Subtidal benthic megafauna in a productive and highly urbanised semi-enclosed bay (Ría de Vigo, NW Iberian Peninsula)

Author’s Accepted Manuscript Subtidal benthic megafauna in a productive and highly urbanized semi-enclosed bay (Ría DE Vigo, NW Iberian peninsula) Fernando Aneiros, Marcos Rubal, Jesús S. Troncoso, Rafael Bañón www.elsevier.com/locate/csr

PII: DOI: Reference:

S0278-4343(15)30069-8 http://dx.doi.org/10.1016/j.csr.2015.09.018 CSR3287

To appear in: Continental Shelf Research Received date: 12 April 2015 Revised date: 11 September 2015 Accepted date: 21 September 2015 Cite this article as: Fernando Aneiros, Marcos Rubal, Jesús S. Troncoso and Rafael Bañón, Subtidal benthic megafauna in a productive and highly urbanized semi-enclosed bay (Ría DE Vigo, NW Iberian peninsula), Continental Shelf Research, http://dx.doi.org/10.1016/j.csr.2015.09.018 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting galley proof before it is published in its final citable form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

Subtidal benthic megafauna in a productive and highly urbanized semi-enclosed bay (Ría de Vigo, NW Iberian Peninsula) Fernando Aneirosa,b, Marcos Rubala,b,c,d, Jesús S. Troncosoa,b, Rafael Bañóne a

Departamento de Ecoloxía e Bioloxía Animal, Facultade de Ciencias do Mar,

Universidade de Vigo. Campus Universitario Lagoas-Marcosende, 36310, Vigo, Pontevedra, Spain. b

ECIMAT, Estación de Ciencias Mariñas de Toralla, Universidade de Vigo. Illa de

Toralla, 36331, Vigo, Pontevedra, Spain. c

CIIMAR/CIMAR, Centro Interdisciplinar de Investigação Marinha e Ambiental,

University of Porto. Rua dos Bragas, 289, 4050-123, Porto, Portugal. d

Department of Biology, Faculty of Sciences, University of Porto. Via Panoramica, 36,

4150-564, Porto, Portugal e

Servizo de Planificación, Dirección Xeral de Desenvolvemento Pesqueiro, Consellería

do Mar e Medio Rural, Xunta de Galicia. Santiago de Compostela, A Coruña, Spain. E-mail Addresses: Fernando Aneiros (Corresponding author): [email protected]. Marcos Rubal: [email protected]. Jesús S. Troncoso: [email protected]. Rafael Bañón: [email protected].

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Abstract The Ría de Vigo is a semi-enclosed bay with high primary productivity due to the influence of coastal upwelling-downwelling dynamics. The area is heavily populated and affected by numerous human activities, which lead to sediment modification. Epibenthic megafauna from the non-estuarine zones of this bay has been studied in order to describe its spatial distribution, testing possible differences between inner and outer areas. With that purpose, 75 sites have been sampled by means of a towing dredge. Megafauna was identified to the lowest taxonomic level possible, and each taxon counted and weighted. 113 different taxa were identified and a high spatial heterogeneity was observed in terms of abundance, biomass, taxa richness, diversity and evenness. Suspension-feeding molluscs dominated the innermost part of the studied area, and were substituted by echinoderms towards the external zones; this spatial pattern was also reflected in the results of multivariate analyses. These shifts in taxonomic and trophic guild composition of the assemblages have been tentatively related to differences in pollution levels and primary productivity along the main axis of the bay. Keywords: East Atlantic, Ría de Vigo, Epibenthic megafauna, Echinodermata, Mollusca, Spatial distribution

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1. Introduction The Galician Rías Baixas are a set of semi-enclosed elongated bays of tectonic origin (Barton et al., 2015; Nombela et al., 1987; von Richthofen, 1886) located in the NW coast of the Iberian Peninsula. Their bottoms include both hard and soft substrata, which host a great faunal biodiversity (e.g. Cacabelos et al., 2008; López-Jamar, 1978a; López-Jamar and Cal, 1990; Lourido et al., 2010; Moreira et al., 2010). The Rías Baixas are in the northern limit of the NW Africa upwelling system and subjected to the influence of wind-driven coastal upwelling (Fraga, 1981), which is considered to dominate their hydrodynamic regime (Álvarez-Salgado et al., 1996). The Rías Baixas show differential topographic features when compared to other bays subjected to upwelling (Barton et al., 2015). In general terms, upwelling conditions are dominant from mid spring to early autumn, while downwelling conditions prevail during the rest of the year (Varela et al., 2005). During upwelling events, the nutrientrich Eastern North Atlantic Central Water enters the Rías (Alvarez-Salgado et al., 1996; Fraga, 1981); these events enhance phytoplankton primary production, especially during spring, thus the Rías become very productive ecosystems (Figueiras et al., 2002). Due to their high productivity and their role as natural harbours, the Rías Baixas are heavily populated and affected by numerous human activities, including several kinds of fishing and aquaculture (especially mussel culture in rafts) (Figueiras et al., 2002). This leads to noticeable changes in the granulometric composition and organic content of the sediment (Alejo and Vilas, 1987; Nombela et al., 1987; Prego et al., 2008; Rubio et al., 2000; Vilas et al., 1995), which may therefore have an effect on the composition and structure of benthic faunal assemblages (e.g. Abella et al., 1996; López-Jamar, 1978b; Mora et al., 1989; Ysebaert et al., 2009).

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Recent research has revealed important differences between the internal and external zones of the Rías regarding circulation patterns, which are mainly controlled by the alternation between upwelling and downwelling conditions (Barton et al., 2015). According to Barton et al. (2015), water circulation in the internal zone is controlled by wind forcing and density-driven exchange, while the external zone is subjected to through-flow of shelf currents with much higher water transport rates. Higher organic matter content and the prevalence of the finest grain size fractions (silt and clay) are characteristic of the sediment in the innermost part of the Rías due to hydrodynamic factors (e.g. Vilas et al., 2005). Other reasons contribute to the organic enrichment of these areas, namely suspended mussel culture (Alonso-Pérez et al., 2010), organic pollution due to industrial activities (Mora et al., 1989) and higher phytoplankton productivity (Margalef, 1958). Higher levels of heavy metals in the sediments are also found in these areas, partly because the characteristics of the sediment favour the accumulation of these elements, but also because of anthropogenic pollution and freshwater inputs (Prego et al., 2008; Rubio et al., 2000). The Ría de Vigo, the southernmost of the Rías Baixas, is 35 km long with maximum depths of 50 m and a total surface of 183 km2; its main axis is oriented towards the SW (Nombela et al., 1987). Its coastline is densely populated and therefore highly urbanized, and maritime traffic is very intense due to the harbour of the city of Vigo. The Ría de Vigo can be divided into three main zones: 1) Ensenada de San Simón, which is located in the innermost part of the Ría, is an estuarine area and receives the main freshwater inputs; its outer boundary is the Rande Strait (Nombela et al., 1987); 2) the Main Channel, which is funnel-shaped, about 14 km in length and subjected to anthropogenic impacts derived from the largest city in the Ría (Vigo) and intense mussel raft culture (Prego et al., 2008; Rubio et al., 2000; Ysebaert et al., 2009); and 3) 4

the External Zone, which is about 10 km wide and sheltered from the open sea by the Cíes Islands. Several studies have addressed different topics related to benthic fauna in Ría de Vigo (e.g. Abella et al., 1996; Ardré et al., 1958; Navaz, 1942), but few of them have dealt with the composition and structure of subtidal soft-substrate assemblages (Ensenada de San Simón: Cacabelos et al., 2008; rest of the Ria: López-Jamar and Cal, 1990; Margalef, 1958). To date, studies on epibenthic megafaunal assemblages were focused on commercial or very relevant species, with poor taxonomic resolution for the rest of the fauna, and they did not address patterns of fauna distribution (see Guerra et al., 1984; Guerra and Pérez-Gándaras, 1987). The importance of studying epibenthic megafauna relies partly on the presence of numerous species of commercial interest (Guerra et al., 1984; Guerra and PérezGándaras, 1987), as well as harmful invasive species (Bañón et al., 2008; Blanchard, 1997). These assemblages are also relevant because they constitute a food resource for other species supporting commercial fisheries (de Leo et al., 2010), their bioturbatory activities influence nutrient fluxes in the water sediment interface (e.g. Sandnes et al., 2000; Smallwood et al., 1999; Smith et al., 1997; Turnewitsch et al., 2000; Vardaro et al., 2009), and even their usefulness as proxies for biogeochemical processes has been suggested (Smith et al., 1997). The relevance of ecosystem engineers for biodiversity conservation (Coleman & Williams, 2002) is another issue that requires deep insight on the structure of epibenthic megafaunal assemblages. In this study, we attempt to fill this gap of knowledge through a semi-quantitative study of the benthic megafauna inhabiting the subtidal soft bottoms of the non-estuarine areas of the Ría de Vigo. The main objective of this study was to describe the composition of

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the assemblages in terms of both abundance and biomass, to test the hypotheses predicting that (1) there will be differences between the internal (Main Channel) and external zones of the Ría in terms of both composition and diversity, and (2) the internal zone will be characterized by lower diversity and the dominance of more resistant species. 2. Material and Methods 2.1. Sampling The present study has been carried out in the two non-estuarine zones of Ría de Vigo, the Main Channel (MC) and the External Zone (EZ) (Fig. 1). These zones were defined considering geographic and demographic criteria. On one hand, there is a change in the shape of the Ría, which becomes narrower in the Main Channel; this morphological feature is of great relevance for circulation patterns as the influence of oceanic water is lower in the inner zone (Barton et al., 2015). On the other hand, the main pollution sources in the Ría (i.e., biggest population nuclei, main harbours and higher density of rafts) are located on the Main Channel. The area was divided into a grid of 119 squares of 1.05x1.05 km, which was stablished before the rest of the details of the study were arranged, so there were a discrete number of places where the division between the two zones could be placed (i.e., between two “columns” of sampling points). Among the possible locations, we selected the one which, according to the criteria mentioned above and our knowledge of the area, best represented the differentiation between the two zones (Fig. 1). Due to time constraints, 75 sampling sites out of the initial 119 in the grid were randomly selected in order to reduce sampling effort (Fig. 1). During June 2014, the selected sites were sampled by means of a towing dredge. The dredge used was a fishing gear named “rastro”, normally 6

used in commercial fishing to harvest pectinid bivalves, namely Aequipecten opercularis (Linnaeus, 1758). The opening of the gear was a rigid rectangular frame 1.5 m wide and 0.41 m high, with a chain attached to its lower part in order to reach into the first centimetres of sediment. The average values of the mesh size (stretched) were 67.8 mm in the upper part and 75.2 mm in the lower part of the net, which was 2.18 m long. The gear was towed at each site for a time period between 2 and 4 minutes after it had reached the bottom. The length of the tows ranged from 134.8 to 426.5 m, so the estimated sampled area varied from 202.2 to 639.75 m2. Sampling areas were considered as estimates because there was no way of checking the contact between the gear and the bottom during the tows. Samplings in this study will be therefore considered semi-quantitative (Eleftheriou and McIntyre, 2005). Initial position, final position and depth were registered for each tow. Megafauna was identified to species level, counted and weighed (wet weight) on-board. When identification to species level was not possible in situ, individuals were preserved in 10% buffered formalin and identification to the lowest taxonomic level possible was carried out later in the laboratory. Furthermore, when the size of the catch made it impossible to carry out the counting during the time between two consecutive tows, total weight was measured and a subsample was weighted and counted. Subsamples composed by different species difficult to sort in situ were preserved in formalin; sorting and counting of fauna were carried out later. 2.2. Data Analysis Taxa excluded from the analyses because they were not accurately sampled a priori included small-sized species, infaunal species which inhabit layers of the sediment deeper than the dredge was able to sample, as well as pelagic fishes and other highly 7

motile species. Abundance (N) and biomass (B) were calculated for each sample, both for the total assemblage and for the main taxonomic groups (poriferans, cnidarians, polychaetes, decapods, molluscs, echinoderms, ascidiaceans and fishes). Species richness (S), Shannon-Wiener diversity index (H’, log2) and Pielou’s evenness index (J’) were also calculated for each site. Values of these variables were represented using gvSIG v.1.12.0 GIS software (http://www.gvsig.org). Multivariate analyses were carried out considering presence/absence of the taxa due to the semi-quantitative nature of the data. Prior to the analyses, a resemblance matrix was constructed using Bray-Curtis similarity index. Similarity Percentages Test (SIMPER) was used to test the hypothesis about differences in the composition of the assemblages between the studied zones and to identify the taxa which most accounted for the similarities within each zone and the differences between them. Non-metric dimensional scaling (MDS) was used to obtain a visual representation of the ordination of the samples in a bidimensional space in order to contrast SIMPER results. Both MDS and SIMPER analyses were performed using the PRIMER v.6 software package (Clarke and Gorley, 2006). 3. Results 3.1. Univariate Measures Total abundance (N) ranged from 2 to 7,532 individuals, with an average value of 629 ± 1,320 individuals/sample (here and from here onwards, average values are expressed as average ± standard deviation); total biomass (B) ranged from 30 g to 64,274 g, with an average value 9,930 ± 14,751 g/sample. In all, 113 different taxa were identified (Table 1); species richness (S) in a single tow varied between 1 and 29 species, with an average of 10 ± 6. J’ index (when its calculation was possible) ranged from 0.03 to 1 and 8

H’(log2) index from 0 to 3.18, with average values of 0.55 ± 0.27 and 1.54 ± 0.85, respectively. On average, MC outscored EZ in all these variables, although deviation values were very high in both zones, especially for abundance and biomass (Table 2, Fig. 2). 3.2. Taxonomic Composition Echinoderms were on average the dominant taxonomic group in EZ, in both abundance and biomass. They were also the dominant group in MC in terms of abundance, with molluscs as a close second which outscored them in terms of biomass (Figs. 3 and 4). The high dominance of echinoderms in EZ was due to several species, and the main contributors were different when considering either abundance or biomass. On one side, the ophiuroids Amphiura chiajei Forbes, 1843, Ophiothrix fragilis (Abildgaard in O.F. Müller, 1789) and Ophiocomina nigra (Abildgaard in O.F. Müller, 1789) were the dominant species in number of individuals, with average abundances of 191 ± 2,076, 136 ± 1,793 and 122 ± 1,377 ind./sample, respectively. On the other side, the echinoid Spatangus purpureus O.F. Müller, 1776 was the dominant species in terms of biomass, with an average of 5,682 ± 16,761 g/sample. Apart from echinoderms, the most relevant species in EZ was the pennatulacean anthozoan Veretillum cynomorium (Pallas, 1766), showing in average 14 ± 71 ind./sample and 295 ± 1,175 g/sample. In MC, the exotic gastropod Crepidula fornicata (Linnaeus, 1758) was the most important species in terms of both abundance and biomass, with average values of 357 ± 843 ind./sample and 3,809 ± 8,998 g/sample, respectively. Other numerically important species were the ophiuroid O. fragilis (69 ± 204 ind./sample) and the bivalve A. opercularis (59 ± 231 ind./sample). In terms of biomass, the bivalves A. opercularis (2,451 ± 11,984 g/sample) and Pecten maximus (Linnaeus, 1758) (2,429 ± 7,604 9

g/sample), and the holothuroid Aslia lefevrii (Barrois, 1882) (1,075 ± 2,282 g/sample) were also relevant. Although faunal composition was rather variable in both zones, a higher heterogeneity could be noticed in MC (Fig. 4), with echinoderms dominating in the outermost sites and molluscs substituting them in the innermost area, the former representing a transition area to EZ. Numerically important species in that area were the ophiuroids A. chiajei, O. fragilis and O. nigra, while V. cynomorium, the ascidian Phallusia mammillata (Cuvier, 1815) and again O. fragilis were important in terms of biomass. In the innermost part of MC, C. fornicata and A. opercularis were the most relevant species in terms of both abundance and biomass, jointly with P. maximus for the latter. 3.3. Multivariate Analyses SIMPER results indicated that four species of echinoderms (Astropecten irregularis (Pennant, 1777), Ophiura ophiura (Linnaeus, 1758), S. purpureus and Marthasterias glacialis (Linnaeus, 1758)) accounted for almost 75% of the similarity among samples from EZ (Table 3). In MC, contribution to similarity among samples was more widely distributed, with the four first species (Asterias rubens Linnaeus, 1758, A. lefevrii, M. glacialis and V. cynomorium) making a cumulative contribution of only 36.03% (Table 3). Despite this, the average similarity among samples in MC was higher than that in EZ (31.55 and 25.28, respectively). The average dissimilarity between samples of the two zones was 84.05 and was mostly defined by the very same species which dominated the similarities within each of them (Table 3). MDS bidimensional plot reflected the existence of differences between EZ and MC in terms of faunal composition (Fig. 5). Samples from each zone did not form defined groups but they stood on opposite sides of the plot not mixing with each other. 10

4. Discussion Results of the multivariate analyses showed differences between the epibenthic megafaunal assemblages in EZ and MC in the Ría de Vigo, as predicted by the first hypothesis. Samples from the two zones formed a single group in the MDS analysis, but this was probably due to the high spatial resolution of the sampling and the progressive changes in the faunal composition of the assemblages. As sampling sites were close to each other and there are no environmental or physical borders which cause abrupt changes in faunal composition, the formation of differentiated groups of samples is prevented. Samples from the two zones, however, did not appear mixed with each other, supporting the existence of differences in faunal composition between the two considered zones, as it was highlighted by the results of SIMPER analysis. When considering the composition of the assemblages in terms of main taxonomic groups, a transition was observed within MC, with assemblages changing from molluscdominated in the innermost part to echinoderm-dominated in the outermost one, the latter being more similar to those in EZ. This pattern resembles to that described by López-Jamar and Cal (1990) and Margalef (1958) for infaunal macrobenthos, i.e., assemblages differed between the outer and inner Ría, and those in the innermost area (inner MC) were the most different from the rest (López-Jamar and Cal, 1990). Values of univariate measures suggest rejection of the second hypothesis, as higher abundance, biomass, species richness and diversity were found in MC; great disparity in values within each zone, however, highlights higher small-scale variation instead. In any case, these results contrast with those obtained for the infaunal macrobenthos of Ría de Vigo (López-Jamar and Cal, 1990; Margalef, 1958); previous work found lower values of both abundance and species richness in the innermost part of the Ría, where higher levels of organic enrichment of the sediment were measured (López-Jamar and 11

Cal, 1990; Vilas et al., 2005). It was proposed that organic enrichment could be triggering hypoxic conditions in the sediments there, leading to an impoverishment of the infaunal assemblages (López-Jamar and Cal, 1990; Margalef, 1958). Cacabelos et al. (2008) also suggest organic enrichment as a possible cause for low densities of infaunal macrobenthos across Ensenada de San Simón, and Aneiros et al. (2014) found a negative correlation between organic matter content and mollusc diversity in muddy sediment at the nearby Ría de Aldán. However, spatio-temporal relations between organic enrichment, hypoxia and benthic assemblages are complex and conditioned by many factors (e.g. Gray et al., 2002; Levin, 2003; Pearson and Rosenberg, 1978). For example, appropriate levels of organic matter serving as food source may contribute to enhanced benthic diversity (Rosenberg, 1995). On the other hand, the effects of hypoxia can be different for infauna and epifauna (Montagna and Froeschke, 2009), because the former are more dependent on sediment features. In the case of the Ría de Vigo, hypoxic conditions are not likely to occur in bottom water (Doval et al., 1998) but just within the sediment; this can explain why epibenthic megafaunal assemblages in MC do not seem to reflect the impoverishment observed on infaunal assemblages in the same area. When feeding guilds of the dominant species are considered, there were no numerically dominant suspension-feeders within the assemblages of EZ in terms of either abundance or biomass, while most of the dominant species in MC are known to be suspensionfeeders, namely C. fornicata, A. opercularis, P. maximus and A. lefevrii.

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suspension-feeding species are able to inhabit organically enriched sediments (Rosenberg, 1995) and their success in MC may be explained by the higher planktonic productivity in this part of the Ría as reported by Margalef (1958) and Doval et al. (1998). The ability of the Rías to support large populations of suspension-feeding 12

organisms thanks to high primary production is reflected on the intensive mussel culture in rafts, as discussed by Figueiras et al. (2002). Anthropogenic impacts may also be partly responsible for the spatial pattern observed in the Ría de Vigo. Molluscs are one of the groups with a highest resistance to different kinds of pollution, while echinoderms are usually more sensitive (e.g.: Genotoxins: Canty et al., 2009; Sewage: Stabili et al., 2013; Trace metals: Zhang et al., 2013). In this sense, the drastic reduction in echinoderm abundance and biomass observed towards the inner part of MC may be a result of the higher levels of metal pollution present there (Prego et al., 2008; Rubio et al., 2000; Vilas et al., 1995); in addition, this area is surrounded by the main population nuclei and connected, in turn, to the Ensenada de San Simón which receives the main freshwater inputs, and therefore more exposed to several kinds of pollutants. The gastropod C. fornicata is an alien species, native from eastern North America where it is found from the Gulf of St. Lawrence to the Caribbean Sea (Blanchard, 1997). It was introduced into Galician waters in the mid- 1970s (Blanchard, 1997) and first cited here by Rolán (1983), who only found a few individuals in Ría de Vigo, all of them in Ensenada de San Simón. Guerra et al. (1984) and Guerra and Pérez-Gándaras (1987) do not mention this species in the Ría but it is possible that it was present in their samples in small numbers, as they explain that rare species were not identified unless they were of commercial interest. Our findings show the great spread experimented by this species during the last 30 years, becoming the dominant species of the epibenthic megafaunal assemblages in MC in terms of both abundance and biomass. The reproductive success of C. fornicata is known to improve with higher seawater temperature, so its spread in recent years may have been favoured by rising seawater temperatures due to global warming, and this may be enhanced in the near future 13

(Bashevkin and Pechenik, 2015; Valdizan et al., 2011). The high abundance of C. fornicata does not seem to have affected epifaunal diversity in the area, despite the expected ecological impacts as described by Blanchard (1997). Highest densities of C. fornicata are found in sheltered areas (Blanchard, 1997) and this might partly explain the higher densities found at MC when compared to EZ. However, this part of the Ría is still too exposed for C. fornicata to reach its expected maximum levels of density, and therefore its potential impact on the assemblage is less substantial. Diversity values were, in general, lower than those found for the infaunal macrobenthos by López-Jamar and Cal (1990) in the same area and by others in adjacent areas (Cacabelos et al., 2008; Lourido et al., 2010; Moreira et al., 2010). Nephin et al. (2014) found an opposite pattern in the shelf and slope of Beaufort Sea, with epibenthic megafauna showing higher values of diversity than infaunal macrobenthos; this was attributed to sampling methodology because a larger area may be covered by trawling than by using grabs or cores (in Nephin et al. (2014), infauna was studied using 0.125 m2 box corer samples and epifauna by means of 450 m2 Agassiz trawl samples). However, the high diversity of the infauna in shallower bottoms prevents this to happen in coastal areas and epibenthic diversity is likely to be lower than infaunal one despite sampling effects, as happened in the normoxic area of the coastal bay studied by Montagna and Froeschke (2009). To a great extent, the higher diversity of infauna when compared to epifauna is a result of the habitat heterogeneity that shallow sedimentary bottoms provide for small-sized infaunal invertebrates. In general terms, our results showed that spatial distribution of epibenthic megafauna in the non-estuarine areas of the Ría de Vigo was mainly defined by a longitudinal pattern, changing from mollusc-dominated assemblages in the innermost part to echinodermdominated ones in its outer part. This general pattern resembled to that found for 14

infaunal macrobenthos in previous studies, and was consistent with the environmental factors pointed out by them. According to these results, changes on epibenthic megafaunal assemblages are dependent on environmental factors. While variations in macrobenthic infauna are usually related to sediment characteristics such as grain size or organic matter content (e.g. Aneiros et al., 2014; Lourido et al., 2010), epibenthic megafauna seems to be more affected by factors related to the water column, such as planktonic productivity or the presence of pollutants. The higher mobility of these organisms means that their distribution reflects larger-scale spatial changes in the mentioned environmental factors, while their bigger size and longer lifespan imply that they respond to longer-scale temporal variations in those environmental variables. Therefore, when shifts in epibenthic megafaunal assemblages are detected, they are likely to be related to high-scale shifts in water-column environmental variables. Further research should focus on understanding how these environmental factors affect both macrofaunal and megafaunal assemblages. In this sense, further similar studies in equivalent coastal areas would help to understand how the relations and patterns observed interact with each other. Additionally, better insight on the effects of different biotic or abiotic factors over distribution patterns of benthic assemblages could be achieved through experimental studies involving the most relevant species. Acknowledgements This study was partially supported by Consellería do Mar e Medio Rural, Xunta de Galicia (regional government of Galicia). The authors want to express their gratitude to the Fishermen’s Guild of Cangas Do Morrazo and the crew of the fishing boat Rucaba for assistance with sampling. The authors also want to acknowledge the very significant help of Mª Berta Ríos (Technical Assistant of the Fishermen’s Guild of Cangas Do

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Morrazo), Jesús Fernández (Consellería do Mar e Medio Rural, Delegación Comarcal de Vilagarcía, Xunta de Galicia) and Alexandra Chrysochoou during sampling. We are also grateful for the comments of two anonymous reviewers, which have been helpful to improve an earlier version of this manuscript. F. Aneiros and M. Rubal were funded by FPU program (Spanish Education Ministry) and a postdoctoral grant (programa posdoutoral Xunta de Galicia, POS-A/2012/221), respectively. Additionally, partial funding for this study was obtained from FEDER GRC2013-004 project (Xunta de Galicia). References Abella, F.E., Parada, J.M., Mora, J., 1996. Relationship between the macrobenthic community structure and the presence of mussel rafts culture in the Ría de Vigo (NW Iberian Peninsula). Crangon 1, 111-118. Alejo, I., Vilas, F., 1987. Dinámica litoral y evolución histórica de la Ensenada de Bayona (Pontevedra). Thalassas 5, 21-32. Alonso-Pérez, A., Ysebaert, T., Castro, G.C., 2010.Effects of suspended mussel culture on benthic-pelagic coupling in a coastal upwelling system (Ría de Vigo, NW Iberian Peninsula). J. Exp. Mar. Biol. Ecol. 382, 96-107. Alvarez-Salgado, X.A., Rosón, G., Pérez, F.F., Figueiras, F.G., Pazos, Y., 1996. Nitrogen cycling in an estuarine upwelling system, the Ría de Arousa (NW Spain). I. Short-time-scale patterns of hydrodynamic and biogeochemical circulation. Mar. Ecol. Prog. Ser. 135, 259-273.

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Aneiros, F., Moreira, J., Troncoso, J.S., 2014. A functional approach to the seasonal variation of benthic molluscs assemblages in an estuarine-like system. J. Sea Res. 85, 73-84. Ardré, F., Cabañas Ruesgas, F., Fischer-Piette, E., Seoane, J., 1958. Petite Contribution à une Monographie Bionomique de la Ria de Vigo. Bulletin de l’Institut Océanographique 1127. Monaco. 56pp. Bañón, R., Rolán, E., García-Tasende, M., 2008. First record of the purple dye murex Bolinusbrandaris (Gastropoda: Muricidae) and a revised list of non native molluscs from Galician waters (Spain, NE Atlantic). Aquat. Invasions 3(3), 331-334. Barton, E.D., Largier, J.L., Torres, R., Sheridan, M., Trasviña, A., Souza, A., Pazos, Y., Valle-Levinson, A., 2015. Coastal upwelling and downwelling forcing of circulation in a semi-enclosed bay: Ria de Vigo. Prog. Oceanogr. 134, 173-189. Bashevkin, S.M., Pechenik, J.A., 2015. The interactive influence of temperature and salinity on larval and juvenile growth in the gastropod Crepidula fornicata (L.). J. Exp. Mar. Biol. Ecol. 470, 78-91. Blanchard, M., 1997. Spread of the slipper limpet Crepidula fornicata (L. 1758) in Europe. Current state and consequences. Sci. Mar. 61(Sup2), 109-118. Cacabelos, E., Gestoso, L., Troncoso, J.S., 2008. Macrobenthic fauna in the Ensenada de San Simón (Galicia, north-western Spain). J. Mar. Biol. Ass. U.K. 88(02), 237245. Canty, M.N., Hutchinson, T.H., Brown, R.J., Jones, M.B., Jha, A.N., 2009. Linking genotoxic responses with cytotoxic and behavioural or physiological consequences:

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Differential sensitivity of echinoderms (Asteriasrubens) and marine molluscs (Mytilusedulis). Aquat.Toxicol. 94, 68-76. Clarke, K.R., Gorley, R.N., 2006. PRIMER V6: user manual/tutorial. PRIMER-E Ltd, Plymouth. Coleman, F.C., Williams, S.L., 2002. Overexploiting marine ecosystem engineers: potential consequences for biodiversity. Trends Ecol. Evol. 17(1), 40-44. de Leo, F.C., Smith, C.R., Rowden, A.A., Bowden, D.A., Clark, M.R., 2010. Submarine canyons: hotspots of benthic biomass and productivity in the deep sea. Proc. R. Soc. B 277, 2783-2792. Doval, M.D., Nogueira, E., Pérez, F.F., 1998. Spatio-temporal variability of the thermohaline and biogeochemical properties and dissolved organic carbon in a coastal embayment affected by upwelling: the Ría de Vigo (NW Spain). J. Mar. Sys. 14, 135-150. Eleftheriou, E., McIntyre, A., 2005. Methods for the Study of Marine Benthos. Blackwell Science, Oxford. Figueiras, F.G., Labarta, U.,Fern, M.J., 2002. Coastal upwelling, primary production and mussel growth in the Rías Baixas of Galicia. Hydrobiologia 484, 121-131. Fraga, F., 1981.Upwelling off the Galician Coast, NorthWest Spain, in: Richards, F.A. (Ed.), Coastal Upwelling, vol. 1. American Geophysical Union, Washington D.C., pp. 176-182. Gray, J.S., Wu, R.S., Or, Y.Y., 2002. Effects of hypoxia and organic enrichment on the coastal marine environment. Mar. Ecol. Prog. Ser. 238, 249-279.

18

Guerra, A., Alonso-Allende, J.M., Pérez-Gándaras, G., Ferreiro, M.J., Figueras, A.J., Labarta, U., 1984.Especies bentónicas y demersales de la Ría de Vigo. Pescas de arrastre de fondo (1982-1984). Instituto de Investigaciones Pesqueras de Vigo. Vigo. 188pp. Guerra, A., Pérez-Gándaras, G., 1987. Especies demersales y bentónicas de la Ría de Vigo. Resultados preliminares. Instituto de Investigaciones Marinas. Vigo. 246-288. Levin, L.A., 2003. Oxygen mínimum zone bentos: adaptation and community response to hipoxia. Oceanogr. Mar. Biol. 41, 1-45. Levin, L.A., Ekau, W., Gooday, A.J., Jorissen, F., Middelburg, J.J., Naqvi, S.W.A., Neira, C., Rabalais, N.N., Zhang, J., 2009. Effects of natural and human-induced hypoxia on coastal benthos. Biogeosciences 6, 2063-2098. López-Jamar, E., 1978a. Macrobentos infaunal de la Ría de Pontevedra. Bol. Inst. Esp. Oceanogr. 4, 113-130. López-Jamar, E., 1978b. Primeros datos sobre la biomasa y la composición del bentos infaunal de la Ría de Pontevedra, en relación con el contenido en materia orgánica del sedimento. Bol. Inst. Esp. Oceanogr. 4, 57-69. López-Jamar, E., Cal, R.M., 1990. El sistema bentónico de la zona submareal de la Ría de Vigo. Macroinfauna y microbiología del sedimento. Bol. Inst. Esp. Oceanogr. 6 (2), 49-60. Lourido, A., Moreira, J., Troncoso, J.S., 2010. Spatial distribution of benthic macrofauna in subtidal sediments of the Ría de Aldán (Galicia, northwest Spain). Sci. Mar. 74(4), 705-715.

19

Margalef, R., 1958. La sedimentación orgánica y la vida en los fondos fangosos de la Ría de Vigo. Inv. Pesq. 11, 67-100. Montagna, P.A., Froeschke, J., 2009. Long-term biological effects of coastal hypoxia in Corpus Christi Bay, Texas, USA. J. Exp. Mar. Biol. Ecol. 381, S21-S30. Mora, J., Planas, M., Silva, R., 1989. Impacto de la contaminación orgánica en la Ensenada de Lourizán (Proyecto Escorp) I – El medio físico y la macrofauna bentónica. Cah. Biol. Mar. 30, 181-199. Moreira, J., Lourido, A., Troncoso, J.S., 2010. Temporal dynamics of the benthic assemblage in the muddy sediments of the Harbour of Baiona (Galicia, NW Iberian Peninsula). Thalassas 26(2), 9-22. Navaz, J.M., 1942. Estudio de los yacimientos de moluscos comestibles de la Ría de Vigo. Instituto Español de Oceanografía, Trabajo Nº16. Madrid. 106pp. Nephin, J., Juniper, S.K., Archambault, P., 2014. Diversity, abundance and community structure of benthic macro- and megafauna on the Beaufort Shelf and Slope. PLoS ONE 9(7), 1-11. Nombela, M.A., Vilas, F., Rodríguez, M.D., Ares, J.C., 1987. Estudio sedimentológico del litoral gallego: III, resultados previos sobre los sedimentos de los fondos de la Ría de Vigo. Thalassas 5(1), 7-19. Pearson, T.H., Rosenberg, R., 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine environment. Oceanogr. Mar. Biol. Ann. Rev. 16, 229-311.

20

Prego, R., Filgueiras, A.V., Santos-Echeandía, J., 2008. Temporal and spatial changes of total and labile metal concentration in the surface sediments of the Vigo Ria (NW Iberian Peninsula): Influence of anthropogenic sources. Mar. Pollut. Bull. 56, 10311042. Rolán, E., 1983. Moluscos de la Ría de Vigo 1. Gasterópodos. Thalassas Anexo 1, 1383. Rosenberg, R., 1995. Benthic marine fauna structured by hydrodynamic processes and food availability. Neth. J. Sea Res. 34(4), 303-317. Rubio, B., Nombela, M.A., Vilas, F., 2000. Geochemistry of Major and Trace Elements in Sediments of the Ria de Vigo (NW Spain): an Assessment of Metal Pollution. Mar. Pollut. Bull. 40, 968-980. Sandnes, J., Forbes, T., Hansen, R., Sandnes, B., Rygg, B., 2000. Bioturbation and irrigation in natural sediments, described by animal-community parameters. Mar. Ecol. Prog. Ser. 197, 169-179. Smallwood, B.J., Wolff, G.A., Bett, B.J., Smith, C.R., Hoover, D., Gage, J.D., Patience, A., 1999. Megafauna can control the quality of organic matter in marine sediments. Naturwissenschaften 86, 320-324. Smith, C.R., Berelson, W., Demaster, D.J., Dobbs, F.C., Hammond, D., Hoover, D.J., Pope, R.H., Stephens, M., 1997. Latitudinal variations in benthic processes in the abyssal equatorial Pacific: control by biogenic particle flux. Deep-Sea Res. Pt. II 44(9-10), 2295-2317.

21

Stabili, L., Terlizzi, A., Cavallo, R.A., 2013. Sewage-exposed marine invertebrates: survival rates and microbiological accumulation. Environ. Sci. Pollut. Res. 20, 1606-1616. Turnewitsch, R., Witte, U., Graf, G., 2000. Bioturbation in the abyssal Arabian Sea: influence of fauna and food supply. Deep-Sea Res. Pt. II 47, 2877-2911. Valdizan, A., Beninger, P.G., Decottignies, P., Chantrel, M., Cognie, B., 2011. Evidence that rising coastal seawater temperatures increase reproductive output of the invasive gastropod Crepidula fornicata. Mar. Ecol. Prog. Ser. 438, 153-165. Vardaro, M.F., Ruhl, H.A., Smith Jr., K.L., 2009. Climate variation, carbon flux, and bioturbation in the abyssal North Pacific. Limnol. Oceanogr. 54(6), 2081-2088. Varela, R.A., Rosón, G., Herrera, J.L., Torres-López, S., Fernández-Romero, A., 2005. A general view of the hydrographic and dynamical patterns of the Rías Baixas adjacent sea area. J. Mar. Sys. 54, 97-113. Vilas, F., Nombela, M.A., García-Gil, E., García-Gil, S., Alejo, I., Rubio, B., Pazos, O., 1995. Cartografía de sedimentos submarinos, Ría de Vigo, Escala 1:50000 (Memoria y Mapa). 40pp. Vilas, F., Bernabeu, A.M., Méndez, G., 2005. Sediment distribution pattern in the Rias Baixas (NW Spain): main facies and hydrodynamic dependence. J. Mar. Sys. 54, 261-276. vonRichthofen, F., 1886. Führer für Forschungsreisende. Oppenheim, Berlin.

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Ysebaert, T., Hart, M., Herman, P.M.J., 2009. Impacts of bottom and suspended cultures of mussels Mytilus spp. on the surrounding sedimentary environment and macrobenthic biodiversity. Helgol. Mar. Res. 63(1), 59-74. Zhang, Y., Lv, Z., Guan, B., Liu, Y., Li, F., Li, S., Ma, Y., Yu, J., Li, Y., 2013. Status of macrobenthic community and its relationships to trace metals and natural sediment characteristics. Clean-Soil Air Water 41(10), 1027-1034.

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Figure Captions Figure 1. Location of the Ría de Vigo and the sampling sites at the External Zone (EZ; black circles) and the Main Channel (MC; grey circles). Figure 2. Sampling sites showing: A – total abundance (N), B – total biomass (B), C – taxa richness (S), D – Shannon-Wiener Diversity Index, log2 (H’) and E – Pielou’s Evenness Index (J’). Vertical black lines represent the limit between EZ and MC. Figure 3. Average values for EZ (A and C) and MC (B and D) of the contribution of each of the main taxonomic groups to A-B – total abundance (N) and C-D – total biomass (B). Figure 4. Sampling sites showing the contribution of the main taxonomic groups to A – total abundance (N) and B – total biomass (B). Vertical black lines represent the limit between EZ and MC. Figure 5. Non-metric multidimensional scaling (MDS) ordination of the assemblages in Ría de Vigo. Assemblages from EZ are represented as solid circles (●) and those from MC are represented as open circles (○). One sample from EZ and two samples from MC have been excluded for a better visualization of the ordination, due to their high dissimilarity with the rest of the samples.

Highlights: - Ría de Vigo is a biologically productive and socioeconomically relevant coastal area. - Benthic megafauna diversity and distribution are studied in detail for the first time. - A shift in assemblage composition along the main axis of the bay is spotted. - Possible environmental and anthropogenic factors affecting distribution are discussed.

24

Table 1. List of taxa identified from the 75 sites sampled in Ría de Vigo. PORIFERA Cliona celata Grant, 1826 Demospongia Indet. 1 Demospongia Indet. 2 Porifera Indet. 1 Porifera Indet. 2 Porifera Indet. 3 Porifera Indet. 4 Porifera Indet. 5 Suberites ficus (Johnston, 1842) CNIDARIA Actiniaria Indet. 1 Actiniaria Indet. 2 Adamsia carciniopados (Otto, 1823) Alcyonium digitatum Linnaeus, 1758 Anthozoa Indet. 1 Calliactis parasitica (Couch, 1842) Pennatulidae Indet. 1 Sagartiogeton sp. Veretillum cynomorium (Pallas, 1766) ANNELIDA Aphrodita aculeata Linnaeus, 1758 Lanice conchilega (Pallas, 1766) Maldane glebifex Grube, 1860 ARTHROPODA Alpheus macrocheles (Hailstone, 1835) Atelecyclus rotundatus (Olivi, 1792) Atelecyclus undecimdentatus (Herbst, 1783) Ebalia tuberosa (Pennant, 1777) Eurynome aspera (Pennant, 1777) Eurynome spinosa Hailstone, 1835 Goneplax rhomboides (Linnaeus, 1758) Inachus dorsettensis (Pennant, 1777) Liocarcinus corrugatus (Pennant, 1777) Liocarcinus depurator (Linnaeus, 1758) Liocarcinus marmoreus (Leach, 1814) Liocarcinus pusillus (Leach, 1816) Macropodia rostrata (Linnaeus, 1761) Maja brachydactyla Balss, 1922 Necora puber (Linnaeus, 1767) Pagurus bernhardus (Linnaeus, 1758) Pilumnus hirtellus (Linnaeus, 1761) Pisidia longicornis (Linnaeus, 1767) Polybius henslowii Leach, 1820 MOLLUSCA Abra alba (W. Wood, 1802) Acanthocardia paucicostata (G. B. Sowerby II, 1834) Aequipecten opercularis (Linnaeus, 1758) 25

Antalis novemcostata (Lamarck, 1818) Aporrhais pespelecani (Linnaeus, 1758) Calliostoma zizyphinum (Linnaeus, 1758) Chaetopleura (Chaetopleura) angulata (Spengler, 1797) Chamelea striatula (da Costa, 1778) Charonia lampas (Linnaeus, 1758) Clausinella fasciata (da Costa, 1778) Crepidula fornicata (Linnaeus, 1758) Doris pseudoargus Rapp, 1827 Gibbula magus (Linnaeus, 1758) Glycymeris glycymeris (Linnaeus, 1758) Laevicardium crassum (Gmelin, 1791) Mimachlamys varia (Linnaeus, 1758) Mysia undata (Pennant, 1777) Mytilus galloprovincialis Lamarck, 1819 Nassarius ovoideus (Locard, 1886) Nassarius pygmaeus (Lamarck, 1822) Nassarius reticulatus (Linnaeus, 1758) Nucula nitidosa Winckworth, 1930 Ocenebra erinaceus (Linnaeus, 1758) Octopus vulgaris Cuvier, 1797 Ostrea edulis Linnaeus, 1758 Palliolum tigerinum (O. F. Müller, 1776) Pecten maximus (Linnaeus, 1758) Scaphander lignarius (Linnaeus, 1758) Sepia officinalis Linnaeus, 1758 Turritella communis Risso, 1826 Turritella turbona Monterosato, 1877 Venus cassinaeformis (Yokoyama, 1926) Venus verrucosa Linnaeus, 1758 ECHINODERMATA Amphiura chiajei Forbes, 1843 Antedon bifida (Pennant, 1777) Aslia lefevrii (Barrois, 1882) Asterias rubens Linnaeus, 1758 Astropecten irregularis (Pennant, 1777) Echinaster (Echinaster) sepositus (Retzius, 1783) Echinocardium cordatum (Pennant, 1777) Echinocucumis hispida (Barrett, 1857) Echinus esculentus Linnaeus, 1758 Holothuria (Panningothuria) forskali Delle Chiaje, 1823 Luidia ciliaris (Philippi, 1837) Marthasterias glacialis (Linnaeus, 1758) Oestergrenia digitata (Montagu, 1815) Ophiocomina nigra (Abildgaard, in O.F. Müller, 1789) Ophiothrix fragilis (Abildgaard, in O.F. Müller, 1789) Ophiothrix luetkeni Wyville Thomson, 1873 Ophiura ophiura (Linnaeus, 1758) Paracentrotus lividus (Lamarck, 1816) Psammechinus miliaris (P.L.S. Müller, 1771) 26

Spatangus purpureus O.F. Müller, 1776 Sphaerechinus granularis (Lamarck, 1816) Thyone sp. CHORDATA Apletodon dentatus (Facciolà, 1887) Aplidium sp. Arnoglossus laterna (Walbaum, 1792) Ascidia mentula Müller, 1776 Ascidia sp. Buglossidium luteum (Risso, 1810) Callionymus lyra Linnaeus, 1758 Ciona intestinalis (Linnaeus, 1767) Gobius niger Linnaeus, 1758 Lepadogaster candolii Risso, 1810 Molgula sp. Phallusia mammillata (Cuvier, 1815) Raja undulata Lacepède, 1802 Scophthalmus rhombus (Linnaeus, 1758) Solea senegalensis Kaup, 1858 Styela clava Herdman, 1881 Torpedo marmorata Risso, 1810 Zeugopterus regius (Bonnaterre, 1788)

Table 2. Summary of the results of the univariate measures of the assemblages (average ± standard deviation) for the two studied zones.

Total N Total B (g) S H’ J’

External Zone 550 ± 1,565 8,021 ± 12,950 8±6 1.38 ± 0.89 0.55 ± 0.28

Main Channel 748 ± 838 12,793 ± 16,933 13 ± 6 1.77 ± 0.75 0.55 ± 0.25

27

Table 3. Results of the SIMPER analysis (summarized to the top four species for each case).

Species

are

ranked

according

to

their

average

contributions

to

similarity/dissimilarity (AS/AD) within/between sites. Average abundance (AA), ratio value (R: similarity or dissimilarity/standard deviation) and percentage of cumulative similarity/dissimilarity (%Cum) are also included.

Similarities Dissimilarities Species AA AS R %Cum Species AA AA AD EZ (Average similarity: 25.28) EZ / MC (Average dissimilarity: 84.05) Astropecten 0.80 10.24 0.95 40.51 EZ MC irregularis Astropecten Ophiura ophiura 0.56 3.53 0.58 54.46 0.80 0.03 4.59 irregularis Spatangus purpureus 0.44 3.01 0.42 66.37 Asterias rubens 0.13 0.70 3.57 Marthasterias Marthasterias 0.42 2.09 0.41 74.64 0.42 0.57 2.97 glacialis glacialis Veretillum MC (Average similarity: 31.55) 0.31 0.60 2.91 cynomorium Asterias rubens 0.70 3.79 0.87 12.02 Aslia lefevrii 0.63 2.64 0.78 20.39 Marthasterias 0.57 2.54 0.62 28.43 glacialis Veretillum 0.60 2.40 0.68 36.03 cynomorium

R

%Cum

1.15

5.46

1.05

9.72

0.81 13.25 0.83 16.71

28

29

30

31