P Bustamante

A F González

F Rocha

P Miramand

A Guerra

Metal and metalloid concentrations in the giant squid Architeuthis dux from Iberian waters

Keywords: trace element, bioaccumulation, cephalopod, giant squid, Architeuthis

This study investigated 14 trace elements (Ag, As, Cd, Co, Cr, Cu, Fe, Hg, Mn, Ni, Pb, Se, V and Zn) in the tissues of the giant squid Architeuthis dux from the Mediterranean and Atlantic Spanish waters. As for other families of cephalopods, the digestive gland and the branchial hearts of Architeuthis showed the highest concentrations of Ag, Cd, Co, Cu, Fe, Ni, Se, V and Zn, highlighting their major role in the bioaccumulation and detoxification processes. With the exception of Hg, the muscles showed relatively low trace element concentrations. Nevertheless, this tissue contained the main proportion of the total As, Cr, Hg, Mn, Ni, and Zn body burden because muscles represent the main proportion of the squid mass. These findings suggest that the metal metabolism is overall the same as other cephalopod families from neritic waters. In females, Zn concentrations increased in the digestive gland with the squid's weight likely reflecting physiological changes during sexual maturation. Comparing the trace element concentrations in the tissues of Architeuthis, higher Ag, Cu, Hg and Zn concentrations in the squid from the Mediterranean reflected different exposure conditions. In comparison to other meso-pelagic squids from the Bay of Biscay, Cd concentrations recorded in the digestive gland suggest that Architeuthis might feed on more contaminated prey or that it displays a longer life span that other cephalopods.

Introduction

Cephalopods play a key role in marine ecosystems both as predators and prey. They constitute a class of marine molluscs which are found in a great variety of habitats from coastal waters to very deep-ocean environments [START_REF] Boyle | Cephalopods: Ecology and Fisheries[END_REF]. Independently of the species, the habitat or the life span they display, cephalopods share the ability to accumulate inorganic and organic pollutants such as metals, PCBs or organochlorine pesticides (e.g. [START_REF] Martin | High copper concentrations in squid livers in association with elevated levels of silver, cadmium, and zinc[END_REF][START_REF] Tanabe | Polychlorobiphenyls, DDT, and hexachlorocyclohexane isomers in the western North Pacific ecosystem[END_REF][START_REF] Miramand | Concentration and distribution of heavy metals in tissues of two cephalopods, Eledone cirrhosa and Sepia officinalis, from the French coast of the English Channel[END_REF][START_REF] Yamada | Organotin compounds and polychlorinated biphenyls of livers in squid collected from coastal waters and Open Ocean[END_REF][START_REF] Bustamante | Bioaccumulation of 12 trace elements in the tissues of the nautilus Nautilus macromphalus from New Caledonia[END_REF]Bustamante et al., , 2006a;;[START_REF] Ueno | Specific accumulation of polychlorinated biphenyls and organochlorine pesticides in Japanese common squid as a bioindicator[END_REF][START_REF] Storelli | Concentrations and composition of organochlorine contaminants in different species of cephalopod molluscs from the Italian waters (Adriatic Sea)[END_REF].

Consequently, they were reported to constitute a significant vector of contaminants to the species feeding on them, in particular seabirds and marine mammals (e.g. [START_REF] Honda | Heavy metal concentrations in muscle, liver and kidney tissue of striped dolphin, Stenella coeruleoalba, and their variations with body length, weight, age and sex[END_REF][START_REF] Muirhead | Heavy metal concentrations in the tissues of seabirds from Gough Island, South Atlantic Ocean[END_REF][START_REF] Bustamante | Cephalopods as a vector for the transfer of cadmium to top marine predators in the north-east Atlantic Ocean[END_REF][START_REF] Lahaye | Long-term dietary preferences of common dolphins in the Bay of Biscay using a metallic tracer[END_REF]. However, most studies focused on commercially targeted species which are 1) easy to sample and 2) of high economic and health interest concerning human consumption. Many of these cephalopod species are also consumed by top marine predators, but there is a gap in the information concerning non-targeted species. This lack of data on bioaccumulation of contaminants is particularly obvious for oceanic and deep-sea species, like the giant squid Architeuthis.

Overall, the biology, behaviour and life cycle of Architeuthis are still poorly known in many aspects even if this squid has received considerable attention over the last decade (see [START_REF] González | Recent findings of the giant squid Architeuthis in northern Spanish waters[END_REF][START_REF] Guerra | A review of records of giant squid in the north-eastern Atlantic, with a note on the two first records of males Architeuthis sp. off the Iberian Peninsula[END_REF][START_REF] Guerra | Enigmas de la ciencia: el calamar gigante[END_REF][START_REF] Kubodera | First-ever observations of a live giant squid in the wild[END_REF]. Most of the information available on this squid comes from dead stranded animals and from predator trophic ecology studies. Indeed, squid flesh and beaks are often recorded in the stomach of sperm whales, but also of seabirds and sharks [START_REF] Roper | The giant squid[END_REF][START_REF] Clarke | Cephalopods as prey. III. Cetaceans[END_REF][START_REF] Santos | Additional notes on stomach contents of sperm whales Physeter macrocephalus stranded in the north-east Atlantic[END_REF][START_REF] Cherel | Stable isotopes, beaks and predators: a new tool to study the trophic ecology of cephalopods, including giant and colossal squids[END_REF]. Giant squids are also increasingly captured by trawling nets because of the development of deep-sea fisheries [START_REF] Guerra | Enigmas de la ciencia: el calamar gigante[END_REF]. In deep water conditions, the giant squid would have a particular diet and exposure conditions to trace elements and metals. For example, Hg bioavailability seems to be enhanced in these deep environments because the absence of solar radiation and the low oxygen concentrations in the deep environment favors a high methylation rate by bacteria [START_REF] Monteiro | Mercury concentrations in prey fish indicate enhanced bioaccumulation in mesopelagic environments[END_REF]. Also, Cd is enriched in mesopelagic waters while depleted in the surface ocean because of its regeneration from sinking biological debris from epipelagic zone and its uptake by organisms at the surface [START_REF] Boyle | On the marine geochemistry of cadmium[END_REF].

In this framework, the objectives of this study were to provide baseline data on a wide range of trace elements in the giant squid Architeuthis dux from the Spanish waters. To this end, the concentrations and tissue distribution of 12 metals (Ag, Cd, Co, Cr, Cu, Fe, Hg, Mn, Ni, Pb, V and Zn) and 2 metalloids (As and Se) were determined in the tissues and organs of fished and stranded specimens. The recorded values were then compared to the data from the current literature for other cephalopod species.

Material and methods

Sampling and sample preparation

Six specimens of giant squid were collected between 2001 and 2005. Table 1 shows the main data of the specimens as well as the sampling site, date, and mode of capture. Five specimens were caught in the Bay of Biscay (Asturias, North Spain) and one in the western Mediterranean Sea. Two of them were mature males and four were immature or maturing females.

Collected specimens were immediately frozen prior to dissection. In the laboratory, each individual was weighed and measured (mantle length ML, total length TL), and the sex and maturity stage determined. The digestive gland, gills, ink sack, branchial hearts and their appendages, systemic heart and brain were totally removed. In addition, pieces of muscle, skin, digestive and genital tissues (i.e. oviduct gland, ovary and testis), were sampled to determine trace element concentrations. As it was not possible to separate the different tissues and to weigh them, the total concentrations in the whole Architeuthis specimens were estimated according to the measured concentrations in the different tissues and to their relative weight in fishery targeted squids.

Analytical procedure

All samples were freeze-dried for several days then grounded. Aliquots of the samples ranging from 50 to 300 mg were digested using a 3:1 v:v nitric-hydrochloric acid mixture with 65% HNO 3 (Merck, suprapur quality) and 70% HCl (Merck, suprapur quality). Acidic digestion was performed overnight under ambient temperature and then heated in a microwave during 30 min with increasing temperature until 105°C, and 15 min at 105°C (1200 W). After the mineralization process, each sample was diluted to 30 or 50 ml with milli-Q quality water, according to the volume of acid added to the mineralization (3.0 ml or 4.5 ml).

Ag, As, Cd, Co, Cr, Cu, Fe, Mn, Ni Pb, Se, V and Zn were analysed either by Inductively Coupled Plasma-Optical Emission Spectrometry (Varian ® Vista-Pro) or Inductively Coupled Plasma-Mass Spectrometry (Varian ® Ultra Mass 700). For Hg, two aliquots ranging from 10 to 50 mg of dried material were directly analysed in an Advanced Mercury Analyser spectrophotometer (Altec ® AMA 254).

Reference tissues (dogfish liver DOLT-3, NRCC, and lobster hepatopancreas TORT-2, NRCC) were treated and analysed in the same way as the samples. Results were in good agreement with the certified values, and the standard deviations were low, proving good repeatability of the method (Table 2). The results for standard reference materials displayed recoveries of the elements ranging from 88% to 116% (n=10).

The detection limits (µg.g -1 dry wt) for ICP-OES were 8.3 (As, Fe, Zn), 3.3 (Ag, Se), 1.67 (Pb, V), 0.83 (Cd, Co, Cr, Cu, Mn, Ni), and were 0.167 (Ni, V), 0.083 (Cd, Co, Cr, Cu, Mn, Pb), 0.033 (Ag) for ICP-MS. Trace element concentrations are given relative to the dry weight (µg.g -1 dry wt) while the distribution percentages are calculated for wet weight.

Results

Trace element concentrations in soft tissues

Trace element concentrations in the tissues and organs of Architeuthis are reported in Figure 1. Among the sampled tissues, the digestive gland was the major site of concentration for Ag, Cd, Co, Cu, Fe and Se with 1.13-14.0 µg Ag.g -1 , 27.2-134 µg Cd.g -1 , 2.0-4.8 µg Co.g -1 , 64-1218 µg Cu.g -1 , 52-1862 µg Fe.g -1 and 9.7-19.5 µg Se.g -1 (Figure 1). The digestive gland also concentrated Ni, V and Zn at concentrations closed to the highest concentrations recorded in the other tissues with 0.11-1.42 µg Ni.g -1 , 0.60-4.34 µg V.g -1 , and 34-219 µg Zn.g -1 .

Interestingly, the digestive gland also exhibited the lowest concentrations of As, Cr and Hg with 31-65 µg As.g -1 , 0.41-1.40 µg Cr.g -1 , and 0.32-1.56 µg Hg.g -1 (Figure 1).

The concentrations of As, Co, Hg, Ni, Se and V were also remarkable in branchial hearts, which play an important excretory role in cephalopods. In this tissue, the concentrations of trace elements were the highest for Ni or very close to the highest for As, Co, Hg, Se and V.

Branchial hearts concentrated As from 86 to 111 µg.g -1 , Co from 3.15 to 3.22 µg.g -1 , Hg from 0.71 to 4.37 µg.g -1 , Ni from 0.74 to 2.60 µg.g -1 , Se from 8.2 to 9.3 µg.g -1 and V from 1.50 to 2.46 µg.g -1 (Figure 1).

Muscles exhibited generally among the lowest concentrations for all trace elements except for Hg which exhibited the highest values in this tissue (1.86-3.32 µg Hg.g -1 dry weight; Figure 1). Cr and Zn were highly concentrated in the ovary with concentrations ranging from 0.65 to 6.12 µg.g -1 and 131 to 149 µg.g -1 , respectively. The oviduct gland also displayed high concentrations of As and Mn as well as the digestive gland appendages had high concentrations of Cr and Pb (Figure 1).

Distribution of the trace elements in soft tissues

The proportions of the whole body burden of the trace elements contained in each organ and tissue are shown in Figure 2. With the exception of As, Cr, Hg and Mn which were mainly found in the body muscular parts (69 ± 9%, 68 ± 20%, 87 ± 5%, and 55 ± 22%, respectively), the digestive gland contained the largest quantities of all trace elements : 98 ± 1% of Ag, 99 ± 1% of Cd, 97 ± 2% of Co, 92 ± 5% of Cu, 85 ± 16 % of Fe, 53 ± 28% of Ni, 69 ± 22% of Pb, 77 ± 5% of Se, 71 ± 19% of V, and 53 ± 17% of Zn (Figure 2).

Although the concentrations of some trace elements were high in the branchial hearts or in the gills, these tissues contained in fact low amounts of the considered elements because of their small masses relative to the whole body weight (Figure 2).

Influence of the size/weight and origin

The size/weight only influenced Zn concentrations in the digestive gland (R²=0.934, p=0.020) and Cr and Ni concentrations in the gills (R²=0.968 p=0.007 and R²=0.969 p=0.007).

No statistical tests were performed to compare the influence of the origin on the accumulated trace elements because our sampling only included one specimen from the Mediterranean.

However, this giant squid clearly displayed higher concentrations of Hg than any of those from the Bay of Biscay with 1.56 vs 0.47 ± 0.13 µg.g -1 in the digestive gland (see Table 3), 2.97 vs 1.38 ± 0.34 µg.g -1 in the gills, and 3.32 vs 2.07 ± 0.19 µg.g -1 in the mantle muscle (data not shown). Important differences also appeared for Ag, Cu and Zn with 14.0 vs 1.90 ± 0.47 µg Ag.g -1 , 1218 vs 108 ± 83 µg Cu.g -1 , and 219 vs 103 ± 51 µg Zn.g -1 in the digestive gland (Table 3), and 4.80 vs 0.31 ± 0.11 µg Ag.g -1 , 206 vs 31 ± 24 µg Cu.g -1 and 111 vs 60 ± 35 µg Zn.g -1 in the gills (data not shown).

Discussion

Previous studies have demonstrated the ability of cephalopods to accumulate high concentrations of trace elements in their tissues but very little data is available in the current literature for non-targeted and/or deep-waters species such as the giant squid Architeuthis.

Although globally poorly documented, metal and metalloid concentrations in cephalopod tissues have received increasing interest over the last decades, particularly in Europe and Japan, as these molluscs play a major role both as predators and food items in marine ecosystems (see the reviews by [START_REF] Clarke | Cephalopods as prey. III. Cetaceans[END_REF][START_REF] Croxall | Cephalopods as prey. I. Seabirds[END_REF][START_REF] Klages | Cephalopods as prey. II. Seals[END_REF][START_REF] Smale | Cephalopods as prey. IV. Fishes[END_REF][START_REF] Boyle | Cephalopods: Ecology and Fisheries[END_REF]. The central role of the digestive gland in trace element bioaccumulation and detoxification has been highlighted many times, particularly for toxic metals such as Ag and Cd (e.g. [START_REF] Martin | High copper concentrations in squid livers in association with elevated levels of silver, cadmium, and zinc[END_REF][START_REF] Miramand | Concentration and distribution of heavy metals in tissues of two cephalopods, Eledone cirrhosa and Sepia officinalis, from the French coast of the English Channel[END_REF][START_REF] Bustamante | Biokinetics of zinc and cadmium accumulation and depuration at different stages in the life cycle of the cuttlefish Sepia officinalis[END_REF][START_REF] Bustamante | Uptake, transfer and distribution of silver and cobalt in the tissues of the common cuttlefish Sepia officinalis at different stages of its life cycle[END_REF]Ichiashi et al., 2001a;[START_REF] Miramand | Variation of heavy metal concentrations (Ag, Cd, Co, Cu, Fe, Pb, V, Zn) during the life cycle of the common cuttlefish Sepia officinalis[END_REF]. Beside the digestive gland, which also plays a major role in the energetic metabolism of cephalopods (e.g. [START_REF] Rosa | Biochemical composition of cephalopods with different life strategies, with special reference to a giant squid, Architeuthis sp[END_REF][START_REF] Moltschaniwskyj | Evidence that lipid can be digested by the dumpling squid Euprymna tasmanica, but is not stored in the digestive gland[END_REF], the branchial hearts and their appendages are involved in trace element excretion processes, allowing the depuration and/or the storage of various metals and radionuclides (e.g. [START_REF] Nardi | Isolation and distribution of adenochrome(s) in Octopus vulgaris[END_REF][START_REF] Miramand | High concentrations of some heavy metals in tissues of the Mediterranean octopus[END_REF][START_REF] Guary | High concentrations of transuranic and natural radioactive elements in the branchial hearts of the cephalopods Octopus vulgaris[END_REF][START_REF] Miramand | Concentration and distribution of heavy metals in tissues of two cephalopods, Eledone cirrhosa and Sepia officinalis, from the French coast of the English Channel[END_REF][START_REF] González | Vulcanoctopus hydrothermalis gen. et sp. nov. (Mollusca, Cephalopoda): a new octopod from a deep sea hydrothermal vent[END_REF][START_REF] Bustamante | Biokinetics of zinc and cadmium accumulation and depuration at different stages in the life cycle of the cuttlefish Sepia officinalis[END_REF]Bustamante et al., , 2006b)). As for coastal and/or oceanic targeted cephalopods, the digestive gland and the branchial hearts of Architeuthis generally contained the highest concentrations of most of the considered elements, i.e. Ag, Cd, Co, Cu, Fe, Ni, Se, V and Zn (Figure 1). This finding strongly suggests that the metabolism of trace elements in Architeuthis is very close to, or even the same as other families of cephalopods. This is supported by the fact that trace element concentrations in the digestive gland of Architeuthis closely fall within the same range than for other cephalopod species (Table 3), indicating that the potential of Architeuthis for their bioaccumulation is relatively similar. Moreover, according to the elevated proportions of the total element body burden, the digestive gland of Architeuthis might play a central role in the detoxification and storage of most of the analysed elements, i.e. Ag, Cd, Co, Cu, Fe, Ni, Pb, Se, V and Zn (Figure 2). It would be a great interest to investigate the detoxification strategies in the digestive gland of Architeuthis in comparison to that of other families of cephalopods.

Besides the digestive gland, muscles contained very high proportions of total body burdens of As, Cr, Hg, Mn, Ni, and Zn (Figure 2). With the exception of Hg, trace element concentrations in Architeuthis muscles were relatively low and these high proportions resulted from the elevated muscular mass respective to the whole body weight. For Hg, the concentrations recorded in the muscles were the highest among the different organs and tissues (Figure 1). Previous studies have reported relatively similar Hg concentrations between the different tissues of different squid species from the Northern Atlantic waters (Bustamante et al., 2006a;[START_REF] Pierce | Geographic, seasonal and ontogenetic variation in cadmium and mercury concentrations in squid (Cephalopoda: Teuthoidea) from UK waters[END_REF]. It is therefore noteworthy that muscular Hg concentrations in Architeuthis were 2 to 4 times higher than in the digestive gland. In comparison with Ag, Cd, Co, Cu, Fe, Ni, Pb, Se, V and Zn, the role of the digestive gland in the storage of Hg appeared to be relatively limited in Architeuthis. This may be due to an excretion function of Hg by the digestive gland, and/or a preferential redistribution of Hg to muscular tissues where it binds to the sulphydryl groups of proteins [START_REF] Bloom | On the chemical form of mercury in edible fish and marine invertebrate tissue[END_REF]Bustamante et al., 2006a). Such a redistribution might indicate that most of the Hg ingested from the prey would be in the organic form such as fish in which Hg content is virtually 100% in the methylated form [START_REF] Bloom | On the chemical form of mercury in edible fish and marine invertebrate tissue[END_REF]. Little is known about the diet of Architeuthis, it includes other cephalopods (Pérez-Gándaras and [START_REF] Pérez-Gándaras | Nueva cita de Architeuthis (Cephalopoda: Teuthoidea): descripción y alimentación[END_REF][START_REF] Pérez-Gándaras | Architeuthis de Sudafrica: nuevas citas y consideraciones biológicas[END_REF], crustaceans (e.g.

Nephros norvergicus) as well as a large proportion of fish of different families accordingly to the available prey in the area (for instance Trachurus trachurus, Maurolicus muelleri and Micromesistius poutassou in Ireland waters and equivalent species from Namibia and New Zealand) (see [START_REF] Guerra | Enigmas de la ciencia: el calamar gigante[END_REF] for a review). Further studies therefore should focus on trace elements in the typical prey of Architeuthis to provide insights on this aspect.

As in other cephalopod species, trace element concentrations in Architeuthis may vary with biological and environmental factors such as age (size/weight), sex, and geographical origin (e.g. Monteiro et al., 1992;[START_REF] Bustamante | Cephalopods as a vector for the transfer of cadmium to top marine predators in the north-east Atlantic Ocean[END_REF][START_REF] Raimundo | Geographical variation and partition of metals in tissues of Octopus vulgaris along the Portuguese coast[END_REF][START_REF] Pierce | Geographic, seasonal and ontogenetic variation in cadmium and mercury concentrations in squid (Cephalopoda: Teuthoidea) from UK waters[END_REF]. Our limited sampling did not allow making comparisons for all these factors. However, considering the 5 specimens from the Bay of Biscay, it appears that size/weight poorly influenced trace element concentrations in Architeuthis tissues. The increase of Zn concentrations from 34 to 160 µg.g -1 dwt in the digestive gland might be related to metal physiological changes related to the sexual maturation as reported for other cephalopod species like Sepia officinalis from the English Channel [START_REF] Miramand | Variation of heavy metal concentrations (Ag, Cd, Co, Cu, Fe, Pb, V, Zn) during the life cycle of the common cuttlefish Sepia officinalis[END_REF]. In females, high concentrations of Zn in the ovary of Architeuthis (120 ± 38 µg.g -1 dwt) were close to that in the genital tract of Sepia officinalis (123 ± 3 µg.g -1 dwt; [START_REF] Miramand | Concentration and distribution of heavy metals in tissues of two cephalopods, Eledone cirrhosa and Sepia officinalis, from the French coast of the English Channel[END_REF]. Within the ovary, essential elements such as Zn are stored in metal-containing enzymes and metalloproteins [START_REF] Gerpe | Cadmium, zinc and copper accumulation in the squid Illex argentinus from the Southwest Atlantic Ocean[END_REF] and transferred to the yolk of the eggs [START_REF] Villanueva | Composition in essential and non-essential elements of early stages of cephalopods and dietary effects on the elemental profiles of Octopus vulgaris paralarvae[END_REF]Bustamante, 2006, Lacoue-Labarthe et al., 2008). Cr and Ni in the gills also displayed a significant increase with size/weight. In the current literature, very little data is available on the variation of Cr and Ni concentrations in cephalopod tissues. For example, in the squid Sthenoteuthis oualaniensis Cr concentrations were higher in juveniles than in adults, whereas juveniles displayed lower Ni concentrations than adults (Ichihashi et al., 2001a). Such a difference for S. oualaniensis was explained by the evolution of food habits between the juvenile and adult stages, juvenile feeding more on crustaceans while adults primarily preyed on fish. According to the lack of significant variation in the digestive gland, such a switch is not likely to occur in the size range of Architeuthis we analysed. Furthermore, even if the diet could represent the main pathway for many elements -as experimentally shown for Am, Cd, Co and Zn [START_REF] Koyama | Bioaccumulation of waterborne and dietary cadmium by oval squid Sepioteuthis lessoniana, and its distribution among organs[END_REF][START_REF] Bustamante | Biokinetics of zinc and cadmium accumulation and depuration at different stages in the life cycle of the cuttlefish Sepia officinalis[END_REF][START_REF] Bustamante | Uptake, transfer and distribution of silver and cobalt in the tissues of the common cuttlefish Sepia officinalis at different stages of its life cycle[END_REF]Bustamante et al., , 2006b) -seawater could also be an important uptake pathway, as elements pass through the skin and through the gills. For instance, seawater represents the main pathway for Ag in Sepia officinalis [START_REF] Bustamante | Uptake, transfer and distribution of silver and cobalt in the tissues of the common cuttlefish Sepia officinalis at different stages of its life cycle[END_REF]. Therefore, Cr and Ni bioaccumulation in gills might also result from a direct uptake from seawater all along the lifespan of

Architeuthis.

Trace element concentrations in cephalopods could also vary according the location where individuals were captured [START_REF] Bustamante | Cephalopods as a vector for the transfer of cadmium to top marine predators in the north-east Atlantic Ocean[END_REF]Seixas et al., 2005ab;[START_REF] Pierce | Geographic, seasonal and ontogenetic variation in cadmium and mercury concentrations in squid (Cephalopoda: Teuthoidea) from UK waters[END_REF].

In the case of Architeuthis, this is clearly exemplified by the much higher Hg concentrations in the tissue of the specimen from the Mediterranean. Higher Hg concentrations in Mediterranean organisms have been highlighted many times and are typically explained by high temperatures and absence of solar radiation in the deep environment. These conditions favor a high methylation rate of the metal, methyl-Hg being highly bioavailable for marine biota, which consistently biomagnifies through the food chain. Moreover, natural sources of Hg in the Mediterranean Sea may contribute to Hg enrichment through the benthic food webs, as it constitutes the richest natural reserve of this element [START_REF] Bacci | Mercury in the Mediterranean[END_REF]. Higher Ag, Cu and Zn concentrations in the digestive gland and in the gills also indicated different exposure conditions of this specimen compared to those from the Bay of Biscay. In cephalopods, Ag bioaccumulation in the digestive gland might reflect the global contamination of the surrounding waters (e.g. [START_REF] Martin | High copper concentrations in squid livers in association with elevated levels of silver, cadmium, and zinc[END_REF][START_REF] Miramand | Variation of heavy metal concentrations (Ag, Cd, Co, Cu, Fe, Pb, V, Zn) during the life cycle of the common cuttlefish Sepia officinalis[END_REF], seawater being the main pathway of exposure and Ag having a fast turn-over in the tissues [START_REF] Bustamante | Uptake, transfer and distribution of silver and cobalt in the tissues of the common cuttlefish Sepia officinalis at different stages of its life cycle[END_REF]. Interestingly, Ag, Cu, Hg and Zn are metals that bind to metallothionein proteins (MTs), which play a role in the homeostasis of the essential metals (i.e. Cu and Zn) and in the detoxification of non-essential metals (i.e. Ag, Cd and Hg) [START_REF] Roesijadi | Metallothionein in metal regulation and toxicity in aquatic animals[END_REF][START_REF] Roesijadi | Metallothionein and its role in toxic metal regulation[END_REF][START_REF] Viarengo | Mini-review. Mechanisms of heavy metal cation homeostasis in marine invertebrates[END_REF]. The role of MTs in metal sequestration in cephalopods is not completely clear (Bustamante et al., 2006c) and this issue clearly deserves further investigation.

Even though trace elements are generally considered for their potential toxicity in ecotoxicological studies and biomonitoring surveys, there is increasing interest in their use for providing information on life history and trophic ecology of cephalopods [START_REF] Jackson | Applying new tools to cephalopod trophic dynamics and ecology: perspectives from the Southern Ocean Cephalopod Workshop, February 2-3, 2006[END_REF]. Thus, Cd is of particular interest because it is highly bioaccumulated by cephalopods. Indeed, Cd is efficiently absorbed and strongly retained in the digestive gland [START_REF] Bustamante | Cephalopods as a vector for the transfer of cadmium to top marine predators in the north-east Atlantic Ocean[END_REF][START_REF] Bustamante | Biokinetics of zinc and cadmium accumulation and depuration at different stages in the life cycle of the cuttlefish Sepia officinalis[END_REF]. Even if most of cephalopod species display short life spans i.e. typically less than 2 years, they can accumulate very high Cd concentrations in their digestive gland reaching up 1000 µg g -1 wet wt in the squid Illex argentinus [START_REF] Dorneles | Cephalopods and cetaceans as indicators of offshore bioavailability of cadmium off Central South Brazil Bight[END_REF]. Because of Cd incorporation by organisms in epipelagic waters and its regeneration from sinking biological debris in the mesopelagic environment [START_REF] Boyle | On the marine geochemistry of cadmium[END_REF], deep-water cephalopods might show relatively high Cd concentrations. In the Bay of Biscay, Cd concentrations reached 9.1 and 33.1 µg.g -1 dry wt in the digestive gland of the mesopelagic squids Histioteuthis reversa and Teuthowenia megalops, respectively (unpublished data) and would be due to the consumption of prey highly contaminated with Cd. Even elevated compared to neritic squid species [START_REF] Bustamante | Cephalopods as a vector for the transfer of cadmium to top marine predators in the north-east Atlantic Ocean[END_REF], these concentrations are lower than those measured in the digestive gland of Architeuthis, suggesting that 1) the giant squid feed on more contaminated prey than Histioteuthidae and Cranchidae, or 2) it displays a much longer life span that other cephalopods. Age estimation and growth rates of giant squids are still open questions. Indeed, isotopic analysis indicated that the age for the giant squid Architeuthis sanctipauli from Tasmania was 14 years for specimens ranging from 191 to 240 cm ML [START_REF] Landman | Habitat and age of the giant squid (Architeuthis sanctipauli) inferred from isotopic analyses[END_REF]. This completely disagrees with the age estimated from growth increment counts in statoliths of Architeuthis dux and Architeuthis sp. from the Atlantic and caught off New Zealand. Specimens ranging from 43 to 161 cm ML had between 153 and 435 increments [START_REF] Jackson | Growth rings within the statolith microstructure of the giant squid Architeuthis[END_REF][START_REF] Gauldie | Statocyst, statolith, and age estimation of the giant squid, Architeuthis kirki[END_REF][START_REF] Lordan | Observations on morphology, age and diet of three Architeuthis caught off the west coast of Ireland in 1995[END_REF][START_REF] González | Recent findings of the giant squid Architeuthis in northern Spanish waters[END_REF]. If the increments were daily deposited, as it occurs in other cephalopods, the age of these animals will not exceed two years. This finding implies that Architeuthis would have a very fast growth rate with intense food intakes that in turn would lead to the bioaccumulation of relatively high Cd concentrations in its digestive gland.

Caption to figures

Figure 1 .

 1 Figure 1. Mean trace element concentrations (µg.g -1 dry wt) in the tissues of Architeuthis dux from Spanish waters. Scale bars represent 1 standard deviation (n=6). B: Brain; BH: branchial hearts; BHA: branchial heart appendages; DG: digestive gland; DGA: digestive gland appendages; DT: digestive tract; G: gills; IS: ink sack; M: muscles; O: ovary; OG: oviduct gland; S: skin; SH: systemic heart; T: testis.

Figure 2 .

 2 Figure 2. Percentage distribution of trace elements in the tissues of Architeuthis dux from Spanish waters. Scale bars represent 1 standard deviation. BH: branchial hearts; DG: digestive gland; DGA: digestive gland appendages; DT: digestive tract; G: gills; IS: ink sack; M: muscles; O: ovary; S: skin; SH: systemic heart.

Table 1 .

 1 Sampling information and squid biological characteristics. T: trawling; S: Stranded; F: Floating at the surface; ML: Dorsal mantle length; TL:

	Total length.
	N°	Sampling site	Sampling date	Mode of capture	Weight	ML	TL	Sex / Maturation
					(kg)	(cm)	(cm)
	1	Off Luarca (Asturias, North Spain)	12 September 2001	T	90	127	710	F / Immature
	2	Ribadesella (Asturias, North Spain)	23 October 2001	S	104	150	800	F / Maturing
	3	Colunga (Asturias, North Spain)	15 September 2003	S	80	152	1200	F / Immature
	4	Off Gijón (Asturias, North Spain)	16 September 2003	F	66	122	620	M / Mature
	5	Off Gandía (Valencia, Western Mediterranean)	19 July 2005	T	50	107	600	M / Mature
	6	Off Gijón (Asturias, North Spain)	22 July 2005	T	139	146	820	F/ Maturing

Table 2 .

 2 Comparison of certified trace elements concentrations (µg g -1 dry weight) in reference materials (n=10) with the values determined in the present study (nc: not certified value, nd: not determined). ICP-MS -Inductively Coupled Plasma Mass Spectrometry; ICP-OES -Inductively Coupled Plasma Optical Emission Spectrometry; AMA -Advanced Mercury Analyser.

	TORT-2

Table 3 .

 3 Reported metal concentrations (Mean ± SD, µg g -1 dry weight) in the digestive gland of different cephalopod species.

	Species	Ag	As	Cd	Co	Cr	Cu	Fe	Hg	Mn	Ni	Pb	V	Zn	Reference

Acknowledgements

We thank Luis Laria and other members of CEPESMA who collected the squids, Dr. Alberto Fernández Costa for his technical assistance in the dissections, and the staff of the Centre Commun d'Analyses (University of La Rochelle) for their help in the analytical process. This study was supported financially by LIENSs (University of La Rochelle).


L De Paz

J M Neto

J C Marques

A Laborda

De Paz

Response of intertidal macrobenthic communities

Keywords: Benthos, Long-term, Shellfish culture, Ecosystem disturbance, Environmental impact, Eo estuary, Cantabric Sea

published or not. The documents may come

INTRODUCTION

The estuarine faunal composition may show considerable variation, over spatial scales from metres to kilometres, and temporal scales of days to years (Morrisey et al., 1992a, b). Such variations in diversity or abundance of benthos in estuarine systems have been often correlated with environmental variables (e.g. salinity; Sanders, 1965; tidal level; [START_REF] Warwick | Distribution of benthic macrofauna associations in the Bristol Channel in relation to tidal stress[END_REF]sediment characteristics;[START_REF] Gray | Animal-sediment relationships[END_REF]Chapman and Tolhust, 2004), density of seagrass plants (e.g. [START_REF] Boström | Community structure and spatial variation of benthic invertebrates associated with Zostera marina (L.) beds in the northern Baltic Sea[END_REF], biotic interactions (e.g. [START_REF] Wilson | Competition and predation in marine soft-sediment communities[END_REF], or perturbations (e.g. [START_REF] Pearson | Macrobenthic succession in relation to organic enrichment and pollution of the marine environment[END_REF][START_REF] Llansó | Effects of hypoxia on estuarine benthos: the lower Rappahannock River (Chesapeake Bay), a case study[END_REF][START_REF] Diaz | Marine benthic hypoxia: a review of its ecological effects and the behavioural responses of benthic macrofauna[END_REF][START_REF] Belan | Marine environmental quality assessment using polychaete taxocene characteristics in Vancouver Harbour[END_REF]. Therefore, the estuarine community structure can change as a continuum on various spatial and temporal scales in relation to both natural and anthropogenic gradients [START_REF] Pearson | Macrobenthic succession in relation to organic enrichment and pollution of the marine environment[END_REF][START_REF] Rakocinski | Macrobenthic Responses to Natural and Contaminant-Related Gradients in Northern Gulf of Mexico Estuaries[END_REF]. Sampling can capture only a segment of a community at temporally or spatially discrete intervals, often missing critical transition points in between, and it may become difficult to establish causal relationships between stressors and effects on marine resources. Long-term surveys are a useful tool to assess changes in marine ecosystems, because they characterize environments and provide a basis to understand temporal and spatial variations in community structure and natural sources of stress that might mask a response to anthropogenic disturbance.

In recent years, aquaculture production has increased worldwide, as well as the need of assessing its environmental effects. Although depositional effects from shellfish farming culture can be negligible (e.g., [START_REF] Crawford | Qualitative risk assessment of the effects of shellfish farming on the environment in Tasmania, Australia[END_REF][START_REF] Grant | A multidisciplinary approach to evaluating impacts of shellfish aquaculture on benthic communities[END_REF], most studies describe a degree of sediment enrichment beneath culture areas and an alteration in the benthic community [START_REF] Castel | Influence of seagrass beds and oyster parks on the abundance and biomass patterns of meio-and macrobenthos in tidal flats[END_REF]Harstein and Rowden, 2004;[START_REF] Forrest | Benthic impacts of intertidal oyster culture, with consideration of taxonomic sufficiency[END_REF]. However, additional impacts, related to intertidal culture, have also been observed, such as topographic changes resulting from altered patterns of sediment accretion or erosion (e.g., [START_REF] Ottmann | Relationship between marine bottom elevation and various types of marine culture[END_REF][START_REF] Everett | Effects of oyster mariculture on submerged aquatic vegetation: An experimental test in a Pacific northwest estuary[END_REF], accumulation of shell litter, destruction of seagrasses beds, and physical disturbance from operational activities. The increase in shellfish culture, mainly oyster culture, in estuarine intertidal areas around the world, and the high economic and ecologic importance of estuaries, requires a more intensive survey of the impacts of this activity on benthic habitats and communities.

The Eo estuary, located in the north coast of Spain, has been a centre of bivalve farming since the 1970s [START_REF] Guerra | Primeras experiencias de cultivo de almeja japonesa (Tapes semidecussatus) y ostra japonesa (Crassostrea gigas) en la ría de Ribadeo (N.W. de España)[END_REF]. Shellfish are off-bottom cultivated, in culture bags placed onto racks. In the first years, the main species cultivated was Ostrea edulis, but since 1978 this oyster culture collapsed due to infestation by Bonamia ostreae, a parasite from California, which caused severe mortalities (Carrasco et al., 1990). In late 1980s and early 1990s, shellfish farming in the estuary received a new impulse with the introduction of the Pacific oyster (Crassostrea gigas) and manila clam (Ruditapes philippinarum) (Gigarría et al., 1995). A total area of 0.56 km 2 was retained for shellfish farming within an inlet located in the eastern side of the estuary (Linera inlet), representing 37% of the whole inlet area. Since 2000 all the farming area has been occupied with oyster racks. The production was estimated at 185 tonnes in 2001, increasing to 368 tonnes in 2004, and since then has become stabilized at approximately 300 tonnes per year. Recently, new plans aiming at increasing the bivalve production to 400 tonnes per year have been developed.

In addition to the development of oyster farming, the building of a new harbour in the western part of the estuary in 1994 created an additional human pressure. Although we do not have data on currents, it might have induced changes in hydrodynamics, which may have caused the observed increase in sand deposition in the navigation channels.

Furthermore, it has probably been necessary to increase the frequency of dredging activities to maintain the main navigation channels operational, but we do not have quantitative data on this activity.

To evaluate the possible impacts resulting from these anthropogenic pressures on the Eo estuary macrobenthic communities, we examined the spatial and temporal changes in the system response before and after the introduction of intensive shellfish farming and the harbour construction and operation [START_REF] Hill | Changes in Irish Sea Benthos: Possible Effects of 40 years of Dedging[END_REF]). An intensive survey of the intertidal benthic fauna was carried out in 1990 at the outer part of the Eo estuary [START_REF] Laborda | Cartografía de los sedimentos superficiales de la Ría del Eo (Asturias). VII Simposio Ibérico de Estudios del Bentos Marino[END_REF][START_REF] Laborda | Informe ambiental de la Ría del Eo. Conserjería de Pesca y Ordenación del Territorio[END_REF], and the results from that study were considered as representing a reference situation prior to the observed changes.

MATERIAL AND METHODS

Study area

The Eo estuary is a shallow mesotidal system (tidal range = 1.20 to 4.80 m) located on the northern Cantabric coast of Spain (43º 28', 33'N; 7º 00', 03'W). It is 10 km in length, with an average width of 800 m, and a total area of 9.6 km 2 . Depth varies between 2 and 7 m, and the intertidal area represents approximately 83% of the total estuarine area. According to the Water Framework Directive (WFD) typology established for the Basque Country (Spain) [START_REF] Borja | Implementation of the European Water Framework Directive from the Basque country (northern Spain): a methodological approach[END_REF], the Eo estuary is classified as type TWII. This type includes mesotidal (tidal range 1-3 m) polyhalineeuhaline estuaries with extensive intertidal areas and permanent stratification.

Sampling stations (Fig. 1) were located in the outer part of the estuary, which comprised two sectors: a) a sheltered inlet, the Linera inlet, at the eastern side of the Estuary; and b) a sand bare area consisting of sand banks in the main channel.

The Linera inlet has an area of 1.5 km 2 and represents a largely flat bottom mainly covered by Zostera noltii and Zostera marina seagrass beds. Zostera noltii occurs on muddy bottoms, at higher tidal levels, in the inner parts of the inlet, while a mixed meadow consisting of Z. noltii and Z. marina is present on sandy-mud bottoms from lower tidal levels, closer to the inlet's mouth.

The shape of the sand banks has been changing continuously and their extension increased over time (Encinar y Flor, 1983). Similar to other estuaries from the northern coast of Spain (Currás and Mora, 1991), the Eo is undergoing a gradual process of infilling with marine and estuarine deposits.

In both sectors, salinity close to the bottom usually ranges from 34 to 36 ppt.

2.2 Reference situation [START_REF] Laborda | Informe ambiental de la Ría del Eo. Conserjería de Pesca y Ordenación del Territorio[END_REF]Cimas, 1991, Laborda, 1995) The analysis of 100 samples collected in 1990 (25 randomly selected sampling stations were seasonally sampled) allowed the characterization of three different habitat types based on sediment's granulometry, organic matter content, presence of seagrasses, and macrobenthic communities' distribution (Table 1) [START_REF] Laborda | Cartografía de los sedimentos superficiales de la Ría del Eo (Asturias). VII Simposio Ibérico de Estudios del Bentos Marino[END_REF][START_REF] Laborda | Informe ambiental de la Ría del Eo. Conserjería de Pesca y Ordenación del Territorio[END_REF][START_REF] Laborda | Distribución de las praderas de Zostera noltii Hornem., 1832 en el litoral del norte de España[END_REF].

Within each habitat type, different zones were identifiable (Table 1):

• Three different zones within the bare sediment (BS) habitat (Fig. 1), which may be explained as a function of hydrodynamic conditions (Table 1).

• Four different zones within the apparently homogeneous habitat of mudflats covered by Z. noltii (MZ) (Fig. 1), explainable as a function of spatial variations in the density of seagrass shoots (Table 1).

• Two slightly different zones were recognized within the mixed Z. noltii and Z. marina meadows (SMZ) (Fig. 1), which may be explained by differences in sediments composition (Table 1).

Sampling programme and laboratory procedures

In 2000 and 2005, 9 sampling stations were selected as representative of the different habitats and zones previously characterized, coinciding with points sampled in the 1990 survey [START_REF] Laborda | Cartografía de los sedimentos superficiales de la Ría del Eo (Asturias). VII Simposio Ibérico de Estudios del Bentos Marino[END_REF][START_REF] Laborda | Informe ambiental de la Ría del Eo. Conserjería de Pesca y Ordenación del Territorio[END_REF] (Fig. 1). Samples were seasonally collected during low spring tides, in February (winter), May (spring), August (summer) and October (autumn).

Sampling procedures were nevertheless different from the 1990 campaigns. In fact, in 1990, following a common procedure at that time (e.g. Junoy and Vieitez, 1990), a minimum representative area of 0.16 m 2 and 0.40 m in depth was selected, and single samples were taken at each one of the sampling stations. On the contrary, in 2000 and 2005 samples consisted of eight replicates of 0.0123 m 2 and 0.2 m in depth, collected with a manual corer, which was considered enough to stabilize spatial variability [START_REF] Frontier | Stratégies d'échantillonage en Écologie[END_REF]. It was therefore necessary to legitimize the comparison of the different data sets. A test comparing two sets of samples collected at five stations using both sampling strategies, carried out in 2005, showed that differences regarding the number of species and their abundances were not significant (nMDS plot Figure 2; ANOSIM test: R = -0.125, p = 0.77).

In all cases, samples were sieved in situ through a 0.5 mm calibrated mesh size bag and preserved in 4% buffered formalin. In the laboratory, after sorting, invertebrates were identified to the species level, whenever possible, counted and stocked in 70% alcohol.

Organisms' biomass was determined as ash free dry weight (AFDW) after oven drying at 60 ºC for 72 h and combustion at 450 ºC for 8 h.

Together with biological sampling, salinity, temperature and dissolved oxygen were measured in situ from water retained in intertidal water pools. Additionally, sediment samples were collected for laboratory analysis of organic matter content and granulometry. The organic matter content was determined by loss on ignition (24 h at 450 °C), and the sediment particles size fractions were determined following a standard mechanic sieving procedure (Buchanan, 1984) and classified according to the Also, the area occupied by the different habitats at the different years was calculated using GPS field data and through the analysis of aerial and satellite photographs (with OziExplorer GPS mapping software and ArcGIS software).

Data analysis

Spatial and inter-annual variations of the Eo estuary benthic communities were assessed through the application of multivariate statistical techniques and ecological indicators.

Multivariate analysis of both physicochemical and biological data was performed by using the PRIMER software package [START_REF] Clarke | Change in Marine Communities. An Approach to Statistical Analysis and Interpretation[END_REF][START_REF] Clarke | PRIMER v5: User manual/tutorial[END_REF]. Regarding data on benthic macrofauna species abundance (ind. m -2 ), a similarity matrix was calculated using the Bray-Curtis coefficient after a four root transformation. Benthic assemblages were examined using a non-parametric multidimensional scaling (n-MDS) ordination. In addition, differences in species abundance among different habitats and years were analysed through a similarity test (ANOSIM), a non-parametric procedure analogous to ANOVA. The relative contribution of each species to average similarities between years was estimated applying SIMPER analysis. Finally, relationships between benthic distributional patterns and physicochemical variables were analysed through the BIOENV (biological environmental gradients) procedure.

Data were also analysed to estimate several ecological indicators: total abundance (ind.

m -2 ) (N), total species richness (S), Margalef index (d), Pielou evenness index (J ), Shannon-Wiener, log 2 (H ), W-statistic index [START_REF] Clarke | Comparisons of dominance curves[END_REF], based on the ABC method proposed by [START_REF] Warwick | A new method for detecting pollution effects on marine macrobenthic communities[END_REF], and the AZTI Marine Biotic Index (AMBI) [START_REF] Borja | A marine biotic index to establish the ecological quality of soft-bottom benthos within European estuarine and coastal environments[END_REF]. AMBI was computed using AMBI© software [START_REF] Borja | The application of a marine biotic index to different impact sources affecting soft-bottom benthic communities along European coasts[END_REF], which is freely available at http://www.azti.es.

The statistical significance of temporal differences observed between the chosen indices was further examined using one-way ANOVA. Data were transformed prior to analysis with a double square root transformation in order to improve homoscedasticity. Post hoc Tukey tests were employed to identify sampling dates differing significantly. These statistical analyses were performed with the SPSS software package.

RESULTS

A total of 53949 individuals belonging to 289 macrobenthic taxa were identified from 0.369, p<0.01). Differences between habitats were therefore obvious, confirming results

previously obtained [START_REF] Laborda | Cartografía de los sedimentos superficiales de la Ría del Eo (Asturias). VII Simposio Ibérico de Estudios del Bentos Marino[END_REF][START_REF] Laborda | Informe ambiental de la Ría del Eo. Conserjería de Pesca y Ordenación del Territorio[END_REF], although a few sampling stations might have exhibited intermediate characteristics.

In general, BIOENV results, for all samples, show weak correlations between fourthroot transformed sediment physicochemical parameters and macrobenthos abundances.

Correlations were nevertheless comparatively stronger with very fine sand and silt fractions (r = 0.353). Besides, correlations were also higher when different years were analysed separately. In 1990, macrobenthic abundances exhibited the highest correlation with the very fine sand fraction (r = 0.643), in 2000 with all the fine fractions (fine sand to silt) and organic matter content (r = 0.604), and in 2005 with the silt fraction (r = 0.669). This suggests that sediment variables, although important, might not be sufficient to explain inter-annual variability.

The nMDS plot comparing two sets of samples collected at five stations using both sampling strategies (Figure 2), shows that the results obtained from different sampling procedures were very similar. Once established that different sampling strategies would not introduce significant changes in results (ANOSIM; R=-0.125, p=0.77), samples from different years were compared.

Spatial changes due to differences between habitats, as well as seasonal variations, if all the samples were considered together, could make the analysis of inter-annual changes complex. Therefore, inter-annual changes taking place at different types of habitat were analysed separately for each season. Regarding the Z. marina -Z. noltii mixed meadow, stations MZ4 and BS1, which in 1990 presented transitional characteristics, were also included in the analysis. This allowed us to examine whether these stations maintained their intermediate characteristics, or became more similar to any of the nearby habitats.

The nMDS ordination analysis (Fig. 6) shows a comparable seasonal pattern of organisation between different years. In general, samples from 1990 and 2005 tended to form distinct groups, while samples from 2000 appear more dispersed in the projection.

Nevertheless, this pattern was not so clear regarding samples from the bare sand habitat, which was due to a lower similarity between different stations, even within the same year.

Taking into account the whole set of samples, ANOSIM did not reveal significant differences between years (R = 0.1, p< 0.01), but differences became clearly higher when each habitat was considered separately. In fact, significant differences in species composition between years were found for both the MZ and SMZ zones (R=0.435, p<0.001, zone MZ; R= 0.588, p<0.01, zone SMZ), which is particularly evident in the case of samples from 1990 and 2005 (R= 0.702, p<0.01, zone MZ; R= 0.861, p<0.01, zone SMZ). Regarding the bare sand habitat stations, ANOSIM did not show significant inter-annual differences (R=0.137, p=0.003), probably because of the great spatial variability in this zone.

Variation in species composition

Species contributing the most to the observed inter-annual differences in community structure were identified applying SIMPER to calculate their relative contributions to the overall average dissimilarity (Table 2).

In the mud flats covered by Z. noltii and in the Z. noltii -Z. marina mixed meadows, with the exception of B. reticulatum and L. lucinalis, all the species primarily responsible for inter-annual differences, increased in abundance in 2000, and again, even more, in 2005 (Table 2).

In the bare sand areas, the species that contributed the most to inter-annual variations (Table 2) The spatial patterns of distribution of individuals from shallow-water benthic species are reported to be largely controlled by abiotic factors at the broader scale and by biotic factors at finer scales (see review by [START_REF] Snelgrove | Getting to the bottom of marine biodiversity: sedimentary habitats[END_REF]. In particular, the distribution and abundance of soft-sediment benthos in natural habitats have been related to the sediment particle size and structure [START_REF] Seiderer | Analysis of the relationship between sediment composition and benthic community structure in coastal deposits: Implications for marine aggregate dredging[END_REF][START_REF] Currie | Macrobenthic community responses to long-term environmental change in an east Australian sub-tropical estuary[END_REF], and also to the presence and coverage of seagrasses [START_REF] Boström | Community structure and spatial variation of benthic invertebrates associated with Zostera marina (L.) beds in the northern Baltic Sea[END_REF][START_REF] Blanchet | Heterogeneity of macrozoobenthic assemblages within a Zostera noltii bed: diversity, abundance, biomass and structuring factors[END_REF]).

Our results for the Eo estuary are consistent with these findings, elucidating how important sediment grain size and seagrasses beds are in controlling the number of species and the abundance of benthic macro-invertebrates. In fact, as observed in other comparative studies [START_REF] Edgar | Comparisons of species richness, size-structure and production of benthos in vegetated and unvegetated habitats in Western Port, Victoria[END_REF]Bröstrom and Bonsdorff, 1997), species richness and abundance were found to be higher in seagrass beds than on the bare sand areas. On the other hand, the Z. marina -Z. noltii mixed meadow (SMZ zone) also presented higher biological diversity than mud flats covered by Z. noltii (MZ zone). This is probably related to differences in tidal level and variations in the sediment composition, and also due to differences regarding the structural characteristics of the macrophytes' vegetation [START_REF] Blanchet | Heterogeneity of macrozoobenthic assemblages within a Zostera noltii bed: diversity, abundance, biomass and structuring factors[END_REF]. In fact, the SMZ zone is located at a lower tidal level as compared to the MZ zone, and consequently, due to shorter emersion periods, is less constrained in terms of abiotic stress.

Macrobenthic diversity appeared to be negatively correlated with the proportion of fine particles, which was particularly obvious at the Linera inlet. On the other hand, in the sand banks habitat, species diversity decreased from more sheltered (stations BS1, BS2)

to more exposed areas (station BS3), demonstrating how diversity can be markedly depressed by sediment instability [START_REF] Warwick | Analysis of macrobenthic and meiobenthic community structure in relation to pollution and disturbance in Hamilton Harbour, Bermuda[END_REF][START_REF] Cooper | Cumulative impacts of aggregate extraction on seabed macro-invertebrate communities in an area off the east coast of the United Kingdom[END_REF]. Station BS1, located closer to Z. marina meadows presented the highest biological diversity of all stations in the bare sand zone. In fact, the presence of Z. marina determines higher sediment stability [START_REF] Fonseca | A comparison of canopy friction and sediment movements between four species of seagrass with reference to their ecology and restoration[END_REF], and provides important food and refuge supply to neighbour populations [START_REF] Edgar | Experimental analysis of structural versus trophic importance of seagrass beds. I. Effects on macrofaunal and meiofaunal invertebrates[END_REF].

The different habitats identified and characterized in 1990 remained present in the estuary over 2000 and 2005. Nevertheless, there was a substantial reduction in the area of the Z. noltii -Z. marina mixed meadow during this fifteen years period (from 40 ha in 1990 to 8 ha in 2005), which was associated with the installation of the oyster culture racks in this zone. In fact, just to prepare the racks' installation, seagrasses were physically removed from the bottom. Additionally, green macroalgal blooms (Ulva spp.) have been recorded since 2000, mainly on mud flats covered by Z. noltii, which may constitute a symptom of eutrophication in estuarine environments [START_REF] Marques | Analysis of the properties of exergy and biodiversity along an estuarine gradient of eutrophication[END_REF][START_REF] Belan | Marine environmental quality assessment using polychaete taxocene characteristics in Vancouver Harbour[END_REF].

From 1990 to 2005, with the exception of station MZ3, the macrobenthic communities showed an increase in number of species and abundance in the areas near the oyster culture racks (zones SMZ and MZ). This was essentially due to the presence of opportunistic, pioneer or tolerant species such as Oligochaeta spp., C. capitata, N. latericius, S. benedicti, P. antennata or H. ulvae, as a function of organic matter enrichment in the system [START_REF] Pearson | Macrobenthic succession in relation to organic enrichment and pollution of the marine environment[END_REF][START_REF] Rhoads | Characterization of organism-sediment relations using sediment profile imaging -an efficient Method of Remote Ecological Monitoring of the Seafloor (Remots TM System)[END_REF].

Moreover, the whole SMZ zone, as well as stations MZ1 and MZ4, exhibited the highest values of biological diversity in 2005. These results suggest that there has been an improvement of macrobenthic communities in these zones in 2005 with respect the reference situation. This type of response to perturbation has been observed in other case studies (e. g. [START_REF] Marques | Impact of eutrophication and river management within a framework of ecosystem theories[END_REF][START_REF] Patrício | Ascendency as an ecological indicator: a case study of estuarine pulse eutrophication[END_REF]), and can be interpreted at the light of the Intermediate Disturbance Hypothesis (IDH) [START_REF] Connel | Diversity in tropical rain forests and coral reefs[END_REF].

The W-statistic index also suggests an improvement of macrobenthic communities in this area in 2000 in comparison to 1990, which can be mostly explained by the dominance of B. reticulatum in 1990, and its absence in 2000. Apparently, the dominance of certain species in the community, small in size although characteristic of non-polluted habitats, may lead to erroneous evaluations when applying this indicator [START_REF] Salas | Evaluation of the applicability of a marine biotic index to characterize the status of estuarine ecosystems: the case of Mondego estuary (Portugal)[END_REF].

On the contrary, AMBI values revealed some degradation of the macrobenthic communities at the whole SMZ and MZ zones in 2000 and 2005 when compared to 1990, which suggests that this index is more sensitive to intermediate perturbations than diversity measures. This is due to AMBI architecture, since it uses species as indicators of different degrees of organic enrichment [START_REF] Borja | A marine biotic index to establish the ecological quality of soft-bottom benthos within European estuarine and coastal environments[END_REF]. In this case, the presence of pollution indicator species, classified as tolerant to organic enrichment (group III) or opportunistic (group IV and V) [START_REF] Borja | The application of a marine biotic index to different impact sources affecting soft-bottom benthic communities along European coasts[END_REF], was able to capture the system deterioration, although there was an increase in biological diversity and abundance.

Organic enrichment did not notably affect the macrobenthic community on the sand banks area, where interannual changes were minor and species composition remained very similar throughout time. In fact, this zone (mainly stations BS2 and BS3) can be considered as naturally physically disturbed because of sedimentary instability caused by tidal currents and wave action. In this type of environment, soft sediment faunal assemblages typically occupy an early succession stage [START_REF] Newell | The impact of dredging works in coastal waters: a review of the sensitivity to disturbance and subsequent recovery of biological resources on the sea bed[END_REF][START_REF] Bolam | Minimizing impacts of maintenance dredged material disposal in the coastal environment: a habitat approach[END_REF], with dominance of species displaying 'r selected' characteristics.

Nevertheless, a certain reduction in biological diversity and biomass, as well as a slight increase in AMBI values from 1990 to 2000, and again to 2005, suggest an increase in the disturbance level. It may be related to an increase in physical disturbance, as the Eo estuary, like other systems from the northern coast of Spain, has been going through a gradual process of in-filling by marine and estuarine deposits (Currás and Mora, 1991).

Furthermore, sediment deposition increased after the construction of a new harbour (in 1994) in the western part of the Eo estuary, and an increase of 17 ha in the area of sand banks took place from 1990 to 2005. Such pressures may partially explain the observed inter-annual variations. In response, there was an increase in abundance of small benthic opportunistic species, such as S. squamata, S. filicornis and A. latreillei, and a slight reduction in biomass and size of individuals in comparison to 1990.

4.2 What has been the impact of oyster culture farming?

Since 1990 the anthropogenic pressure due to oyster culturing in the area has been increasing, and this activity may have become the major driving force behind the organic enrichment observed in the system. In fact, the accumulation of organic matter by bivalve biodeposition is usually the major impact of shellfish farming [START_REF] Kaiser | Environmental impacts of bivalve mariculture[END_REF].

Most of the studies about the effects of bivalve farming were based on post-impact assessment concerning differences in environmental variables from the farm and from adjacent control/reference sites in the same temporal scale [START_REF] Grant | A multidisciplinary approach to evaluating impacts of shellfish aquaculture on benthic communities[END_REF][START_REF] Crawford | Qualitative risk assessment of the effects of shellfish farming on the environment in Tasmania, Australia[END_REF][START_REF] Forrest | Benthic impacts of intertidal oyster culture, with consideration of taxonomic sufficiency[END_REF]. Some studies revealed significant differences between reference zones and cultivation areas, and eutrophication symptoms were detected in macrobenthic communities beneath farms [START_REF] Stenton-Dozey | Impact of mussel culture on macrobenthic community structure in Saldanha Bay, South Africa[END_REF][START_REF] Mirto | Microbial and Meifaunal response to intensive mussel-farm bideposition in coastal sediments of the Western Mediterranean[END_REF]Harstein and Rowden, 2004). However, in other studies, observed effects of shellfish farming were minimal [START_REF] Grant | A multidisciplinary approach to evaluating impacts of shellfish aquaculture on benthic communities[END_REF][START_REF] Crawford | Qualitative risk assessment of the effects of shellfish farming on the environment in Tasmania, Australia[END_REF].

In the present study all the sampling stations were located outside of the farms, and the control was temporal, not spatial. We compared the benthic conditions at the study site prior to the start of intensive oyster cultivation (1990), during the installation of the oyster racks (2000), and after some years of shellfish farming (2005). Our results show that a) changes in the macrobenthic communities at the Linera inlet may be explained by organic enrichment, and b) such organic enrichment affects not only the area beneath a farm, but also sites located at a certain distance from it.

It has been demonstrated that differences in the extent and magnitude of environmental impacts are likely to depend on farm type (e.g. farm size, culture method, stocking density, and age of the operation) and on the site conditions (e.g. bathymetry and hydrodynamic conditions (Harstein and Rowden, 2004). Although the oysters stocking densities in the Eo estuary is relatively low compared to other bivalve culture areas (e.g. [START_REF] Grant | A multidisciplinary approach to evaluating impacts of shellfish aquaculture on benthic communities[END_REF][START_REF] Crawford | Qualitative risk assessment of the effects of shellfish farming on the environment in Tasmania, Australia[END_REF], an important increase in organic matter has been detected in the system. Some studies illustrate that local current patterns are the most important factors influencing the effect of bivalve cultivation on the seabed. Due to dispersion, effects of shellfish biodeposits from farms may be minor in sites that experience high hydrodynamic energy [START_REF] Hartstein | Effect of biodeposits from mussel culture on macroinvertebrate assemblages at sites of different hydrodynamic regime[END_REF]). In the Eo estuary case, shellfish farms are located at the entrance of a shallow sheltered inlet. Oyster biodeposits originating from the racks, combined with a naturally reduced hydrodynamics in that part of the system, can clearly explain the increase in organic matter [START_REF] Castel | Influence of seagrass beds and oyster parks on the abundance and biomass patterns of meio-and macrobenthos in tidal flats[END_REF]. On the other hand, the presence of seagrass beds in this area also favours biodeposition by reducing currents velocity [START_REF] Fonseca | A comparison of canopy friction and sediment movements between four species of seagrass with reference to their ecology and restoration[END_REF].

Besides, Zostera debris also tends to accumulate since they can hardly be transported out of the system.

The supply of organic matter provided by oyster farms may be responsible for the increase in sedimentary organic matter, which has been observed especially in stations closer to the racks (essentially the SMZ zone and MZ4 station). The diversity and richness indices suggest that the increase in organic matter has caused an improvement in macrobenthic communities from the Linera inlet in 2005 with respect the reference situation (1990). Thus, in stations with higher increases in organic matter (essentially the SMZ zone) values of the diversity and richness indices have raised significantly.

This effect of bivalve culture has been pointed out in other studies (e.g. [START_REF] Grant | A multidisciplinary approach to evaluating impacts of shellfish aquaculture on benthic communities[END_REF].

However, the macrobenthic community changes at MZ1, MZ2 and MZ3 stations may be examined further. At MZ1 the diversity indices suggest also an improvement in benthic community, but this station has not been affected by organic enrichment. On the contrary, the sedimentary organic matter content has diminuished since 1990. On the other hand, at stations MZ2 and MZ3, although the organic matter increase was lower than in SMZ and MZ4 station, the indices show degradation in macrobenthic communities.

These results show that trends in macrobenthic community changes may depend on its initial conditions. Hence, the more diverse macrobenthic communities inhabiting in Z.

noltii and Z. marina mixed meadows and transitional station MZ4 presented capability to absorb increases in organic matter, and its response was consistent with the IDH, which is thought to be realised in systems with a certain capacity to absorb changing environmental conditions over a larger time scale [START_REF] Patrício | Ascendency as an ecological indicator: a case study of estuarine pulse eutrophication[END_REF]. Instead, the MZ2 and MZ3 stations, from mud flat covered by Z. noltii, responded negatively to the organic enrichment, even though the increase in organic matter observed in this area was minor. Thus, the benthic communities' response to perturbations acting in a system depends in a large extent of its intrinsic buffer capacity.

Changes in macrobenthic communities on different habitats observed since 1990 in the Linera inlet could be interpreted following the approach suggested by [START_REF] Marques | Impact of eutrophication and river management within a framework of ecosystem theories[END_REF] for long time series studies. We assumed the situation in 1990 as an initial status, a kind of climax society with minimal anthropogenic pressure. In 2000, all the indicators suggest an imbalanced situation of the macrobenthic communities, with changes exceeding the expected natural variation. Such a situation seems to reflect a severe perturbation, resulting from an extensive development of shellfish farming, with the installation of many new oyster racks. Physical disturbance was probably the main factor affecting the communities in 2000. In 2005, the macrobenthic community appears more structured again, similar to 1990, which suggests a recovery, probably as a function of the cessation of new racks installation. Nevertheless, the community did not return to a state similar to that observed in 1990. This suggests that a) the system might have reached a new state of dynamic equilibrium corresponding to a new specific set of environmental conditions, and b) the major driving force behind this new state has been the supply of organic matter provided by bivalve farms.

CONCLUSIONS AND FUTURE CONCERNS

Although the anthropogenic pressure (e.g. oyster racks' implementation, harbour construction) were present throughout the study period in the Eo estuary it was between 1990 and 2000 that the macrobenthic community revealed its worst status.

Eutrophication symptoms were visible in the system until 2005, (e.g. increase in biomass of green macroalgae, reduction of Zostera coverage, increase in sedimentary organic matter) but the macrobenthic community showed a higher disorganisation in 2000 followed by an apparent recovery to 2005.

Results suggested that the additional stimulation of biomass and diversity that may happen as a benthic community responds to anthropogenic induced perturbations (depending on systems' intrinsic buffer capacity), usually referred to as Intermediate Disturbance Hypothesis (IDH) [START_REF] Connel | Diversity in tropical rain forests and coral reefs[END_REF], could describe the developmental pattern found in the system (especially at the Z. noltii -Z. marina mixed meadows). All the original habitats were still present in 2000 and 2005, but the latest improvements in the benthic quality followed a different direction, rather than a simple return to the original macrobenthic community.

The use of a set of selected indices, rather than a single one, provided an effective approach to analysing the changes. Although indices like the 'Number of Species', 'Margalef' or the 'Shannon-Wiener' could indicate a status improvement, the values of AMBI revealed a general deterioration in the system in 2000 and 2005 relatively to 1990, confirming the different direction that the macrobenthic community followed to find a new equilibrium point.

Therefore, to keep shellfish farming at a sustainable level without undesirable impacts, the disturbance intensity must be kept below the system carrying capacity (see for instance [START_REF] Mckindsey | Review of recent carrying capacity models for bivalve culture and recommendations for research and management[END_REF]. This precaution will allow natural communities to cope with pressures, avoiding ecological quality deterioration, which is a key issue in terms of sustainable ecosystem management.

Figures caption

Fig.

  Subsequently, in 2000 and 2005, only the most representative sampling stations utilised in 1990 sampling were re-sampled. The results from the three sampling campaigns were compared with three main objectives: i) To determine the extent to which the Eo estuary macrobenthic communities have changed over the last fifteen years; ii) To characterize such changes relating it to the system's physical conditions; iii) To assess to what extent these changes occurred as a response to increasing human pressures, with emphasis on oyster farming.

  250mm>Ø>0.125mm), very fine sand (0.125mm>Ø>0.062mm), and silt (Ø <0.062 mm).

  the samples collected in1990, 2000, and 2005. Only 84 species (29%) were found in all the campaigns, while 44 (15%) were collected exclusively in 1990, 27 (9%) in 2000, and 52 (18%) in 2005.3.1 Habitat parametersThe estuarine intertidal area increased over the study period(from 193 ha in 1990 to 212 ha in 2005), especially the bare sand areas (from 53 ha in 1990 to 70 ha in 2005).On the other hand, it was observed an important reduction in the extension of the Zostera seagrass beds, namely of the mixed Z. noltii -Z. marina meadow (from 40 ha in 1990 to 8 ha in 2005). This is actually the area where most of the culture racks have been installed, and before putting them in place seagrasses were directly removed from the site, leaving the sediments beneath the farms without vegetation.Results for sediment particle size are showed in Figure3. Only the relative percentages of the main fractions have been presented, which were medium, fine and very fine sand, and silt. With the exception of MZ1, a gradual increase in the proportion of fine sediments was observed in all samples from 1990 to 2005. Similarly, the percentage of organic matter in the sediment increased in all zones, with the exception of station MZ1 (Figure4). The strongest increase was found in the stations of Z. noltii-Z. marina mixed meadows.3.2 Spatial structure and temporal changes in macrobenthic communitiesStations located at bare sandy bottoms, mudflats covered by Zostera noltii, and Zostera noltii-Zostera marina mixed meadows on sandy and sandy-muddy sediments are clearly differentiated in the nMDS analysis (Fig.5), ranging along a gradient from the inner part of the Linera inlet to the sand banks. ANOSIM shows that there are significant differences in the macrofauna community composition among different habitats (R = 0.547, p< 0.01), especially between mudflats with Z. noltii (MZ) and bare sand areas (BS) (R= 0.717, p< 0.01). The Z. noltii -Z. marina mixed meadows (SMZ) appear to have intermediate characteristics (R= 0.391, p< 0.01), as compared to MZ and BS (R=

  , exhibited strong seasonal variability, namely in 2005. As a trend, the average abundance of polychaetes increased from 1990 to 2000, decreasing in 2005. On the other hand, bivalves suffered a decrease both in number of species and average abundance from 1990 to 2000 and although densities recovered after 2000, it was observed a reduction in the size of individuals. Finally, the crustaceans' average abundance increased from 1990 to 2000, suffering a reduction in 2005. Indeed, most of the seasonal and inter-annual variations observed in this habitat were due to isolated species.3.4 Ecological indicatorsValues of the ecological indicators seasonally estimated for the different habitats during the study period are shown in Fig.7. With the exception of AMBI, all indices showed a strong seasonal variability which makes it difficult to interpret inter-annual variations (Fig.7). Actually, these results are consistent with the findings ofReiss and Kröncke (2005), who demonstrated that univariate indices show higher seasonal variability than multimetric indices such as AMBI. In general, if an index shows higher/lower values than in a previous campaign, at least in three of the four seasons, we assumed that it increased/decreased in that station during that year. Independently of the year, values of the chosen indices showed a strong variability between the different zones. The SMZ zone presented the highest values in terms of species richness and other diversity measures, while the BS zone showed the highest values of W-statistic and Pielou evenness. On the other hand, the maximum values of AMBI were always found in the MZ zone (Table3).Values of the different indices are consistent with results regarding the macrobenthic community structure, illustrating the intermediate characteristics of transitional stations between the different zones. In fact, in the MZ zone, station MZ4 showed the highest values regarding species richness and diversity indices, analogous to those obtained in the SMZ zone, and in the BS zone the highest diversity measures were found at the BS1 station (Fig.7).In 1990, despite the fact that this year was assumed as the reference situation, the mudflat stations showed relatively high AMBI values, especially at the MZ1 station(Fig 4). Nevertheless, at all stations and in every season, AMBI values were higher in 2000, and even more in 2005. Therefore, there was an apparent degradation of the macrobenthic community at the different habitats from the 1990s' until the present. This degradation is also illustrated by the diversity indices and the W-statistic index both at the sandy stations and at stations with larger cover of Z. noltii (MZ2 and MZ3).Nevertheless, with regard to the SMZ zone and MZ1 and MZ4 stations, we found some inconsistencies between the assessments provided by AMBI, the W-statistic, and the other indices. In fact, according to the diversity indices the ecological quality of these stations appears to have improved in 2005 as compared to the reference situation (table3), while AMBI values suggest the contrary. Actually, the species number and abundance increased at these zones in 2005, especially at SMZ zone, but most of these species were opportunists or indicators of organic enrichment. Consequently, despite the increase in diversity values, AMBI showed a deterioration of the system throughout time. On the other hand, at the SMZ, the W-statistic values were higher in 2000 than in 1990 and 2005, indicating a possible disturbance situation in those years. The Pielou index provides a similar picture, showing an improvement at the SMZ zone between 1990 and 2000. Nevertheless, the variation in the values of these two indices was not statistically significant (see Table 4

  ).One way ANOVA, taking into account the values of the different indices, showed that the mean values were significantly different between sampling years, with the exception of the W-statistic and Pielou evenness (Table4). The Tukey's test showed a statistically significant decrease (p<0.05) in the values of the Margalef and Shannon-Wiener between 1990 and 2000, and a concomitant increase in the values of AMBI and abundance (p<0.001). On the other hand, in 2005, abundance increased significantly (p<0.005) as compared to both 1990 (the reference situation) and 2000, while the number of species increased significantly only in relation to 2000 (p<0.005), and AMBI showed a significant (p<0.001) increase in comparison to the reference situation.In general, taking into account the mean values for each habitat values, all the selected indices suggest a worsening in the system between 1990 and 2000. Afterwards, the diversity indices express a slight recovery of the macrobenthic community in 2005 in comparison to 2000, while AMBI, on the contrary, suggests that degradation continued.Similarly, regarding the SMZ and MZ zones, diversity indices showed an improvement on the macrobenthic communities in 2005 as compared to the reference situation, while AMBI, the W-statistic and Pielou indices showed aggravation. Finally, all the indices showed degradation of the macrobenthic communities in the bare sand zone from have macrobenthic communities been changing in the Eo estuary?

Fig. 3 .

 3 Fig. 1. Terminal part of Eo estuary, showing sampling areas and location of oyster farms. Fig. 2. Non-metric MDS plot of faunal density data comparing the two sampling procedures. Grey triangles represent corer samples and white circles represent spade samples. Fig. 3. Relative percentages of medium sand (white), fine sand (clear grey), very fine sand (dark grey) and silt (black). From left to right 1990, 2000 and 2005.

Fig 4 .

 4 Fig 4. Percentage of sedimentary organic matter content (means+s.d.). Grey bars represent 1990 samples, clear bars represent 2000 samples and black bars represent 2005 samples.Fig. 5. Non-metric MDS plots of benthic community structure at 9 stations sampled in Eo estuary in Febrary, May, August, October of 1990, 2000 and 2005. Grey triangles represent mud flat covered by Z. noltii samples (MZ area), black triangles represent Z. marina and Z. noltii mixed meadow samples (SMZ area) and empty squares represent bare sand samples (BS area).

Fig. 6 .

 6 Fig. 6. Non-metric MDS ordination plots for the macrobenthos assemblages arranged by habitat (columns) and season (rows). Grey triangles represent 1990 samples, empty inverse triangles represent 2000 samples and black squares represent 2005 samples.Fig. 7. Temporal and spatial variation of the richness (A), Margalef index (B), Shannon-Wiener index (C), Pielou index (D), density (ind. m-2) (E), AMBI index (F) and Wstatistic index (G) at the 9 stations in the Eo estuary (1990, 2000 and 2005 campaigns). Grey bars represent mud flat covered by Z. noltii samples (from left to right MZ1, MZ2, MZ3 and MZ4), black bars represent Z. marina and Z. noltii mixed meadow (from left to right SMZ1 and SMZ2) samples and clear bars represent bare sand samples (from left to right BS1, BS2 and BS3).

  Fig.1 748

Table 1 .

 1 Description of the Reference Situation.

	762
	HABITAT	CHARACTERISTICS		ZONES
	Bare Sand	LOCATION (Fig.1)		BS1
	(BS)	Sand banks and beaches	The edge of a channel bottom covered by
					Z.marina. Sheltered area.
		SEDIMENT			P. foetida; N. cirrosa; A. tenuis; Thracia
					papyracea
		Sand
		MACROBENTHIC COMMUNITY	BS2
		Cerastoderma edule; Nephtys	Relatively sheltered area
		cirrosa; Scolelepis squamata;	A. tenuis; C. edule
		Angulus tenuis; Phylo foetida;
		Urothoe poseidonis		BS3
					Exposed area
					S. squamata
	Mudflats	LOCATION (Fig.1)		MZ1
	covered by	Inner part of the Linera inlet	Isolated patches of Z. noltii.
	Zostera noltii (MZ)				H. ulvae
		SEDIMENT			MZ2
		Mud			Continuous Z. noltii bed.
					H. ulvae; B. reticulatum; L. lucinalis; N.
					hombergii
		MACROBENTHIC COMMUNITY	MZ3
		Hydrobia	ulvae;	Bittium	Continuous Z. noltii bed; higher shoot
		reticulatum			density than in MZ2.
		Loripes	lucinalis;	Nephtys	B. reticulatum; H. ulvae; L. lucinalis; N.
		hombergii			hombergii

Table 3 .

 3 Values of biotic indices estimated at the nine stations of theEo estuary;[START_REF] Clarke | Comparisons of dominance curves[END_REF], 2000 and 2005 sampling campaings (mean±s.d.)

	SMZ2	BS1	BS2	BS3
	0.76 ± 0.48	1.20 ± 0.29	1.22 ± 0.35	1.40 ± 0.91
	1.65 ± 0.58	0.79 ± 0.28	2.51 ± 0.35	2.50 ± 0.46
	2.60 ± 0.34	1.80 ± 0.81	2.15 ± 0.70	1.26 ± 0.28
	0.11 ± 0.08	0.39 ± 0.09	0.48 ± 0.23	0.36 ± 0.06
	0.31 ± 0.09	0.52 ± 0.13	0.29 ± 0.18	0.25 ± 0.06
	0.16 ± 0.07	0.56 ± 0.08	0.41 ± 0.29	0.26 ± 0.03
	36.00 ± 10.86	22.75 ± 8.18	16.75 ± 6.50	30.75 ± 13.50
	24.00 ± 15.60	9.00 ± 6.63	13.75 ± 8.26	22.00 ± 6.48
	59.75 ± 10.40	16.00 ± 15.60	9.00 ± 1.63	30.50 ± 5.51
	1477.35 ± 471.34	160.94 ± 115.25	115.63 ± 71.53	229.82 ± 128.71
	950.00 ± 979.29	162.50 ± 145.86	320.31 ± 154.90	821.88 ± 388.83
	11865.97 ± 5024.10	1274.13 ± 2294.53	388.60 ± 419.69	1964.86 ± 1525.06
	4.80 ± 1.37	4.37 ± 1.11	3.34 ± 1.09	5.45 ± 1.93
	3.36 ± 1.75	1.55 ± 1.00	2.24 ± 1.34	3.14 ± 0.74
	6.29 ± 0.84	2.26 ± 1.55	1.47 ± 0.30	4.01 ± 0.48
	0.60 ± 0.15	0.85 ± 0.04	0.91 ± 0.06	0.84 ± 0.04
	0.76 ± 0.05	0.86 ± 0.10	0.80 ± 0.12	0.73 ± 0.08
	0.71 ± 0.08	0.88 ± 0.07	0.77 ± 0.28	0.77 ± 0.05
	3.09 ± 0.92	3.77 ± 0.55	3.58 ± 0.63	4.02 ± 0.51
	3.35 ± 0.60	2.42 ± 0.76	2.90 ± 1.17	3.21 ± 0.37
	4.20 ± 0.45	3.04 ± 0.84	2.41 ± 0.85	3.80 ± 0.35
	mean±s.d.)

Sanders, H. L., Mangelsdorf, P. C. and Hampson, G. R., 1965. Salinity and faunal distribution in the Pocasset River, Massachusetts. Limnology and Oceanography, 10,

ACKNOWLEDGEMENTS

Laura de Paz was supported by a doctoral grant of the Universidad de León. Parts of this work were funded through the Universidad de León project (ULE-AG 85) and the Consejería de Agricultura y Pesca of Principado de Asturias contract. We are grateful to E. Cimas, M., Alonso, M. A., Mayoral and especially J. C. González Ruiz for their help in collecting samples and benthic sorting and to the staff of the Centro de Recursos Pesqueros de Castropol and R. de Paz for logistic support during the field work. The work of the Portuguese team was partially supported in the scope of the project RECONNECT ((PTDC/MAR/64627/2006). We thank the two anonymous reviewers who made a number of valuable comments to improve this manuscript. REFERENCES