSD Tukey's multiple comparison test.

The significance level for statistical analyses was always set at α =0.05.

RESULTS

Sediment exposure

Regular measurements of 241 Am concentration in sediment did not show any significant variation during the experimental time course (14.5 ± 1.8 Bq g -1 wet wt) while 134 Cs activities decreased from 12.4 ± 0.1 to 7.0 ± 0.4 Bq g -1 wet wt.

Very low 241 Am and 134 Cs activities were recorded in juveniles cuttlefish even after 29 d of exposure, and transfer factors (TF) were lower than 0.5 for both elements. Dissection of 3 individuals after 29 d of exposure showed that for both radionuclides the digestive gland contained the highest proportion of the whole-body burden, i.e. 47 ± 28 % of 241 Am and 49 ± 12 % of 134 Cs (Table 1).

Seawater exposure

Regular monitoring of the radionuclide concentrations in seawater allowed calculation of time-integrated radioactivities, viz. 6.4 ± 0.3 and 8.6 ± 0.7 kBq L -1 for 241m Am and 134 Cs, respectively.

Juveniles. The whole-body activities measured after 36 h exposure in spiked seawater were 38 ± 10 and 37 ± 1 Bq g -1 wet wt for 241 Am and 134 Cs, respectively, giving relatively low mean calculated whole-body CFs of 6 ± 2 and 4 ± 1 for these radionuclides.

Following transfer to non-contaminated seawater, loss kinetics of 241 Am in juvenile cuttlefish were best fitted by a single-component exponential model whereas loss of 134 Cs was best described by a two-component model (Figs 1A and1B; Table 2). Loss kinetics were characterised by a biological half-life (T b½ ) of 2 wk for 241 Am and 1 wk for 134 Cs.

At the end of the depuration period, 134 Cs was mainly associated with the digestive gland of the young cuttlefish (61 ± 4% of whole-body activity) whereas 241 Am was principally retained in the remaining tissues (61 ± 13%) (Table 1). The lowest fraction of both radiotracers was found in the cuttlebone (< 15% of the total activity).

Adults. 241 Am and 134 Cs activities recorded in whole-body as well as in the different organs and tissues of adult cuttlefish after 8 h of exposure and corresponding CFs are presented in Table 3. The highest activities of 241 Am were found in the branchial hearts and their appendages (264 ± 85 and 103 ± 66 Bq g -1 wet wt, respectively). In the case of 134 Cs, the branchial hearts, their appendages, gills and digestive tract displayed the highest activities, ranging from 9 to 13 Bq g -1 wet wt.

When considering the tissue distribution of the radionuclides, muscle and skin of adults (i.e. the sum of the mantle muscles, skin and head) contained the highest proportion of 241 Am and 134 Cs, viz. 68 and 85%, respectively (Table 3). A somewhat lesser 241 Am fraction was found in the branchial hearts and digestive gland (10 ± 2% for both tissues). The radionuclide distribution among the tissues did not vary significantly (G test, p > 0.05) between the beginning and the end of the depuration period (Table 3).

Food exposure

In these experiments, juveniles (n = 8) were fed ad libitum radiolabelled adult brine shrimp for 1 h and adult cuttlefish (n = 18) ingested a total of 123 radiolabelled mussels during a 2-h feeding period. Immediately after feeding, all cuttlefish were γ-counted for determination of their radionuclide content.

Juveniles. The loss kinetics of ingested 241 Am and 134 Cs were best fitted by a 2-component exponential model composed of one rapid loss component followed by a single slow component (Figs 1C and1D; Table 2). The short-lived component was derived from 40% and 70% of the initially ingested 241 Am and 134 Cs activities, respectively (Table 2), and was characterised by a T b½s < 1 d for both radionuclides. The long-lived component, which represents the fraction of the radionuclides actually absorbed by cuttlefish, displayed a T b½l of 5 d for 241 Am and 66 d for 134 Cs (Table 2). The same long-lived component allowed estimation of the assimilation efficiencies (AE) of the ingested nuclides. Results showed that 241 Am was readily assimilated in juveniles with an AE of 60% whereas the AE of 134 Cs was much lower, viz. 29% (Table 2). Dissections performed 29 d after feeding indicated that the highest proportion of remaining activity of both nuclides occurred in the digestive gland (ca. 60% of the whole-body activity; Table 1).

Adults. The loss kinetics of both radionuclides ingested with food by adult cuttlefish were best described by a 2-component exponential model. As shown Figs 1E and 1F and in Table 2, 69 and 78% of the ingested activity of 241 Am and 134 Cs, respectively, were rapidly lost with a T b½s of 4 and 13 h, respectively. The assimilated fraction of ingested 241 Am was much lower in adults than in juveniles (AE = 31 vs 60%) but was lost at a slower rate from adults with a T b½l of 28 d compared to 5 d in juveniles. For 134 Cs, AEs were nearly similar for both age groups (AE = 23 vs 29% in adults and juveniles, respectively), however the radionuclide was depurated much faster in adults (T b½l = 16 d) than in juveniles (T b½l = 66 d).

The tissue distribution of ingested radionuclides was determined on several occasions after feeding (Table 4). At the end of the depuration period, both 241 Am and 134 Cs were predominantly distributed in the digestive gland (viz. 98 and 54%, respectively). The distribution of 241 Am among tissues remained unchanged for 29 d of observation; in contrast, some significant changes were observed for 134 Cs (G-test, p ≤ 0.01). For example, the proportion of 134 Cs activity decreased in the muscular tissues (mantle muscles and head), whereas between 1 and 18 days of excretion it increased in the digestive gland (Table 4).

DISCUSSION

Cephalopods are an important resource of marine food and are fished and consumed in large quantities all around the world (Amaratunga 1983). Hence, the intake of contaminants such as radionuclides by humans through cephalopod consumption is a matter of potential concern.

Cephalopods have been reported to concentrate natural and anthropogenic radionuclides such as 210 Po, 210 Pb, 137 Cs, and 239+240 Pu in their tissues (e.g. [START_REF] Smith | Concentrations of the elements Ag, Al, Ca, Cd, Cu, Fe, Mg, Pb and Zn, and the radionuclides 210 Pb and 210 Po in the digestive gland of the squid Nototodarus gouldi[END_REF][START_REF] Finger | Molecular association of Cu, Zn, Cd and 210 Po in the digestive gland of the squid Nototodarus gouldi[END_REF], Yamada et al. 1999); however, little is known about the behaviour of radionuclides in these higher trophic level molluscs. To the best of our knowledge, only two species of cephalopods, viz. the octopus Octopus vulgaris and the squid Doryteuthis bleerkeri, have been used experimentally for investigating Am, Cs, and Pu biokinetics [START_REF] Suzuki | Accumulation of cesium-137 by useful Mollusca[END_REF][START_REF] Guary | Experimental studies on the biokinetics of plutonium and americium in the cephalopod Octopus vulgaris[END_REF]. These works were limited to measuring the uptake from seawater (i.e. [START_REF] Suzuki | Accumulation of cesium-137 by useful Mollusca[END_REF] or used a less than optimal experimental approach such as injecting the prey with radionuclides for the feeding experiments [START_REF] Guary | Experimental studies on the biokinetics of plutonium and americium in the cephalopod Octopus vulgaris[END_REF].

Cephalopods are found in a great variety of habitats from coastal waters to very deep ocean environments, with some living in direct contact with bottom sediments and others experiencing different environments during their life cycle (e.g. demersal species becoming temporarily pelagic during migration). Therefore, there is an obvious need to specifically determine 1) the uptake and retention of radionuclides at different stages of the life cycle of cephalopods, and 2) to assess the relative importance of the different pathways of exposure to radionuclides (sediments, seawater and food). In this context, the common cuttlefish Sepia officinalis appeared to be a good model cephalopod for such experiments as it spends part of its time buried in the sediment and is easy to rear and manipulate under laboratory conditions.

After a 1 month exposure to 241 Am and 134 Cs through sediments, juvenile cuttlefish still exhibited very low transfer factors (TF < 0.5), indicating that direct contamination due to burying into sediments is a minor uptake pathway for these radionuclides in cephalopods. The occurrence of a substantial fraction of both nuclides in internal tissues (viz. digestive gland and cuttlebone), which have no direct contact with the sediment, suggests that both radionuclides were progressively translocated from the tissues in direct contact with sediment and pore water to the digestive gland and, to a lesser extent, the cuttlebone (see Table 1). Such a translocation of elements to the cuttlebone was observed in a previous study on bioaccumulation of Cd in S. officinalis (Bustamante et al. 2002).

Following an acute contamination of adults via seawater, activities recorded in the whole cuttlefish suggest that they do not efficiently accumulate 241 Am and 134 Cs directly from the dissolved phase. Indeed, both elements displayed low whole-body CFs (CF = 2 for 241 Am and CF = 1 for 134 Cs). Nevertheless the 8 hr acute contamination time was relatively short and CFs are likely to be higher after a longer period of exposure. Activities of 134 Cs measured in the different organs and tissues were all of the same order of magnitude. In contrast, for 241 Am the organs involved in respiration (the branchial hearts, their appendages and the gills) and digestion (digestive gland) displayed higher activities compared to other body compartments (see Table 3). However, in terms of their relative distribution in the whole body, both radionuclides were mainly located in muscular tissues which represent the main fraction (viz. 75%) of the total body weight: muscles and head contained 65% and 82% of the total 241 Am and 134 Cs, respectively. A longer exposure (14 d) of Octopus vulgaris to 137 Cs in sea water gave a similar distribution (i.e. 88%) of the radioisotope in the edible parts [START_REF] Suzuki | Accumulation of cesium-137 by useful Mollusca[END_REF]. In contrast, a 15-d exposure of the same species in seawater spiked with 241 Am resulted in only ca. 20% of the retained radioactivity being found in the muscular parts with most of the 241 Am concentrated in the branchial hearts and their appendages [START_REF] Guary | Experimental studies on the biokinetics of plutonium and americium in the cephalopod Octopus vulgaris[END_REF]. In our experiments with S. officinalis, these tissues contained low percentages of the total 241 Am, most probably because of the short duration of the exposure in sea water.

Nevertheless, even after 8 hours they significantly concentrated the radionuclide with CFs as high as 42 in the branchial hearts and 16 in the appendages.

Both field and laboratory investigations with cephalopods have demonstrated the ability of branchial hearts to concentrate transuranic elements to fairly high levels [START_REF] Guary | High concentrations of transuranic and natural radioactive elements in the branchial hearts of the cephalopods Octopus vulgaris[END_REF][START_REF] Guary | Experimental studies on the biokinetics of plutonium and americium in the cephalopod Octopus vulgaris[END_REF]. This ability could be related to the presence of polyhedral cells containing granular, Fe-rich, pigment concretions (adenochromes) (e.g., [START_REF] Fox | Adenochrome a glandular pigment in the branchial hearts of the octopus[END_REF][START_REF] Nardi | Isolation and distribution of adenochrome(s) in Octopus vulgaris[END_REF]. The affinity of 241 Am for adenochromes in the branchial hearts has been demonstrated using autoradioradiographic techniques (Miramand & Guary 1981); however, adenochromes have not been found in the appendages of the branchial hearts [START_REF] Nardi | Isolation and distribution of adenochrome(s) in Octopus vulgaris[END_REF], an observation which suggests that they serve as an excretion pathway for 241 Am rather than as storage sites.

Following exposure of juveniles in contaminated seawater, subsequent 241 Am and 134 Cs elimination over a one month period displayed a one-and a two-component exponential loss model, respectively. Whole-body loss was relatively rapid for both nuclides with a mean T b1/2 of 14 and 6 d, respectively. After 29 d of depuration, residual 241 Am was mainly located in the remaining tissues (comprising the branchial hearts) of juveniles. However, as the juvenile branchial heart was not fully developed, additional work is needed to examine its role as a preferential storage organ such as occurs in adults.

In the case of dietary exposure, 31 ± 3% of the ingested 241 Am was assimilated into the tissues of adult cuttlefish, whereas in contrast 241 Am was absorbed to a much greater extent in juveniles (AE = 60 ± 10%). This difference between AEs could be due to differences in efficiency of digestion between juveniles and adults, since digestive metabolism is thought to decrease with age in cephalopods [START_REF] Mangold | Association of americium-241 with adenochromes in the branchial hearts of the cephalopod Octopus vulgaris[END_REF]). More likely, however, the difference could also be partly due to variations in the bioavailability of 241 Am in the food used for juveniles (brine shrimp) compared to that ingested by adults (i.e. mussels). Indeed, different storage mechanisms in prey can determine metal bioavailability to higher trophic levels [START_REF] Wallace | Bioavailability of biologically sequestered cadmium and the implications of metal detoxification[END_REF][START_REF] Wallace | Subcellular compartmentalization of Cd and Zn in two bivalves. II. Significance of trophically available metal (TAM)[END_REF][START_REF] Seebaugh | Bioenhancement of cadmium transfer along a multi-level food chain[END_REF], which in a similar fashion could lead to different proportions of transferable 241 Am. Overall, such very high AEs for 241 Am in the common cuttlefish are rather unique whereas in herbivorous bivalves, many crustaceans, echinoids, and fish, assimilation of particle-reactive transuranic elements is typically very low (e.g. [START_REF] Fowler | Accumulation and retention of plutonium by marine zooplankton[END_REF]Pentreath 1977Pentreath , 1981;;[START_REF] Fisher | Interaction of marine plankton with transuranic elements. 3. Biokinetics of neptunium, plutonium, americium, and californium in phytoplankton[END_REF][START_REF] Carvalho | Biokinetics of plutonium, americium and californium in the marine isopod Cirolana borealis, with observations on its feeding and molting behavior[END_REF][START_REF] Warnau | Biokinetics of selected heavy metals and radionuclides in the common Mediterranean echinoid Paracentrotus lividus: sea water and food exposures[END_REF]. Such a difference could be related to the organism's feeding regime since cephalopods are strict carnivores. For instance, unexpected high AEs (up to 60%) of plutonium have also been found in carnivorous crustaceans, viz. the crabs Carcinus maenas and Cancer pagurus [START_REF] Fowler | High absorption efficiency for ingested plutonium in crabs[END_REF]. Hence, the contribution of the radionuclide from the trophic pathway is very likely to be strongly enhanced in certain carnivorous invertebrates.

Once assimilated, 241 Am was retained to a much greater degree in adults, with a half-life approximately 6 times longer than in juveniles (i.e. 28 d vs 5 d), which suggests that different processes govern 241 Am elimination/retention in the two age groups. In other molluscs such as mussels, 241 Am has been reported to be strongly retained in the digestive gland (Bjerregaard et al. 1985[START_REF] Fisher | Influence of food composition on the biokinetics and tissue distribution of zinc and americium in mussels[END_REF], a finding which is in agreement with our own observations. Indeed, after 29 d of depuration, the major fraction of residual 241 Am was in the digestive gland, with a much higher fraction in adults than in juvenile cuttlefish (98% vs 59%). In the digestive gland of Octopus vulgaris, [START_REF] Guary | Experimental studies on the biokinetics of plutonium and americium in the cephalopod Octopus vulgaris[END_REF] reported that 241 Am is likely associated with the cellular waste products such as brown bodies. Considering this hypothesis together with our experimental observations, the longer retention of 241 Am observed in adult S. officinalis could be due to a more rapid turnover of digestive cells in juveniles, thus resulting in a higher 241 Am excretion rate.

In contrast to 241 Am, ingested 134 Cs was assimilated to a similar extent in juveniles (29%) and adults (23%), and the depuration rate constant was 4 times higher in adults resulting in a significantly much shorter 134 Cs half-life in adults (16 d) than in juveniles (66 d) (Table 2).

The longer retention time of 134 Cs in juveniles is difficult to explain since, for certain transition elements (Ag, Cd, Co and Zn) previously investigated in cuttlefish (Bustamante et al. 2002(Bustamante et al. , 2004) ) as well for 241 Am (our study), early juveniles displayed shorter retention halftimes than adults. The main difference in tissue distribution of 134 Cs between adults and juveniles was the higher proportion present in the cuttlebone (22 ± 21% in juveniles vs 2 ± 0% in adults; see Tables 1 and4). This higher skeleton-associated fraction is most likely tightly bound and hence results in the high retention half-time observed. Although, our results clearly indicate that 134 Cs does not follow the same excretion pathway as 241 Am, the above interpretation should be considered with caution since to the best of our knowledge, calcareous skeletons have not been shown to act as a particularly efficient sink for cesium in contrast with other elements such as, e.g., 241 Am or Pb (see e.g. [START_REF] Grillo | Comparative studies on transuranium nuclide biokinetics in sediment-dwelling invertebrates[END_REF][START_REF] Warnau | Distribution of heavy metals in the echinoid Paracentrotus lividus (Lmk) from the Mediterranean Posidonia oceanica ecosystem: seasonal and geographical variations[END_REF]). Furthermore, in our feeding experiment the very low activities measured in minute organs such as juvenile cuttlebone were frequently associated with low counting accuracy, which in turn can lead to a rather poor estimation of radioactivities and hence radionuclide distribution (as indicated by the elevated SD value of the cuttlebone-associated fraction of 134 Cs). Clearly, further study is needed to better understand the observed differences in the fate of 134 Cs and 241 Am once taken up in young and adult cephalopod tissues.

Captions to Figure Parameters for the best fitting equations are given in Table 3.

Fig. 1 .

 1 Fig. 1. Sepia officinalis. Whole-body loss kinetics of 241 Am and 134 Cs (% of remaining activity; mean ± SD): (A, B) juvenile cuttlefish previously exposed to spiked seawater for 36 h (n = 8 from day 0 to 20 and n = 4 on day 29); (C, D) juvenile cuttlefish previously fed radiolabelled brine shrimp (n = 8 from day 0 to 22 and n = 5 on day 29);(E, F) adult cuttlefish previously fed radiolabelled mussels (n = 18 on day 0, n = 15 from day 1 to 18, n = 12 from day 19 to 29).

  Fig. 1

Table 1 .

 1 Sepia officinalis. Distribution (%; mean ± SD) of 241 Am and 134 Cs among three body compartments of juvenile cuttlefish (1) after a 29-d exposure to spiked sediments, (2) after a 29-d depuration following a 36-h exposure to spiked seawater, and (3) after a 29-d depuration following ingestion of spiked food (brine shrimp).

	Exposure pathway	N	Body compartment
		Digestive gland	Cuttlebone	Remaining tissues
	1. Sediments (29-d exposure)	3
	241 Am	49 ± 12	12 ± 3	39 ± 15
	134 Cs	47 ± 28	17 ± 4	36 ± 24
	2. Seawater (29-d depuration)	4
	241 Am	27 ± 13	13 ± 0	61 ± 13
	134 Cs	61 ± 4	5 ± 0	34 ± 4
	3. Feeding (29-d depuration)	5
	241 Am	59 ± 23	12 ± 10	29 ± 16
	134 Cs	60 ± 27	22 ± 21	18 ± 14

Table 2 .

 2 Sepia officinalis. Parameters of the equations best fitting the whole-body loss kinetics of 241 Am and 134 Cs in cuttlefish previously exposed to the radionuclides via different pathways: (1) juveniles previously exposed for 36 h via seawater; (2) juveniles previously fed radiolabelled brine shrimp; (3) adults previously fed radiolabelled mussels. O and T: one-and two-exponential loss equations, respectively; A 0l : assimilation efficiency (AE); ASE: asymptotic standard error; R²: determination coefficient; p: probability of the model adjustment.

	Pathway	Model	A 0s (ASE)	λs (ASE)	T b1/2s (days) A 0l (ASE)	λl (ASE)	T b1/2l (days)	R²	p
	1. Loss in juveniles after seawater exposure
	241 Am	O	87.7 (2.8)	0.048 (0.005)	14	-	-	-	0.96	< 0.001
	134 Cs	T	74.6 (7.1)	1.015 (0.163)	0.7	25.6 (6.8)	0.114 (0.036)	6.1	0.97	< 0.001
	2.

Loss in juveniles after a single feeding on brine shrimp

	241 Am	T	39.6 (10.5) 1.282 (0.654)	0.5	60.3 (10.1)	0.137 (0.029)	5.1	0.95	< 0.001
	134 Cs	T	70.3 (4.4)	0.972 (0.153)	0.7	29.2 (3.6)	0.011 (0.008)	66	0.98	< 0.001
	3.

Loss in adults after a single feeding on mussels

	Remaining activity (%)			Remaining activity (%)
	Time (d)			Time (d)
	241 Am Remaining activity (%)	T	68.6 (3.8)	4.125 (3.683) Remaining activity (%) 0.17	31.4 (2.5)	0.025 (0.009)	28	0.95	< 0.001
	134 Cs	T	77.6 (4.4)	1.310 (0.197)	0.53	22.5 (3.7)	0.045 (0.019)	16	0.95	< 0.001
	Time

Acknowledgements. We thank N. Tevenin and P. Gilles (Musée Océanographique, Monaco) for providing us with the organisms. We are also grateful to E. Boucaud-Camou (Université de Caen, France) for her advice on cuttlefish rearing. MW is an Honorary Research Associate of the National Fund for Scientific Research (NFSR, Belgium). The International Atomic Energy Agency is grateful for the support provided to its Marine Environment Laboratory by the Government of the Principality of Monaco. LITERATURE CITED Amaratunga, T., 1983. The role of cephalopods in the marine ecosystem. In: IF Caddy (ed.). Advances in assessment of world cephalopod resources. FAO Fish Tech Pap 231:379-415 Bjerregaard, P., Topçuoğlu, S., Fisher, N.S., Fowler, S.W., 1985. Biokinetics of americium and plutonium in the mussel Mytilus edulis. Mar Ecol Prog Ser 21:99-111 Bustamante, P., Teyssié, J-L., Fowler, S.W., Cotret, O., Danis, B., Miramand, P., Warnau, M., 2002. Biokinetics of zinc and cadmium accumulation and depuration at different stages in the life cycle of the cuttlefish Sepia officinalis. Mar Ecol Prog Ser 231:167-177 Bustamante, P., Teyssié, J-L., Danis, B., Fowler, S.W., Miramand, P., Cotret, O., Warnau, M., 2004. Uptake, transfer and distribution of silver and cobalt in tissues of the common cuttlefish Sepia officinalis at different stages of its life cycle. Mar Ecol Prog Ser 269:185-195

 (CFs, mean), radionuclide activities (Bq g -1 wet wt; mean ± SD) and tissue distribution of radioactivity (%; mean ± SD) in adult cuttlefish after 8 h of exposure via seawater (n = 3) and after 6 d of depuration (n = 2).