Implementation of the MSFD to the Deep Mediterranean Sea IDEM

Project Coordinator: Roberto Danovaro

Deliverable 2.2. Report on the first assessment of the deep Mediterranean environmental status

Leader: UNIVPM Participants: CNR, CSIC, DFMR, ENEA, TAU, UB, UM, UNIVPM

SUBMISSION DATE 26th October| 2018

Deliverable 2.2 Table of Content

1. INTRODUCTION ...... 3 2. DESCRIPTOR 1: BIODIVERSITY ...... 5 3. DESCRIPTOR 2: NON-INDIGENOUS SPECIES ...... 37 4. DESCRIPTOR 3: POPULATIONS OF ALL COMMERCIALLY EXPLOITED FISH AND SHELLFISH ...... 38 5. DESCRIPTOR 4: ECOSYSTEMS, INCLUDING FOOD WEBS ...... 54 6. DESCRIPTOR 5: EUTROPHICATION ...... 67 7. DESCRIPTOR 6: SEAFLOOR INTEGRITY ...... 68 8. DESCRIPTOR 7: PERMANENT ALTERATION OF HYDROGRAPHICAL CONDITIONS ...... 78 9. DESCRIPTOR 8 AND 9: CONCENTRATIONS OF CONTAMINANTS/CONTAMINANTS IN FISH AND OTHER SEAFOOD FOR HUMAN COMSUMPTION ...... 106 9.1. Contaminants in water ...... 106 9.2. Contaminants in sediments ...... 125 9.3 Contaminants in the biota ...... 141 10. DESCRIPTOR 10: MARINE LITTER ...... 180 11. DESCRIPTOR 11: INTRODUCTION OF ENERGY ...... 193 12. CONCLUSIONS AND FUTURE WORK ...... 194

www.msfd-idem.eu Deliverable 2.2 1. INTRODUCTION

The main aim of the present Deliverable is to provide an overview of the state, pressures and impacts on the deep Mediterranean Sea. We here report trends in the status, levels of pressures and impacts for those deep-sea areas for which data are available. This Deliverable is the product of the work carried out within the task 2.2 and is also based on the results of previous task 2.1 and Deliverable 2.1. During the last months of the project we focused on the collection and analysis of existing datasets obtained from scientific literature, open-access repositories, public datasets and partners’ own datasets related to monitoring programs from the different Mediterranean sub-regions. To achieve one of the aim of the IDEM project, the first assessment of deep Mediterranean status, here we analysed the available datasets and reported the results of a proper meta-analysis or a semi- quantitative analysis, depending on the availability of the data. A meta-analysis refers to a process of integration of the results of many studies to arrive at evidence synthesis (Normand, 1999). Meta-analysis is essentially a systematic review; however, in addition to narrative summary that is conducted in systematic review, in meta analysis, the analysts also numerically pool the results of the studies and arrive at a summary estimate. In the case of the overall datasets, this numerical estimation cannot be performed and we carried out a systematic review, due to the constrains highlighted below. Additionally, a systematic review has been done for those MSFD Descriptors for which a sufficient amount of data is available. The previous report (Deliverable 2.1) has evidenced indeed several gaps concerning data availability for some Descriptors. Specifically, Descriptors 2 (non-indigenous species), 5 (eutrophication) and 11 (introduction of energy) presented very few available data. More in detail, the records of deep-living non-indigenous species (Descriptor 2) are reported in few papers (11) and limited to the Eastern Mediterranean and the south-eastern part of Aegean Sea. Regarding eutrophication (Descriptor 5), most of the papers do not cover the criteria defined by the MSFD. Even a lower amount of papers (5) have been published regarding the presence and the effects of noise on deep Mediterranean ecosystems and organisms and the available data are essentially provided by deep-sea cabled observatories (i.e. ANTARES in the North-western Mediterranean and KM3NET off Cape Passero in southern Sicily). Thus, for the purpose of this report we focused on D1, 3, 4, 6, 7, 8, 9 and 10. As previously reported, also for these Descriptors we have a fragmented knowledge with spatial and temporal gaps and lack of long- term data.

For Descriptor 1, the amount of data available varied depending on species groups and habitat types. Thus, we carried out a proper meta-analysis on mega, macro and meio-fauna inhabiting different habitats (both soft and hard bottom substrates).

Regarding the Descriptor 3, most of the assessed stocks are in the Western Mediterranean Sea. In this report we analysed the trend of the first two criteria of D3, i.e., the ratio between fishing mortality and Maximum Sustainable Yield and the spawning stock biomass, in order to assess if commercially exploited stocks, inhabiting deep-sea bottoms are in a healthy state and if exploitation should be sustainable. Further, in the case of descriptor 3, regarding those species which inhabit both the shelf and the slope (e.g. red mullets, hakes etc.), it was not possible to disentangle data from shallow and deep-sea bottoms, as stock assessments considered data complessively.

Marine food webs were analysed on the basis of the stable isotope signatures of the species analysed in the three different sub-basins of the deep Mediterranean, based on the datasets provided by Report 2.1 3

www.msfd-idem.eu Deliverable 2.2 and updated with recent publications/partners’ data, if any. As stable isotope analysis (SIA) is considered a promising approach in depicting food webs and data are easily comparable among basins, when a baseline of reference is taken into account, this dataset was considered the most appropriate for the meta-analysis approach. SIA may comply with one of the primary criterion (D4C1: diversity of trophic guilds) and in part with one of the secondary criteria (D4C3: size distribution of individuals across trophic guilds) established by COMM DEC 848/2017.

Regarding seafloor integrity (Descriptor 6), we carried out a semi-quantitative analysis reporting the main pressures identified in the deep Mediterranean Sea, quantitative data on bottom trawling and waste disposal and their impacts on habitats and ecosystems. For Descriptor 7 we here reviewed information regarding data distribution/availability, reporting long-term variations of hydrological conditions (i.e., alterations of temperature, salinity, nutrients, Chlorophyll a and acidification level) and the impacts on habitat and ecosystems.

Concerning descriptors 8 and 9, presence of contaminants was described in deep Mediterranean waters, sediments and biota. The main toxic effects were also highlighted.

Finally, for marine litter (Descriptor 10) a semi-quantitative analysis was performed focusing on the four D10 criteria.

www.msfd-idem.eu Deliverable 2.2 2. DESCRIPTOR 1: BIODIVERSITY

The deep sea is characterized by outstanding features from a geological, biological, and oceanographic perspective. The deep sea hosts focal points for fisheries and threatened ecosystems, particularly where peculiar habitats occur (such as canyons) along the continental slope. The Mediterranean Sea is estimated to contain more than 500 canyons that are characterized by peculiar features: they are indeed steeper, more closely spaced, and are amongst the most dendritic of the world (Harris and Whiteway, 2011; Würtz, 2012). In the Mediterranean Sea, the alternation of different habitats along the continental margins is responsible for a high spatial heterogeneity and complex hydrographic patterns, which result in exchange of organisms, organic matter and sediments from the shelf to the deep sea (Vetter and Dayton, 1998; Tyler et al., 2009; Vetter et al., 2010; Thomsen et al., 2017), enhancing faunal abundance and diversity (De Leo et al., 2010; Ramirez-Llodra et al., 2010). The high spatial heterogeneity and habitat complexity at both local and regional scale (De Leo et al., 2010; Huvenne et al., 2011) are factors expected to contribute to the high meio- macro- and megafaunal biodiversity (reviewed by Fernandez-Arcaya et al., 2017). In the Mediterranean Sea, the low primary productivity, limited freshwater inputs and high and nearly constant temperatures (13-14°C, below 200 m), which causes rapid degradation of particulate organic matter (POM) exported from the photic zone, have detrimental effects on the food availability for the deep-sea benthic fauna (Danovaro et al., 1999; Pusceddu et al., 2009; 2010; Luna et al., 2012). Several studies indicated a clear and strong zonation of fauna by depth, with decreasing abundance, biomass, and diversity, especially below 1500 m (Pérès, 1985; Company et al., 2004; D’Onghia et al., 2004; Tecchio et al., 2013). The Mediterranean shows also an increasing food limitation (oligotrophy) moving eastward, with a consequent decrease of organic matter availability on the seafloor, which is mirrored by deep-sea benthic assemblages (Azov, 1991; Danovaro et al., 1999; Gambi et al., 2017). However, this is apparently not the case for fish assemblages. Indeed, an extensive study, based on MEDITS shelf and slope data (Granger et al., 2015) showed the absence of such a decreasing trend for demersal fish assemblages sampled by trawling, and suggested that other factors than primary production (or more generally food availability) may explain large scale patterns of their species richness. Other studies suggested that differences in deep-water masses, can play a fundamental role in structuring deep-sea bentho-pelagic communities (Tecchio et al., 2013), together with habitat heterogeneity (i.e. changes in topography and other physical conditions) being the major driving factor in deep-benthic areas (Levin et al., 2001). In this context, an extensive review of the literature of the last 20 years and a comparison of these results with a new datasets, dealing with meiofaunal taxa and megafaunal species inhabiting different habitats were performed (Fanelli et al. 2018). The authors considered data collected or recorded with different gears, from baited cameras, mounted on aluminium frame (Jones et al. 2003) and landers (D’Onghia et al., 2015) and baited traps (Della Croce & Albertelli, 1995; Albertelli et al., 1992; our own data), to commercial trawl gears, in order to update available knowledge on the biodiversity and assemblage structure of mobile megafaunal species in such complex habitats, and to explore the role of canyons, as highly heterogeneous habitats, in structuring also the communities inhabiting the adjacent slope. In this report, we analysed data available in the literature on meiofaunal abundance, biomass, richness of taxa and taxonomic composition, in the Western, Central and Eastern Mediterranean sub-basins, in different habitats (specifically canyons and open slopes, and compared with data from Portuguese margin, NE Atlantic sector). Similarly, we report the results from Fanelli et al. (2018), in which all available literature on deep-sea mobile megafauna occurring in canyons and adjacent slope were explored. Only papers

www.msfd-idem.eu Deliverable 2.2 containing public datasets where data extraction was possible have been considered for the subsequent meta-analysis. For meiofauna, 200 samples within canyons and 84 along the open slopes were considered. For megafauna, 20 public data sets were used (Supplementary Table S1, Fanelli et al., 2018) on megafaunal assemblages, in terms of abundance, richness of species, and taxonomic composition. The datasets utilized span from the Western to the Central and Eastern Mediterranean sub-basins, from Catalan, Balearic, Ligurian seas, Western Mediterranean abyssal plain (Western Mediterranean sub- basin), Western Ionian Sea, Central Mediterranean abyssal plain (Central Mediterranean sub-basin), and Eastern Ionian, Aegean and Cretan Seas (Eastern Mediterranean sub-basin) (Figure 2.1). All the data were analyzed by means of uni- and multivariate permutational analyses of variance (PERMANOVA), in order to assess differences in the faunal abundance, biomass and diversity among basins and depth ranges. Multivariate multiple regression analysis (DistLM) were also performed in order to assess whether the environmental variables were responsible for the observed faunal variability (for further details, see Fanelli et al., 2018). Moreover, the meta-analysis has also been conducted on hard-substrates communities. In this case, the meta-analysis was based on 13 papers (D'Onghia et al., 2001; 2012; 2016; Bo et al., 2012; 2014; 2015; Cartes et al., 2013; Gori et al., 2013; Giusti et al., 2015; 2017; Sandulli et al., 2015; Grinyó et al., 2016; Santin et al., 2017). The collected data span from the Western Mediterranean to the Central sector of the basin but dealt with different biotic components (such as corals but also benthic and bento-nectonic associated fauna).

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Figure 2.1. A) Location of the sampling sites in the deep Mediterranean Sea and Portuguese margin (NE Atlantic Ocean) canyons and open slopes. B) Location of the sampling sites in the deep Mediterranean Sea. Bottom trawls areas are reported in light blue, traps and baited cameras are reported in red. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

Meiofaunal abundance and biomass: basin, habitat and bathymetric effects Higher values of meiofaunal abundance in N Atlantic Ocean than in Mediterranean Sea were observed below 3000 m water depth for canyons, whereas higher values of biomass were observed below 500 m water depth for canyons and at all the bathymetric ranges for the open slope. In the Mediterranean Sea, meiofaunal abundance in canyons was higher than in the open slopes at 150- 500 m water depth in the Ligurian margin (more specifically in the Calvi canyon). In all the considered margins, a decreasing pattern of meiofaunal abundance with increasing water depth was observed for both canyons (except for Adriatic, and more specifically along Cap de Creus, Lacaze-Duthiers, Petit Rhône and Samaria canyons) and open slopes (more specifically along slopes adjacent to Blanes, Cap de Creus and Calvi canyons). Higher values of meiofaunal biomass were observed in canyons than in open slopes only at 500-1000 m water depth in the Adriatic margin. In Catalan and Adriatic margins a decreasing pattern of meiofaunal biomass with increasing water depth was observed for both canyons (more specifically along Cap de Creus and C) and open slopes (along slopes adjacent to Lacaze-Duthiers and Adriatic canyons).

Meiofaunal richness of taxa and taxonomic composition: basin, habitat and bathymetric effects Similar values of expected richness of meiofaunal taxa in Mediterranean than in N Atlantic Ocean were observed for canyons and slopes at all the bathymetric ranges. Significant differences in the taxonomic composition between N Atlantic Ocean and Mediterranean Sea canyons were observed between 1000 and 3000 m or below 2000 m water depth when data were presence/absence transformed, whereas no differences were observed for open slopes (Figure 2.2). The percentage dissimilarity in the taxonomic composition (Table 2.1) between N Atlantic and Mediterranean ranged from 87 to 98% in canyons and from 82 to 99% in open slopes. When data are presence/absence transformed, the percentage dissimilarity ranged from 41 to 64% in canyons and from 41 to 62% in open 7

www.msfd-idem.eu Deliverable 2.2 slopes. The percentage dissimilarity increased with increasing bathymetric range in both canyons and open slopes (Table 2.1).

In the Mediterranean Sea, over a total of 16 available comparisons between canyons and adjacent open slopes at the same depth range, the expected richness of taxa was higher in the canyons than in the adjacent open slopes in 5 cases, was the same in 8 cases and in 3 cases was higher in the adjacent open slope than in the canyon. The cumulative expected richness of taxa reported within each margin was higher in canyons than in the open slopes at similar depth ranges except for Ligurian and Cretan margins, where the richness of taxa was higher in slopes. In the Mediterranean Sea, differences between canyons and open slopes in the meiofaunal taxonomic composition were observed at 150-500 m (in Ligurian margin, Calvi canyon) and 1000-2000 m (in Catalan margin, Cap de Creus and Blanes canyons). Significant differences were also observed among bathymetric ranges along canyons and slopes in Catalan region (along Blanes, Lacaze-Duthiers and Petit-Rhône canyons, and open slopes adjacent to Cap de Creus). When data were presence/absence transformed, significant differences were also observed Calvi slope. In the Mediterranean Sea, the overall percentage dissimilarity between canyons and adjacent slopes was 34-45% (at 500-1000 and 2000-3000 m water depth, respectively) and 25-46% when data were presence/absence transformed (2000-3000 and 150-500 m, respectively). The highest dissimilarity between canyons and adjacent slopes was observed in the Catalan margin (Table 2.1). In all margins, the dissimilarity was highest when the data were presence/absence transformed (except for Adriatic at 500- 1000 m and Cretan at 2000-3000 m water depth).

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Figure 2.2. CAP analysis output reporting the taxonomic composition of meiofaunal assemblages at different bathymetric ranges in NE Atlantic ocean vs Mediterranean canyons (A) and opens slopes (B), and in canyons vs open slopes in the Mediterranean Sea (C).

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Table 2.1. Dissimilarity % in the meiofaunal taxonomic composition, also after presence/absence transformation, among basins and between canyons and open slopes in investigated Mediterranean continental margin. Bathymetric % % dissimilarity range (m) dissimilarity (presence/absence) N Atlantic vs Mediterranean canyon 150-500 88.9 40.8 500-1000 86.6 38.7 1000-2000 91.8 43.6 2000-3000 89.2 42.7 3000-4000 97.8 64.4 open slope 150-500 82.1 41.11 500-1000 86.9 48.38 1000-2000 89.9 43.86 2000-3000 99.6 62.39 Mediterranean canyon vs Catalan 150-500 34.5 44.81 slope

500-1000 33.1 41.99

1000-2000 32.8 40.19 Ligurian 150-500 28.4 31.54 Adriatic 150-500 25.5 26.54 500-1000 26.7 18.62 Cretan 1000-2000 28.9 36.51 2000-3000 15.8 0 Mediterranean canyon vs 150-500 38.2 46.02 slope (avg % dissimilarity) 500-1000 34.2 42.56 1000-2000 34.8 41.81 2000-3000 45.8 25.08

Environmental drivers of meiofaunal assemblages in N Atlantic and Mediterranean deep canyons Meiofaunal abundance, biomass and richness of taxa showed a significant negative log-linear relationship with the water depth in the Mediterranean Sea canyons and open slopes (Figure 2.3). Conversely, no significant relationships were observed in the N Atlantic canyons and open slopes (Figure 2.3). Meiofaunal abundance, biomass and richness of taxa showed a significant negative log-linear relationship with increasing longitude in the Mediterranean Sea canyons and open slopes (Figure 2.4). The results of the multivariate multiple regression analyses (DistLM forward) conducted on the variance of the meiofaunal abundance, biomass, expected richness of taxa and taxonomic composition in canyons indicated that canyons geomorphological characteristics significantly explained the variance of meiofaunal assemblages, cumulatively accounting for ca. 42% for the abundance, 50% for the biomass and 40% taxonomic composition (also after presence/absence transformation) and more than 56% for the richness of taxa. dbRDA analyses indicate that the geomorphological characteristics explained, cumulatively for the first 2 axis, 39% of the variance in the meiofaunal taxonomic composition and 35% when the data are presence/absence transformed (Figure 2.5).

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Figure 2.3. Regression analyses of meiofaunal abundance (A), biomass (B) and richness of higher taxa (C) against depth in Mediterranean and NE Atlantic canyons and open slopes. Only significant regressions are indicated.

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Figure 2.4. Regression analyses of meiofaunal abundance (A), biomass (B) and richness of higher taxa (C) against longitude in Mediterranean canyons and open slopes. Significant regressions are indicated.

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Figure 2.5. dbRDA ordination after DistLM forward analysis, describing the relationship between the canyons geomorphological characteristics and meiofaunal taxonomic composition (A), also after presence/absence transformation (B).

www.msfd-idem.eu Deliverable 2.2 Megafaunal biodiversity and assemblage composition The analyses of published datasets and our own data allow identifying a total of 400 species in the whole Mediterranean area. Overall, the most abundant fraction of deep-sea megafauna was accounted by the following species number: 127 teleost fishes, 27 cartilaginous fishes, 64 decapod crustaceans and 28 cephalopods. Deep-sea fishes were dominated by myctophids (12 species) and macrourids (10 species), while among sharks the two catsharks Galeus melastomus and Scyliorhinus canicula and the lanternshark Etmopterus spinax were the most frequently observed species. The two red shrimps Aristeus antennatus and Aristaemorpha foliacea, the oplophorid Acanthephyra eximia and the sergestid Sergia robusta were the dominant decapod species, while among cephalopods, the jewel squid Histioteuthis reversa and the two sepiolids Heteroteuthis dispar and Neorossia caroli were the most commonly encountered species. Other megafaunal species included in the sampling were echinoderms with 25 species, and anthozoans (23 species). Although the analysis of macrofauna was out of the scope of the present study, the gears/instruments here reviewed, collected also some macrofaunal taxa represented mostly by gastropods (24 species), bivalves (28 species), and polychaetes (21 species). At sub-basin level 20 cephalopod species were collected in the western basin, 61 decapods, 12 elasmobranchs, and 96 fishes. In the central basin, 15 cephalopods, 44 decapods, 6 elasmobranchs and 57 fishes were caught during the surveys. Finally, in the Eastern Mediterranean 14 cephalopod species were recorded, together with 30 decapod, 23 elasmobranch and 89 fish species (Figure 2.6).

Figure 2.6. Percentage of the main megafaunal taxa recorded in the three Mediterranean sub-basins.

Effects of the sampling equipment Overall the ANOSIM test showed a significant effect of factor “gear” (R=0.29, p<0.001). The pair-wise comparisons proved significant differences among all kinds of trawl gears used (Maireta Trawl System, OTSB-14, and commercial trawl-gear) and those collected/recorded by baited traps or cameras (all pairwise p<0.001, with an average R=0.57). Conversely, pairwise comparisons did not show any significant difference among all trawl gears and also with samples collected with Agassiz dredge (p>0.05). Samples recorded by ROV significantly separated from all the others (all pairwise p<0.001, with an average R=0.69). Thus the following analyses were performed considering two separate datasets: 1) samples collected by trawl gears and Agassiz and 2) samples collected/recorded by traps and baited cameras. ROV samples, being only from a single survey carried out in the Western Mediterranean (Ayma et al. 2016), for a total of 10 dives, were not used in the subsequent analyses.

www.msfd-idem.eu Deliverable 2.2 Megafauna collected by traps/baited cameras and recorded by ROV Traps and baited cameras attracted mostly scavenger species and 44 taxa were identified. Among them, crustaceans accounted for the largest number of species, with decapods (i.e. deep-sea crab Geryon longipes and the ophlophorid shrimp Acanthephyra eximia, accounting for 40% of the total number of individuals recorded), being dominant together with lyssianassid amphipods (i.e. Scopelocheirus spp., Orchomene grimaldi, Hippomedon bidentatus, 31% of the total number). Among fishes, sharks were frequently observed with the gulper shark Centrophorus granulosus and the Portuguese dogfish Centroscymnus coelolepis as the most abundant recorded species, observed together in the 38% of the samples. The bluntnose sixgill shark Hexanchus griseus was encountered in 17% of the samples (10 specimens). The macrourid Coryphaenoides mediterraneus was the most abundant bony fish, being recorded in 27% of the samples in high numbers (up to 26 individuals). The CAP ordination carried out on the Jaccard resemblance matrix of traps/baited cameras data, showed a clear separation of samples as a function of the geographic area/sub-basin (Figure 2.7A). The PERMANOVA test carried out on a two-factor nested design, showed that both factors were significant (sub-basin: pseudo-F2,46=2.63, p=0.009; bathymetric range(sub-basin): pseudo-F4,46=3.19, p=0.0001).

Figure 2.7. Canonical Analysis of Principal coordinates (CAP) output showing the taxonomic composition of megafaunal assemblages sampled by traps and baited cameras (A) and trawls and Agassiz dredge (B) at different bathymetric ranges and in the Mediterranean Sea sub-basins.

www.msfd-idem.eu Deliverable 2.2 The SIMPER analysis showed that decapod crustaceans were the species that allowed to distinguish the deep-sea assemblages collected/observed by traps/baited cameras at all depths (see Fanelli et al., 2018; Table 1A). Deep-sea decapods mostly dominated the lower slope assemblage down to 3000 m of depth, while the macrourid Coryphaenoides mediterraneus were observed at the abyssal stations (i.e. 3000-4500 m; see Fanelli et al., 2018; Supplementary Table 1 for sampling maximum depth limit). Different species characterized the deep-sea assemblages in the three basins (see Fanelli et al., 2018; Table 1B), with C. mediterraneus essentially observed only in the eastern basin and C. coelolepis in the western basin, the total dissimilarity between the two datasets accounted for more than 90%. The megafaunal assemblages of the central Mediterranean were essentially dominated by the European conger Conger conger. High dissimilarities (97-100%) were also observed with the assemblages recorded in the central Mediterranean, due to the shallower deployment of the baited cameras here (450-1000 m). Overall the α-diversity was similar in the Eastern and Western sub-basins (on average H’=1.26±0.45 in the WMED vs. 1.25±0.50 in the EMED), with a similar increasing trend from 1000-2000 m (on average H’=0.96±0.26 in the WMED vs. 1.44±0.23 in the EMED) to 2001-3000 m of depth (on average H’=1.33±0.49 in the WMED vs. 1.63±0.23 in the EMED). In the central Mediterranean, the Shannon diversity (H´) of the megafaunal assemblages was 1.31 (±0.88) at 450-1000 m of depth. The univariate PERMANOVA test provided significant differences for factor depth (pseudo-F4,46=6.82, p<0.001). The PERMDISP test showed greater multivariate dispersion of samples in the western Mediterranean than in the eastern (t=4.86, p<0.001) and the central Mediterranean (t=4.74, p<0.01). No differences existed between the central and the eastern Mediterranean basin in terms of multivariate dispersion (t=1.90, p=0.34). The average distance of centroid for samples from the western Mediterranean is higher than that of samples from the central and the eastern Mediterranean (in the case of Jaccard distance: WMED=60.60±1.31 vs. CMED=43.52±5.12 vs. EMED=51.06±1.41). Thirty-four taxa were identified by ROV images in the three canyons (Blanes, Cap de Creus and La Fonera) of the Western Mediterranean, among them of ecological relevance resulted the occurrence of the deep- sea carnivore ascidian Dicopia anthirrinhum, in one dive at 1240 m.

Assemblage composition of megafauna collected by trawls The PERMANOVA test, carried out on the resemblance matrix of species collected by both trawls and Agassiz and excluding the abyssal stations (see Fanelli et al., 2018, for further details), provided significant differences for both bathymetric range and sub-basin and the interaction term (see Fanelli et al., 2018; Table 2). The CAP ordination showed a clear separation of samples as function of depth (Figure 2.8B). The SIMPER analysis carried out separately for both factors showed the most typifying species for each basin and depth ranges (see Fanelli et al., 2018; Table 3A and B). Particularly Sergestes arcticus, Zeus faber, Plesionika acanthonotus and Aristeus antennatus were the most typifying species in the Western basin contributed 22% to the similarity, while in the Central Mediterranean A. antennatus, the sergestid Sergia robusta and A. eximia accounted for 37% of the total similarity. In the Eastern Mediterranean still A. antennatus, Polycheles typhlops and the blackmouth catshark G. melastomus were dominant. These three species accounted for 28% of the total similarity. Overall decapods were the dominant species throughout the wide bathymetric range explored in the whole basin (200-4000 m of depth, Table 3B), with only few fish species contributing to the similarity in the upper slope assemblage (200-700 m), such as Z. faber, Trachurus mediterraneus and Scyliorhinus canicula in the upper slope (200-700 m) and Lampanyctus crocodilus in the upper-middle slope (700-900 m). The deepest assemblage (3300-4000 m) was characterized by the two ophlophorid shrimps A. eximia and A. pelagica and S. robusta, accounting for 72% of the similarity.

www.msfd-idem.eu Deliverable 2.2 Overall the α-diversity (in terms of ES(200)) was greater on the upper slope (Figure 2.8), decreasing with depth. In the western basin high α -diversity were also observed on the middle slope (900-1200 m). No significant differences were observed at sub-basin level (p>0.05), while did for factor “bathymetric range” (pseudo-F3,159=1.98, p=0.045).

Figure 2.8. Megafaunal α-diversity reported as Expected Species number ES(200) (A) and β-diversity reported as PERMDISP results (B) at different bathymetric ranges and in different Mediterranean Sea sub- basins. In (B), also the dissimilarity % among the Mediterranean Sea sub-basins is reported.

PERMDISP test evidenced greater dispersion of samples in the Western Mediterranean than in the other sub-basins (Jaccard resemblance, F2,157=26.75, p<0.01; pairwise comparisons: WMED≠CMED=EMED). The average distance from centroids followed a Western Mediterranean>Eastern Mediterranean>Central Mediterranean trend.

Influence of canyon features/attributes on assemblage composition The DISTLM run on the overall assemblage showed that some canyon attributes may contribute to the observed pattern of species composition (Table 4). Four variables, i.e. canyon slope, dendricity, the distance to closest canyon centroid (measured in Km) and canyon sinuosity explained 17% of the total variance. At sub-basin level, any geomorphological variables explained the assemblage composition at CMED, while at EMED the distance to closest canyon centroid explained 28% of the total variance. At WMED canyon slope explained 5% of the total variance.

www.msfd-idem.eu Deliverable 2.2 Hard substrates communities The results of the meta-analysis indicated up to more than 10 taxonomic groups in the considered habitats (Figure 2.9).

Figure 2.9. Number of taxa observed in each hard-substrate habitat.

The PERMANOVA analysis showed significant differences among habitats and water depth ranges, however the nMDS (Figure 2.10) clearly indicated that such differences may be attributed to the different biotic components analyzed in each scientific paper.

Transform: Presence/absence Resemblance: S7 Jaccard 2D Stress: 0.01 Habitat rocky ground muddy coral pockmark open slope Chondrichthyes + Osteichthyes canyon corals DendrophylliaMadrepora oculata cornigera canyon

Lophelia pertusa

TurbellariaOstracoda Dictyonella sp. 1 Nematoda HexadellaPoecillastraHamacanthaRhabderemia cf. compressa dedritifera falcula sp. CopepodaKinorhyncha + nauplii Cephalopods, crustaceans and fishes ChondrichthyesOsteichthyesDecapoda EunicellaCephalopodaC. verticillata cavoliniiS.pallidaB.mollis Annelida HelicolemusViminellaPagellusGaleusPhycis flagellum dactylopterus melastomus bogeraveoblennoides MerlucciusConger merluccius conger

Savalia savaglia

Figure 2.10. MDS conducted on the taxonomic composition of hard-substrate communities. Data have been presence/absence transformed before the analyses. 18

www.msfd-idem.eu Deliverable 2.2 Conclusions Meiofauna Meiofaunal assemblages and biodiversity in deep Mediterranean canyons The results of our meta-analysis indicate weak differences in meiofaunal abundance and biomass between canyons and slopes at all depth ranges in each investigated Mediterranean continental margin. Moreover, our data report similar or even higher values of expected richness of taxa in canyons than in the adjacent slopes in most cases and at all depth ranges, confirming that habitat diversification (the alternation of canyons and open slopes along continental margins) allows diversified assemblages, increasing the overall diversity at regional and basin scale (Bianchelli et al., 2010; Zeppilli et al., 2016). In this regard, we also observed similar expected richness of taxa in Mediterranean Sea and in the NE Atlantic margins at all bathymetric ranges, both in canyons and open slopes, confirming that the deep Mediterranean Sea is not biodiversity depleted (Bianchelli et al., 2010; Coll et al., 2010; Danovaro et al., 2010). The Mediterranean Sea, indeed, is recognized as hot spot of biodiversity, with a highly heterogeneous distribution of taxonomic groups in the different regions, comprising meiobenthic assemblages in deep-sea habitats (Coll et al., 2010). The meta-analysis conducted on the meiofaunal taxonomic composition revealed high dissimilarity levels in the meiofaunal assemblages between canyons and open slopes in the Mediterranean Sea, with % dissimilarity even higher when data were presence/absence transformed. These results suggest that in Mediterranean canyons specific faunal assemblages may be preferentially observed, and this is true not only for macro- and mega- (e.g., benthic megafauna, corals, crustaceans, fishes, mammals, Özturk et al., 2012; Vella and Vella, 2012; Watremez, 2012; David and Di-Méglio, 2012; Madurell et al., 2012; Company et al., 2012; Baro et al., 2012), but also for meiofauna. This specificity in assemblages occurrence in the Mediterranean canyons is also confirmed by the high levels of dissimilarity of meiofaunal taxonomic composition between NE Atlantic and Mediterranean canyons. Such differences increase at increasing water depth. Overall, these results suggest that sea floor heterogeneity allows diversified assemblages across different habitats and margins, increasing the overall diversity at regional and basin scale. The specific features of canyon fauna has been previously reported from several Mediterranean margins, in particular on canyons off the Alboran, Catalan, Malta, and Turkey coasts (Özturk et al., 2012; Vella and Vella, 2012; Watremez, 2012; David and Di-Méglio, 2012; Madurell et al., 2012; Company et al., 2012; Baro et al., 2012). In the Mediterranean, one of the most relevant aspects is the relationship between the number of species, the number of individuals of endemic species and the ecological features of submarine canyons (Palanques et al., 2005). Indeed, no individual canyon is identical to another, and this is reflected by differences in fauna between canyons even located along the same continental margin, also for meiofaunal assemblages (Hecker, 1990; Rogers et al., 2002). Regarding the habitat heterogeneity, the most complex Mediterranean region is the Catalan margin, one of the areas of the world oceans with the higher canyon density per 100 km (Würtz, 2012b). In this region, it has been demonstrated an isolation effect related to peculiar hydrodynamic processes, canyon morphology and ecological differences, which has led to high faunal diversification and even speciation processes, particularly for Foix, Lacaze-Duthiers, La Fonera (Palamòs) and Planier canyons (Gili et al., 1999; Palanques et al., 2005; Würtz, 2012b).

Patterns and environmental drivers of meiofaunal assemblages in deep Mediterranean habitats The regression analyses revealed that meiofaunal abundance, biomass and richness of taxa display a significant negative relationship with water depth in both Mediterranean canyons and slopes, whereas no significant relationships were observed in the NE Atlantic margin. We report here that the decline of meiofaunal biomass with increasing water depth is more evident in open slopes than in canyons, suggesting that the environmental constrains related to water depth (e.g., increasing pressure, decrease

www.msfd-idem.eu Deliverable 2.2 of food supply with the increasing water depth) may have a different effect depending on the deep-sea habitat considered. Moreover, the decline of meiofaunal biomass with increasing water depth is more evident than the decline of abundance only in open slopes, suggesting that environmental (or even trophic or anthropogenic) constrains acting inside canyons (as example, massive and pulsing food supply, deep currents and tides characteristics) can determine a shift of the size of individuals, leading to a differential miniaturization at greatest depths in canyons and in open slopes (Pusceddu et al., 2013; 2014; Gambi et al., 2017). At basin scale, our results indicate that meiofaunal assemblages were characterised also by significant differences among the investigated deep Mediterranean regions. Indeed, at basin spatial scale, meiofaunal abundance, biomass and richness of taxa displayed clear decreasing patterns from Western to Eastern deep Mediterranean Sea. This gradient in meiofaunal variables has been repeatedly observed in the deep Mediterranean basin (Danovaro et al., 2008; Gambi and Danovaro, 2006; Gambi et al., 2014) and reflects the differences in the trophic conditions already observed in previous studies, also in the deep-sea sediments (Danovaro et al., 1999; Gambi and Danovaro, 2006; Bianchelli et al., 2010; Pusceddu et al., 2010). Here, the decreasing patterns of meiofaunal abundance, biomass and richness of taxa from Western to Eastern Mediterranean Sea has been observed both for canyons and open slopes, suggesting that, at basin scale, one of the main drivers for deep-sea benthic assemblages variables is the trophic status of the region, whatever the considered habitat (Bianchelli et al., 2010; Pusceddu et al., 2010).

Influence of canyons geomorphological characteristics on meiofaunal assemblages The multivariate multiple regression analysis showed that in canyon systems, their geomorphological characteristics are responsible for the observed variability among such complex and peculiar habitats. During the last decade the intense exploration of the deep ocean, along with the refinement of the available technologies has revealed the presence of a wide variety of different geomorphological features of the submarine canyons, resulting in a wide variety of topographic structures also at small spatial scales (i.e., within each canyon). In this regard, due to their high spatial and temporal variability in their morphological, hydrographic and sedimentological characteristics, canyons have been recently recognized as extreme environments (Zeppilli et al., 2018), since they comprise complex, highly heterogeneous environments that encompasses a patchwork of environmental and trophic conditions with different degrees of stability (Tyler et al. 2009; Amaro et al. 2016). Faunas inhabiting such complex environments are thus influenced by multiple factors, which include hydrodynamic conditions and current regimes, topography and habitat heterogeneity, amount, origin and quality of sedimentary organic matter, sedimentation processes and turbidity events and anthropogenic impacts (Danovaro et al., 1999; Baguley et al., 2006; Ingels et al., 2009; 2011a; b; 2013; Pusceddu et al., 2014; Ramalho et al., 2014; Amaro et al., 2016; Román et al., 2016; Thistle et al., 2017). Among the investigated drivers, previous studies revealed the strong influence of trophic resources in shaping meiobenthic assemblages in deep-sea canyons (Soltwedel et al., 2005; Ingels et al., 2011a; b; 2013) and indicated that up to the 30% of the variability in assemblages traits was explained by environmental variables, including the amount and nutritional quality of sedimentary organic matter (Bianchelli et al., 2010; Román et al., 2016). The results of the present study indicate that ca. 40% of the variability in the meiofaunal taxonomic composition was explained by the geomorphological characteristics of canyons. In particular, beside the water depth, the characteristics more influencing the meiofaunal taxonomic composition are class (shelf- incising with correlation - or not - to river systems or blind canyons), density (number of canyons in the same area) and status (thalweg or tributary systems) of the canyons. Conversely, other characteristics (such as sinuosity, top and low height), even having a significant effect, explained lower % of the

www.msfd-idem.eu Deliverable 2.2 taxonomic composition variability. Even more in details, different geomorphological characteristic are responsible for the presence/absence of the taxa, whereas others are responsible for their composition. Overall, our data suggest that the complex combination of geomorphological characteristics is responsible for high % of the taxonomic composition variability but also suggest that most of the variance depends on variables not considered in the present study (supposedly, sediment features and trophic resources; Soltwedel et al., 2005; Bianchelli et al., 2010; Ingels et al., 2011a; b; 2013; Román et al., 2016). The role of the seafloor heterogeneity in shaping meiobenthic assemblages has also been observed for other deep seabed morphologies, which are inhabited by different meiofaunal assemblages in term of taxonomic composition and may influence faunal distribution more than other factors as trophic resources (Zeppilli et al., 2016). Deep-sea canyons have been repeatedly proposed as biodiversity hotspots (Vetter and Dayton, 1998; Curdia et al., 2004; Ingels et al., 2009; Bianchelli et al., 2010; Amaro et al., 2016; Román et al., 2016). Data presented here indicate that this is particularly true for the Mediterranean Sea, where canyons are typically different from any other marine region worldwide from a geomorphological point of view (Harris and Whiteway, 2011). Our meta-analysis also demonstrates that deep-sea Mediterranean canyons contribute significantly to enhance the deep-sea biodiversity at both regional and whole-basin scale thus representing crucially important deep-sea habitats deserving appropriate protection.

Megafauna Overall deep-sea megafaunal composition Until recently, our knowledge on deep-sea megafaunal communities was still scant, due extremely high heterogeneity of continental margins, characterized by the presence of seamounts, deep-sea coral banks and submarine canyons. Among these, the rough topography of canyons has affected our ability to collect samples and data (Company et al., 2012). Only recently, the coupling of different traditional gears such as otter and Agassiz trawls and new technologies such as ROVs (Remote Operated Vehicles), allowed to better understand the biodiversity of deep-sea megafaunal assemblages (Company et al., 2012; Tecchio et al., 2011a). However, the recent studies were confined in specific areas, such as the Catalan margin, due to their geomorphological, hydrodynamic, ecological and economic importance (Company et al., 2008; Tecchio et al., 2011a; 2013; Mecho et al., 2017; Ayma et al., 2018). In this study, for the first time, we compiled the information on deep-sea megafauna collected in the last 20 years in different surveys in the Mediterranean Sea, at broad basin level, integrating information collected using different gears/instruments (bottom trawls, Agassiz, traps, baited cameras and ROV) and testing the effect of different longitudinal geographic scales, bathymetric gradients and habitats geomorphological characteristics in structuring the megafaunal assemblages. Overall, the results of our analysis highlighted the high levels of biodiversity characterizing the megafaunal assemblages in the Mediterranean Sea, with a total of more than 400 species identified, and specifically 28 species of cephalopods, 64 of decapod crustaceans, 27 elasmobranches, 127 fishes and 25 echinoderms. Moreover, the Mediterranean Sea differs from other deep-sea oceans also in its species composition, due to its historical, chemical-physical and trophic characteristics which allow/prevent colonization depending on the geological period (Coll et al., 2010). In general, fish species richness was lower in the Mediterranean than in the deep Atlantic (Massuti et al., 2004), and this has also been observed for other megafaunal groups such as asteroids (Sibuet, 1979). Conversely, decapod crustaceans diversity was similar to that observed in the North-Eastern Atlantic, displaying also greater abundance and biomass values than those reported for eutrophic regions in the North Atlantic (Markle et al., 1988). For these reasons, the Mediterranean Sea, indeed, is recognized as hot spot of biodiversity, at least for some taxa, with a highly heterogeneous distribution of taxonomic groups in the different sub-basins (Coll

www.msfd-idem.eu Deliverable 2.2 et al., 2010), comprising megafaunal assemblages in deep-sea habitats. Indeed, our results indicated that, beside the high levels of biodiversity, megafaunal assemblages of the Mediterranean Sea are also characterized by high variability at different spatial scales, depending on the habitat complexity and heterogeneity as well as geographical, trophic and bathymetric characteristics of the Mediterranean Sea (Gambi et al., 2017).

Longitudinal and bathymetric gradients of deep-sea megafauna biodiversity As expected, there is a strong “gear” effects on the estimate of deep-sea biodiversity, with samples obtained by trawl-surveys and also employed the Agassiz dredge, clearly separated from those collected by traps or recorded by baited cameras. A part from the lower diversity of these latter samples, the faunistic composition was remarkably different with the exception of few common species, such as the ophlophorid shrimp Acanthephyra eximia. This species was also dominant in the abyssal assemblages and has been thought to have functionally replaced the giant amphipod Eurythenes gryllus, the dominant deep-sea scavenging crustacean throughout most of the world's oceans (Christiansen, 1989). Overall traps and baited cameras allowed recording some uncommon species, difficult to collect with other methods, such as trawls. This is the case of some deep-sea sharks such as the Portuguese dogfish Centroscymnus coelolepis, the gulper shark Centrophorus granulosus and the bluntnose six-gill shark Hexanchus griseus. As the majority of deep-sea sharks, they are considered highly vulnerable species due to their K-selected life-history characteristics and very low capacity for recovery from overfishing (Stevens et al., 2000), thus the use of less-invasive methods can be improved in the future to monitor deep-sea sharks and to analyze their spatial distribution, in particular in the Mediterranean Sea (McLean et al., 2015). Specifically, baited cameras and ROV can be very effective, as the majority of deep-sea species were not influenced by the presence of the ROV (Ayma et al., 2016). In general, we observed significant longitudinal differences in assemblage composition, with species substitution throughout the Mediterranean basin. At a higher taxonomic level, while the number of species within some groups, such as bony fishes or cephalopods, were similar along the west-east gradient, others, such as decapods and sharks changed drastically, showing greater species diversity of decapods vs. sharks in the western basin and the opposite trend in the Eastern basin, where 24 elasmobranchs species were recorded. As deep-sea sharks are usually caught as by-catch by different type of fishing gears, i.e. trawl-nets long-lines, the higher diversity of elasmobranchs recorded here, is attributable to the lower fishing pressure in the Eastern Mediterranean (D’Onghia et al., 2005), as in many areas, deep-sea fisheries targeting red shrimps is in its infancy (Farrag, 2016). Changes in species composition could be linked to the different food availability existing throughout the Mediterranean basin, with a longitudinal west-to-east decrease in productivity, being the Eastern Mediterranean one of the most oligotrophic ocean areas in the World (Azov, 1991), with extreme scarcity of sinking organic matter in the water column. Still, the influence of grain size composition with finer sediments in the Western Mediterranean has been also invoked as responsible of such differences (Tecchio et al., 2011a), as they present greater percentages of particulate organic matter and thus higher food availability, which favours the settlement of a more abundant and diverse infauna (Levin et al. 2001), as echinoderms or bivalves. However, the β -diversity remained quite constant throughout the basin, as species replacement, more than a west-to-east decrease in species abundance seemed to be the main pattern of deep-sea biodiversity of demersal assemblages. In terms of β -diversity, this was considerably greater in the western Mediterranean, comparing with the other two sub-basins. This suggests that the factors determining species turnover in deep Mediterranean Sea act at local and regional, more than at large spatial scales (western vs. central and eastern basins). The higher food supply of the western Mediterranean (Danovaro

www.msfd-idem.eu Deliverable 2.2 et al., 1999; Tecchio et al. 2011), together with greater habitat heterogeneity related in turn to different water masses on the bottom can be the cause of this greater β -diversity. The western Mediterranean is affected by periodic cascading events of shelf dense waters (Canals et al. 2006), while the hydrodynamics of the eastern basin are less variable and intense (Bethoux et al. 1999). The high spatial heterogeneity of canyons (see below) can also be responsible for such a high species turnover (β -diversity). In addition, although it is expected that deep-sea β-diversity across areas at the same depth is lower than that occurring with increasing bathymetric depth, an exception was found in the upper bathyal zone under strong coastal influence (McClain and Rex, 2015). The north-western Mediterranean area is, by far, the most densely populated (UN/MAP/BP/RAC 2005), industrialized and intensively impacted area of the whole Mediterranean basin (Danovaro et al., 2010), and this can have contributed to the patterns observed here. Overall, our analysis confirmed that water depth was the main factor influencing the structure of the deep-sea macro- and megafaunal demersal assemblages collected using trawls and Agassiz gears (Stefanescu et al., 1993; Moranta et al., 1998; Fanelli et al., 2013). Patterns of species distributions with increasing bathymetric of the NW Mediterranean depth followed a hump-shaped curve, a pattern repeatedly observed in other deep-sea oceanic regions (Gage and Tyler, 1991; Rex and Etter, 2010) as well as in the Balearic basin (Fanelli et al., 2013). Conversely, in both the central and the eastern Mediterranean, megafaunal biodiversity decreased from the upper slope to the lower slope and the abyssal plain (200-700 m, below 1200 m and 3300-4000.m depth respectively), which seems to be a common feature in the deep ocean, but only below 1600-2500m depth (Gage and Tyler,1991; Atlantic: Haedrich et al., 1980, Gordon and Duncan, 1985; Pacific waters: Pearcy et al.,1982). According to different previous studies conducted in the Mediterranean (Stefanescu et al., 1993), our results confirmed that the depth range 1200-1600 m is a transitional zone where an important species turnover occurs. A transitional zone at depths comprised from 1200-1300 m to 1200-1600 m was also reported in the north Atlantic, reflecting differences in local topography and geomorphology (Hecker, 1900). However, according to previous studies (Cartes and Sardà, 1993), in the Mediterranean Sea such a transitional zone appears uplifted in its deepest bathymetry by ca. 200 m (i.e., ranging from 1200-1400 m). Considering the very limited variability of the most environmental variables in the deep water of the Mediterranean Sea, the patterns of megafaunal abundance and diversity, especially for the most abundant groups such as fishes and decapods, has been attributed to changes in quantity and quality of the trophic resources, above and below the depth of 1200 m. Our results indicate that this holds true also for the distribution of mesopelagic organisms and benthic/suprabenthic fauna, which represent the main prey for the dominant predators up to deep-sea sharks (Fanelli & Cartes, 2008; Fanelli et al., 2009; Cartes et al., 2014; Morales-Nin et al., 1996; Carrassón and Matallanas, 2001).

Influence of canyon features on deep-sea community structure Canyons morphologic and oceanographic features are assumed to influence faunal characteristics (Ramirez-Llodra et al., 2009; Danovaro et al., 2010), in particular fishery resources (Stefanescu et al., 1994; Sardà et al., 2009), thus providing us important ecosystem goods and services (Fernandez-Arcaya et al., 2017). Canyon-driven upwelling can enhance the local primary productivity, thus increasing the amount food resources available to euphausiids, mesopelagic fishes, shrimps and squids, which in turn attract a variety of top pelagic and bentho-pelagic predators (Würtz, 2012). Submarine canyons are important conveyors of organic matter and promote bentho-pelagic coupling (Thomsen et al., 2017), which feeds diversified benthic macro- and megafauna assemblages and bento-nekton species of commercial interest (Company et al., 2008; Fernandez-Arcaya et al., 2017). In addition, canyons can provide spawning and

www.msfd-idem.eu Deliverable 2.2 recruitment areas for several types of fauna (Sardà et al., 1994; 2009). As a result, the pelagic food chain is extended and intensive pelagic and demersal fisheries are commonly located at the heads of submarine canyons, down to 800 m depth (Würtz, 2012). Moreover, recent studies have highlighted the so-called “canyon effect”, i.e. a spill-over of species on the adjacent slope (Ramirez-Llodra et al., 2009). In this study, the multivariate multiple regression analysis showed that canyons geomorphological characteristics (canyon slope, dendricity, sinuosity and distance from adjacent canyons) explained a significant portion of the variance of the taxonomic composition of megafaunal assemblages. Our results, also indicate that the relative importance of canyon morphology changed across Mediterranean sub- basins, explaining cumulatively only 5% of variance in the Western Mediterranean, whereas reached 28% in the Eastern Mediterranean. The low percentage of variance explained in the Western Mediterranean Sea, which is one of the areas of the world oceans with the higher canyon density per 100 km (Würtz, 2012), suggests that other environmental and anthropogenic constrains can play here a more important role in shaping the structure of megafaunal assemblages (Puig et al., 2012).

Conclusions During the last decade, the increasing and expanding deep-ocean exploration has revealed the presence of a wide variety of different geomorphological features within deep-sea habitats, resulting in a wide variety of topographic features at different spatial scales. This study confirmed that the high biodiversity of meio- macro- and megafaunal assemblages is associated to canyons and is linked to their topographic complexity thus confirming to the so-called “canyon effect” for the Mediterranean continental margin. We also indirectly highlighted the importance of the food availability in relation to water depth as a major structuring factor for these macro- megafaunal assemblages in the deep Mediterranean Sea. Further, the hump-shaped curve of the bathymetric patterns of macro-megafaunal diversity here found, is similar to that reported worldwide, but with highest values at shallower depths (ca. 1200 m depth in spite of the 1600-2500 m depth of the Atlantic Ocean). We also showed that the abundance of some megafaunal species, such as deep-sea sharks, can be completely overlooked by using traditional gears. Less invasive methodologies (i.e., baited cameras, landers, ROVs and AUVs), could be used to preserve these species (killed by using baited traps) and the vulnerable deep-sea ecosystems in which habitat forming species, such as deep-water corals, can be damaged by using trawling. Future surveys and technologies implementation together with rigorous and standardized sampling approaches are desirable to obtain a full picture of Mediterranean deep-sea megafaunal biodiversity and assemblage structure.

www.msfd-idem.eu Deliverable 2.2 References

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www.msfd-idem.eu Deliverable 2.2 slopes of the Portuguese, Catalan, Southern Adriatic and Cretan Sea margins. Deep-Sea Research I 57, 441-457. DOI:10.1016/j.dsr.2009.11.008 Pusceddu, A., Bianchelli, S., Martín, J., Puig, P., Palanques, A., Masqué, P., 2014. Chronic and intensive bottom trawling impairs deep-sea biodiversity and ecosystem functioning. Proceedings of the National Academy of Sciences 111 (24), 8861-8866. DOI: https://doi.org/10.1073/pnas.1405454111 Pusceddu, A., Gambi, C., Zeppilli, D., Bianchelli, S., Danovaro, R., 2009. Organic matter composition, metazoan meiofauna, and nematode biodiversity in sediments of the deep Mediterranean Sea. Deep Sea Research II 56, 755-762. DOI: https://doi.org/10.1016/j.dsr2.2008.10.012 Pusceddu, A., Mea, M., Canals, M., Heussner, S., Durrieu De Madron, X., Sanchez-Vidal, A., Bianchelli, S., Corinaldesi, C., Dell'Anno, A., Thomsen, L., Danovaro, R., 2013. Major consequences of an intense dense shelf water cascading event on deep-sea benthic trophic conditions and meiofaunal biodiversity. Biogeosciences 10, 2659-2670. DOI: https://doi.org/10.5194/bg-10-2659-2013 Ramalho, S.P., Adão, H., Kiriakoulakis, K., Wolff, G.A., Vanreusel, A., Ingels, J., 2014. Temporal and spatial variation in the Nazaré Canyon (Western Iberian margin): inter-annual and canyon heterogeneity effects on meiofauna biomass and diversity. Deep-Sea Research I 83, 102-114. DOI: https://doi.org/10.1016/j.dsr.2013.09.010 Ramirez-Llodra, E., Company, J.B., Sardà, F., Rotllant, G. 2010. Megabenthic diversity patterns and community structure of the Blanes submarine canyon and adjacent slope in the Northwestern Mediterranean: A human overprint? Marine Ecology 31(1), 167-182. DOI:10.1111/j.1439- 0485.2009.00336.x Ramirez-Llodra, E., Company, J.B., Sardà, F., Rotllant, G., 2009. Megabenthic diversity patterns and community structure of the Blanes submarine canyon and adjacent slope in the Northwestern Mediterranean: A human overprint? Marine Ecology 31, 1-16. DOI:10.1111/j.1439-0485.2009.00336.x Ramirez-Llodra, E., Tyler, P.A., Baker, M.C., Bergstad, O.A., Clark, M.R., et al., 2011. Man and the last great wilderness: human impact on the deep sea. PLoS ONE 6(7): e22588. DOI:10.1371/journal.pone.0022588 Rex, M.A., Etter, R.J., 2010. Deep-Sea Biodiversity. In: Ormond, R., Gage, J.D. (Eds.), Marine Biodiversity: Causes and Consequences. Cambridge University Press, pp. 94–121 Harvard University Press. Rex, M.A., Etter, R.J., Morris, J.S., Crouse, J., McClain, C.R., Johnson, N.A., Stuart, C.T., Deming, J.W., Thies, R., Avery, R., 2016. Global bathymetric patterns of standing stock and body size in the deep-sea benthos. Marine Ecology Progress Series 317, 1-8. DOI: 10.3354/meps317001 Riaux-Gobin, C., Dinet, A., Dugué, G., Vétion, G., Maria, E., Grémare, A., 2004. Phytodetritus at the sediment-water interface, NW Mediterranean Basin: spatial repartition, living cells signatures, meiofaunal relationships. Scientia Marina 68 (1), 7-21. DOI: https://doi.org/10.3989/scimar.2004.68n17 Ricklefs, R.E., Schluter, D. (eds.), 1993. Species Diversity in Ecological Communities: Historical and Geographical Perspectives. University of Chicago Press Rogers, A., Billett, D.S.M., Berger, W., Flach, E., Freiwald, A., Gage, J., Hebbeln, D., Heip, C., Pfannkuche, O., Ramirez-Llodra, E., Medlin, L., Sibuet, M., Soetaert, K., Tendal, O., Vanreusel, A., Wlodarska- Kowalczuk, M., 2003. Life at the edge: Achieving prediction from environmental variability and biological variety. In: Wefer, G., Billett, D.S.M., Hebbeln, D., Jørgensen, B.B., Schlüter, M., van Weering, T.C.E. (eds.), Ocean Margin Systems. Springer Verlag, Berlin, 387-404. Román, S., Vanreusel, A., Romano, C., Ingels, J., Puig, P., Company, J.B., Martin, D., 2016. High spatiotemporal variability in meiofaunal assemblages in Blanes Canyon (NW Mediterranean) subject to anthropogenic and natural disturbances. Deep Sea Research I 117, 70-83. DOI: https://doi.org/10.1016/j.dsr.2016.10.004

www.msfd-idem.eu Deliverable 2.2 Romano, C., Coenjaerts, J., Flexas, M.M., Zúñiga, D., Vanreusel, A., Company, J.B., Martin, D., 2013. Spatial and temporal variability of meiobenthic density in the Blanes submarine canyon (NW Mediterranean). Progress in Oceanography 118, 144-158. DOI: https://doi.org/10.1016/j.pocean.2013.07.026 Rumolo, P., Cartes, J.E., Fanelli, E., Papiol, V., Sprovieri, M., Mirto, S., Gherardi, S., Bonanno, A., 2015. Seasonal variations in the source of sea bottom organic matter off Catalonia coasts (western Mediterranean): links with hydrography and biological response. Journal of Oceanography 71, 325- 343. DOI: 10.1007/s10872-015-0291-7 Sardà, F., Cartes, J.E., Company, J.B., 1994. Spatio-temporal variations in megabenthos abundance in three different habitats of the Catalan deep-sea (Western Mediterranean). Marine Biology 120(2), 211-219. DOI: 10.1007/BF00349681 Sardà, F., Cartes, J.E., Company, J.B., Albiol, A., 1998. A modified commercial trawl used to sample the deep-sea megabenthos. Fisheries Science 64: 492-493. DOI: https://doi.org/10.2331/fishsci.64.492 Sardà, F., Company, J.B., Bahamon, N., Rotllant, G., Flexas, M.M., et al., 2009. Relationship between environment and the occurrence of the deep-water rose shrimp Aristeus antennatus (Risso, 1816) in the Blanes submarine canyon (NW Mediterranean). Progress in Oceanography 82, 227-238. DOI: https://doi.org/10.1016/j.pocean.2009.07.001 Sibuet, M., 1979. Distribution and diversity of Asteroids in Atlantic abyssal basins. Sarsia 64: 85–91. Snelgrove, P.V.R., Smith, C.R., 2002. A riot of species in an environmental calm: the paradox of the species- rich deep-sea floor. Oceanography and Marine Biology: an Annual Review 40, 211-242. Soetaert, K., Heip, C., Vincx, M., 1991. The meiobenthos along a Mediterranean deep-sea transect off Calvi (Corsica) and in an adjacent canyon. Marine Ecology 12, 227-242. DOI: 10.1111/j.1439- 0485.1991.tb00255.x Soltwedel, T., 2000. Metazoan meiobenthos along continental margins: a review. Progress in Oceanography 46, 59-84. DOI: https://doi.org/10.1016/S0079-6611(00)00030-6 Soltwedel, T., Hasemann, C., Quéric, N.V., von Juterzenka, K., 2005. Gradients in activity and biomass of the small benthic biota along a channel system in the deep Western Greenland Sea. Deep Sea Research I 52, 815-835. DOI: https://doi.org/10.1016/j.dsr.2004.11.011 Soyer, J., Bodiou, J.Y., de Bovée, F., Guidi, L., 1978. Evolution quantitative du meiobenthos sur le plateau continental et la marge de la cote catalane francaise. Coll. Intern. Oceanol., Perpignan, C.I.E.S.M. (1987). Stefanescu, C., Lloris, D., Rucabado, J., 1993. Deep-sea fish assemblages in the Catalan Sea (western Mediterranean) below a depth of 1000 m. Deep-Sea Research I 40, 695-707. DOI: https://doi.org/10.1016/0967-0637(93)90066-C Stefanescu, C., Nin-Morales, B., Massuti, E., 1994. Fish assemblages on the slope in the Catalan Sea (western Mediterranean): influence of a submarine canyon. Journal of the Marine Biological Association of the United Kingdom 74 (3), 499-512. DOI: https://doi.org/10.1017/S0025315400047627 Stevens, J.D., Bonfil, R., Dulvy, N.K., Walker, P.A., 2000. The effects of fishing on sharks, rays, and chimaeras (chondrichthyans), and the implications for marine ecosystems. ICES Journal of Marine Science, 57: 476–494. DOI:10.1006/jmsc.2000.0724 Taviani, M., Angeletti, L., Canese, S., Cannas, R., Cardone, F., Cau, A., Cau, A.B., Follesa, M.C., Marchese, F., Montagna, P., Tessarolo, C. 2017. The “Sardinian cold-water coral province” in the context of the Mediterranean coral ecosystems. Deep-Sea Research Part II 145, 61-78. DOI: 10.1016/j.dsr2.2015.12.008 Tecchio, S., Ramírez-Llodra, E., Aguzzi, J., Sanchez-Vidal, A., Flexas, M.M., Sardà, F., Company, J.B., 2013. Seasonal fluctuations of deep megabenthos: Finding evidence of standing stock accumulation in a flux-

www.msfd-idem.eu Deliverable 2.2 rich continental slope. Progress in Oceanography 118, 188-198. DOI: http://dx.doi.org/10.1016/j.pocean.2013.07.015 Tecchio, S., Ramírez-Llodra, E., Sardà, F., Company, J.B., Palomera, I., Mechó, A., Pedrosa-Pàmies, R., Sanchez-Vidal, A., 2011a. Drivers of deep Mediterranean megabenthos communities along longitudinal and bathymetric gradients. Marine Ecology Progress Series 439, 181-192. DOI: 10.3354/meps09333 Tecchio, S., Ramírez-Llodra, E., Sardà, F., Company, J.B., 2011b. Biodiversity of deep-sea demersal megafauna in western and central Mediterranean basins. Scientia Marina 75(2), 341-350. DOI: 10.3989/scimar.201175n2341 Thistle, D., Sedlacek, L., Carman, K.R., Barry, J.P., 2017. Influence of habitat heterogeneity on the community structure of deep-sea harpacticoid communities from a canyon and an escarpment site on the continental rise off California. Deep Sea Research I 123, 56-61. DOI: https://doi.org/10.1016/j.dsr.2017.03.005 Thomsen L., Aguzzi J., Costa C., De Leo F., Ogston A., Purser A. 2017. The oceanic biological pump: Rapid carbon transfer to the Deep Sea during winter. Sci. Rep. 7: 10763. Tyler, P., Amaro, T., Arzola, R., Cunha, M.R., de Stigter, H., Gooday, A.J., Huvenne, V., Ingels, J., Kiriakoulakis, K., Lastras, G., Masson, D., Oliveira, A., Pattenden, A., Vanreusel, A., van Weering, T., Vitorino, J., Witte, U., Wolff, G., 2009. Europe’s grand canyon: Nazaré submarine canyon. Oceanography 22, 46–57. DOI: https://doi.org/10.5670/oceanog.2009.05 Vanreusel, A., Fonseca, G., Danovaro, R., et al., 2010. The contribution of deep-sea macrohabitat heterogeneity to global nematode diversity. Marine Ecology 31, 6-20. DOI:10.1111/j.1439- 0485.2009.00352.x Vella, A., Vella, J., 2012. Central-southern Mediterranean submarine canyons and steep slopes: role played in the distribution of cetaceans, bluefin tunas, and elasmobranchs. In: Würtz, M. (ed.), Mediterranean submarine canyons: ecology and governance. Gland, Switzerland and Málaga, Spain: IUCN, 73-88. ISBN: 978-2-8317-1469-1 Vetter, E.W., Dayton, P.K., 1998. Macrofaunal communities within and adjacent to a detritus-rich submarine canyon system. Deep Sea Research II 45, 25-54. DOI: https://doi.org/10.1016/S0967- 0645(97)00048-9 Vetter, E.W., Dayton, P.K., 1999 Organic enrichment by macrophyte detritus, and abundance patterns of megafaunal populations in submarine canyons. Marine Ecology Progress Series 186:137-148. Vetter, E.W., Smith C.R., De Leo F.C., 2010. Hawaiian hotspots: enhanced megafaunal abundance and diversity in submarine canyons on the oceanic islands of Hawaii. Marine Ecology 31, 183-199. DOI:10.1111/j.1439-0485.2009.00351.x Watremez, P., 2012. Canyon heads in the French Mediterranean Overview of results from the MEDSEACAN and CORSEACAN campaigns (2008-2010). In: Würtz, M. (ed.), Mediterranean submarine canyons: ecology and governance. Gland, Switzerland and Málaga, Spain: IUCN, 105-112. ISBN: 978-2- 8317-1469-1 Weaver, P.P.E., Billett, D.S.M., Boetius, A., Danovaro, R., Freiwald, A., Sibuet, M., 2004. Hotspot ecosystem research on Europe’s deep-ocean Margins. Oceanography 17(4), 132-143. DOI: http://dx.doi.org/10.5670/oceanog.2004.10 Würtz, M. (ed.), 2012a. Mediterranean Submarine Canyons: Ecology and Governance. Gland, Switzerland and Málaga, Spain: IUCN. 216 pages. ISBN: 978-2-8317-1469-1 Würtz, M., 2012b. Submarine canyons and their role in the Mediterranean ecosystem. In: Würtz, M. (ed.), Mediterranean submarine canyons: ecology and governance. Gland, Switzerland and Málaga, Spain: IUCN, pp. 11-26. ISBN: 978-2-8317-1469-1

www.msfd-idem.eu Deliverable 2.2 Zeppilli, D., Leduc, D., Fontanier, C., Fontaneto, D., Fuchs, S., Gooday, A.J., Goineau, A., Ingels, J., Ivanenko, V.N., Kristensen, R.M., Cardoso Neves, R., Sanchez, N., Sandulli, R., Sarrazin, J., Sørensen, M.V., Tasiemski, A., Vanreusel, A., Autret, M., Bourdonnay, L., Claireaux, M., Coquillé, V., De Wever, L., Rachel, D., Marchant, J., Toomey, L., Fernandes, D., 2018. Characteristics of meiofauna in extreme marine ecosystems: a review. Marine Biodiversity, 48, 35–71. DOI: https://doi.org/10.1007/s12526- 017-0815-z Zeppilli, D., Pusceddu, A., Trincardi, F., Danovaro, R, 2016. Seafloor heterogeneity influences the biodiversity–ecosystem functioning relationships in the deep sea. Scientific Reports 6, 26352. DOI: 10.1038/srep26352.

www.msfd-idem.eu Deliverable 2.2 3. DESCRIPTOR 2: NON-INDIGENOUS SPECIES

As task 2.1 and the related deliverable (report 2.1) evidenced the existence of very few datasets (11 papers in total and any open access repositories at the moment) for this descriptor, a meta-analysis was not carried out. Moreover, of the eleven species listed in table 3.1. of report 2.1. only two can be strictly considered deep-sea species (i.e. the swimming crabs Charybdis longicollis and Gonioinfradens giardi) collected up to 250 m and 200 m, respectively. A recent published paper (Galil et al., 2018), reported the presence at 200-m depth of individuals of three Erythraean species, collected off Ashdod during October- December 2017. More in detail authors reported the recovery of 8 specimens of the crocodile toothfish Champsodon nudivittis (typically of the Indo-Pacific Ocean), of a single individual of Golani’s round herring Etrumeus golanii (typically occurring in the upper shelf of the northern Red Sea), and of the burrowing goby, Trypauchen vagina (typically of the coastal Indo-west Pacific Ocean). This study evidenced the need of monitoring deep dwelling bioinvasions and their harm to sensitive habitats (i.e., slope) and of establishing appropriate protocols to define GES for Descriptor 2 in mesophotic habitats.

References Galil, B. S., Danovaro, R., Rothman, S. B. S., Gevili, R., & Goren, M. 2018. Invasive biota in the deep-sea Mediterranean: an emerging issue in marine conservation and management. Biological Invasions, 1-8.

www.msfd-idem.eu Deliverable 2.2 4. DESCRIPTOR 3: POPULATIONS OF ALL COMMERCIALLY EXPLOITED FISH AND SHELLFISH

Commercially exploited fish and shellfish are all living marine resources targeted for economic profit such as the bony fish, sharks and rays (known as elasmobranchs), crustacean such as lobsters and shrimps, and molluscs (including bivalves, cuttlefish and squid). It also includes other creatures such as jellyfish and starfish. In scientific terms, Descriptor 3 has various implications. Stocks should be, (1) exploited sustainably consistent with high long-term yields, (2) have full reproductive capacity in order to maintain stock biomass, and (3) the proportion of older and larger fish/shellfish should be maintained (or increased) being an indicator of a healthy stock. Good Environmental Status is achieved for a particular stock only if all of the three attributes are fulfilled. This implies that all commercially exploited stocks should be in a healthy state and that exploitation should be sustainable, yielding the Maximum Sustainable Yield (MSY). MSY is the maximum annual catch, which can be taken year after year without reducing the productivity of the fish stock. Heavy fishing pressures, such as overexploitation or overfishing, can have very negative environmental impacts. They can result in the loss of significant potential yield of the stocks being fished and can even precede severe stock depletion and fisheries collapse. Because of overfishing, fish stocks can reduce dramatically to the point where they lose internal diversity and with it, their capacity to adapt to environmental changes. Fish communities can be altered in a number of ways, for example they can decrease if particular-sized individuals of a species are targeted, as this affects predator and prey dynamics (the question of trophic relationships and marine food webs is the focus of Descriptor 4). In the Mediterranean Sea the enforcement of MSFD with the definition of GES and targets in each Member State (MS), and the concurrent application of the Common Fisheries Policy (CFP), is still far from achieving its objectives for the marine resources (e.g., Colloca et al., 2013; Vasilakopoulos et al., 2014). Notwithstanding, the enforcement of the EU-Data Collection Regulation (EU, 2000) in the early 2000s by all EU MSs, and the rapid increase in the number of assessed stocks by the General Fisheries Commission for the Mediterranean (GFCM) and the European Scientific, Technical and Economic Committee for Fisheries (STECF), Mediterranean Sea marine resources are still exploited above the levels that deliver the maximum sustainable yield and no signs of recovery are evident (Vasilakopoulos et al., 2014; Cardinale and Scarcella, 2017). Particularly, in deeper areas of Mediterranean Sea, the achievements of MSFD targets is at risk to be further delayed by the management systems currently enforced both at the national and international level (e.g., GFCM), taking into account that often the deep-water resources are distributed outside the 12 nautical miles. In the present deliverable we collated information on the Mediterranean fish stocks from relevant reports of STECF (https://stecf.jrc.ec.europa.eu/reports/medbs) and GFCM SAC (http://www.fao.org/gfcm/reports/statutory-meetings/en/), published over the period 2008–2017. Here we focus on criteria 1 and 2, since criterion 3 is still under evaluation, because of the difficulties encountered to interpret it in terms of GES.

www.msfd-idem.eu Deliverable 2.2 These reports were used to extract estimates of fishing mortality (F), fishing mortality which corresponds 1 to MSY (FMSY) and SSB (spawning stock biomass) for each demersal stock, assumed as the combination of species and GFCM Geographical Sub Areas (GSA; Figure 3.1).

Figure 3.1. Subregions of the Mediterranean according to MSFD, with the indication of GSAs as defined by the GFCM.

2 For such stocks, the ratio F/FMSY (Figures 3.2-9) and the trends of SSB (3.10-19) are presented below. In particular, F/FMSY can be considered a clear indicator of criterion 1 (exploited sustainably consistent with high long-term yields) and SSB of criterion 2 (have full reproductive capacity in order to maintain stock biomass). In term of spatial coverage, it is clear that in the Aegean-Levantine subregion the availability of stock assessments is lower than the other areas (Table 3.1). In addition, a clear overfishing (F>FMSY) for most of the stocks assessed emerged, especially in the Western and Central Mediterranean Sea and in the Ionian Sea.

1 SSB- The spawning stock biomass (SSB) is the combined weight of all individuals in a fish stock that are capable of reproducing. The SSB is an indication of the status of the stock and the reproductive capacity of the stock 2 F/FMSY- the level of fishing mortality that achieves maximum sustainable yield (MSY) over the long term based on growth and natural mortality rates, the selection pattern of the fishery and recruitment changes associated with the level of adult biomass (stock-recruitment relationship)

www.msfd-idem.eu Deliverable 2.2 Table 3.1. Number of available stock assessments and stocks in overfishing in MFSD subregion

MSFD Sub-Region N. of stock assessed N. of stocks in overfishing (F>FMSY) Aegean Levantine Sea 3 1 Adriatic Sea 7 3 Ionian Sea and Central Mediterranean 7 7 Western Mediterranean Sea 37 31

When interpreting the information presented in Table 3.1 above it is important to bear in mind that the number of stocks benefiting from an assessment in the Mediterranean is in fact very low. Only 48 of the 235 stocks exploited in the western Mediterranean benefit from a scientific evaluation (Foucher and Delaunay, 2017), and the proportion is even lower in the eastern Mediterranean.

www.msfd-idem.eu Deliverable 2.2 Criterion 1 - F/FMSY ratios

M. barbatus 8,0 7,0 6,0

5,0 MSY

4,0 F/F 3,0 2,0 1,0 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA6 GSA7 GSA9 GSA10

M. barbatus 6,0 5,0 4,0

MSY 3,0

F/F 2,0 1,0 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA15-16 GSA17-18 GSA17 GSA18 GSA19 GSA22 GSA25

Figure 3.2. Trends of F/FMSY for the red mullet Mullus barbatus for the different GSAs. Top: trends for the GSA of the MSFD subregion “Western Mediterranean”; bottom: trends for the MSFD sub-regions “Central Mediterranean and Ionian Sea” and “Aegean-Levantine basin”.

www.msfd-idem.eu Deliverable 2.2 M. merluccius 16,0 14,0 12,0 10,0

MSY 8,0

F/F 6,0 4,0 2,0 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA1-3 GSA5 GSA6 GSA7 GSA9 GSA1-5-6-7 GSA2-3-4-5 GSA9-10-11

M. merluccius 14,0 12,0 10,0

MSY 8,0

F/F 6,0 4,0 2,0 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA12-13-14-15-16 GSA17-18 GSA19 GSA22

Figure 3.3. Trends of F/FMSY for the European hake Merluccius merluccius for the different GSAs. Top: trends for the GSA of the MSFD sub-region “Western Mediterranean”; bottom: trends for the MSFD sub- regions “Central Mediterranean and Ionian Sea” and “Aegean-Levantine basin”.

www.msfd-idem.eu Deliverable 2.2 L. piscatorius 14,0 12,0 10,0

MSY 8,0

F/F 6,0 4,0 2,0 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA5 GSA6

Figure 3.4. Trends of F/FMSY for the anglerfish Lophius piscatorius for the different GSAs.

M. poutassou 14,0 12,0 10,0

MSY 8,0 F 6,0 4,0 2,0 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA6 GSA9

Figure 3.5. Trends of F/FMSY for the blue whiting Micromesistius poutassou for the different GSAs.

www.msfd-idem.eu Deliverable 2.2 A. foliacea 3,5 3,0 2,5

MSY 2,0

F/F 1,5 1,0 0,5 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA9 GSA10 GSA11 GSA18-19

Figure 3.6. Trends of F/FMSY for the giant red shrimp Aristaeomorpha foliacea for the different GSAs.

A. antennatus 9,0 8,0 7,0 6,0

MSY 5,0

F/F 4,0 3,0 2,0 1,0 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA5 GSA6 GSA9

Figure 3.7. Trends of F/FMSY for the blue and red shrimp Aristaeus antennatus for the different GSAs.

www.msfd-idem.eu Deliverable 2.2 P. longirostris 4,5 4,0 3,5 3,0

2,5 MSY

2,0 F/F 1,5 1,0 0,5 0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA5 GSA6 GSA9 GSA10 GSA19 GSA12-13-14-15-16 GSA17-18

Figure 3.8. Trends of F/FMSY for the deep-water rose shrimp Parapenaeus longirostris for the different GSAs.

N. norvegicus 10,0

8,0

6,0

MSY F/F 4,0

2,0

0,0 2008 2009 2010 2011 2012 2013 2014 2015 2016

GSA5 GSA6 GSA9 GSA11 GSA17-18

Figure 3.9. Trends of F/FMSY for the Norway lobster Nephrops norvegicus for the different GSAs. A conditional rule in MSFD assessments is the 'one-out-all-out (OOAO) rule, which specifies that all variables have to achieve good status. Based on the data shown in Figures 3.2-3.9 it is clear that most of the assessed stocks are not in good status since they are not being exploited sustainably consistent with high long-term yields. Moreover since at least one overexploited species is present in each of the MSFD sub- regions the various sub-regions are not in a good status either when considering commercially exploited species of fish and shellfish.

www.msfd-idem.eu Deliverable 2.2 Criterion 2 - Spawning stock biomass trends

Spawning stock biomass data extracted from STECF and GFCM assessments (Figures 3.10 - 3.19 below) reveal that SSB trends either do not show trends or show a decreasing trend. This indicates that reproductive capacity is either not improving or declining, which is particularly concerning for stocks which are overexploited according to Criterion 1. However the present analysis shows that significant knowledge gaps remain; hardly any of the species for which data was extracted from stock assessments had SSB information available for the entire 2009-2016 timeseries.

The precautionary limit MSY B trigger, below which there is a high risk that reproductive capacity is impaired, has been widely used in MSFD assessments in Northern Europe. Since MSY B trigger reference points are generally not available in the Mediterranean Sea, this criterion cannot be fully applied to Mediterranean stocks.

www.msfd-idem.eu Deliverable 2.2 M. barbatus

2500

2000

1500 SSB

1000

500

0 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA5 GSA6 GSA7 GSA9 GSA11

M. barbatus 14000 12000 10000

8000 SSB 6000 4000 2000 0 2009 2010 2011 2012 2013 2014 2015 2016

GSA19 GSA15-16 GSA17-18

Figure 3.10. Trends of SSB for the red mullet Mullus barbatus for the different GSAs. Top: trends for the GSA of the MSFD subregion “Western Mediterranean”; bottom: trends for the MSFD sub-region “Central Mediterranean and Ionian Sea”.

www.msfd-idem.eu Deliverable 2.2 M. merluccius 50000

40000

30000 SSB

20000

10000

0 2009 2010 2011 2012 2013 2014 2015 2016

GSA1-5-6-7 GSA9-10-11 GSA17-18 GSA19

Figure 3.11. Trends of SSB for the European hake Merluccius merluccius for the different GSAs.

M. surmuletus 6000

5000

4000

SSB 3000

2000

1000

0 2009 2010 2011 2012 2013 2014 2015 2016

GSA5 GSA9 GSA15-16

Figure 3.12. Trends of SSB for the striped red mullet Mullus surmuletus for the different GSAs.

www.msfd-idem.eu Deliverable 2.2 L. piscatorius and L. budegassa (LB) 2500

2000

1500 SSB

1000

500

0 2009 2010 2011 2012 2013 2014 2015 2016

GSA1-5-6-7 GSA5 LB GSA6 LB GSA15-16 LB

Figure 3.13. Trends of SSB for the anglerfish Lophius piscatorius and for the blackbellied angler L. budegassa for the different GSAs.

M. poutassou 1200

1000

800

SSB 600

400

200

0 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA6 GSA9

Figure 3.14. Trends of SSB for the blue whiting Micromesistius poutassou for the different GSAs.

www.msfd-idem.eu Deliverable 2.2 P. erythrinus (PE) and T. minutus (TM) 1400 1200 1000

800 SSB 600 400 200 0 2009 2010 2011 2012 2013 2014 2015 2016

PE_GSA15-16 TM_GSA9

Figure 3.15. Trends of SSB for the common Pandora Pagellus erythrinus (PE) and the poor cod Trisopterus minutus (TM) for the different GSAs.

A. foliacea 7000 6000 5000

4000 SSB 3000 2000 1000 0 2009 2010 2011 2012 2013 2014 2015 2016

GSA9 GSA10 GSA11 GSA12-13-14-15-16 GSA18-19

Figure 3.16. Trends of SSB for the giant red shrimp Aristaeomorpha foliacea for the different GSAs.

www.msfd-idem.eu Deliverable 2.2 A. antennatus 5000

4000

3000 SSB

2000

1000

0 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA6

Figure 3.17. Trends of SSB for the blue and red shrimp Aristaeus antennatus for the different GSAs.

P. longirostris 12000

10000

8000 SSB 6000

4000

2000

0 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA5 GSA6 GSA9-10-11 GSA17-18-19

Figure 3.18. Trends of SSB for the deep-water rose shrimp Parapenaeus longirostris for the different GSAs.

www.msfd-idem.eu Deliverable 2.2 N. norvegicus 4000 3500 3000 2500

SSB 2000 1500 1000 500 0 2009 2010 2011 2012 2013 2014 2015 2016

GSA1 GSA5 GSA9 GSA11 GSA15-16

Figure 3.19. Trends of SSB for the Norway lobster Nephrops norvegicus for the different GSAs.

Criterion 3 - Healthy age and size structure

The assumption behind this criterion is that a stock with sufficient large and therefore old fish is healthy and thus in a good status. The more large/oder fish there are in a stock the healthier the stock is considered to be. This criterion has not yet been well developed and no accepted threshold for GES has to date been accepted by the scientific community for this criterion (European Environment Agency, 2018). Moreover no work on thresholds which would be suitable for the deep-sea in the Mediterranean has been conducted to date.

Way Forward

The results of this first assessment of environmental status based on Descriptor 3 indicators and criteria for the deep-sea shows that none of the MSFD sub-regions can currently be considered to be in a Good Environmental Satus (GES). However many knowledge gaps and challenges need to be addressed to improve the accuracy of future Descriptor 3 assessments of the deep Mediterranean Sea:

• More stock assessments are urgently needed for deep-water Mediterranean species. A list of exploited deep-sea stocks should be compiled at sub-region level, and assessments covering a significant proportion of landings (e.g. >70%) should be carried out. • The timeseries coverage of SSB trend data should be improved, and suitable proxies to qualitatively estimate biomass against reference point proxies available in the Mediterranean Sea should be investigated. For species where SSB estimates based on stock assessments are not available biomass indices based for instance on MEDITS survey data could be analysed instead. • Suitable thresholds for the population and age size distribution indicator should be identified and specifically tested for the deep-sea.

www.msfd-idem.eu Deliverable 2.2 References

Cardinale M and Scarcella G (2017) Mediterranean Sea: A Failure of the European Fisheries Management System. Front. Mar. Sci. 4:72. doi: 10.3389/fmars.2017.00072 Colloca, F., Cardinale, M., Maynou, F., Giannoulaki, M., Scarcella, G., Jenko, K., et al. (2013). Rebuilding Mediterranean fisheries: a new paradigm for ecological sustainability. Fish Fish. 14, 89–109. doi: 10.1111/j.1467-2979.2011.00453.x EEA (2018). European Environment Agency. Status of marine fish and shellfish in European seas. Indicator Assessment Prod-ID: IND-13-en. CSI 032 , MAR 007. Accessed 14/10/2018 at: https://www.eea.europa.eu/data-and-maps/indicators/status-of-marine-fish-stocks-3/assessment Foucher, E., Delaunay, D., 2017. Evaluation 2018 du bon état écologique des espèces exploitées à des fins commerciales au titre du descripteur 3 de la DCSMM. EU (2000). Council Regulation (EC) No 1543/2000 of 29 June 2000 Establishing a Community Framework for the Collection and Management of the Data Needed to Conduct the Common Fisheries Policy, EU. Vasilakopoulos, P., Maravelias, C. D., Tserpes, G. (2014). The alarming decline of mediterranean fish stocks. Curr. Biol. 24, 1643–1648. doi: 10.1016/j.cub.2014.05.070

www.msfd-idem.eu Deliverable 2.2 5. DESCRIPTOR 4: ECOSYSTEMS, INCLUDING FOOD WEBS

Descriptor 4 refers to marine food webs. The descriptor “concerns important functional aspects such as energy flows and the structure of food webs (size and abundance)” (European Commission, 2010; 2010/477/EU). However, this descriptor is probably the most challenging as food webs are complex and difficult to depict especially at their low trophic levels. Among the eleven descriptors of GES, Descriptor 4 is generally categorized as a “state” descriptor (vs. the “Pressure” descriptors D5, D6, D8-D11), while food webs better describe a function of the ecosystem rather then its state. This controversial issue makes the definition of indicators for marine food webs complicate and even more for deep-sea food webs where the availability of data is rather fragmented, with several both bathymetric and latitudinal/longitudinal gaps. Most of the datasets concerning Descriptor 4 are related to two different approaches to the study of marine food webs, as already highlighted in Report 2.1: (1) data from stomach contents, stable isotope analyses and/or fatty acid trophic markers (this latter to a lesser extent) and (2) data from modelling techniques, such as Ecopath with Ecosim (EwE). The present meta-analysis considered mostly the first datasets and specifically data from stable isotope analysis (SIA), which may comply with at least two of the criteria established by the COMM DEC 848/2017, i.e. the primary criterion D4C1 (diversity of trophic guilds) and the secondary criterion D4C3: distribution of individuals across the trophic guild (D4C3). Datasets regarding modelling approaches (essentially EwE) were discarded as this kind of approach considers metadata and produces discrete outputs which can be integrated with other data (see figure 4.1) to, such as biomasses of the species/trophic guilds, data from SIA or stomach contents analysis, to provide useful outputs for D4, but not used per se in fulfilling MSFD D4 criteria.

Figure 4.1. Integration and analysis of different data (biomass, stomach contents data, SIA) through modelling, to generate useful inputs for D4 (Courtesy of Simone Libralato, OGS, Trieste, Italy). 54

www.msfd-idem.eu Deliverable 2.2 A statistical meta-analysis was not performed in the case of D4 because a comparison between a “treatment”, i.e. a pressure acting on the ecosystem, such as bottom trawling, and a “control” situation was not possible. Although the overall dataset encompassed both data from fishing grounds (i.e. up to 1000 m of depth) and “pristine areas” (at least from trawling activities) below 1000 m of depth, thus virtually a somewhat comparison can be done, the own nature of isotopic enrichment with depth prevent from such an evaluation. Indeed, different authors observed an increase in 15N 1 and 13C values of POM, and in turn of species, with depth (Mintenbeck et al., 2005; Fanelli et al., 2013). A total of 22 papers (see appendix 4.1) were considered for the meta-analysis, among them 16 concerns specimens and potential food sources collected in the western Mediterranean, essentially from the Balearic-Catalan basin. Only four and three papers, presented isotopic data of species collected in the Central and the Eastern Mediterranean, respectively. Deep-sea food webs, with few exceptions represented by chemosynthetic environments (i.e., hydrothermal vents, cold seeps, whale and wood falls) sustained by in situ primary production by chemoautotroph organisms (), are mostly supported by inputs of organic matter from the surface, via vertical flux, or from the continental shelf, through lateral advection (Figure 4.2). Further to the 'passive' flux of sinking particles (mostly detritus, faecal pellets and phytodetritus), a portion of organic matter reaches the depths through the daily vertical migrations of macrozooplankton and micronekton, with few mesozooplanktonic species also contributing to deep-sea food webs through excretion and mortality (Miquel et al. 1994) or being prey of deep-sea species (Vinogradov 1970, Vinogradov and Tseitlin 1983).

Figure 4.2. Schematic representation of the main food inputs to the bathyal food webs in the Mediterranean (Courtesy of Vanesa Papiol, UNAM Merida, Mexico).

www.msfd-idem.eu Deliverable 2.2 The advective flux, especially through submarine canyons, transports terrigenous and shelf material at bathyal depths, generally with a seasonal periodicity (Fanelli et al., 2011) or enhanced by oceanographic processes such as the dense shelf water cascading (Company et al. 2008).

Not all the datasets presented data on potential food sources for deep-sea food webs, i.e. terrestrial materials or POM in water/sediments as the baseline of the food webs, however at least at basin level it was possible to identify a baseline of the food web for comparisons among basins.

All values were entered in an excel matrix, ordered by sub-basin (Western Mediterranean – WMED, Central Mediterranean –CMED and Eastern Mediterranean –EMED), depth range (i.e. upper –US, middle -MS, lower slope -LS, and abyssal plane -AP) and group of species according to high level taxonomy (fish, decapods, cephalopods etc.) or behaviour (i.e. benthopelagic fish, nektobenthic decapods etc.). The trophic level (TL) of deep-sea species was estimated based on their 15N data, using a POM feeder (i.e., Pyrosoma atlanticum) as reference material (Post, 2002). This was selected, among other POM feeders, being analysed in the three sub-basins. 15N values were converted to TL on the basisof the assumption that there is a fractionation of 3.4‰ per trophic level and that the base materialhas a TL 2: 15 15 TLi =( Ni - Nref)/3.4)+2 15 15 15 15 where TLi is the trophic level of species i,  Ni is the mean  N of species i, and  Nref is the mean  N of the reference material.

Overall a total of 183 species, 7 genera, 10 families and 5 other taxonomic groups (i.e. copepod, cumaceans as a whole) were analysed, along with bulk micro-, meso- and macrozooplankton and suprabenthos in the three sub-basins at different depths from the middle slope to the abyssal plain (see Appendix 4.2).

These data allowed to generate two kinds of 13C-15N scatterplots: one focused on depth assemblages (Figure 4.3) and the second based the different groups of species/behaviour (Figure 4.4a-c).

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Figure 4.3. 13C-15N scatterplots of species and their potential food sources analysed in each sub-basin, according to their depth range of sampling. The first graph offered an overview of data distribution, with the large amount of data coming from the western Mediterranean. However, in the three sub-basins, although to a different extent, all depths were covered by stable isotopes data and the typical distribution of 13C-15N along the food webs were observed starting from potential food sources. Overall deep-sea Mediterranean food webs, despite the evident differences in the number of species analysed and also compartments (i.e. any studies reported stable isotopes data for benthic species from the Eastern Mediterranean), showed similar structure throughout the three sub-basins (Figure 4.4). Both nektobenthic and cartilaginous fishes occupied the top of the food web with zooplankton (from micro- to macro) located at the base.

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Figure 4.4. 13C-15N scatterplots of species categorized according to their taxonomy (fishes, decapods, cephalopods), behaviour (nektobenthic, mesopelagic) and position with respect to the seafloor (zooplankton, suprabenthos and benthos) and their potential food sources divided by type of source (POM from riverine inputs – POM river - or collected in the deep – POMwat deep, POM analysed from shelf and deep-sea sediments – POMsed_shelf and POMsed_deep or SOM, leaves of Posidonia oceanica or terrestrial remains, both pieces of wood or leaves, collected from deep-sea bottoms, used as proxy of marine or terrestrial inputs) collected in the Western (a), Central (b) and Eastern (c) Mediterranean sub- basins.

Nektobenthic and benthopelagic decapods (i.e. essentially of the family Sergestidae together with Gennadas elegans), and cephalopods occupy intermediate trophic levels, confirming previous observations related to the north-western basin (Papiol et al., 2013; Fanelli et al., 2013). Benthic species are located at different trophic levels, as “benthos” per se is a very large category and here encompasses both sessile and mobile species (i.e. anthozoans and sponges vs. echinoderms), but also species belonging to different trophic guilds, from subsurface deposit feeders, which generally displayed very high 15N and 13C values (Fanelli et al., 2011), due to high remineralization of the organic matter they ingest, to suspension feeders (such as bivalves or ophiuroids with low 15N and 13C values). The only exception to this general pattern is represented by some samples of deep-sea zooplankton from the Eastern basin, which displayed high 15N values coupled to quite low 13C (Figure 4.4c). These samples were collected in the abyssal plain, at 3000-4550 m of depths (Tecchio et al., 2013), and the high values observed are consistent with the enrichment of POM with depth found by Mintebeck et al. (2005).

www.msfd-idem.eu Deliverable 2.2 However, it is not clear, why such high 15N values was not incorporated through the food webs and further investigations would need in this sense. The PERMANOVA analysis carried out on a two-factors crossed design (sub-basin, with three levels, WMED, CMED and EMED, and depth, with five levels, continental shelf -CS, upper -US, middle –MS and lower slope - LS- and abyssal plain -AP), showed significant differences in the isotopic composition of species across sub-basins, depths and for the interaction term (Table 4.1).

Table 4.1. Results of the PERMANOVA carried out on the stable isotopic composition (15N and 13C) of species collected in the three sub-basins at different depth ranges. Source df MS Pseudo-F sub-basin 2 291.18 43.12*** depth 4 24.54 3.63** sub-basinxdepth 8 58.23 8.62*** Res 700 6.75 Total 714

In the case of the western Mediterranean (Figure 4.4a) for which more data on potential food sources are available, also from riverine inputs (Darnaude et al., 2004), some considerations on potential food sources for deep-sea food webs can be done. Despite this area is characterised by several submarine canyons, and thus it would be strongly affected from river discharge and the transport of terrestrial material though these conduits, as observed at mesoscale (i.e. the Catalan sea: Fanelli et al., 2011), when all the data are pooled together, the main contributors seemed to be represented by vertical fluxes (measured by POM in both waters and sediment). Terrestrial inputs represented by river POM and remains of terrestrial plants, seemed not sustain this food web.

Conclusions The meta-analysis fully complied with the first D4-MSFD ctiterion, i.e. D4C1 - diversity of trophic guilds, as the outputs of the stable isotope analysis carried out on micro- (essentially microzooplankton as a whole), macro- (both macrozooplankton and macrobenthos) and megafauna (mostly demersal species with few mesopelagic species analysed and limited to the western basin) allows to define different trophic guilds, essentially distributed in four trophic levels. The lowest trophic level (TL=2) comprised mostly small zooplanktonic and micronektonic species, such as the hatchet fish Argyropelecus hemigymnus, which rely on the pelagic food webs (i.e. the vertical flux, based on the low 13C values, spanning from ca. -24‰ to -20‰). At the same trophic level also below some suprabenthic species (i.e. the mysid Boreomysis arctica or the amphipod Lepechinella manco, which are known as omnivores species with a preference for phytodetritus; Fanelli et al., 2009). The second trophic level (TL=3) encompassed the bulk of the species, including benthic suspension-feeders, i.e. sponges, cold-water corals and bivalves, low trophic level carnivores, such as mesopelagic fishes (i.e. Myctophidae and Stomiiformes), nektonbenthic and benthopelagic decapod crustaceans, mostly feedeing on small benthic and zooplannktonic prey, respectively. The bulk of suprabenthic species analysed in the western Mediterranean belonged to this trophic level and comprised omnivores and low-trophic level carnivores, as well as euphausiids (i.e. Meganycthiphanes norvegica, Euphausia kronii and Nematoscelis megalops, alternatively acting as omnivores or carnivores depending on food availability. Demersal fishes, sharks, rays and Chimaera monstruosa occupied the third and fourth trophic levels (TL=4 and 5). All these species essentially feed on large benthic prey or other fish, being macrourids, i.e. Nezumia spp. and 59

www.msfd-idem.eu Deliverable 2.2 Coryphenoides mediterraneus, together with the Greater forkbeard Phycis blennoides, which typically showed a benthic diet (Papiol et al., 2014) located at the highest trophic level (TL=4.5-5). Also some carnivore polychaetes of the family Oenonidae and Nephthydae occupied the highest trophic level, confirming the high complexity of benthic food web (Fanelli et al., 2011). Surprisingly some surface deposit feeders such as the holothuroid Mesothuria intestinalis from the western sub-basin and microzooplankton (e.g. POM feeders) from the abyssal plain of the eastern Mediterranean belonged to TL 4. This apparent discrepancy has been already discussed by Fanelli et al. (2011), where the authors related the high 15N values pbserved for this species with the depth range of collection between 1400 and 1600m, which in turn is linked with higher 15N of POM (Polunin et al., 2001; Mintenbeck et al., 2007). Similarly the high 15N values observed for micro- and mesozooplankton at the abyssal plain and lower slope of the Eastern Mediterranean can be related to the excretion of nitrogen depleted in 15N with the corresponding enrichment in 15N of the residual material during bacterial degradation of POM (Macko et al., 1986). Microbial consumption is thus reflected in an increase of POM 15N with depth.

References

Fanelli E., Cartes J.E., Rumolo P., Sprovieri M. (2009) Food web structure and trophodynamics of mesopelagic-suprabenthic deep sea macrofauna of the Algerian basin (Western Mediterranean) based on stable isotopes of carbon and nitrogen. Deep Sea Research I, 56: 1504-1520. Fanelli E, Papiol V., Cartes J.E., Rumolo P., Brunet C., Sprovieri M. (2011). Food web structure of the megabenthic, invertebrate epifauna on the Catalan slope (NW Mediterranean): evidence from 13C and 15N analysis. Deep-sea Research I 58: 98-109. Macko, S.A., Fogel Estep, M.L., Engel, M.H., Hare, P.E., 1986. Kinetic fractionation of stable nitrogen isotopes during amino acid transamination. Geochimica et Cosmochimica Acta 50, 2143. Mintenbeck, K., Jacob, U., Knust,R., Arntz, W.E., Brey,T., 2007. Depth-dependence in stable isotope ratio 15N of benthic POM consumers: the role of particle dynamics and organism trophic guild. Deep ea Research I 54(6), 1015–1023. Papiol V., Cartes J.E., Fanelli, E. (2014). Regulation of the reproductive cycles of benthopelagic fish on northwest Mediterranean continental slopes by food availability and prey partitioning among species. Limnology & Oceanography 59(5): 1779–1794. Polunin, N.V.C., Morales-Nin, B., Herod, W., Cartes, J.E., Pinnegar, J.K., Moranta, J., 2001. Feeding relationships in Mediterranean bathyal assemblages elucidated by carbon and nitrogen stable-isotope data. Marine Ecology Progress Series 220, 13–23. Post, D.M., 2002. Using stable isotopes to estimate trophic position: models, methods, and assumptions. Ecology 83 (3), 703–718.

www.msfd-idem.eu Deliverable 2.2

Appendix 4.1. List of the papers used for the meta-analysis, the indication of the sub-basin of reference and of the specific area covered by the study is also shown: W= Western, CI= Central/Ionian, AL= Aegean/Levantine, Med=Whole Mediterranean. For further details on the D4 dataset, please refer to Report 2.1.

Papers Sub- Basin Area covered

Carlier, A., Le Guilloux, E., Olu, K., Sarrazin, J., Mastrototaro, F., Taviani, M., Clavier, J., 2009. Trophic relationships in a deep CI Santa maria di Leuca CWC province Mediterranean cold-water coral bank (Santa Maria di Leuca, Ionian Sea), Marine Ecology-Progress Series 397, 125-137, 361. Cartes, J.E., Fanelli, E., Kapiris, K., Bayhan, Y.K., Ligas, A., López-Pérez, C., Murenu, M., Papiol, V., Rumolo, P., Scarcella, G., 2014. Balearic basin, Algerian basin, Tyrrhenian Spatial variability in the trophic ecology and biology of the deep-sea shrimp Aristaeomorpha foliacea in the Mediterranean Sea, Deep Sea Med Sea (off Rome and Livorno), Gulf of Research Part I: Oceanographic Research Papers 87, Supplement C, 1-13, Cagliari, Sicily channel, Ionian Sea (off Cresson, P., Fabri, M.C., Bouchoucha, M., Brach Papa, C., Chavanon, F., Jadaud, A., Knoery, J., Miralles, F., Cossa, D., 2014. Mercury in W Gulf of Lyon organisms from the Northwestern Mediterranean slope: Importance of food sources, Science of the Total Environment 497-498, 229-238, Cresson, P., Fabri, M.C., Miralles, F.M., Dufour, J.-L., Elleboode, R., Sevin, K., Mahé, K., Bouchoucha, M., 2016. Variability of PCB burden W - in 5 fish and sharks species of the French Mediterranean continental slope, Environmental Pollution 212, 374-381, Darnaude A.M., Salen-Picard C., Harmelin-Vivien M.L., 2004. Depth variation in terrestrial particulate organic matter exploitation by marine W Rhone river delta coastal benthic communities off the Rhone River delta (NW Mediterranean)MEPS 275:47-57 Fanelli, E., Cartes, J.E., 2008. Spatio-temporal changes in gut contents and stable isotopes in two deep Mediterranean pandalids: Balearic basin (Soller and Cabrera island W influence on the reproductive cycle, Marine Ecology Progress Series 355, 219-233, off Mallorca) Fanelli, E., Cartes, J.E., Rumolo, P., Sprovieri, M., 2009a. Food-web structure and trophodynamics of mesopelagic–suprabenthic bathyal macrofauna of the Algerian Basin based on stable isotopes of carbon and nitrogen, Deep Sea Research Part I: Oceanographic Research W Balearic basin (Archipelago Cabrera) Papers 56, 9, 1504-1520, Fanelli E., Cartes, J.E., 2010. Temporal variations in the feeding habits and trophic levels of three deep-sea demersal fishes from the Algerian Basin off the Archipelago of western Mediterranean Sea, based on stomach contents and stable isotope analyses, Marine Ecology Progress Series 402, 213–232 W Cabrera Fanelli, E., Cartes, J.E., Papiol, V., 2011a. Food web structure of deep-sea macrozooplankton and micronekton off the Catalan slope: W Catalan sea Insight from stable isotopes, Journal of Marine Systems 87, 1, 79-89. Fanelli, E., Papiol, V., Cartes, J.E., Rumolo, P., Brunet, C., Sprovieri, M., 2011b. Food web structure of the epibenthic and infaunal invertebrates on the Catalan slope (NW Mediterranean): Evidence from delta(13)C and delta(15)N analysis, Deep-Sea Research I 58, 1, W Catalan Sea 98-109, Fanelli, E., Cartes, J.E., Papiol, V., 2012. Assemblage structure and trophic ecology of deep-sea demersal cephalopods in the Balearic W Catalan-Balearic basin basin (NW Mediterranean), Marine and Freshwater Research 63, 3, 264-274, Fanelli, E., Papiol, V., Cartes, J.E., Rumolo, P., Lopez-Perez, C., 2013b. Trophic webs of deep-sea megafauna on mainland and insular W Catalan Sea and Algerian basin slopes of the NW Mediterranean: a comparison by stable isotope analysis, Marine Ecology Progress Series 490, 199-221, Koppelmann, R., Bottger-Schnack, R., Mobius, J., Weikert, H., 2009. Trophic relationships of zooplankton in the eastern Mediterranean 4 stations within a transect from Crete to AL based on stable isotope measurements, Journal of Plankton Research 31, 6, 669-686, Cyprus Madurell, T., Fanelli, E., Cartes, J.E., 2008. Isotopic composition of carbon and nitrogen of suprabenthic fauna in the NW Balearic Islands W off Soller, NW of Mallorca (Balearic Islands) (western Mediterranean), Journal of Marine Systems 71, 3, 336-345, Papiol, V., Cartes, J.E., Fanelli, E., Rumolo, P., 2013. Food web structure and seasonality of slope megafauna in the NW Mediterranean W Catalan sea elucidated by stable isotopes: Relationship with available food sources, Journal of Sea Research 77, 53-69, Papiol, V., Cartes, J.E., Fanelli, E., 2014. The role of food availability in regulating the feeding dynamics and reproductive cycles of bathyal W Catalan sea benthopelagic fish in the northwest Mediterranean slope, Limnology and Oceanography 59, 5, 1779-1794, Rumolo, P., Cartes, J.E., Fanelli, E., Papiol, V., Sprovieri, M., Mirto, S., Gherardi, S., Bonanno, A., 2015. Seasonal variations in the source of sea bottom organic matter off Catalonia coasts (western Mediterranean): links with hydrography and biological response, Journal of W Catalan Sea Oceanography 71, 4, 325-343, Sinopoli M., Fanelli E., D’Anna G., Badalamenti F., Pipitone C. (2012) Assessing the effects of a trawling ban on diet and trophic level of CI Northern Sicily hake, Merluccius merluccius , in the southern Tyrrhenian Sea. Scientia marina 76(4): 677-690 Tecchio, S., van Oevelen, D., Soetaert, K., Navarro, J., Ramirez-Llodra, E., 2013b. Trophic Dynamics of Deep-Sea Megabenthos Are Med three sampling sites one for each sub-basin Mediated by Surface Productivity, Plos One 8, 5, e63796 Valls, Maria & Olivar, M & Fernández de Puelles, Mª & Molı,́ Balbina & Bernal, A & Sweeting, Christopher. (2014). Trophic structure of mesopelagic fishes in the western Mediterranean based on stable isotopes of carbon and nitrogen. Journal of Marine Systems. 138. W 10.1016/j.jmarsys.2014.04.007. Valls, M., Rueda, L., Quetglas, A., 2017. Feeding strategies and resource partitioning among elasmobranchs and cephalopods in W Algerian-Balearic basin Mediterranean deep-sea ecosystems, Deep-Sea Research Part I-Oceanographic Research Papers 128, 28-41,

www.msfd-idem.eu Deliverable 2.2 Appendix 4.2. List of the taxon (species, genus or family level) analysed for stable isotopes in each sub- basin. The code indicates the category they belong to as shown in figure 4.4. WMED=Western Mediterranean; CMED=Central Mediterranean; EMED=Eastern Mediterranean. B=benthos; Z=zooplankton (from micro to macro); S=suprabenthos; CEPH=cephalopods; NBD= nektobenthic decapods; BPD=benthopelagic decapods; NBF=nektobenthic fishes; MF= mesopelagic fishes; CF=cartilagineous fishes. Taxon analysed per COD WME CME EM Taxon analysed per COD WME EME sub-basin E D D ED sub-basin E D CMED D PORIFERA ARTHROPODA Hyalinema thompsoni B X Mysida Pachastrella monilifera B X Anchialina agilis S X Poecillastra compressa B X Boreomysis arctica S X Polymastia tissieri B X Boreomysis megalops S X CNIDARIA Eucopia henseni S X Hydrozoa Z Lophogaster typicus S X Abylopsis tetragona Z X Mysideis parva S X Chelophyes appendiculata Z X Euphausiacea Scyphozoa Euphausia kronii Z X Meganycthiphanes Pelagia noctiluca Z X norvegica Z X Nematoscelis Peryphilla peryphilla Z X megalops Z X Anthozoa Euphausiidae Z X Desmophyllum dianthus B X Decapoda Hormathia alba B X Acantephyra eximia NBD X X X Isidella elongata B X Acantephyra pelagica NBD X X Leiopathes glaberrima B X Alpheus glaber NBD X Aristaeomorpha Lophelia pertusa B X foliacea NBD X X X Madrepora oculata B X Aristeus antennatus NBD X X X Paramuricea cf. macrospina B X Calocarides coronatus NBD X MOLLUSCA Calocaris macandreae B X Chaceon Bivalvia mediterraneus NBD X X Abra longicallus B X Gennadas elegans BPD X Asperarca nodulosa B X Geryon longipes NBD X Delectopecten vitreus B X Monodaeus couchii NBD X Gastropoda Munida tenuimana NBD X Aporrhais serresianus B X Nematocarcinus exilis NBD X X X 62

www.msfd-idem.eu Deliverable 2.2 Cymbulia peroni Z X Nephrops norvegicus NBD X Cephalopoda Paromola cuvieri NBD X CEP Pasiphaea Abralia verany H X multidentata BPD X X Ancistroteuthis CEP lichsteini H X Pasiphaea spp. NBD X Bathypolipus CEP Plesionika sponsalis H X acanthonotus NBD X X CEP Plesionika Eledone cirrhosa H X heterocarpus NBD X CEP Heteroteuthis dispar H X Plesionika martia NBD X CEP Histioteuthis bonelli H X Polycheles typhlops NBD X X X CEP Pontophilus Histioteuthis reversa H X norvegicus NBD X CEP Illex coindetii H X Rochinia rissoana NBD X CEP Loligo forbesi H X Sergestes arcticus BPD X CEP Neorossia caroli H X Sergestes corniculum BPD X X CEP Octopus salutii H X Sergestes henseni BPD X Opisthoteuthis CEP calypso H X Sergia robusta BPD X X Pteroctopus CEP tetracirrhus H X Stereomastis sculpta NBD X CEP Rondeletiola minor H X Sergestidae BPD X CEP Rossia macrosoma H X ECHINODERMATA CEP Scaeurgus unicirrhus H X Asteroidea CEP Ceramaster Sepia orbignyana H X grenadiensis B X X CEP Sepietta oweniana H X Ophiuroidea CEP Amphipholis Tetradopsis eblanae H X squamata B X CEP Todarodes sagittatus H X Amphiura chiajei B X NEMERTEA Brisingella coronata B X Cerebratulus sp. B X Echinoidea Nemertea X Bryssopsis lyrifera B X ANELLIDA Cidaris cidaris B X 63

www.msfd-idem.eu Deliverable 2.2 Bonellia viridis B X Holothuroidea Mesothuria Chirimia biceps B X intestinalis B X Eunice norvegica B X Molpadia musculus B X Pseudostichopus Labioleanira hyleni B X oculatus B X Lumbrinereis sp. B X Ypsilothuria bilineata B X Neolanira tetragona B X CHORDATA Nephtys hystricis B X Tunicata Nephtys hombergi B X Pyrosoma atlanticum Z X X Nephtys incise B X Salpa sp. Z X Nephtys paradoxa B X Vertebrata Serpula cf. vermicularis B X X Chondrichthyes Tomopteris sp. Z X Chimaera monstrosa CF X Aphroditidae B X Dalatias licha CF X X Capitellidae B X Dipturus oxyrhinchus CF X Maldanidae B X Etmopterus spinax CF X X Oenonidae B X Galeus melastomus CF X X Polynoidae B X Hexanchus griseus CF X SIPUNCULA Raja clavata CF X Aspidosiphon muelleri B X Scyliorhinus canicula CF X Sipunculus nudus B X Osteichthyes ARTHROPODA Arctozenus risso MF X X Argyropelecus Chelicerata hemigymnus MF X X X Bathypterois Pycnogonida mediterraneus NBF X X X Bathypallenopsis scoparia B X Benthosema glaciale MF X Crustacea Cataethyx alleni NBF X Ostracoda Cataethyx laticeps NBF X X X Ceratoscopelus Philine sp. Z X maderensis MF X Chalinura Cirripedia mediterranea MF X X X Scalpellum scalpellum B Chauliodus sloani MF X X Coelorhynchus Copepoda labiatus NBF X Coelorhynchus Candacia tenuimana Z X mediterraneus NBF X Coryphaenoides Candacia sp. Z X guentheri NBF X X Calanus helgolandicus Z X Cyclothone braueri MF X 64

www.msfd-idem.eu Deliverable 2.2 Neoscolecithrix sp. Z X Cyclothone pygmea MF X Copepoda Z X Diaphus holti MF X Cumacea Electrona risso MF X Campylaspis sulcata S X Evermanella balbo MF X Campylaspis Helycolenus verrucosa S X dactylopterus NBF X Cyclaspis Hoplostethus longicaudata S X mediterraneus NBF X Platysimphus typicus S X Hygophum benoiti MF X Procampylaspis armata S X Hygophum hygomi MF X Hymenocephalus Cumacea S X italicus NBF X Lampanyctus Amphipoda crocodilus NBF X X Gammaridea Lampanyctus pusillus NBF X Andaniexis mimonectes S X Lepidion lepidion NBF X X X Bathymedon longirostris S X Lepidorhombus boscii NBF X Bruzelia typica S X Lobianchia dofleini MF X Epimeria parasitica S X Lobianchia gemellari MF X Harpinia spp. S X Maurolicus muelleri NBF X Melanostigma Lepechinella manco S X atlanticum NBF X Nicippe tumida S X Merluccius merluccius NBF X X Micromesistius Rhachotropis caeca S X poutassou NBF X Rhachotropis grimaldii S X Mora moro NBF X X Rhachotropis integricauda S X Myctophidae MF X Myctophum Rhachotropis rostrata S X punctatum MF X Nettastoma Scopelocheirus hopei S X melanurum NBF X X X Stegocephaloides christianensis S X Nezumia aequalis NBF X Nezumia Syrrhoe affinis S X sclerorhynchus NBF X X X Notacanthus Tryphosites longipes S X bonapartei NBF X X Tryphosites alleni S X Notolepis rissoi NBF X Westwoodilla Notoscopelus rectirostris S X elongatus NBF X 65

www.msfd-idem.eu Deliverable 2.2 Hyperiidea Phycis blennoides NBF X X Polyacanthonotus Phonima sedentaria Z X X rissoanus NBF X Phosina semilunata Z X Stomia boa MF X X Symbolophorus Vibilia armata Z X veranyi NBF X Trachyrhynchus Phrosinidae Z X scabrus NBF X Platyscelidae Z X Leptocephalus larvae Z X Isopoda Gnathia sp. S X Munnopsurus atlanticus S X Natatolana borealis S X Isopoda unid. S X

www.msfd-idem.eu Deliverable 2.2 6. DESCRIPTOR 5: EUTROPHICATION

The work done during task 2.1 and reported in the related Deliverable (report 2.1) evidenced the existence of very few datasets related to Descriptor 5, which are included in 16 papers in total. However such papers do not cover most of the criteria defined by the MSFD. Member States have collected extensive datasets on eutrophication acquired through national monitoring programs in the framework of WFD implementation or the Regional Sea Conventions Most, but mostly cover coastal waters. Regarding deep Mediterranean Sea only data on benthic trophic status are available in the scientific literature (Pusceddu et al., 2010; 2014). Consequences of changes in the trophic status and oxygen availability on the structure of the deep-sea communities, food web and carbon fluxes are still poorly studied (Crise et al., 2015). In addition, very few information regarding pressures (monthly/seasonal variation, natural/anthropogenic sources) on nutrient dynamics and deep ocean circulation, related but not equivalent to eutrophication, are supported by a sufficient data base (Crise et al., 2015). Thus at the moment for this descriptor, a meta- analysis was not carried out.

References

Crise A., Kaberi H., Ruiz J., 2015. A MSFD complementary approach for the assessment of pressures, knowledge and data gaps in Southern European Seas: The PERSEUS experience. Marine Pollution Bulletin 95, 28–39. Pusceddu, A., Bianchelli, S., Canals, M., Sanchez-Vidal, A., Durrieu De Madron, X., Heussner, S., Lykousis, V., de Stigter, H., Trincardi, F., Danovaro, R., 2010. Organic matter in sediments of canyons and open slopes of the Portuguese, Catalan, Southern Adriatic and Cretan Sea margins. Deep-Sea Research I 57, 441–457. Pusceddu, A., Bianchelli, S., Martín, J., Puig, P., Palanques, A., Masqué, P., Danovaro, R., 2014. Chronic and intensive bottom trawling impairs deep-sea biodiversity and ecosystem functioning. Proceedings of the National Academy of Sciences 111, 8861–8866.

www.msfd-idem.eu Deliverable 2.2 7. DESCRIPTOR 6: SEAFLOOR INTEGRITY

Descriptor 6 refers to seafloor integrity. The descriptor examines the impacts of human activities on the seabed by providing an evaluation of five criteria, two concerning physical loss and disturbance, and three focused on the adverse effects caused to benthic habitats. In practice, most of the datasets referred to the deep Mediterranean Sea target only one main pressure: bottom trawling. Waste disposal in the deep Mediterranean is partly described in the literature, although most studies are out-dated. Other pressures remain almost unrevised. It should be noticed that seafloor massive sulphide (SMS) deposits, though illustrated in the figures, do not represent a pressure nowadays. They have been identified in view of the precautionary approach given the devastating effects that mining of SMS deposits could cause on the deep-sea.

Bibliography on bottom trawling is quite diverse in character, hindering direct comparisons. The meta- analysis performed within Task 2.2 requires at least 10-15 papers providing the same measures per pressure. Datasets compiled for D6 assessment do not fulfil this condition, thus preventing a sound meta- analysis. Therefore, a semi-quantitative analysis has been carried out. Results are illustrated and summarized in three figures. Figure 7.1 shows the main pressures identified in the deep Mediterranean Sea. Two boxes within Figure 7.1 show the available references on pressures and the main impacts identified as revised in different articles. The histogram about references on pressures highlights the disproportion in the number of articles between topics. It is to be noted that although ammunition dumpsters are not identified per se in the map of Figure 7.1, they have been accounted in the cables and pipelines and SMS deposits histograms, following UNEP-MAP (2009). Quantitative data on bottom trawling and waste disposal is described in Figure 7.2 following the introduction on their respective impacts in the right box of Figure 7.1.

Few D6-related impacts are identified and even less are quantitatively analyzed in the relevant literature. It should be noted that literature background information is not available for all potential pressures, either existing or future. This is the case, for instance, of ammunition dumpsters, seafloor cables or eventually intended deep-sea mining activities. Therefore, any actual or potential impact remains essentially unknown, thus preventing an accurate assessment within the frame of IDEM.

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Figure 7.1. Summary chart describing the geographical allocation of D6-related pressures and the references available concerning both pressures and impacts. The abbreviation (P) refers to pressures, with colours indicating each pressure type as presented in the map legend. The same colour code is used to illustrate Impacts with the letter (I) instead. The histogram and the percentages have been calculated using the literature database compiled for Task 2.1 D6. Ammunition dumpsters are not identified in the map but they are accounted for in the histogram. See UNEP-MAP (2009) for further information. The upper left map was adapted from IDEM Deliverable 1.1 (30/06/2018). The images illustrating different impacts were obtained from: Puig et al. (2012), Mechó et al. (2017), Fontanier et al. (2012) and Martín et al. (2014). AL: Aegean-Levantine Basin; CI: Central-Ionian Basin); W: Western Mediterranean Basin.

www.msfd-idem.eu Deliverable 2.2 Bottom trawling and waste disposal are the two pressures to which most of the available data refer to, which allow assessing their impacts on deep-sea ecosystems. Some study cases including quantitative data are illustrated in Figure 7.2. Impacts of bottom trawling are the most investigated, even though a complete evaluation of the whole Mediterranean Basin is currently lacking. Apart from one article encompassing most of the CI geographical area (Eigaard et al., 2017), most articles focus in Western Mediterranean canyons. The local character of most of these studies highlights the spatial fragmentation of the data used so far. Decreasing seabed integrity due to bottom trawling results in major ecosystemic consequences starting by physical disturbance and the alteration of seafloor properties. Direct physical loss and permanent change of seabed morphology due to abrasion has been demonstrated in submarine canyons (Martín et al., 2014; Puig et al., 2012). Physical disturbance of the seafloor alters sediment fluxes, accumulation rates and sediment resuspension pools (Martín et al., 2014; Paradís et al., 2017a). Consequently, biogeochemical properties of the sediment also change because of mixing and remobilization (Sañé et al., 2013). Information on waste disposal and its impacts in the deep Mediterranean Sea is limited, as it refers to only three cases, two of which are almost 20 years old. Consequently, part of the analyses and methods applied are out-dated. Other waste disposal activities, such as munitions, remain unstudied. Altogether bottom trawling, waste disposal and likely other undocumented pressures, impact the benthic communities and their functions, likely leading to a cascade of effects with unknown consequences. Habitat and ecosystem impacts are presented in Figure 7.3.

Descriptor 6 encompasses three criteria on the impacts on habitats, including extent of habitat loss and adverse effects altering the habitat conditions. The IDEM D6 spreadsheet in Task 2.1 contains 15 articles addressing D6 pressures and related effects on ecosystems. These papers represent the 45% of the total number of papers compiled. A brief description of their features is shown in Figure 7.3.

Alteration of deep-sea communities may result from a wide range of impacts, of which the vast majority affects the physical and biogeochemical conditions of the habitats. Feedbacks between different descriptors are worth considering as well. Also, impacts relevant for descriptor 6 do influence targets addressed by other descriptors, in the same way that other descriptor-related impacts have an effect on seafloor integrity. Whereas geographical fragmentation is not as important as for other descriptors, the Western Mediterranean Basin still gathers the majority of the studies on D6 impacts. However, monitoring sustained through time, which is essential for a reliable analysis of ecosystem change and evolution, is only addressed in 60% of the articles. Finally, besides some few research projects that intend to address the whole ecosystem, most of the investigations only concern specific groups of organisms. Therefore, systemic impacts are understudied.

Short descriptions of selected investigations provide examples of impacts on seafloor integrity, focusing on deep-sea ecosystems. The impacts refer to the two more studied pressures: bottom trawling and waste disposal.

www.msfd-idem.eu Deliverable 2.2

Figure 7.2. Summary of quantitative data available from case studies on bottom trawling and on waste disposal pressures in the deep Mediterranean Sea. The bottom trawling section presents data from selected studies in order to exemplify the main impacts investigated. Impact abbreviation (I) follows the same colour code as in Figure 7.1. A, B and C on the upper left map refer to the three waste disposal study cases cited in the accompanying text. Rectangles drawn on the map identify the regions where studies have been conducted. The grey one points out the Central-Ionian Region and neighbouring sub-basins, whereas the blue one covers the submarine canyons in the Northwestern Mediterranean Sea. Histograms in the lower half compare samples from places impacted by bottom trawling (black outlined bars) with samples from non-impacted locations. The case study on alteration of sediment fluxes shows data from three samples from Besòs (B1, B2 and B3) and Arenys (A1, A2, A3) submarine canyons, offshore Barcelona. The upper left map is from Ferrà et al. (2018), while the upper right data and map are from Eigaard et al. (2017). The two figures from the left lower corner are from Puig et al. (2012). The alteration of sediment fluxes and the biogeochemical changes plots have been produced with data from Paradis et al. (2017b) and Sañé et al. (2013), respectively. Finally, the figure in the right lower corner is from Martín et al. (2014). VMS: Vessel Monitoring System; DW: dry weight.(UNEP-MAP, 2009)

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Figure 7.3. Summary of the main features of the bibliographic references on descriptor 6 criteria 4, 5 and 6. Fifteen articles, representing the 45% of the total number of articles from the Task 2.1 IDEM D6 spreadsheet, refer to impacts of bottom trawling and waste disposal on habitats and ecosystems. Impact abbreviation (I) follows the same colour code as in Figure 7.1. A brief description of the impacts that alter deep-sea communities and the feedback with other descriptors is provided in the upper right box. AL: Aegean-Levantine Basin; CI: Central-Ionian Basin); W: Western Mediterranean Basin.

Fontanier et al. (2012, 2015) analyzed samples from the Cassidaigne Canyon (NW Mediterranean Sea) contaminated by bauxite red mud waste from a coastal aluminium processing plant. Figure 7.4 displays observations of a sediment core obtained very close to the disposal pipe outlet. The foraminifer community there shows low diversity and the living component is dominated by unusual species. Although an impact of the aluminium waste on the benthic community could be suggested, the authors state that no crystal-clear impacts could be confirmed on the benthic foraminifer community (Fontanier et al. 2012).

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Figure 7.4. Analysis of the red mud waste disposal on the Cassidaigne Canyon (NW Mediterranean Sea). The results showed belong to the shallower core (725m depth), the one closest to the disposal pipe outlet. A) Sampling of the red mud deposit with a blade corer. B) Stereoscopic view of sieve residues (>150µm) from the sediment collected. C) Pie chart illustrating the composition of the benthic live and dead foraminiferal community. Only major species (at least >2.5% in the living or dead faunas) are showed. The figures are adapted from Fontanier et al. (2012).

Research papers on impacts of bottom trawling on the benthic communities are somehow more abundant. In order to assess the magnitude of this pressure and its impacts, Eigaard et al. (2017) examined the extent of disturbance by bottom trawling on different habitats of the deep Mediterranean Sea. The map showing habitat distribution and the percentage affected in each habitat is in Figure 7.5.

Two further case studies describe direct impacts on different species and groups of organisms. Dimech et al. (2012) compare biomass, density, individual weight and length of species living in trawled and untrawled sites, thus demonstrating the impact that trawling (and overfishing) may cause on populations. Figure 7.6 shows two measures (length frequency distribution and total biomass) as examples of the results presented in that paper. The main outcome of the study is that lower biomass, density and diversity indices occur in trawled locations. However, high resilience to trawling activities was also observed for some species that are indeed fishing targets.

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Figure 7.5. Summary of the footprint of bottom trawling in habitats of the deep Mediterranean Sea. The EUNIS habitat classification is shown in the lower panel. A) Indicators of trawling pressure per EUNIS habitat. Trawling footprint metric (grey bars) corresponds to the percentage of all grid cells trawled, independently of intensity (%). Trawled seabed area (white bars) refers to the area of the seabed trawled at least once per year. The trawled seabed area is plotted alongside the relative surface area of each habitat type, thus defining the habitat area affected (black bars). B) Habitat distribution based on EUNIS Habitat level 3 (adapted from Eigaard et al., 2017).

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Figure 7.6. Effects of bottom trawling on species and populations from the Central Mediterranean Sea after comparison between trawled and non-trawled locations. A) Length frequency distribution in number of individuals (no. of ind.·km-2) for Helicolenus dactylopterus dactylopterus and Nephrops norvegicus at trawled and non-trawled locations. Lm: length at first maturity; Lopt: theoretical optimal size. B) Biomass size spectra (total biomass in g·km-2) expressed in logarithmic scale for different taxonomic groups (adapted from Dimech et al., 2012).

Pusceddu et al. (2014) revise the habitat impacts of trawling from sediment properties and composition to changes in meiofaunal communities. They demonstrate that sediments from trawled areas display decreased organic matter contents and slower carbon turnover. Additionally, meiofauna is reduced in abundance and biodiversity. Figure 7.7 illustrates two measures regarding sediment biogeochemical properties (organic matter and phytopigment content) and two measures of meiofaunal communities’ components (abundance and richness). Further data, analyses and figures can be found in the paper. Degradation of deep-sea sedimentary habitats due to trawling is suggested, highlighting the role of bottom trawling as a major threat for the deep-sea environments.

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Figure 7.7. Histograms showing the impacts of bottom trawling on sediments and meiofaunal communities from the deep La Fonera Canyon, in the NW Mediterranean Sea. A) Sedimentary organic matter (mg biopolymeric C· g-1) and total phytopigment concentration (in the top 1 cm of sediment, µg·g- 1) in samples from trawled and untrawled grounds at different depths. B) Meiofauna abundance (n individuals · 10cm-2) and biodiversity richness (in number of taxa · 10cm-2). Error bars represent standard errors between stations at similar depths and level of impact. Standard deviations in the data from 2000 m depth samples were calculated among replicates. *P <0.05; **P <0.01; ***P <0.001; NS, not significant vs. untrawled samples from 500, 800, and 2,000 m (adapted from Pusceddu et al., 2014).

www.msfd-idem.eu Deliverable 2.2 References Dimech, M., Kaiser, MJ., Ragonese, S., & Schembri, P. (2012). Ecosystem effects of fishing on the continental slope in the Central Mediterranean Sea. Marine Ecology Progress Series, (449), 41–54. Doi: 10.3354/meps09475. Eigaard, O.R., Bastardie,F., Hintzen, N.T., Buhl-Mortensen, L., Buhl-Mortensen, P., Catarino, R., Dinesen, GE., Egekvist, J., Fock, HO., Geitner, K., Gerritsen, HD., Martín González, M., Jonsson, P., Kavadas, S., Laffargue, P., Lundy, M., Gonzalez-Mirelis, G., Nielsen, R., Papadopoulou, N., Posen, PE., Pulcinella, J., Russo, T., Sala, A., Silva, C., Christopher, JS., Vanelslander, B., & Rijnsdorp, A., (2017). The footprint of bottom trawling in European waters: Distribution, intensity, and seabed integrity. ICES Journal of Marine Science, 74(3), 847–865. Doi: 10.1093/icesjms/fsw194. Ferrà, C., Tassetti, A.N., Grati, F., Pellini, G., Polidori, P., Scarcella, G. & Fabi, G. (2018). Mapping change in bottom trawling activity in the Mediterranean Sea through AIS data. Marine Policy, 94(December 2017), 275–281. Doi: 10.1016/j.marpol.2017.12.013. Fontanier, C., Fabri, M.C., Buscail, R., Biscara, L., Koho, K., Reichart, G.J., Cossa, D., Galaup, S., Chabaud, G., & Pigot, L. (2012). Deep-sea foraminifera from the Cassidaigne Canyon (NW Mediterranean): Assessing the environmental impact of bauxite red mud disposal. Marine Pollution Bulletin, 64(9), 1895–1910. Doi: 10.1016/j.marpolbul.2012.06.016. Fontanier, C., Biscara, L., Mamo, B., & Delord, E. (2015). Deep-sea benthic foraminifera in an area around the Cassidaigne Canyon (NW Mediterranean) affected by bauxite discharges. Marine Biodiversity, (45), 371–382. Doi: 10.1007/s12526-014-0281-9. Implementation of the MSFD to the Deep Mediterranean Sea (IDEM). Report 2.1. Review and collection of the available datasets on indicators and human pressures/impacts on Mediterranean deep-sea ecosystems. (30th June of 2018). DG Environment programme. 37pages. Martín, J., Puig, P., Palanques, A. & Giamportone, A. (2014). Commercial bottom trawling as a driver of sediment dynamics and deep seascape evolution in the Anthropocene. Anthropocene, 7(2014), 1–15. Doi: 10.1016/j.ancene.2015.01.002. Mecho, A., Aguzzi, J., De Mol, B., Lastras, G., Ramirez-Llodra, E., Bahamon, N., Company, JB., & Canals, M. (2017). Visual faunistic exploration of geomorphological human-impacted deep-sea areas of the northwestern Mediterranean Sea. Journal of the Marine Biological Association of the United Kingdom, 1– 12. Doi: 10.1017/S0025315417000431. Paradis, S., Puig, P., Masqué, P., Juan-Díaz, X., Martín, J. & Palanques, A. (2017). Bottom-trawling along submarine canyons impacts deep sedimentary regimes. Scientific Reports, 7(July 2016), 1–12. Doi: 10.1038/srep43332. Puig, P., Canals, M., Company, J.B., Martín, J., Amblas, D., Lastras, G., Palanques, A., & Calafat, A.M. (2012). Ploughing the deep sea floor. Nature, 489(7415), 286–289. Doi: 10.1038/nature11410. Pusceddu, A., Bianchellia, S., Martín, J., Puig, P., Palanques, A., Masqué, P., & Danovaro, R. (2014). Chronic and intensive bottom trawling impairs deep-sea biodiversity and ecosystem functioning. PNAS 111(24), 8861-8866. Doi: 10.1073/pnas.1405454111 Sañé, E., Martín, J., Puig, P. & Palanques, A. (2013). Organic biomarkers in deep-sea regions affected by bottom trawling: Pigments, fatty acids, amino acids and carbohydrates in surface sediments from the la Fonera (Palamós) Canyon, NW Mediterranean Sea. Biogeosciences, 10(12), 8093–8108. Doi: 10.5194/bg-10-8093-2013 UNEP, IMO/MED POL (2009). Ammunition dumping sites in the m Mediterranean Sea. Athens. (22th May of 2009). UNEP(DEPI)/MED WG. 334/Inf. 9. 53pages

www.msfd-idem.eu Deliverable 2.2 8. DESCRIPTOR 7: PERMANENT ALTERATION OF HYDROGRAPHICAL CONDITIONS

Descriptor 7 focuses on permanent alterations of hydrological conditions. As originally formulated, it appears to address essentially coastal shallow waters. Therefore, it needs to be addressed differently before it could be applied to the deep-sea. In particular, during the last 50 years plenty of literature and data on hydrological properties (mostly T and S) have been collected in the deep open Mediterranean Sea. However, the lack of specific information clearly establishing the connections between property change and ecosystem impacts for the deep Mediterranean Sea has prevented a meta-analysis equivalent to the ones performed for other descriptors within IDEM. The main bibliography used as reference has been presented in Table 8.1, according to the area of concern and described topic (e.g. Eastern Mediterranean Transient EMT, dense shelf water cascading DSWC). The attachment DATASET_D7.xlsx presented in Report 2.1 has been used as primary source of information. We decided to focus first on articles compiling long-term monitoring data (MAIN ARTICLES). However, some articles have been added in order to complement the time-series by providing data of episodic events (COMPLEMENT). We decided to collect mainly temperature and current speed data (for describing dense shelf water cascading events) and salinity, reporting long-term data series and trends, for which standard deviations/relative errors were available. These three variables are the most common ones in the D7-literature. Results have been presented in Attachment 1.

Table 8.1. Selected paper for the D7 meta-analysis, as extracted from the review activity conducted for Report 2.1 of the IDEM Project. REFERENCE LOC TOPIC MAIN ARTICLE (Bensi et al., 2016) Eastern DSWC and EMT Mediterranean (Emed) MAIN ARTICLE (Bethoux et al., 2002) Western DSWC Mediterranean (Wmed) MAIN ARTICLE (Borghini et al., 2014) Wmed Salinity and T trends MAIN ARTICLE (de Fommervault et Wmed Nutrients trends al., 2015) Complement (Durrieu de Madron et Wmed DSWC al., 2005) Complement (Durrieu de Madron et Wmed DSWC and al., 2013) Convection Complement (Gasparini et al., 2005) Wmed EMT MAIN ARTICLE (Herrmann et al., Wmed Deep Water 2017) Formation (DWF) and Convection MAIN ARTICLE (Heussner et al., 2006) Wmed DSWC, currents transport MAIN ARTICLE (Kress et al., 2003) Emed T, salinity, density, O2, nutrients

www.msfd-idem.eu Deliverable 2.2 Complement (Kress et al., 2014) Emed T, salinity, density, O2, nutrients, Chl-a, EMT MAIN ARTICLE (Meccia et al., 2016) Whole Salinity and T Mediterranean trends (WholeMed) MAIN ARTICLE (Ozer et al., 2017) Emed T, salinity, nutrients, Chl-a trends Complement (Puig et al., 2013) Wmed DWF, DSWC

Complement (Sanchez-Vidal et al., Wmed DSWC 2008) MAIN ARTICLE (Schroeder et al., W-Central Salinity and T 2016) Med trends MAIN ARTICLE (Schroeder et al., WholeMed Salinity and T 2017) trends MAIN ARTICLE (Shaltout & Omstedt, Wmed and Salinity and T 2015) Emed trends MAIN ARTICLE (Touratier & Goyet, WholeMed CO2 trends 2011) MAIN ARTICLE (Vargas-Yáñez et al., Wmed Heat content, 2010a) Salinity and T trends MAIN ARTICLE (Vargas-Yáñez et al., WholeMed Salinity and T 2010b) trends In the following, three descriptive subsections will be presented: the first one more introductory reflecting data distribution/availability. The second one with data directly related to the pressure as described by long term time series or/and episodic events, and finally a third Section representing the impact of the pressure on the habitats and on ecosystems.

Data Distribution

A set of figures and diagrams are thus provided in order to describe and exemplify the available data and the current knowledge related to D7 for the deep Mediterranean Sea. Figure 8.1 summarizes data availability and distribution. The analysis was performed using the tools and datasets displayed in two large databases: EMODnet (http://www.emodnet-physics.eu/map/) and SeaDataNet (http://seadatanet.maris2.nl/v_cdi_v3/search.asp). and SeaDataNet. Information regarding the deep-sea was obtained by applying water depth and instrument depth filters. In addition, a duration of at least 6 months was also demanded in a second search in order to identify long-term data series from the deep- sea. Finally, the amount of data provided by different devices along the 1953-2016 period is also reported as a bar graph, also shown in IDEM Deliverable 2.1.

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Figure 8.1. Summary of data availability and distribution for D7 considering temperature and salinity. Data distribution is analyzed by water depth, sampling device, geography and dataset duration. The blue box in the upper-left part emphasizes the two main databases used for the analysis (EMODnet and SeaDataNet). Each of the orange dots from the SeaDataNet maps corresponds to a dataset. The rectangles and arrows in the map of the lower right corner point to the datasets retrieved from the SeaDataNet database, since their low number hinders an easy spotting (zooming in is recommended to better observe the information in that map). The graph in the bottom left corner was also shown in IDEM Deliverable 2.1.

Pressures - Long-term monitoring data

As concerns D7, changes in thermohaline properties, dissolved oxygen and pH of seawater are the identified pressures with the highest impact on marine ecosystems and the focus of the majority of the reviewed datasets (see Report 2.1 for more details).

Modelling has been widely used to describe the oceanic characteristics of the Mediterranean, allowing long-term runs. Shaltout and Omstedt (2015) modelled MEDAR data of Mediterranean basins (http://www.ifremer.fr/medar/), dealing with annual temperatures and salinities for the surface (0-150 m), intermediate (150-600 m), and deep (>600 m) layers (Figure 8.2) for the 1958-2010 period, obtaining

www.msfd-idem.eu Deliverable 2.2 the long-term mean values reported in Table 8.2. Modelled surface water in the Eastern Mediterranean (Emed) had a higher mean temperature (by approximately 1.68 °C) and was more saline (by approximately 0.87 g kg-1) than in the Western Mediterranean (Wmed) over the studied period.

Table 8.2. Long-term mean values for temperatures and salinities for the surface and deep waters Temperature Salinity °C g kg-1 Wmed surface-layer 14.7 37.5 Emed surface-layer 16.8 38.2 Wmed deep-layer 13.1 38.5 Emed deep-layer 13.7 38.7

Figure 8.2. Average temperature and salinities of three different layers of the Western and Eastern Mediterranean sub-basins (from Shaltout and Omstedt, 2015).

www.msfd-idem.eu Deliverable 2.2 Meccia et al. (2016) explored the recent decadal variability of some parameters (data from the Copernicus Marine Environmental Monitoring Service) during the 1987–2013 time period to evaluate possible changes in the water mass properties, focusing on sub-basins of the Mediterranean Sea to simplify the analysis of the results. The analysis of θ-S diagrams for the entire Mediterranean Sea, in relation to three distinct periods 1987– 1996, 1997–2006 and 2007–2013, shows that the water mass properties are different (Figure 8.3). In particular, a gradual shift towards higher values of temperature and salinity is clearer for the intermediate and deep waters (see Figures 8.3b and c respectively) than for the surface waters (Figure 8.3a) and can be associated with the various intensive events of intermediate and deep water formation.

Figure 8.3. θ-S diagrams for the deep waters in the a) Levantine, b) Adriatic, c) Ionian and d) Tyrrhenian sub-basins according to the mean fields of 1987–1996 (in green), 1997–2006 (in blue) and 2007–2013 (in red) (from Meccia et al., 2016).

The results for the intermediate waters in the Levantine, Adriatic, Ionian and Tyrrhenian sub-basins are shown in Figure 8.4. Similar features are present in the Levantine and in the Ionian Sea, which seem to be the most sensitive to decadal variability. They show the major increase in salinity (approximately 0.07) during the second decade (1997-2006) (Figure 8.4a and 8.4c) as a consequence of the Levantine Intermediate Water (LIW) entrance in the Ionian at intermediate depths. The penetration of the colder and fresher waters formed in the Adriatic could explain the lower θ and S values in the Ionian than in the Levantine (Figure 8.4a, c). The intermediate waters in the Adriatic exhibit a gradual increase in both salinity and potential temperature (Figure 8.4b). The θ rise in particular may affect the Ionian Sea, which appears warmer during the last period (2007– 2013) as a result of intermediate waters mixed. Finally, the Tyrrhenian Sea (Figure 8.4d) seems to have been subjected to a salinization of approximately 0.07 during both the second and third periods. The intermediate waters in theTyrrhenian might thave been influenced by both the LIW and the Western Mediterranean Deep Water (WMDW).

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Figure 8.4. θ-S diagrams for the intermediate water in the a) Levantine, b) Adriatic with a grid resampling of 0.5°x0.5°, c) Ionian , d) Tyrrhenian sub-basins according to the mean fields of 1987–1996 (in green), 1997–2006 (in blue) and 2007–2013 (in red) (from Meccia et al., 2016).

The θ-S diagrams for the deep waters (~1000m depth) in the Levantine, Ionian, NW Mediterranean and Tyrrhenian seas are shown in Figure 8.5. The salinity seemed to increase with time according to the variability of the deep water properties. As with the intermediate waters, the Ionian and Levantine exhibit similar characteristics (Figures 8.5a and b). This feature can be associated to the EMT, which also caused changes in diffusive process as salt fingering. On the other hand, the Tyrrhenian Sea seems to be correlated with the NW Mediterranean deep water variability, although the temperatures of the former are higher (Figures 8.5c and d).

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Figure 8.5. θ-S diagrams for the deep waters in the a) Levantine, b) Adriatic, c) Ionian and d) Tyrrhenian sub-basins according to the mean fields of 1987–1996 (in green), 1997–2006 (in blue) and 2007–2013 (in red) (from Meccia et al., 2016).

Long term increasing of temperature and salinity in the deep Mediterranean waters have been also documented by Vargas-Yáñez et al. (2010) by analyzing over 100-years time series (from 1900 to 2008) from four selected sectors within the Western Mediterranean. In the study, the water column was divided into three layers, the upper one, from the surface to 200 m depth, the intermediate one, 200 to 600 m, and the deep layer, from 600 m to the sea bottom. The results show the warming and the salinity increase of the three layers from the beginning of the twentieth century until 2008 (Figure 8.6).

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Figure 8.6. Time series for temperature (A) and salinity (B) for the intermediate (1) and deep (2) layers, averaged for the four study areas. The trends show the warming and the salinity increase of the two layers from the beginning of the twentieth century until 2008 (from Vargas-Yáñez et al., 2010a).

An increase in the temperature and salinity values in the deep Mediterranean Sea waters related to the formation of new warmer, saltier denser deep water has been also reported by Borghini et al. (2014) by analyzing the temporal evolution (from 1961 to 2008) of deep-water masses in the Western Mediterranean. The results shows an increase in temperature by 0.080 °C±0.007 and in salinity by 0.035 ± 0.014, from 2004 to 2013 (twice the rates of increase found from 1961 to 2008), in the vertical depth range from 1900 to 2750 dbar. Moreover, the study highlights the role of another process contributing to the increases in salinity and temperature of the deep water: salt finger processes which transport salt and heat steadily downward through the halocline–thermocline. The effects of these two processes are visible in the monthly time series of CTD profiles of salinity and temperature in the Ligurian Sea from 1995 to 2008 (Figure 8.7).

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Figure 8.7. Time series of (a) salinity and (b) potential temperature in the layers 1600–1800m and 1800– 2000m at the station in the Ligurian Sea. Blue dots refer to monthly values while red dots refer to running mean 13-month average values (from Borghini et al., 2014).

In Ozer et al. (2017), a 30-years dataset of CTD casts in the Levantine Basin was analyzed to examine the thermohaline trends of the surface (~0–50 m) and intermediate (~150–350 m) water masses (LSW, LIW), and a 13 years (2002–2014) dataset was used to explore the relations between the physical and nutrient properties in the LIW in the eastern Levantine Basin. LIW core salinity and temperature displayed a significant long term increasing trends of +0.005 ± 0.003 year−1 and +0.03 ± 0.02 °C year−1, respectively (Figure 8.8). The rate of temperature increase is in agreement with the predictions for the period 2016 to 2035 released by the IPCC, in 2014 (+0.7 °C or +0.035 °C year−1).

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Figure 8.8. Temperature and salinity trends for LIW in the south-eastern Mediterranean, obtained by analyzing 1382 CTD (salinity and temperature) profiles from the PERSEUS cast database (http://isramar.ocean.org.il/perseus_data/) as well as from observations in situ, covering the period 1978 to 2014 (from Ozer et al. 2017).

Time series of temperature, salinity, nutrients and Chlorophyll a (0–200 m) in the core of the LIW water mass are shown in Figure 8.9. The results document an evident shift in LIW physical properties. The rise of salinity and temperature values correspond to a decrease of the nitrate + nitrite, phosphate and silicic acid levels from 3.99±0.74 to 0.77 ± 0.15 μmol kg−1, 0.13 ± 0.03 to 0.02 ± 0.01 μmol kg−1 and 4.77 ± 0.93 to 0.87 ± 0.06 μmol kg−1, respectively.

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Figure 8.9. Time series of temperature (red), salinity (yellow), nitrate+nitrite (purple), phosphate (orange), silicic acid (blue) and integrated Chlorophyll a (0–200m) (green) performed in the core of the LIW water mass (ca. 130m b z b 350m). Dots refer to station-specific values, the solid lines represent the moving average. Standard error ranges are indicated with the coloured area (Ozer et al. 2017).

Several other studies are also conducted about temporal evolution of nutrient chemistry in Mediterranean Sea. De Fommervault et al. (2015) over the period 1991–2011 (Figure 8.10) revealed decadal trends in nitrate and phosphate concentrations in the 800-2000 depth layer (+0.23% and –0.62%, respectively) resulting in increasing N:P and Si:P ratios (+1.14% and +0.85% per year, respectively). Such a long-term variability is presumably related to changes in water mass (increase of winter convection events) and/or changes in external sources (atmospheric and riverine inputs), even if it is difficult to assess due to insufficient concomitant data from atmospheric and riverine inputs. The abrupt change occurring in 2005–2006 (linked to the deep mixing event) is believed to strongly contribute to the increasing trend of deep nitrate concentration over the whole period (1991–2011). Opposite trends are observed for P, suggesting that the role of P as limiting factor of primary production may be currently increasing.

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Figure 8.10. Evolution of nutrient concentration (a) nitrate, (b) phosphate, (c) silicate and molar ratio (d) N:P, (e) Si:P, (f) Si:N in deep water at the DYFAMED site from 1991 to 2011. Data are averaged over the 800 m-bottom layer and the error bar are 2 times the standard deviation length. Regression lines are in black and triangles indicates the date of change, determined from the Pettitt test, if significant (from De Fommervault et al., 2015).

Synoptic pictures are available in literature for Total Dissolved Carbon, Alkalinity, pH and Anthropogenic

Carbon (CANT). In particular Touratier and Goyet (2011) defined recent change in acidification of intermediate and deep waters due to the EMT as reported in Figure 8.11.

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Figure 8.11. Distribution of the acidiﬁcation level (ΔpH) reached during the year 2001 along the section shown above (more details in Touratier and Goyet, 2011)

From Figure 8.11 it is clear that all Mediterranean waters are acidified, particularly those in the Western Basin, where deep-water renewal is faster. Long term data on carbonate system are not common but more statistics could be elaborated starting from long time series available at least for the DYFAMED Station in the NW Ligurian Sea (43°25N, 7°52E; measurements started in 1991), supported as a contribution to EMSO (European Multidisciplinary Seafloor and water column Observatories, http://emso.eu/) EU Infrastructure.

www.msfd-idem.eu Deliverable 2.2 Pressures - Episodic events of modification in the Mediterranean deep-sea hydrodynamic conditions: Eastern and Western Mediterranean Transient and cold-water cascading

Until the early 1990s the deep dense water formed in the southern, the Adriatic Adriatic Deep Water (ADW), occupied the deep layer of the Eastern Mediterranean, spreading through the Otranto Strait to fill the deepest parts of the Ionian and Levantine basins. Then, the Aegean Sea became the main deep water formation area, in an event known as the Eastern Mediterranean Transient (EMT). The EMT changed considerably the circulation in the Eastern Mediterranean and the depth distribution of physical and chemical parameters. It was accompanied by a vast rearrangement of the entire water column circulation as a result of shallower Aegean outflow and mixing of mid-depth water, lifting several hundred metres as a consequence of Aegean water addition deeper down. Kress et al. (2003 and 2014) documented the continuing effect of the EMT event on the distribution of physical and chemical parameters in the Easternmost Levantine basin from 2002 to 2010. By mid-2002, the characteristics of the deep waters at the Levantine basin was already modified as a result of the EMT event, and the new deeper, younger, dense water was well defined, as warmer and more saline than the older ADW, raised from the bottom layers to mid depths. From 2002 to 2010, the effects altered also biogeochemical parameters, increasing the nutrient concentrations at the nutricline depth and the supply to the photic zone (Figure 8.12). Figure 8.12 also represents the observed trend for dissolved oxygen: dissolved oxygen is not a conservative property and its depth distribution can be influenced by general circulation (such as the EMT event), and by its utilization in the decomposition of organic matter by bacteria.

Figure 8.12. Temporal evolution of potential temperature, salinity, dissolved oxygen, silicic acid, nitrate and phosphate from two study sites in the Eastern Mediterranean Sea (from Kress et al., 2014).

The consequences of the change of dense-water formation area occurred during EMT extended far beyond the eastern basin, influencing also the western deep layer. Gasparini et al. (2005) reported the alteration in the hydrographic characteristics of deep water of the Strait of Sicily and in the Tyrrhenian Sea. The impact of the EMT on the western basin was maximum between 1992–1994. Starting from 1987, the salinity in the Aegean Sea progressively increased while this increase appeared in the Ionian only since 1991, when a significant volume of saltier water of Aegean origin occupied the deep layer, explaining the progressive salinity and temperature increases (Figure 8.13).

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Figure 8.13. Time evolution of temperature, salinity, and density in the central region of the Strait of Sicily (from Gasparini et al., 2005).

Similar results have been documented for the western basin, where the EMT reached its maximum in 1992–1994, when a great portion of the outflow across the Strait of Sicily sinks into the deep Tyrrhenian basin, contributing to deep temperature and salinity trends (Figure 8.14).

Figure 8.14. Time evolution of temperature, salinity, and density in the central Tyrrhenian Sea (from Gasparini et al., 2005). 92

www.msfd-idem.eu Deliverable 2.2 Beyond the EMT in the Eastern Basin, abrupt hydrological changes, with gradual increases of temperature and salinity, have also been noticed in the Western Mediterranean Sea. The main alarm clock for those changes was the so-called Western Mediterranean Transition (WMT) event, as documented by Schroeder et al. (2016). In this study authors focused their attention on the major deep- water formation event in winter 2004/05, which set the beginning of the WMT, a climate shift which changed the basic structure and properties of the intermediate and deep layers in the Western Mediterranean. The 2004/05 event was the first of a series of similar events that occurred, with varying intensities, in some of the following years, namely in 2009/10, 2011/12 and 2012/13, which led to major reinforced thermohaline variability in the structure of the intermediate and, especially, the deep layers in the Western Mediterranean, producing large amounts of denser water masses than ever before (Figure 8.15).

Figure 8.15. Temporal and geographical evolution of the processes associated to the WMT. The lines indicate the upper interface, divided by the years, of the new WMDW in the Mediterranean basin (from Schroeder et al., 2016).

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Figure 8.16. (A) Deep-water θ-S diagrams from a repeat station (2800 m, 37.98 °N, 4.65 °E) of the southern Western Mediterranean in 2010 (upper panel, light grey points 2005–2009, dark grey points refer to pre- WMT situation in 2004), 2013 (middle panel, light grey 2005–2010, dark grey 2004) and 2015 (lower panel, light grey 2005–2014, dark grey 2004); (B) θ-S diagrams from Sardinia Channel station (1900 m, 38.33 °N, 9.33 °E); left panel shows bottom mooring data and the right one show the repeated CTD profiles (years are colour-coded); (C) θ-S diagrams in a repeat station (3500 m, 38.92 °N, 13.3 °E) of the southern Tyrrhenian Sea in 2005 (pre-WMT situation), 2012 and 2015 (from Schroeder et al., 2016)

After the WMT, inversions in the typical deep θ S diagrams of the deep layer have been documented: in the following winters a new warmer, saltier and denser deep water is formed, leading to a stepwise increase of heat and salt contents in the deep layer (Figure 8.16A). In 2012–2015 the thickness of the modified deep layer increased to almost 1000 m and the WMT signature extended in almost all the Tyrrhenian Sea as shown by typical hooks in the θ S diagram of Figure 8.17.

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Figure 8.17. (A) vertical profiles (> 800 dbar) of salinity (potential temperature is colour-coded); (B) deep-sea (>1500 dbar) potential temperature vertical profiles (upper panel, salinity is colour-coded for 2015, light grey points data from 2012, dark grey points from 2005, the pre-WMT situation) and salinity (lower panel, potential temperature is colour coded for 2015, light grey 2012, dark grey 2005; from Schroeder et al., 2016).

Winter transformation of sea surface water into dense water, and its cascading and spreading as deep water, is one of the main driving force of the water masses circulation. In the Mediterranean Sea we have evidence of cascading processes off the Adriatic shelf (Trincardi et al. 2007), in the Cretan Sea and in the Gulf of Lion (Pusceddu et al. 2010). Since the intensity of the cooling is directly related to climatic variability, in some cases, the resulting water mass cascading process may be stronger than usual. Even if sporadic, these events are known to largely influence the chemical physical properties of the deep Mediterranean water masses. Durrieu de Madron et al. (2005) documented large effects of cold-water cascading in the Gulf of Lion by monitoring temperature and current since 1993 in the lower part of the Lacaze–Duthiers canyon. The results reveals an interannual variation of the shelf water overflow intensity, with an intense shelf cascading, contributing to the renewal of the bottom waters, as consequence of the abnormally cold 1998–1999 winter (Figure 8.18).

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Figure 8.18. Temperature time series at 470 m and 1000 m in the Lacaze–Duthiers Canyon. The arrows indicate significant cascading events (from Durrieu de Madron et al., 2005).

The analysis of deep CTD casts carried out in the same area (Gulf of Lion) by Puig et al. (2013) from 1998 to 2011 has documented cascading events also in 2005, 2006 and 2010. The decrease of temperature remarks the deepening of colder water masses typical of cascading processes (Figure 8.19). Moreover, the study highlights the role of deep dense shelf water cascading off the Gulf of Lions in transporting suspended particulate matter to the deep layer. After the 1999 and 2005–2006 deep cascading events, the WMDW was characterized by a thick bottom nepheloid layer (BNL) generated by both the deep convection in the open sea and by deep cascading.

Figure 8.19. Temperature time series at 480 m, 1000 m, and between 750-1000m in the Gulf of Lion (from Puig et al., 2013).

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Impact of the pressure on the habitats and on ecosystems.

The second criteria of D7 refers to the extent of adversely affected benthic habitats due to permanent alteration of hydrological conditions. Habitats impacts should encompass both physical and biological changes. The analysis of documents in the IDEM D7 spreadsheet generated in Task 2.1 shows that only 13% of the datasets refer to consequences of hydrological changes on habitats and ecosystems. Figure 8.20 illustrates the main characteristics of such documents.

Figure 8.20. Summary of the characteristics of the datasets that refer to alterations on habitats and ecosystems due to hydrological changes. The IDEM D7 spreadsheet generated in Task 2.1 is illustrated in the top left corner, displayed the percentage of datasets referring to the D7C2. The images in the upper central box are from Paolo Montagna and Marco Taviani (https://schmidtocean.org/cruise-log-post/the- secret-world-of-deep-sea-corals/), Canals et al. (2009), Mecho et al. (2014), Dr. Murray Roberts (https://www.marlin.ac.uk/species/detail/1806) and LECOB Chaire UPMC-Foundation Total (https://www.skepticalscience.com/print.php?n=3211). D7C2: Descriptor 7, Criterion 2; AL: Aegean- Levantine Basin; CI: Central-Ionian Basin); W: Western Mediterranean Basin; T: temperature, S: salinity.

www.msfd-idem.eu Deliverable 2.2 Most of the articles did not identify specific anthropogenic actions as pressures for benthic habitats. However, almost all of them included climate change as a driver influencing and altering ocean circulation patterns, physical parameters and deep-water formation events. Knowledge is geographically fragmented, since most of the studies were performed in the Western Mediterranean Basin. Regarding the methodology, only half of the datasets involved a continued tracking of the issue targeted, by monitoring during different time periods. The rest were snapshot studies that did not consider any temporal evolution of the habitat. Although different biological groups appear as the target of the research, analyses focusing on the impacts caused to a complete trophic web, habitat or ecosystem were missing. Finally, when focusing on impacts on habitats, multiple descriptors besides D7 are relevant too. Changes in hydrological conditions could have multiple consequences, partly addressed in other descriptors.

Some selected study cases, which are shortly described below, illustrate how changes of hydrological conditions, eventually combined with atmospheric forcing, other environmental changes and anthropogenic influences, may affect ocean habitats and ecosystem functions in numerous ways.

The article published by Conversi et al. (2010) analyzes the Eastern Mediterranean Transient (EMT) and the hydrographic changes involved by the end of the 1980s. The modification of circulation and mixing patterns, and temperature increases caused by the EMT directly affected biological communities. A summary of the relations between pressures and impacts on marine ecosystems is displayed in Figure 8.21.

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Figure 8.21. Scheme describing the different impacts on marine ecosystems. The figure is from Conversi et al. (2010).

Other hydrographic alterations, such as dense shelf water cascading (DSWC) events, and their ecosystemic impacts have been addressed by several studies looking, for instance, at their impacts on primary production rates and phytoplankton communities, or on carbon export to the deep (Kress et al., 2014; Bosc et al., 2004; Canals et al., 2009). Although phytoplankton does not belong to the deep-sea but to the sea surface, either coastal or open sea, primary production shifts have profound implications for the whole marine ecosystem. Climate change is again presented as a key forcing influencing the recurrence and intensity of such hydrographic alterations.

Deep-water formation due to DSWC and open-ocean convection influences physical and biogeochemical properties of the deep-sea, and thus impacts deep pelagic and benthic habitats and the organisms forming them. Pusceddu et al. (2013) targeted DSWC consequences on deep canyon trophic conditions and meiofaunal communities. They concluded that labile organic material from the continental shelf was flushed to the deep margin, changing the biochemical properties of the deep sediments. In addition, the cascading event fostered low meiofaunal abundance, biomass and biodiversity values. Figure 8.22 presents two histograms reflecting two measures as examples of the DSWC consequences stated, bioavailable C as sediment property, and meiofaunal abundance as community component. The article suggests hydrodynamic stress and suffocation due to massive sediment deposition as the main reasons of the community change. However, the impact did not last long, since the communities recovered after 99

www.msfd-idem.eu Deliverable 2.2 six months only. The article provides more data and analysis regarding other sediment components and biodiversity indexes (Pusceddu et al. (2013)).

Figure 8.22. Histograms comparing biogeochemical and biological properties of samples exposed to DSWC (red bars) with unexposed ones (blue bars). Both histograms reflect spatial and temporal variations. A) Bioavailable carbon content (mg C· g-1). B) Meiofauna abundance in number of individuals (n.ind · 10 cm2). Graphs from Pusceddu et al. (2013).

Effects and responses caused by cold water cascading were also studied in the Eastern Mediterranean Basin. The article suggests a cascade of effects causing an alteration of biodiversity richness and abundance resulting in the alteration of benthic ecosystem functions. These cascade effects are summarized in Figure 8.23.

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Figure 8.23. Summary of the cascade of effects of the EMT event in the Eastern Mediterranean and how the deep-sea ecosystem responded. The figure is from Pusceddu et al. (2010).

The consequences of DSWC on living resources inhabiting the deep-sea were presented by Company et al. (2008). By crossing data on landings of the seep-sea shrimp Aristeus antennatus and background knowledge they were able to show the effects of DSWC on the shrimp’s population and the catches (Figure 8.24). At first, cascading causes the disappearing of the shrimp population and a severe reduction of landings, temporarily down to a collapse situation. But the shrimp’s population recovers afterwards till a significant increase of juveniles occurs a few years after the event. Once more, climate change influence on DSWC events with consequences for the shrimp population and the fishing industry were inferred.

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Figure 8.24. Temporal evolution in annual landings since 1977 to 2005 of Aristeus antennatus in two Catalan harbours (Blanes and Arenys de Mar). The years were strong cascading events occurred are illustrated in green. The blue arrows remark slow landing minima caused by the previous strong cascading events. Red arrows show the following landing peak three to five years after the cascading event. Figure modified from Canals et al. (2009).

The alteration of deep-sea hydrological conditions by DSWC events (and eventually open sea convection) and their profound impact on matter and energy transfers, and on the properties of the water masses, also hit microbial assemblages and their metabolism. This, in turn, results in an alteration of ecosystem functions and biogeochemical cycling in the deep. Whereas this topic has not been traditionally considered in the study of benthic communities, research on these aspects is on the rise (Tamburini et al., 2013; Packard et al., 2009; Luna et al., 2016).

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Bensi, M., Velaoras, D., Meccia, V., Cardin, V. (2016). Effects of the Eastern Mediterranean Sea circulation on the thermohaline properties as recorded by fixed deep-ocean observatories, Deep Sea Research Part I: Oceanographic Research Papers 112, 1-13. Bethoux, J. P., Durieu de Madron, X., Nyffeler, F., & Tailliez, D. (2002). Deep water in the western Mediterranean: peculiar 1999 and 2000 characteristics, shelf formation hypothesis, variability since 1970 and geochemical inferences. Journal of Marine Systems, 33–34, 117–131. https://doi.org/10.1016/S0924-7963(02)00055-6 Borghini, M., Bryden H., Schroeder, K, Sparnocchia, S., Vetrano, A. (2014). The Mediterranean is becoming saltier, Ocean Sci. 10, 693-700. https://doi.org/10.5194/os-10-693-2014 Bosc, E., Bricaud, A., & Antoine, D. (2004). Seasonal and interannual variability in algal biomass and primary production in the Mediterranean Sea, as derived from 4 years of SeaWiFS observations. Global Biogeochemical Cycles, 18, GB1005. Doi:10.1029/2003gb002034. Canals, M., Danovaro, R., Heussner, S., Lykousis, V., Puig, P., Trincardi, F., Calafat, AM., Durrieu de Madron, X., Palanques, A., & Sànchez-Vidal, A. (2009). Cascades in Mediterranean submarine grand canyons. Oceanography 22(1):26–43. Doi: 10.5670/oceanog.2009.03. Company, J.B., Puig, P., Sardà, F., Palanques, A., Latasa, M., & Renate Scharek. (2008). Climate influence on deep sea populations. PLoS ONE 3(1): e1431. Doi: 10.1371/journal.pone.0001431. Conversi, A., Fonda Umani, S., Peluso, T., Molinero, JC., Santojanni, A., & Edwards, M. (2010) The Mediterranean Sea regime shift at the end of the 1980s, and intriguing parallelisms with other European Basins. PLoS ONE 5(5): e10633. Doi:10.1371/journal.pone.0010633 de Fommervault, O.P., Migon, C., D'Ortenzio, F., d'Alcala, M.R., Coppola, L. (2015). Temporal variability of nutrient concentrations in the northwestern Mediterranean sea (DYFAMED time-series station), Deep-Sea Research Part I-Oceanographic Research Papers 100, 1-12, doi: 10.1016/j.dsr.2015.02.006. Durrieu de Madron, X., Zervakis, V., Theocharis, A., Georgopoulos, D. (2005). Comments on "Cascades of dense water around the world ocean", Progress in Oceanography 64, 1, 83-90, doi: 10.1016/j.pocean.2004.08.004. Durrieu de Madron, X. et al. (2013). Interaction of dense shelf water cascading and open-sea convection in the northwestern Mediterranean during winter 2012. Geophysical Research Letters, 40(7), 1379– 1385. https://doi.org/10.1002/grl.50331 Gasparini, G.P., Ortona, A., Budillon, G., Astraldi, M., Sansone, E. (2005). The effect of the Eastern Mediterranean Transient on the hydrographic characteristics in the Strait of Sicily and in the Tyrrhenian Sea, Deep Sea Research Part I: Oceanographic Research Papers 52, 6, 915-935, doi:https://doi.org/10.1016/j.dsr.2005.01.001. Herrmann, M., Auger, P.-A., Ulses, C., & Estournel, C. (2017). Long-term monitoring of ocean deep convection using multisensors altimetry and ocean color satellite data. Journal of Geophysical Research: Oceans, 122(2), 1457–1475. https://doi.org/10.1002/2016JC011833 Heussner, S., Durrieu de Madron, X., Calafat, A., Canals, M., Carbonne, J., Delsaut, N., & Saragoni, G. (2006). Spatial and temporal variability of downward particle fluxes on a continental slope: Lessons from an 8-yr experiment in the Gulf of Lions (NW Mediterranean). Marine Geology, 234(1–4), 63–92. https://doi.org/10.1016/j.margeo.2006.09.003

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www.msfd-idem.eu Deliverable 2.2 Implementation of the MSFD to the Deep Mediterranean Sea (IDEM). Report 2.1. Review and collection of the available datasets on indicators and human pressures/impacts on Mediterranean deep-sea ecosystems. (30th June of 2018). DG Environment programme. 37pages. Kress, N. (2003). Continuing influence of the changed thermohaline circulation in the eastern Mediterranean on the distribution of dissolved oxygen and nutrients: Physical and chemical characterization of the water masses. Journal of Geophysical Research, 108(C9), 8109. https://doi.org/10.1029/2002JC001397 Kress, N., Gertman, I., & Herut, B. (2014). Temporal evolution of physical and chemical characteristics of the water column in the Easternmost Levantine basin (Eastern Mediterranean Sea) from 2002 to 2010. Journal of Marine Systems, 135, 6–13. https://doi.org/10.1016/j.jmarsys.2013.11.016 Meccia V.L., Simoncelli S. and Sparnocchia S. (2016). Decadal variability of the Turner Angle in the Mediterranean Sea and its implications for double diffusion, Deep-Sea Research Part I, Volume 114, p. 64-77. Luna, GM., Chiggiato, J., Quero, GM., Schroeder. K., Bongiorni, L., Kalenitchenko D, & Galand PE. (2016). Dense water plumes modulate richness and productivity of deep sea microbes. Environmental Microbiology 18(12), 4537–4548. Doi: 10.1111/1462-2920.13510. Mecho, A., Aguzzi, J., Company, JB., Canals, M., Lastras, G., & Turon, X. (2014). First in situ observations of the deep-sea carnivorous ascidian Dicopia antirrhinum Monniot C., 1972 in the Western Mediterranean Sea. Deep-Sea Research I (83) 51–56. Doi: 10.1016/j.dsr.2013.09.007. Ozer, T., Gertman, I., Kress, N., Silverman, J., & Herut, B. (2017). Interannual thermohaline (1979–2014) and nutrient (2002–2014) dynamics in the Levantine surface and intermediate water masses, SE Mediterranean Sea. Global and Planetary Change, 151(October), 60–67. https://doi.org/10.1016/j.gloplacha.2016.04.001 Packard, T., Gómez, M., Christensen, J., (2009). Fueling Western Mediterranean deep metabolism by Deep Water formation and shelf-slope cascading: evidence from 1981. In CIESM, 2009. Dynamics of Mediterranean deep waters. N° 38 in CIESM Workshop Monographs [F. Briand, Ed.], 132 pages, Monaco. Puig, P., Madron, X. D. de, Salat, J., Schroeder, K., Martín, J., Karageorgis, A. P., … Houpert, L. (2013). Thick bottom nepheloid layers in the western Mediterranean generated by deep dense shelf water cascading. Progress in Oceanography, 111, 1–23. https://doi.org/10.1016/j.pocean.2012.10.003 Pusceddu, A., Mea, M., Gambi, C., Bianchelli, S., Canals, M., Sánchez-Vidal, A., Calafat, A.M., Heussner, S., Durrieu De Madron, X., Avril, J., Thomsen, L., García, R., Danovaro, R. (2010). Ecosystem effects of dense water formation on deep Mediterranean Sea ecosystems: an overview, Advances in Oceanography and Limnology 1, 51- 62. Pusceddu, A., Mea, M., Canals, M., Heussner, S., Durrieu de Madron, X., Sanchez-Vidal, A., Bianchelli, S., Corinaldesi, C., Dell'Anno, A., Thomsen, L., & Danovaro, R., (2013). Major consequences of an intense dense shelf water cascading event on deep-sea benthic trophic conditions and meiofaunal biodiversity, Biogeosciences 10: 2659–2670, Doi: 10, 2659-2670, 10.5194/bg-10-2659-2013. Sanchez-Vidal, A. et al. (2008). Impact of dense shelf water cascading on the transfer of organic matter to the deep western Mediterranean basin. Geophysical Research Letters, 35(5). https://doi.org/10.1029/2007GL032825 Schroeder, K., Chiggiato, J., Bryden, H. L., Borghini, M., & Ben Ismail, S. (2016). Abrupt climate shift in the Western Mediterranean Sea. Scientific Reports, 6(1), 23009. https://doi.org/10.1038/srep23009 104

www.msfd-idem.eu Deliverable 2.2 Schroeder, K., Chiggiato, J., Josey, S. A., Borghini, M., Aracri, S., & Sparnocchia, S. (2017). Rapid response to climate change in a marginal sea. Scientific Reports, 7(1), 4065. https://doi.org/10.1038/s41598- 017-04455-5 Shaltout M., Omstedt A. (2015). Modelling the water and heat balances of the Mediterranean Sea using a two-basin model and available meteorological, hydrological, and ocean data. Oceanologia 57, 116-131 Tamburini, C., Canals, M., Durrieu de Madron, X., Houpert, L., Lefèvre, D., Martini, S., D'Ortenzio, F, et al. (2013) Deep-Sea Bioluminescence Blooms after Dense Water Formation at the Ocean Surface. PLoS One. 8(7): e67523. Doi: 10.1371/journal.pone.0067523. Touratier, F., Goyet, C., 2011. Impact of the Eastern Mediterranean Transient on the distribution of anthropogenic CO2 and first estimate of acidification for the Mediterranean Sea, Deep-Sea Research Part I-Oceanographic Research Papers 58, 1, 1-15, doi: 10.1016/j.dsr.2010.10.002. Trincardi, F., Foglini, F., Verdicchio, G., Asioli, A., Correggiari, A., Minisini, D., Piva, A., Remia, A., Ridente, D., Taviani, A. (2007). The impact of cascading currents on the Bari Canyon System, SW-Adriatic Margin (Central Mediterranean), Marine Geology 246, 2-4, 208-230, doi: 10.1016/j.margeo.2007.01.013. Vargas-Yáñez et al. (2010a). Climate change in the Western Mediterranean Sea 1900–2008. Journal of Marine Systems, 82(3), 171–176. https://doi.org/10.1016/j.jmarsys.2010.04.013 Vargas-Yáñez et al. (2010b). How much is the western Mediterranean really warming and salting? Journal of Geophysical Research, 115(C4), 04001. https://doi.org/10.1029/2009JC005816

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www.msfd-idem.eu Deliverable 2.2 9. DESCRIPTOR 8 AND 9: CONCENTRATIONS OF CONTAMINANTS/CONTAMINANTS IN FISH AND OTHER SEAFOOD FOR HUMAN COMSUMPTION

9.1. Contaminants in water Mercury

0 II The main dissolved Hg species in the seawater are elemental Hg in gas form (Hg (g)), complexes of Hg with various organic and inorganic ligands, and organomercurial compounds, mainly monomethyl mercury (MMHg) and dimethyl mercury (DMHg) (Horvat et al. 2003). Most mercury enters marine waters by wet or dry deposition or by river discharges, with a significant fraction in oxidized form (Fitzgerald et al. 2007).

Figure 9.1. Total mercury dissolved in water. Data corresponding to several depth sections are represented. Compiled from data reported in Cinnirella et al. (2013), Cossa et al. (1997), Ferrara et al. (1992), Horvat et al. (2003), Kotnik et al. (2007; 2017).

Elemental mercury in marine waters originates from transformations of HgII (Costa and Liss, 1999, 2000; Amyot et al. 1997) and organo-mercury compounds (Mason and Sullivan, 1999). The highest concentrations of total Hg in dissolved form in the Mediterranean Sea waters can be found in the most northerly part of the Adriatic Sea (Figure 9.1; see also Kotnik et al., 2014).

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Figure 9.2. Dissolved gaseous mercury in water. Compiled from data reported in Andersson et al. (2007), Fantozzi et al. (2013), Gardfeldt et al. (2003), Lanzillotta et al. (2002), Vaupotic et al. (2008).

0 Elemental Hg (Hg (g)) is a species whose concentration is usually supersaturated (Schroeder and Munthe, 0 1998), especially in surface waters. The concentrations of Hg (g) are rather uniform all along the water column (Figure 9.2). Despite the geological mercury anomaly in the Mediterranean Basin, the concentrations of various Hg compounds in water are generally lower than those found in Atlantic and Pacific waters (Mason et al., 2001a,b; Cossa et al., 2004; Mason and Sullivan, 1999; Laurier et al., 2003; Kotnik et al., 2007).

Laboratory and field studies indicate that photoproduction and bacterial activity are probably the main sources of volatile Hg in surface waters (Gardfeldt et al., 2001; Lanzillotta and Ferrara, 2001; Lanzillotta et 0 al., 2002). In the deeper water layers, the Hg (g) distribution shows a correlation with oxygen concentration, which points to redox processes related with microorganism activity as one of the main drivers of the generation of this mercury species. Accordingly, microorganism production and diffusion from sediment may be major sources of volatile Hg in the deep waters. However, tectonic activity, e.g. in the S Adriatic Pit, which is tectonically very active, and in other locations, could also be a main source of volatile Hg in the deep Mediterranean Sea (Kotnik et al., 2014). In any case, the origin of this bottom sea gaseous Hg is still an open question to be elucidated for clarification of the main sources of this metal in the open waters of the Mediterranean Sea. 107

www.msfd-idem.eu Deliverable 2.2 The concentrations of Hg associated to particles are highest in the Adriatic Sea (Figures 9.3A and B), which is consistent with the above indicated maximum concentrations of this metal in dissolved form also in this Mediterranean area (Figure 9.1). The particulate phase mercury is found in highest concentrations in the shallower waters, which again is consistent with inputs from heavily polluted rivers and other direct or indirect natural or anthropogenic Hg loads, especially in the northern and central parts of the Adriatic Sea (Kotnik et al., 2014).

Figure 9.3A. Mercury in the particulate phase (0-500 m). Compiled from data reported in Cossa and Coquery (2005), Ferrara and Maserti (1992), Ferrara et al. (1989).

High abundance of oxidised Hg compounds is observed over the water surface in the Mediterranean Sea, with concentrations significantly higher than those found in northern Europe (Sprovieri et al. 2003; Wängberg et al., 2001). The high solar radiation, humidity and temperature characteristics of this semi- enclosed basin may play an active role in the generation of these oxidized Hg forms (Sprovieri et al., 2003).

Although atmospheric deposition represents the major mercury input to the Mediterranean, the amount of this metal emitted from the surface waters is actually far greater than the fallout fluxes, and the Mediterranean Sea is a net Hg source to the atmosphere (Hedgecock et al., 2006).

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Figure 9.3B. Mercury in the particulate phase, below 500m. Compiled from data reported in Cossa and Coquery (2005), Ferrara and Maserti (1992), Ferrara et al. (1989).

Organochlorine compounds

Specific reviews of PCBs have been published (Tolosa et al., 1997). In general, the PCB concentrations for all investigated areas in the Mediterranean Sea were similar (0.1–2.5 ng L–1Aroclor eq), except in the Ligurian Sea where concentrations were higher, because of the influence of urban and industrial waste waters as well as river discharges. Decreasing concentration gradients were also found in transects offshore from continental sources. PCB concentrations in NW Mediterranean deep open waters in the 1980s were 100–500 pg L–1 Aroclor 1254, and in 2005 the concentrations of the ΣICES-7 congeners were 0.57 and 1.2 ng L–1 in the particulate and dissolved phases, respectively (Bayona et al., 1990). These concentrations were in the same range as those detected in the North Sea (14–574 pg L–1) and North Atlantic (2–21 pg L–1 for 18 congeners), where decreasing temporal trends were observed (Schulz et al., 1991).

Lindane concentrations in deep open waters were in the range of 0.06–0.12 ng L–1 (Albaiges et al., 2005).

DDT levels in deep open seawater collected in 1993-94 in the Western Mediterranean were of 0.1–0.7 pg L–1 and 0.4–2.8 pg L–1in the particulate and dissolved phases, respectively (Yilmaz et al., 1998). In the

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www.msfd-idem.eu Deliverable 2.2 continental shelf, the values were 1 and 4 pg L–1, respectively. The concentrations of dissolved PCBs (28– 63 pg L–1 as Σ12 congeners) were almost one order of magnitude higher than in the corresponding particulate phase. These concentrations exhibited a decreasing gradient from the continental shelf (3.5– 26.6 pg L–1) towards the deep open sea (1.7–6.6 pg L–1), with a relatively important enrichment in deep open sea stations located in higher productivity frontal zones. The concentrations observed in the Western Mediterranean were of the same order of magnitude as those reported in other marine regions (Tolosa et al., 1997)

Polycyclic aromatic hydrocarbons

The concentrations of 16 PAHs in samples from the North Aegean and Marmara Seas are in the range of 10 to 30 ng L–1and 0.2–7.4 ng L–1, respectively (Yilmaz et al., 1998; Telli-Karakoç et al., 2002). Data on individual PAHs in the water column of the Western Mediterranean have also been reported (Bouloubassi and Saliot, 1991; Dachs et al., 1997). Particulate PAHs (Σ16) are evenly distributed in subsurface waters, and their concentrations range from 200 to 750 ng L–1. The vertical profiles exhibited a decreasing concentration with depth, with a relative enrichment of the more condensed compounds, derived from pyrolytic sources and probably more refractory to degradation. The average concentrations of particulate PAHs in subsurface waters (Σ14PAHs) were higher off Barcelona (dissolved: 24.5 ng L–1, particulate: 22.0 ng L–1) than in Banyuls (France) (dissolved: 9.71 ng L–1and particulate: 10.6 ng L–1), consistent with the local pollution sources.

Alkyltins

Contamination of surface waters by tributyltin (TBT) is as high as 0.47 ng L–1, 20 km off-shore, and lower than 0.08 ng L–1 in the open sea. Contamination of abyssal deep waters reach a maximum of 0.04 ng L–1 at a depth of 1200 m and is always significant at 2500 m (Michel and Averty, 1999).

Radionuclides

The general aspects of radioactivity distribution in the Mediterranean Sea have been summarized in 2010 by Delfanti and Papucci. The conclusion of their work states that the Mediterranean Sea is a ‘‘clean environment’’, from the radiological point of view. In fact it received fallout only from nuclear weapon testing and the Chernobyl accident. Local sources, such as discharges from the nuclear industry are small and usually reach the sea through rivers, where they are retained for some time. They affect the sediments in relatively small areas on the shelf, close to the river mouth. Concentrations in biota are presently undistinguishable from those in areas without point sources. Hence the relevance of these contaminants lie on their usefulness as process tracers more then on their negligible impact on the state of environment. The radioactivity in seawater in deep Mediterranean Sea comes from two different origins: anthropogenic release (mainly 137Cs, 90Sr, 239,240Pu and 241Am) and Natural Occurring Radioactivity Materials (NORM) (Betti et al., 2004). Anthropogenic radionuclides entered the Mediterranean Sea mainly through the atmosphere and through rivers. When in surface water, depending on their biogeochemical behaviour, they can follow the water masses (soluble, ‘‘conservative’’ radionuclides like 3H, 137Cs, and 90Sr) or be scavenged by the particulate matter in suspension or taken in the biological cycle (‘‘nonconservative’’ radionuclides like 239,240Pu and 241Am). The main inputs in the Mediterranean Sea are summarized in Table 9.1.

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www.msfd-idem.eu Deliverable 2.2 Table 9.1. Anthropogenic radionuclides delivered to the Mediterranean Sea from different sources (reference time: 2018). Source 137Cs 239,240Pu References (PBq) (TBq) Global fallout up to 2000 3.4 200 Hardy et al. 1973, Delfanti and Papucci 2010 Chernobyl fallout 1.2 0.02 Papucci et al. 1996 Global fallout 2000–2010 0.02 1.2 Lee et al. 2003 Marcoule reprocessing plant 0.02 0.37 Charmasson 2003, Eyrolle et al. 2004 Black Sea up to 1986 0.05 3 UNEP 1992 Black Sea 1986–2010 0.2 0.4 UNEP 1992, Egorov et al. 1999 Exchange with Atlantic Ocean up to 1986 0.7 -40 UNEP 1992

Exchange with Atlantic Ocean 1990-2010 - -6.8 Gascò et al. 2002

Input from rivers 0.004 - Delfanti and Papucci 2010 Total 5.7 158.2

In the last decade and following the Chernobyl accident, there was a significant increase in the number of 137Cs measurements carried out in the water column of the open sea. The 137Cs distribution in intermediate and deep waters along the Mediterranean Sea is presented in Table 9.2 and Figure 9.4.

Figure 9.4. Range of 137Cs in the water column (300-bottom) (reference time: 2018). Compiled from data ENEA (2007) and Tsabaris (2014). 111

www.msfd-idem.eu Deliverable 2.2 In general, the 137Cs vertical profiles are more homogeneous moving westward, and the maximum concentrations are observed in Aegean Sea, where the Chernobyl contribution is higher. The dataset on 239,240Pu and transuranics, in general, is much smaller because of the low concentrations, absence of significant local sources, and cumbersome radiochemical procedures for their determination. Pu chemistry in seawater is complex: so its distribution is different from that of the conservative radionuclides. Table 9.2 shows the inventories in the water column of 137Cs and 239,240Pu in the deep Mediterranean Sea.

Table 9.2. Inventory of 137Cs and 239,240Pu in the intermediate and deep water of the Mediterranean Sea (reference time: 2018) 137Cs 137Cs 239,240Pu 239,240Pu (TBq) (Bq m−2) (PBq) (Bq m−2) Water column 300–1000m 600 9 17 Water column 1000–1500m 800 12 18 Water column 1500–2000m 660 10 12 Water column 2000–3000m 1000 1.0 15 15 Water column 3000–4000m 1300 0.3 20 4.7

The general picture above presented is sufficient to delineate future trends: most anthropogenic radionuclides (90% of 137Cs and 50% of 239,240Pu) reside in the water column. For 239,240Pu, important reservoirs are the shelf and slope sediments, containing another 25% of its delivery. The Mediterranean total inventories are going to decrease in the future since:

• there are no significant sources of anthropogenic radionuclides; • there is a net outflow of 239,240Pu at Gibraltar, not balanced by the Black Sea input; • 137Cs is decreasing due to physical decay (half life = 30,1 yrs), not balanced by the input from the Black Sea.

Recently the distributions of 236U and 129I was assessed both in the Western and in the Eastern Basins of the Mediterranean Sea. These new measurements allow to trace the movement of water masses along the Mediterranean Sea (Castrillejo et al 2017).

Naturally occurring radionuclides that can be found in deep sea environment are 40K and radionuclides from the uranium and thorium decay chains, such as 226Ra, 210Pb and 210Po. The 210Po is the dominant sources of radiation doses through ingestion of seafood. Their natural concentrations may be increased due to anthropic activities, such as off-shore oil production. There is hence a general lack of public measurements near the deep fields of extraction in the Eastern Mediterranean Sea.

Settling fluxes.

The role of submarine canyons for the transport of terrigenous materials accumulated in the continental shelf to the deep marine areas has been outlined (Canals et al., 2006; Puig et al., 2014; Rumin-Caparros et al., 2016). Mediterranean canyons may funnel large volumes of sediment and organic matter from the

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www.msfd-idem.eu Deliverable 2.2 shallow continental shelf to the deep margin and basin (Estournel et al., 2005; Canals et al., 2009; Palanques et al., 2012; Pasqual et al., 2013). They experience occasional sediment gravity flows which may be triggered by mass failure, river flooding or dense shelf water cascading (DSWC), a type of current that is driven by seawater density contrast (Palanques et al., 2006; 2009; Puig et al., 2008). Here we summarize the effects of this transport mode in deep-sea environments using data from the area of the Mediterranean Sea where these processes have been studied experimentally.

Polychlorobiphenyls.

The PCB concentrations (sum of 7 ICES congeners) in the settling particles collected in the northwestern Mediterranean at depths of 300-1900 m between October 2005 and October 2006 range between 2.7 and 160 ng·g-1 (Salvado et al., 2012a). The average concentrations in the different sediment trap stations range between 16 and 26 ng·g-1 (Table 9.3). These concentration values are lower than those found in the same area 20 years ago (31-1500 ng·g-1; Fowler et al., 1990), although in that case they were reported as Aroclor 1254 equivalents.

The congener profiles observed in the Mediterranean Sea (Fowler et al., 1990; Tolosa et al., 1997; Dachs et al., 1996; Mandalakis et al., 2005) are similar to those found in other marine settling particles, such as the Sargasso Sea (Knap et al., 1986) and the Arabian Sea (Dachs et al., 1999). Comparison of the settling fluxes during common sedimentation found in the western Gulf of Lion with those in other sites shows similar values as those in the Arabian Sea (1988-91; 4017 m), 16 ng·m-2·d-1 (13 PCB congeners) (Dachs et al., 1999), the Alboran Sea (1992; 1200 m), 18.5 ng·m-2·d-1 (7 ICES PCB congeners) (Dachs et al., 1996), or in the Bothnian Sea (1991-93; 115 m), 34 ng·m-2·d-1 (68 PCB congeners) (Strandberg et al., 1998). However, the settling fluxes are much lower in other marine areas, e.g. eastern Mediterranean Sea (2001; 186-1426 m), 0.63 ng·m-2·d-1 (54 PCB congeners) (Mandalakis et al., 2005), Sargasso Sea (1978-80; 3200 m), 4.3 ng·m- 2·d-1 (12 PCB congeners) (Knap et al., 1986), Swedish western coast (1999-01; 125-155 m), 5.5 ng·m-2·d-1 (6 PCB congeners) (Palm et al., 2004). The observed PCB settling fluxes during DSWC in the Mediterranean are the highest ever described in marine continental slopes and open sea areas.

DDTs.

The DDT concentrations in the settling particles (sum of 2,4’-DDE, 4,4’-DDE, 2,4’-DDD, 4,4’-DDD, 2,4’-DDT, 4,4’-DDT) collected at depths of 300-1900 m in the NW Mediterranean ranged between 0.8 and 85 ng·g-1 (Table 9.3). These concentrations are lower than those observed in a sediment trap study performed in the Lacaze-Duthiers Canyon (LDC; 3.1-64.5 ng·g-1) in 1985-86 (Fowler et al., 1990) The present concentrations from the NW Mediterranean are higher than those found in sediment traps deployed in the Alboran Sea (1.9 ng·g-1; Dachs et al., 1996), the Baltic Sea (8.1 ng·g-1; Strandberg et al., 1998) and the Arabian Sea (4.1 ng·g-1; Dachs et al., 1999).

The observed DDT settling fluxes during common sedimentation in the Gulf of Lion are in the same order of magnitude than those found in the Bothnian Sea (1991-93; 115 m), 16 ng·m-2·d-1 (Strandberg et al., 1998), and higher than in the Arabian Sea (1988-91; 4017 m), 0.6 ng·m-2·d-1 (Dachs et al., 1999), and the Alboran Sea (1992; 1200 m), 1.4 ng·m-2·d-1 (Dachs et al., 1996). Again, the DDT settling fluxes observed during DSWC in the Mediterranean are the highest ever described in marine continental slopes and pelagic areas.

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www.msfd-idem.eu Deliverable 2.2 Chlorobenzenes.

Total chlorobenzenes (CBzs) encompass the sum of pentachlorobenzene (PeCB) and hexachlorobenzene (HCB). Even though the production of PeCB and HCB was banned globally under the Stockholm Convention (2004), these compounds are still released into the environment as by-products of the synthesis of several organochlorine solvents and as a result of backyard trash burning and municipal waste incineration (Bailey, 2001). The concentrations of these compounds in the settling particles collected at depths of 300- 1900 m in the NW Mediterranean vary between not detected and 21 ng·g-1. The average concentrations range between 3.8 and 6.6 ng·g-1 (Table 9.3). These values correspond to average percentages of 31% and 67 % for PeCB and HCB, respectively, and both compounds show the same temporal pattern in the sediment trap stations. The temporal and geographic distribution of these CBzs is very similar to that observed for the PCBs, with the highest concentrations found in the deepest traps. In comparison to other studies, the HCB concentrations (nd-14.5 ng·g-1; average 3.5 ng·g-1) are similar to those found in the Baltic Sea (0.49 – 3.4 ng·g-1; average 1.9 ng·g-1; Strandberg et al., 1998) and in the LDC in 1985-86 (0.3-3.7 ng·g- 1; average 1.7 ng·g-1; Fowler et al., 1990).

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www.msfd-idem.eu Deliverable 2.2 Table 9.3. Settling fluxes (ng·m-2·d-1) and concentrations (ng·g-1) of PCBs, DDTs, CBzs and HCHs and total mass (TM; g·m-2·d-1) and organic carbon (TOC; %) in particles collected in sediment traps deployed along the Cap de Creus Canyon (CCC), Lacaze-Duthiers Canyon (LDC) and the Southern Open Slope (SOS) (Salvado et al., 2012a). nd = not detected; nq = not quantified. TMc TOC PCBsd DDTs CBzs Lindane Station Flux % Fluxe Conc Flux Conc Flux Conc Flux Conc CCC300a Max 12.5 2.9 320 30 92 7.5 62 5.1 6.1 0.7 Min 3.8 1.4 99 21 28 5.5 5.2 1.2 2.0 0.4 Meanb 7.0 1.8 180 25 50 7.0 30 4.2 3.4 0.5 CCC1000 Max 90 2.9 960 51 630 34 340 5.1 180 7.2 Min 1.1 0.8 28 8.5 11 4.0 nd nd nd nd Mean 12 1.3 190 16 100 8.6 57 5.0 26 2.4 CCC1500* Max 6.3 9.5 120 95 58 36 53 21 18 5.9 Min <0.1 1.4 1.4 12 0.6 6.4 nq nq 0.1 0.6 Mean 1.3 1.9 28 20 14 10 8.8 6.6 2.7 2.0 CCC1900* Max 3.2 6.5 82 81 35 28 20 14 19 7.5 Min <0.1 1.4 2.6 11 1.0 5.5 0.4 1.0 0.1 0.6 Mean 0.7 2.1 18 25 6.2 9.3 3.1 5.0 2.5 4.1 LDC300 Max 18 3.5 460 33 350 25 120 9.7 52 6.5 Min 0.3 1.4 6.4 9.1 3.9 4.4 2.4 2.5 1.7 0.3 Mean 6.7 1.8 125 19 57 9.5 29 4.8 8.9 1.6 LDC1000 Max 11 4.5 280 46 130 19 77 12 51 11 (500mab) Min 0.6 1.5 15 14 5.7 7.2 1.4 1.5 1.0 0.6 Mean 2.9 2.2 80 26 30 10 14 5.2 8.5 3.3 LDC1000* Max 40 2.3 570 35 260 14 140 5.4 96 6.4 Min 2.1 1.2 25 11 13 5.3 6.1 2.5 nq nq Mean 6.7 1.5 105 18 42 8.0 20 3.8 12 2.4 LDC1500 Max 12 5.4 270 160 100 49 61 19 45 11 Min <0.1 1.4 0.8 2.7 0.3 0.8 0.1 0.3 <0.1 0.1 Mean 1.4 1.9 30 21 11 8.2 6.3 4.8 3.9 3.1 SOS1000* Max 34 4.2 640 59 2900 85 100 8.9 44 5.5 Min 0.2 1.2 3.4 6.5 2.0 3.4 0.6 1.8 nd nd Mean 3.4 1.5 70 20 145 41 13 3.9 6.6 1.9 SOS1900* Max 5.7 6.5 64 140 41 44 20 17 13 7.4 Min <0.1 1.2 1.2 9.8 0.4 7.1 0.2 1.5 <0.1 0.8 Mean 0.8 1.9 20 18 7.4 8.7 3.3 4.2 2.2 3.8 aSediment trap material was collected fortnightly from mid October 2005 to late October 2006. The number refers to the depth of the water column. The traps were located at 30 m above sea bottom except LDC1000 (500 m ab). Due to low amounts of collected particles during July 15, 2006 to October 30, 2006 no samples were analyzed in CCC1500, CCC1900 and SOS1900.

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www.msfd-idem.eu Deliverable 2.2 b OC n OC Mean concentrations for each specific OC were calculated by as follows: M = ∑i=1(Ci · n OC TMi)/ ∑i=1 TMi where i is the indicator of each cup, Ci is the concentration of the OC in cup i, TMi the total mass collected in cup i and n the total number of cups. c Total mass fluxes in each cup: TMi = TMi/Di where Di is the number of days in which the cup was deployed. d7 ICES PCB Congeners e OC n OC n OC Mean fluxes for each specific OC were calculated as: MF = ∑i=1(Fi · TMi)/ ∑i=1 TMi where Fi is the flux of the OC in cup i and the other variables are defined as above.

The average deposition fluxes during DSWC range between 6.6 and 200 ng·m-2·d-1 whereas the average fluxes during common sedimentation are 1.3-37 ng·m-2·d-1 (Table 9.4). This flux range during the period without cascading is observed in LDC and is consistent with previous flux measurements, 9.2 ng·m-2·d-1, in this canyon in 1985-8624. These values are higher than those found in the Bothnian Sea (Strandberg et al., 1998), 3.7 ng·m-2·d-1, as for DDTs and PCBs. Again, the deposition fluxes during DSWC are the highest ever reported in the literature.

Lindane.

γ-HCH is the only hexachlorocyclohexane isomer found in most of the samples collected at depths of 300- 1900 m in the NW Mediterranean. The concentrations of lindane in the settling particles vary between not detected and 11 ng·g-1. The average trap concentrations in each mooring range between 0.5 and 4.1 ng·g-1. These concentrations are similar to those found in the Baltic Sea, 3.6 ng·g-1 (Strandberg et al., 1998), and in the LDC 20 years ago, 2.8 ng·g-1 (Fowler et al., 1990), but lower than those found in the Alboran Sea, 24 ng·g-1 (Dachs et al., 1996).

The average fluxes during common sedimentation are higher than the average fluxes observed in the Alboran Sea, 4.2 ng·m-2·d-1 (Dachs et al., 1996), but lower than those found in the Bothnian Sea, 34 ng·m- 2·d-1 (Strandberg et al., 1998), and in the LDC in 1985-86, 24 ng·m-2·d-1 (Fowler et al., 1990). However, the deposition fluxes during DSWC are similar to those from these two last cases and much higher than those recorded in the Alboran Sea (Dachs et al., 1996).

Polycyclic aromatic hydrocarbons

The total PAHs content in the sediment trap samples collected at depths of 300-1900 m in the NW Mediterranean varied between 370 and 2700 ng·g-1 (Salvado et al., 2017). The average concentrations ranged between 720 and 1100 ng·g-1. It is difficult to compare total PAH concentrations with those observed in previous studies because of differences in the individual compounds included in the sum. The concentrations observed in the present study were similar to those found in the Alboran Sea (1200 m; ∑PAH15=100-2800 ng·g-1; Dachs et al., 1996), the Baltic Sea (15-30 m; ∑PAH18=1200-1530 ng·g-1; Broman et al., 1988) and the Ligurian Sea (200-2000 m; ∑PAH12=500-4070 ng·g-1; Lipiatou et al., 1993). They were, however, significantly higher than those found west of the Island of Sardinia (in the western Mediterranean Sea; 2875 m; ∑PAH25=248-874 ng·g-1; Bouloubassi et al., 2006), the eastern

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www.msfd-idem.eu Deliverable 2.2 Mediterranean basin (186-2867 m; ∑PAH35=259-468 ng·g-1; Tsapakis et al., 2006), the western Arabian Sea (4017 m; ∑PAH15=102-719 ng·g-1; Dachs et al., 1999) and the Kaoping submarine canyon (Taiwan) (60-280 m; 200-440 ng·g-1; Fang et al., 2009).

The qualitative composition of the PAH mixtures is dominated by retene (> 150 ng·g-1), phenanthrene (> 100 ng·g-1) and fluoranthene (80 > ng·g-1; Figure 9.5). Other major compounds are chrysene+triphenylene, indeno[1,2,3-cd]pyrene, benzo[ghi]perylene and pyrene (> 50 ng·g-1). In addition, other parent PAH such as benz[a]anthracene, benzo[ghi]fluoranthene, benzo[k]fluoranthene, benzo[e]pyrene, benzo[a]pyrene and perylene are found in average concentrations above 20 ng·g-1. Retene may originate from diterpenoid diagenesis (Alexander et al., 1987; Simoneit et al., 1986; Wakeham et al., 1980) and from combustion of conifer wood (Ramdahl, 1983).

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www.msfd-idem.eu Deliverable 2.2 Table 9.4. Mean settling fluxes (ng·m-2·d-1) and inventories (μg·m-2) of organochlorine compounds during the dense seawater cascading episode (D; January-March 2006) and during the period of common sedimentation (R) (Salvado et al., 2012a). PCBs DDTs CBzs Lindane OCsd Flux Inventorya Flux Inventory Flux Inventory Flux Inventory Flux Inventory CCC1000 D 570 8.8b 51 74c 320 13b 29 81c 200 20b 18 87c 98 41b 8.8 93c 1200 12b 110 80c R 65 18 24 6.6 9.9 2.7 2.4 0.66 100 27.5 CCC1500 D 63.5 4.2 5.7 58 32 4.5 2.9 60 24 9.2 2.2 76 5.5 3.4 0.495 53 125 4.8 11 61 R 15 4.1 7.0 1.9 2.6 0.71 1.6 0.44 26 7.5 CCC1900 D 24 1.5 2.2 33 11 2.3 0.99 45 6.6 3.7 0.59 55 5.5 3.9 0.495 56 47 1.9 4.2 39 R 16 4.4 4.7 1.2 1.8 0.49 1.4 0.385 24 6.6 LDC300 D 24 0.15 2.2 5 12 0.16 1.1 5 8.1 0.21 0.73 7 5.1 0.51 0.46 14 49 0.23 4.4 5 R 160 44 73 20 37 10 10 2.75 290 80 LDC1000 (500) D 135 2.2 12 41 60 3.0 5.4 50 35 5.4 3.15 64 23.5 6.7 2.1 71 250 2.7 23 48 R 62 17 20 5.5 6.5 1.8 3.5 0.85 93 25 LDC1000 D 400 5.3 36 63 190 6.8 17 69 95 7.9 8.55 72 79 33 7.1 91 760 6.3 69 68 R 76 21 28 7.7 12 3.3 2.4 0.66 120 33 LDC1500 D 80 6.7 7.2 69 33 9.2 3.0 75 20 13 1.8 81 13 22 1.8 92 150 8.3 13 73 R 12 3.3 3.6 0.99 1.5 0.41 0.60 0.165 18 4.9 SOS1000 D 220 12 20 80 530 79 48 96 42 14 3.8 82 20 12 1.8 80 810 27 73 90 R 18 4.9 6.7 1.8 3.0 0.82 1.7 0.46 30 8.2 SOS1900 D 31 2.6 2.8 46 17 4.2 1.5 58 8.9 6.8 0.80 69 6.3 9.0 0.57 75 63 3.5 5.7 54 R 12 3.3 4.0 1.1 1.3 0.36 0.70 0.19 18 4.9 aInventories during DSWC and common sedimentation are calculated by multiplication of the average fluxes during these two periods by 90 and 275 days, respectively. bD/R ratio of deposition fluxes. c100·D/(R+D) ratio of inventories. dOCs, sum of all measured organochlorine compounds.

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Deliverable 2.2 Phenanthrene is related to pyrolytic sources, oil spills or diagenetic processes (Sicre et al., 1987). In the NW Mediterranean this compound is mainly related to oil spills as indicated by the high proportion of methylphenanthrenes (MePh/Ph ratios between 0.3-2.4) and dimethylphenanthrenes (DiMePh/Ph ratios between 0.38-2.2). In addition, the distribution of parent PAH dominated by catacondensed structures is characteristic of pyrolytic sources (Simoneit, 1985). Thus, the ratios of fluoranthene/(fluoranthene+ pyrene), benz[a]anthracene/(benz[a]anthracene + chrysene+triphenylene) and indeno[1,2,3- cd]pyrene/(indeno[1,2,3-cd]pyrene+benzo[ghi]perylene) were 0.65-0.73, 0.35-0.43 and 0.5-0.57, respectively. These ratios are consistent with both grass/wood (biomass) and coal combustion as sources (Sicre et al., 1987; Yunker et al., 2002, 2011, 2015). These observed PAH distributions are similar to those found in other areas of the western Mediterranean Sea (Lipiatou et al., 1993; Dachs et al., 1996) and in Chesapeake Bay (Ko et al., 2003).

Figure 9.5. Average distributions of concentrations (ng·g-1) of individual PAHs deposited in the NW Mediterranean during the DSWC and in the absence of cascading (Salvado et al., 2017).

In general, the concentrations of PAH follow a rather uniform profile indicating the dominance of pyrolytic sources in the PAH mixtures. The PAH distributions were rather similar in conditions of DSWC or in the absence of cascading (Figure 9.5) which is consistent with the accumulation of a substantial amount of these hydrocarbons in the mid-shelf mud belt formed from the sediments and subsequent wash out towards the deep shelf and basin during DSWC.

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Deliverable 2.2

The role of DSWC in the transport of pollutants

The DSWC pulses have capacity of OC transport over several tens of kilometers offshore in a few days. Peak fortnightly-averaged settling fluxes of 960 ng·m-2·d-1, 630 ng·m-2·d-1, 340 ng·m-2·d-1 and 180 ng·m-2·d- 1 for PCBs, DDTs, CBzs and lindane, respectively, are observed. In the LDC, peak settling values are also found at 1000 m depth, 280 ng·m-2·d-1, 110 ng·m-2·d-1, 77 ng·m-2·d-1 and 51 ng·m-2·d-1, respectively.

The enhanced settling fluxes of OCs during DSWC are paralleled by increases of total mass flux. Representation of the settling fluxes of OCs and total particulate organic carbon from the fortnightly- averaged samples collected during DSWC and common sedimentation show very strong correlation coefficients (r2 = 0.94, 0.89, 0.92 and 0.79 for PCBs, DDTs, CBzs and lindane, respectively). The changes in deposition fluxes are therefore related to the general dynamics of the settling particles. Accordingly, during common sedimentation, the observed increase in fluxes of PCBs, DDTs and CBzs occur in parallel to increases of total mass flux.

The depositional fluxes of PAHs in the NW Mediterranean range between 23 and 55000 ng·m-2·d-1 (Table 9.5). The average fluxes vary between 660 and 8700 ng·m-2·d-1 (Table 9.5). All traps except one show higher settling PAH fluxes in the period of DSWC than in the absence of cascading (Table 9.5). These differences were also observed for all individual PAHs in all traps with the only exception of fluorene in CCC1900 (Table 9.5). This compound is the most water soluble of the PAHs analyzed. Its distinct behaviour may reflect lower association to particles and therefore lower influence of DSWC, which may be noticed to a higher extent in this trap situated at deepest water depth.

During DSWC episodes a maximum flux of 55000 ng·m-2·d-1 was recorded (Table 9.5). Average PAH deposition fluxes between 9800 and 26000 ng·m-2·d-1 were observed in moorings located at 1000 m of water depth during DSWC. The LDC1000 trap situated close to the bottom (30 mab) exhibited higher fluxes, 12000 ng·m-2·d-1, than the one at shallower water depth (500 mab), 7100 ng·m-2·d-1 (Table 9.5), indicating that the cascading effect essentially encompasses the transfer of particles and PAH to the deeper water column in association with sediment transport.

Comparison of the average settling fluxes showed 12-20 times higher values during DSWC (Table 9.5). The high PAH fluxes recorded during DSWC parallel those of total mass flux (TMF), thus evidencing the close association between these compounds and the sediment load carried by cascading waters. The association between TMF and high molecular weight PAHs of pyrolytic origin has also been observed in other sediment trap studies (Dachs et al., 1996; Raoux et al., 1999). It protects them from degradation (Fernandez et al., 2002) and enhances their transport to the bottom sediments (Baker and Eisenreich, 1990).

Comparison of the PAH settling fluxes found in the NW Mediterranean in the absence of cascading with those observed in other sites show values in the same range as those in the coast of Monaco (1989-90; 200 m), 1300-7000 ng·m-2·d-1 (Raoux et al., 1999), in the Ligurian Sea (1987; 200-2000 m), 500-4100 ng·m- 2·d-1 (Lipiatou et al., 1993), in the Baltic Sea (1989-90), 550-4250 ng·m-2·d-1 (Pettersen et al., 1997), or in

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Deliverable 2.2 the East China Sea (2006-2007; inner shelf), 5205 ng·m-2·d-1 (Lin et al., 2013). However, in other marine areas the observed settling fluxes are lower, e.g. in the Alboran Sea (1992; 1200 m), 100-2800 ng·m-2·d-1 (Dachs et al., 1996), or much lower, e.g. in the Kaoping canyon (Taiwan; 2004; 60-280 m), 200-440 ng·m- 2·d-1 (Fang et al., 2009), west of Sardinia (2001-02; 2875 m), 250-870 ng·m-2·d-1 (Bouloubassi et al., 2006), in the southern Ionian Sea (2001; 186-2867 m), 260-470 ng·m-2·d-1 (Tsapakis et al., 2006), in the Swedish western coast (1999-01; 125-155 m), 200 ng·m-2·d-1 (Palm et al., 2004), in the western Arabian Sea (1988- 91; 4017 m), 100-720 ng·m-2·d 1 (Dachs et al., 1999), and in the southwestern Black Sea (2007-2008; 965 and 1965 m), 10.4-310 ng·m-2·d-1 (Parinos et al., 2013). Thus, the PAH fluxes obtained in this study are the highest ever described in a marine open area.

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Deliverable 2.2 Table 9.5. Average fluxes of total mass flux (TMF; g·m-2·d-1), lipid biomarkers (µg·m-2·d-1), individual PAHs (ng·m-2·d-1) and total PAHs (ng·m-2·d-1) along transects in Cap de Creus (CCC) and Lacaze-Duthiers (LDC) submarine canyons and in the Southern Open Slope (SOS) during the dense shelf water cascading episodes (C; January- March 2006) and during pelagic sedimentation (P). ªC/P ratio of fluxes. (Salvado et al., 2017) CCC1000 CCC1500 CCC1900 LDC300 LDC1000 (500) LDC1000 LDC1500 SOS1000 SOS1900 C P C/Pa C P C/P C P C/P C P C/P C P C/P C P C/P C P C/P C P C/P C P C/P TMF 39 2.6 15 3.5 0.48 7.3 1.3 0.49 2.7 1.5 8.5 0.175 6.3 1.7 3.8 12 3.3 3.8 4.1 0.44 9.2 11 0.6 19 2.0 0.44 4.4

Flu 400 46 8.6 21 10 2.1 7.9 9.4 0.84 5.2 210 0.025 43 41 1.0 92 33 2.8 35 8.5 4.1 15 10 1.5 22 7.5 2.9 Phe 3400 290 12 250 54 4.5 102 40 2.6 60 945 0.063 580 130 4.4 1100 200 5.7 340 51 6.7 800 52 15 250 37 6.7 An 590 36 176 63 6.9 9.1 23 7.4 3.1 19 110 0.18 130 22 5.8 220 31 7.0 57 4.8 12 480 11 44 42 6.4 6.5 Fla 3300 205 16 300 35 8.6 120 35 3.4 120 740 0.16 680 190 3.7 1400 310 4.4 300 29 10 1400 49 29.5 200 38 5.4 Pyr 2400 120 19 210 25 8.5 58 17 3.4 78 435 0.18 300 87 3.5 615 150 4.1 220 16 14 1200 28 42 100 18 5.7 Ret 5100 600 8.5 810 120 6.7 290 160 1.8 230 2700 0.085 2000 385 5.2 2300 500 4.7 1500 77 20 1200 150 8.2 610 83 7.4 Bz(a)A 1400 88 15 145 13 11 73 13 5.4 56 270 0.21 430 67 6.4 820 110 7.2 150 9.2 16 740 18 40.5 130 17 7.7 Chry+Try 1900 160 12 210 25 8.5 98 24 4.1 91 515 0.18 610 130 4.6 1100 225 4.9 220 19 11 885 32 28 160 26 6.3 Bz(b+j)Fla 1400 120 12 150 17 8.8 46 15 3.0 74 410 0.18 320 75 4.3 580 140 4.3 170 14 12 585 24 24 74 16 4.7 Bz(k)Fla 1120 91 12 110 13 8.8 36 12 2.9 54 320 0.17 250 63 4.0 440 110 3.9 120 12 10 470 19 24.5 55 13 4.2 Bz(e)Pyr 900 84 11 95 12 7.8 34 12 2.8 47 290 0.16 230 59 4.0 410 110 3.9 100 11 9.2 405 18 22 51 13 4.0 Bz(a)Pyr 1200 57 21 76 8.5 9.0 26 8.6 3.0 27 170 0.16 130 37 3.5 230 65 3.5 74 7.6 9.7 300 12 24 39 9.5 4.2 Per 990 55 18 55 5.9 9.3 14 5.5 2.5 32 240 0.14 150 41 3.7 320 70 4.5 45 5.8 7.9 310 11 28 21 5.9 3.5 IndPyr 1000 180 5.8 130 25 5.2 87 32 2.7 62 500 0.12 530 180 3.0 870 300 2.9 160 27 5.7 455 34 13.5 130 33 4.0 Bz(ghi)Per 1200 130 9.0 140 21 6.7 82 23 3.5 67 420 0.16 560 130 4.4 900 230 4.0 150 20 7.3 470 32 15 120 27 4.5 DBz(ah)An 310 34 9.1 32 4.2 7.5 19 5.2 3.6 14 97 0.14 110 28 3.9 190 49 3.9 31 4.2 7.4 94 4.3 22 30 5.5 5.5 Total PAHs 26000 2300 12 2800 400 7.0 1100 420 2.7 1000 8300 0.12 7100 1700 4.2 12000 2600 4.4 3700 320 12 9800 500 20 2000 350 5.8 Flu, fluorene; Phe, phenanthrene; An, anthracene; Fla, fluoranthene; Pyr, pyrene; Ret, retene; Bz(a)An, benz[a]anthracene; Chry+Try, chrysene+triphenylene; Bz(b+j)Fla, benzo[b+j]fluoranthene; Bz(k)Fla, benzo(k)fluoranthene; Bz(e)Pyr, benzo(e)pyrene; Bz(a)Pyr, benzo(a)pyrene; Per, perylene; IndPyr, indene[1,2,3-cd]pyrene; Bz(ghi)Per, OC n OC n OC benzo[ghi]perylene; DBz(ah)An, dibenz[a,h]anthracene; Average fluxes for each compound: MF = ∑i=1(Fi · TMi)/ ∑i=1 TMi where Fi is the flux of the compound in cup i, TMi the total mass collected in cup i and n the total number of cups.

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Deliverable 2.2 These values are considerably higher than the observed atmospheric deposition of PAH in central sites of the western Mediterranean Sea, e.g. 130 ng·m-2·d-1 in Mallorca (Lipiatou et al., 1997), which again reinforces the role of the Rhone river as one major source of these hydrocarbons in the NW Mediterranean.

Total PAH fluxes (without retene) were linearly correlated with TMF (r = 0.963, n = 212, p < 0.05) which was consistent with the relatively uniform PAH concentrations observed in the settling particles within the NW Mediterranean. The flux of total benzofluoranthenes, phenanthrene and the other parent PAH also followed the same pattern than total PAHs.

Inventories

The overall significance of the enhanced flux deposition due to DSWC episodes can be assessed by calculation of the inventories by integration of the average fluxes during the time intervals in which DSWC and common sedimentation were predominant. The PCB inventories show that deposition during DSWC is more important than during common sedimentation in the mooring sites at 1000 and 1500 m depth, those with highest settling fluxes during DSWC (Table 9.4). In these sites, the accumulated PCBs during the DSWC episodes represented about 58-80% of the total annual input. In the moorings at 1900 m depth, the settling PCBs during DSWC involved 33-46% of the whole annual budget, whereas in LDC300 they only involved 5% (Table 9.4). The lower rate in LDC1000 (500) than in LDC1000, 41% vs 63%, respectively, was consistent with the physical characteristics of the DSWC episodes involving formation and sinking of dense water close to the continental shelf and main displacement through the slope by the bottom.

Similar figures were obtained for DDTs, in which sedimentation during DSWC at 1000 and 1500 m depth involve 60-96% of the annual input, at 1900 m depth 45-58% and at 300 m depth only 5% (Table 9.4). The inventories of CBzs showed that sedimentation during DSWC was very significant at 1000 m and 1500 m depth, 72-87%, but also at 1900 m depth, 55-69%, and not in LDC300, 7% (Table 9.4). The same was the case for lindane, with 53-93% in the mooring traps at 1000 and 1500 m depth, 56-75% at 1900 m depth and 14% at 300 m depth (Table 9.4). In all these cases higher inventories were found in the trap located at 30 mab depth than at 500 mab depth, e.g. 69% vs 50%, 72% vs 64% and 91% vs 71% for DDTs, CBzs and lindane, respectively, which was consistent with the preferred transport close to the bottom sea as consequence of the formation of dense waters. Representation of the summed inventories of all measured compounds (Table 9.4) showed an area of highest settling of OCs during DSWC that was related to the morphology of the canyons but that also extends outside them, in the open slope region (Figure 9.6).

The PAH inventories showed that deposition during DSWC was higher than in non-cascading conditions in the mooring sites at 1000 m depth, where fluxes were the highest during DSWC. In these sites, the PAHs that accumulated during DSWC events represented about 59-86% of the total annual input. In the moorings at 1900 m depth, the settling PAHs during DSWC involved 46-65% of the whole annual budget and in LDC300 they only involved 3.9%. The differences between these moorings indicated that a substantial proportion of the PAHs carried by these DSWC were sedimented at depths of 1000 m, which is consistent with previous observations on particle dynamics under DSWC (Canals et al., 2006).

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Deliverable 2.2

Figure 9.6. Inventories (µg·m-2) of settling organochlorine compounds (sum of PCBs, DDTs, CBzs and lindane) during DSWC (red) and during common sedimentation (in orange) (Salvado et al., 2012a). The acronyms refer to Cap de Creus (CCC) and Lacaze-Duthiers (LDC) canyons and Southern Open Slope (SOS) sites. Numbers besides the acronyms refer to water column depths.

Radionuclides

Only few data are present in scientific literature regarding anthropogenic radionuclide in the deep environment of the Mediterranean Sea and they are almost limited to the Western basin.

Plutonium and americium experiments have demonstrated the usefulness of these contaminants as tracers for delineating particulate transport processes in the northwestern Mediterranean. Transuranic fluxes in the Lacaze-Duthiers Canyon, the first reported for the Mediterranean, were considerably higher than those previously measured in the North Pacific and result in short, upper layer residence times of 2.5 and 0.14 years for 239,240Pu and 241Am, respectively (Fowler et al, 1990).

The observation that 241Am/239,240Pu activity ratios in unfiltered NW Mediterranean seawater are 6 times lower than those in the north Pacific suggests the existence of a specific mechanism for enhanced scavenging and removal of 241Am from the generally oligotrophic waters of the open Mediterranean. It is proposed that atmospheric inputs of aluminosilicate particles transported by Saharan dust events which frequently occur in the Mediterranean region could enhance the geochemical scavenging and resultant removal of 241Am to the sediments. While deep water column (e.g., 0–2,000 m) inventories have not

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Deliverable 2.2 substantially changed over the 10–15 yr period examined, the fraction of the 239,240Pu inventory at a given depth interval in the deeper layers has increased considerably, demonstrating the slow, downward movement of 239,240Pu in the northwestern Mediterranean. 241Am is removed much more rapidly from the water column than 239,240Pu (Fowler et al, 2000).

Regarding anthropogenic radionuclides in the Eastern Mediterranean Sea, a particular attention is due to the deposition of Chernobyl-derived 137Cs that was particularly heavy onto the Black Sea (1.7e2.4 PBq) and the Aegean e Ionian Seas along the Greek coasts (800 TBq). Moreover, the inflow of surface Black Sea water through the Dardanelles Straits still constitutes a continuous point source of 137Cs into the North Aegean Sea (Tsabaris et al., 2014).

9.2. Contaminants in sediments About 20-30% of the global marine primary production occurs in continental margins (Wollast et al., 1991). Massive amounts of terrestrial organic carbon are also deposited in these areas, which are key sites for the global burial of organic carbon in marine sediments (Hedges and Keil, 1995; Goñi et al., 1997; Sanchez-Vidal et al., 2013). In continental margins the transfer of organic matter to deep environments occurs either through pelagic settling of biogenic particles generated in the photic zone or by lateral transport (advection) of materials accumulated in shallower sites due to storms, convection currents and internal waves (Sleath, 1987; Stastna and Lamb, 2008; Sanchez-Vidal et al., 2012).

The present section is concerned on the influence of these processes for the accumulation of these pollutants in deep-sea marine environments.

In general, the main inputs of organic pollutants into the Mediterranean are due to coastal discharges as consequence of the human activities in the continents or at the coast. The predominance of these inputs leads to major retention of pollutants in the sediments of coastal areas and therefore higher amounts of these compounds are observed in these environments. Accordingly, calculation of the average concentrations of PCBs, DDTs and HCB in harbours, coastal lagoons, coastal areas under urban influence, coastal areas under river influence, continental shelf and open sea sediments show a gradient from higher to lower pollutant concentrations the further away the sediments are from the coast. These differences are shown in Table 9.6. However, the topic of IDEM is concerned with the deep-sea environments and the present chapter will be devoted to these environments.

Atmospheric inputs are also significant, namely in locations far away from the coast. Nevertheless, the summary results of Table 9.6 show that these inputs are overrun by the coastal contributions.

Besides coastal and atmospheric inputs, the physical chemical properties of the pollutants also determine their accumulation patterns. Thus, strong differences are observed between the polychlorobiphenyl congeners PCB 28 and PCB 52 on one hand and PCB 101, PCB 118, PCB 153, PCB 138 and PCB 180 on the other. The separation of these congeners in two groups is consistent with the differences in logKOW, 5.67-

5.84 and 6.62-7.25, respectively, and logKOA, 7.93-8.22 and 8.9-10.1, respectively (Beyer et al., 2002).

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Table 9.6. Concentrations (ng g-1 dry weight) of PCBs, DDTs and HCB in Mediterranenan sediments according to their locations in the basin (Gomez-Gutierrez et al., 2007). Compound n Mean Median Range PCBs 7 ICES Harbours 22 566 151 2-3900 Coastal lagoons 72 42 4 0.5-1800 Urban influenced (<10 km) 41 52 16 0.2-650 River influenced (<10 km) 18 36 13 0.3-180 Others (continental shelf) 88 6 3 0.03-74 Open sea (depth > 1000 m) 8 2 2 1-5

PCBs Arochlor equivalents Harbours 111 824 191 1-17000 Coastal lagoons 225 155 39 0.9-5600 Urban influenced (<10 km) 364 271 30 0.4-16000 River influenced (<10 km) 139 82 28 0.5-1500 Others (continental shelf) 749 37 9 0.05-1200 Open sea (depth > 1000 m) 68 5 2 0.3-48

4,4’-DDT Harbours 38 940 26.5 0.03-21500 Coastal lagoons 59 4.6 0.6 0.05-79 Urban influenced (<10 km) 276 18 2.0 0.001-1600 River influenced (<10 km) 45 34 6.3 0.05-510 Others (continental shelf) 462 2.3 0.4 0.001-360 Open sea (depth > 1000 m) 23 0.8 0.6 0.02-2.9

Sum DDTs Harbours 60 7900 37 0.3-76000 Coastal lagoons 67 41 6 0.1-640 Urban influenced (<10 km) 287 31 5 0.003-1400 River influenced (<10 km) 121 60 11 0.01-830 Others (continental shelf) 582 6 2 0.003-370 Open sea (depth > 1000 m) 29 2 1 0.08-5

Hexachlorobenzene Harbours 11 1600 1900 0.1-2800 Coastal lagoons 14 240 2.7 0.3-2400 Urban influenced (<10 km) 75 5.4 0.2 0.01-60 River influenced (<10 km) 24 10.5 4.5 0.05-39 Others (continental shelf) 202 0.6 0.1 0.01-24.5 Open sea (depth > 1000 m) 13 0.2 0.2 0.04-0.8

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Deliverable 2.2 These properties are also relevant for the association of these pollutants to small grain size particles and sedimentary organic carbon. DDTs, pentachlorobenzene and HCB also show high octanol-water coefficients (logKow = 5.95-6.16, 4.92 and 5.48, respectively; Schoeib and Harner, 2002). There is a good affinity between most organochlorine compounds and organic carbon and the fine size sediment fractions (Pierard et al., 1996).

Accordingly, around 90% of the global sediment burial of organic carbon occurs in the continental margin (Hedges and Keil, 1995) and the persistent organic pollutants tend to strongly associate with organic matter. This converts the continental shelf into an important reservoir of these contaminants in the marine environment (Jonsson et al., 2003).

In addition to these predominant coastal inputs, there is an atmospheric transport of pollutants which involves a background of deposition over the whole Mediterranean involving mercury (Lamborg et al., 2014), organochlorine compounds (van Drooge et al., 2004b; Arellano et al., 2015), organobromine compounds (Arellano et al., 2014), polycyclic aromatic hydrocarbons (van Drooge et al., 2012) and other pollutants. In addition, Saharan dust may also be responsible for the long-range transport of some of these pollutants (Kuzu, 2016; Garrison et al., 2014).

Mercury

Concentrations of total Hg in Mediterranean sediments vary more than 2,000-fold within and among locations, depending on the source and natural or anthropogenic loadings (Kotnik et al., 2014). Consistently with the above mentioned summary of pollution inputs in Mediterranean sediments, the lowest total Hg concentrations are found in regions that are remote from fluvial and coastal point sources, including urbanized areas (Figures 9.7A and 9.7B). Direct atmospheric Hg deposition presumably represents the principal source in these remote areas.

The highest concentrations are observed in N Adriatic Sea, which is severely polluted with Hg originating from former Hg mining and smelting industry in Idrija and heavy industry of N Italy (Po River and Venice Lagoon) (Kotnik et al., 2015). Another important polluted area is the Gulf of Lion, stemming from polluted rivers (i.e. Rhone River) and industry of Toulon and Marseille areas (no data shown in Figures 7A and 7B). In any case, while coastal areas have been extensively studied, data on Hg concentrations and its species in deep-sea cores are still lacking.

It was found that total Hg concentrations in the off-shore marine sediments of the Mediterranean Sea, with an average of 0.1-0.08 µg g−1 (Figure 9.7B), are twice as high as the world-wide natural background (Baldi et al. 1983). This difference may be a consequence of terrestrial anthropogenic and natural underwater tectonic sources and Hg reactivity and biogeochemical behaviour (Kotnik et al., 2017).

The results of studies in the Mediterranean Sea indicate that although much research has already been performed, there are still gaps in our knowledge which require to be filled to fully understand the past and present dynamics of Hg in this marine environment, and thus in urgent need of further investigation (Kotnik et al., 2014).

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Figure 9.7A. Studied sites with sediment mercury concentrations higher than 1 µg/g. Compiled from data reported in Cossa and Coquery (2005), Ferrara and Maserti (1992), Ogrinc et al. (2007).

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Figure 9.7B. Studied sites with sediment mercury concentrations lower than 1 µg/g. Compiled from data reported in Cossa and Coquery (2005), Ferrara and Maserti (1992), Ogrinc et al. (2007). Hexachlorobenzene

Hexachlorobenzene (HCB) is widely distributed in the Western Mediterranean. The HCB concentrations in deep-sea areas of the NW Mediterranean range between not detected and 1.3 ng·g-1 (Figure 9.8; Salvado et al., 2013). Since this compound has a moderate water solubility, deep-sea sediments are not a major sink for it storage. The sediment concentrations observed in the NW Mediterranean are higher than those described in the Gulf of Alaska (Bering Sea), 0.03-0.08 ng·g-1, and in the Caspian Sea (Iran), 0.01-0.2 ng·g- 1. However, they are lower than those described in the Northern Baltic Sea (Table 9.5).

To date, the information available on the occurrence of this compound in deep-sea sediments is essentially restricted to the NW Mediterranean (Figure 9.8). In this area, the Rhone River appears as a major source for the occurrence of this compound. The significance of this coastal source is higher than other river courses such as the Ebro (Grimalt et al., 1988) or urban discharges. Atmospheric precipitation may constitute one main general source since this compound is currently found in the remote atmospheres both in gas and particle phases (van Drooge et al., 2001; 2002; 2004a).

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Figure 9.8. Concentrations of hexachlorobenzene in NW Mediterranean sediments. Compiled from data reported in Gomez-Gutierrez et al. (2007).

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Deliverable 2.2 Polychlorobiphenyls

Most of the reports describing the occurrence of PCBs in marine sediments still refer the concentrations of these compounds to Aroclor equivalent mixtures. Thus, most of the information on the occurrence of these compounds in the Mediterranean sediments is referred to these standards.

Figure 9.9A. Sediments with PCB concentrations (Aroclor equivalents) in the range of 5 and 50 ng g-1. Compiled from data reported in Gomez-Gutierrez et al. (2007).

As shown in Figures 9.9A and 9.9B, the western Mediterranean has been studied more intensively than the eastern basin and the north regions more than the southern areas in which there is virtually no information. Most of the studied deep-sea areas concern sites located not far away from the French and the Italian coasts (Figure 9.9A). In the North Western area there is a rather uniform concentration of PCBs in the order of 1-3 ng g-1 that constitutes a sort of background level in this area. However, as stated before, a more comprehensive description of the concentrations of these pollutants in the Mediterranean Sea would probably show another geographic distribution pattern.

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Deliverable 2.2 Studies on the measurement of these compounds from the composition of the individual congeners in open sea deep water areas show concentrations between 0.5 and 7 ng g-1 (Figure 9.10). Measurements using this analytical method have only been made in the NW Mediterranean. The distributions are consistent with predominant continental inputs, either rivers or urban discharges, and transfer to deep- water environments. In contrast with HCB, the Rhone River does not appear to be the main source in the area. Other inputs such as the Ebro River and urban discharges could also have contributed significantly to the transfer of these compounds to deep-sea sediments.

The measured concentrations of PCBs in the deep basin were about 2.5 ng g–1 dw in 2005 cruises and 1.4- 3.8 ng g–1 dw in 1990 (Table 9.7). These concentrations are also similar to those found during a cruise along the southern Mediterranean in 1975 (Fowler, 1987). Accordingly, no decrease in PCB concentrations is observed in this period.

Figure 9B. Sediments with PCB concentrations (Aroclor equivalents) below 5 ng g-1. Compiled from data reported in Gomez-Gutierrez et al. (2007).

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Table 9.7. Comparison between the mean PCB congener concentrations (ng·g-1) in open sea areas of the NW Mediterranean in 2005 and 1987 and 1991 (range concentrations in brackets). n, number of samples. Zone Sampling N PCB-52 PCB-101 PCB-153 PCB-138 PCB-180 Reference Sete canyon lowermost course period2005 2 0.6 (0.2-1.0) 0.4 (0.1-0.7) 0.3 (0.2-0.4) 0.6 0.2 (0.2-0.3) Salvado et al., 2013 Open slope south of Cap de Creus canyon 2005 3 0.4 (0.2-0.6) 0.2 (0.1-0.2) 0.5 (0.3-0.7) 1.1 (0.8-1.6) 0.3 (0.2-0.5) Salvado et al., 2013 Slope and deep basin 1990-91 8 0.1 (nd-0.1) 0.2 (0.1-0.4) 0.3 (0.2-1.0) 0.5 (0.2-1.3) 0.3 (0.2-0.9) Tolosa et al., 1995 Continental shelf, slope and deep basin 1990-91 6 0.3 (0.1-0.5) 0.4 (0.2-0.6) 1.0 (0.7-1.5) 1.3 (0.9-2.1) 0.8 (0.5-1.1) Tolosa et al., 1995

PCB 138 dominates the congener distribution of PCBs in almost all samples. The dominance of this congener is consistent with the composition of the technical mixtures used in the Mediterranean countries, in which it was a main constituent, and also with its physical-chemical properties since this congener is rather insoluble in water and therefore occurs in association with suspended particles.

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Figure 9.10. Sediments in which PCB concentrations measured by reference to the ICES congeners have been reported. Compiled from data reported in Gomez-Gutierrez et al. (2007).

Hexachlorocyclohexanes

Usually, the technical HCHs mixtures contain isomers in the following percentages: α-HCH: 55-80%, β- HCH: 5-14%, γ-HCH: 8-15%, δ-HCH: 2-16% (Metcalf, 1955). After forty years of extensive use worldwide, there has been a gradual replacement of technical HCHs by lindane, which is the only isomer exhibiting significant insecticide activity. Mixtures of the HCH isomers are found in the NW Mediterranean probably reflecting old HCH usages in the river basin. α- and γ-HCH have been described to be transformed into β- HCH after long aging (Walker et al., 1999; Willett et al., 1998). In the deep water sediments of this area γ- HCH is the predominant compound.

In general, the concentrations of HCHs in the deep basin are below detection limit, probably due to the high water solubility of these compounds. In the NW Mediterranean they range between 0.2-0.5 ng·g-1; Salvado et al., 2013) which is higher than in offshore sediments from the continental shelf in the Gulf of Alaska (0.003-0.048 ng·g-1; Iwata et al., 1994; Table 9.8). Unlike the concentration decreases found for PCBs, DDTs and CBzs over the past 20 years in the Rhone prodelta, comparison of the results in deep-

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Deliverable 2.2 environments (Salvado et al., 2012b; Marchand et al., 1988) show that the concentrations of lindane did not diminish in this area.

DDTs

Most of the information available on total DDT concentrations in deep-sea sediments from the Mediterranean Sea is concerned with the northwestern area. No information is available from the eastern basin or the southern Mediterranean areas. These pollutants range between 0.1 and 12 ng·g-1 (Figure 9.11). According to the data available, the Rhone River is one of the main sources of these compounds in the area which is consistent with the observations with other pollutants, e.g. HCB. However, other sites in the Catalan coast also seem to be relevant sources of these compounds in the deep-sea sediments. As described above, deep-sea canyons, e.g. Cap de Creus canyon, may have acted as preferential transport systems of pollutants accumulated in the continental slope towards deep water environments.

A comparison between the total DDT concentrations in these NW Mediterranean sediments and other deep-sea environments is summarized in Table 9.8. DDTs are the organochlorine compounds found in highest abundance. The concentrations of these compounds in the open continental slope are higher than in offshore sites from the Kara Sea (< 0.002-1.2 ng·g-1; Sericano et al., 2001), the Gulf of Alaska (0.006-0.17 ng·g-1; Iwata et al., 1994) and the western Xiamen Sea (0.01-0.43 ng·g-1; Maskaoui et al., 2005). They are similar to those found in the Northern Baltic Sea, 1.9-5.4 ng·g-1 (Strandberg et al., 1998) and in the East China Sea, nq-6.1 ng·g-1 (Yang et al., 2005).

Table 9.8. Concentrations of HCB, DDTs, HCHs and lindane (γ-HCH) (ng·g-1) in open sea areas of the Northwestern Mediterranean Sea and superficial sediments from the World Ocean. nd = not detected. Area Survey HCB DDTs HCHs γ-HCH References Sete canyon lowermost course year2005 0.25-0.9 0.8-1.9 0.2-0.7 0.2-0.5 Salvado et al., 2013 Open slope south of Cap de Creus canyon 2005 0.27-0.32 2.2-3.3 0.2-0.7 0.2-0.5 Salvado et al., 2013 Slope and deep basin (GoL) 1990-91 0.05-0.5 1.4-5.4 - - Tolosa et al., 1995 Gulf of Lion slope and deep basin 1987-88 0.05-0.5 1.4-5.4 - - Tolosa et al., 1995 Western Xiamen Sea (China) 1999 - 0.01-0.43 0.48-9 0.03-0.1 Maskaoui et al., Northern Baltic Sea 1991-92 0.8-0.9 1.9-5.4 5.0-7.0 1.3-1.8 Strandberg2005 et East China Sea 2002 - nq-6.1 nq-1.45 <0.06-0.13 Yang et al.,1998al.,2005 Caspian Sea, Iran 2001 0.01-0.2 0.06-3.9 0.03-0.6 0.001-0.04 De Mora et al., Kara Sea 1993 - nd-1.2 nd-0.6 0.03-0.11 Sericano et2004 al., Gulf of Alaska, Bering Sea 1990 0.03-0.08 0.006- 0.043- 0.003- Iwata et al., 19942001 0.17 0.25 0.048

4,4’-DDE is the compound in highest abundance in the DDT mixtures, which is indicative of old DDT residues, as the original DDT molecule loses HCl and is transformed into DDE in freshwater and marine environments (Wolfe et al., 1977). In anoxic conditions, DDD is the major DDT degradation product (Zoro et al., 1974). Higher DDD percentage in the DDT mixtures may reflect lower degrees of oxygenation in the shallow-water fine sediments. In any case, the high relative concentrations of DDE in the deep sediments of this area indicate the preferential accumulation of the most stable DDT residue, as commonly observed in many other aquatic environments.

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Deliverable 2.2 Considering that the NOAA ERL for DDTs is 1.6 ng·g-1 (de Mora et al., 2004), DDTs in the NW Mediterranean sediment dispersal system clearly exceed this quality standard in almost all sites.

The concentrations of DDT in the deep basin (Table 9.8) were similar to those found in a cruise along the southern Mediterranean in 1975 (Fowler, 1987). Finally, in the Gulf of Lion western canyons and in the open slope south of Cap de Creus canyon the concentrations (nd-7.3 ng·g-1) are very similar, or even higher, than those found in the continental slope and deep basin of the Gulf of Lion twenty years ago (1.4- 5.4 ng·g-1; Tolosa et al., 1995). These results show that the background concentrations of DDTs have not decreased during these observed time periods, which is in agreement with the results reported above for PCBs.

Thus, despite the dominance of the DDE metabolite, which indicates old mature DDT distributions, the persistence of these compounds all along the studied period suggest that the bans on the use of this insecticide did not stop the continental flows of old DDT residues towards the deep sea sedimentary environments of the Mediterranean Sea.

Figure 9.11. Concentrations of total DDTs in deep-sea sediments of the Mediterranean. Compiled from data reported in Gomez-Gutierrez et al. (2007).

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Deliverable 2.2 Polycyclic aromatic hydrocarbons

Usually, the distributions of these hydrocarbons are dominated by the series of 3 to 6 ring PAHs, considered to be typically combustion derived. This series encompasses from the 3-ring phenanthrene/anthracene to the 6-ring indeno[1,2,3-cd]pyrene and benzo[ghi]perylene, including the benzo[b]+[k]fluoranthenes.

These compounds are strongly adsorbed onto suspended particles or soot, and are more refractory to degradation. Moreover, hydrophobicity, increasing with the number of aromatic rings, controls the PAH scavenging from the water column towards the sediments.

Tolosa et al (1996) found that the PAH concentrations on the Mediterranean continental slope are below 1 μg g–1. Open sea sediments exhibit 0.1–0.6 μg g–1 of total PAHs. An increasing trend of pyrolytic PAHs in transects from the coastal areas towards the open sea indicates the predominance of atmospheric inputs to the latter, which account for 80–90% of the total PAHs in the deep basins.

Polybromodiphenyls (PBDEs)

Polybromodiphenyls (PBDEs) are usually commercialized as congener mixtures of different bromine content. The penta-product contains a mixture of tetra- to hexa-BDEs including BDE 47, BDE 99, BDE 100, BDE 153 and BDE 154, as well as trace amounts of BDE 17 and BDE 28. The octa-product consist primarily of BDE 183, followed by BDE 153 and BDE 154, whereas the deca-product is mostly composed of BDE 209 (>97%) (Alaee et al., 2003; North, 2004). The production and use of penta- and octa-BDE mixtures was banned in Europe in 2004, only deca-BDE was still permitted until 2008. However, there are stocks of all PBDEs in products in service and waste (Guerra et al., 2010). A significant proportion of these compounds reaches the marine environment which acts as PBDE reservoir (Binelli et al., 2007; Chen et al., 2006; Christensen and Platz, 2001; Grant et al., 2011).

All samples analyzed in the NW Mediterranean contained detectable PBDE concentrations in the superficial sediments (collected at 0-1 cm core depth) (Table 9.9), indicating that these pollutants are widespread in marine sediments. ∑PBDEs (sum of all PBDE congeners except BDE-209) varied from non- detectable to 2.6 ng/g with a mean value of 0.43 ng/g. BDE-190 is a congener absent from the technical mixtures (Alaee et al., 2003; La Guardia et al., 2006) but found in the NW Mediterranean samples. The concentrations in Cap de Creus and Lacaze-Duthiers canyons and the Open Slope were higher to those in the Danube Delta (Covaci et al., 2006), but similar or lower to those found in the coast of Korea (Moon et al., 2007a) and the Yangtze River delta, China (Chen et al., 2006).

BDE 209 is found in all sediment samples at concentrations ranging between 0.06 and 4.0 ng/g, two orders of magnitude higher than the sum of the concentrations of tri- to hepta-BDEs analyzed in the same samples (Table 9.9). The concentrations of this congener are similar to those found in the coast of Denmark (Christensen and Platz, 2001) or the coast of Hong Kong, China (Liu et al., 2005). BDE 209 exhibits a westward decreasing concentration gradient that is much stronger than for ∑PBDEs (Table 9.9).

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Deliverable 2.2 The concentrations of BDE 71 increase as those of BDE 209 decrease, which suggests that this congener originate, at least in part, from transformation of the latter (Salvado et al., 2012b) since this congener is absent from the technical mixtures. In general, the specific distributions of the less brominated PBDEs do not match the commercial mixtures of penta- and octa-BDE (Alaee et al., 2003; La Guardia et al., 2006). Their occurrence is related to decomposition of the fully brominated BDE 209. All the potential intermediates in the reductive debromination of BDE 209 are leading to BDE 47 (Salvado et al., 2012b).

Several studies have shown that diverse environmental processes may generate BDE congeners by debromination of BDE 209 in the environment (Bezares-Cruz et al., 2004; Eljarrat et al., 2004; Lacorte et al., 2003; Soderstrom et al., 2004). These processes may involve photooxidation (Bezares-Cruz et al., 2004) which has been reported to generate BDE 28, BDE 47, BDE 66, BDE 71, BDE 85, BDE 99, BDE 100, BDE 138, BDE 153, BDE 154 and BDE 183, or microbial decomposition (Bartrons et al., 2011; Robrock et al., 2008; Tokarz et al., 2008) which may involve the formation of BDE 17, BDE 28, BDE 47, BDE 66, BDE 71, BDE 99, BDE 138, BDE 153, BDE 154, BDE 183 and 190. These transformations often lead to BDE distributions highly enriched in BDE 28 and BDE 47 (Eljarrat et al., 2004; Robrock et al., 2008).

Table 9.9. Sediment concentrations (ng/g) of the individual PBDEs congeners found above limit of detection in deep marine environments of the NW Mediterranean (Salvado et al., 2012b). Sites 17 28 71 47 66 100 99 85 154 153 138 183 190 209 ∑PBDEsa Cap de Creus canyon C1 0.06 0.03 0.03 0.16 0.02 0.05 0.04 0.01 0.02 0.01 0.00 0.02 nd 1.3 0.44 C2 Nd 0.03 0.01 0.17 0.02 0.04 0.06 0.00 0.01 0.00 0.01 0.03 0.01 0.84 0.39 C3 0.07 nd 0.01 0.14 0.02 0.04 0.05 0.01 0.01 0.01 nq 0.01 0.01 1.4 0.39 C4 0.05 nq 0.05 0.10 0.01 0.04 0.08 nq 0.01 0.01 0.01 0.01 nd 0.24 0.38 C5 Nq nd nd nq Nd 0.01 Nd nq 0.00 nd nd nd nq 0.22 0.10 C6 0.05 0.02 nd 0.07 0.01 0.07 Nq nd 0.01 0.01 nd nd 0.01 0.80 0.28 C7 Nq nq nd 0.08 0.01 0.03 0.03 0.01 0.02 0.01 nq nq 0.01 2.0 0.27 C8 0.08 0.02 0.01 0.04 0.00 nd Nd nq 0.00 nd nd nd nq 0.09 0.19 C9 0.06 0.02 0.01 0.10 0.01 0.06 Nq 0.01 0.01 0.01 0.01 0.01 0.00 1.1 0.34 C10 0.14 nq nq 0.10 0.01 0.07 0.03 nq 0.02 0.01 0.00 0.01 0.01 1.0 0.41 C11 Nd 0.03 0.01 0.07 0.01 0.04 Nq 0.00 0.01 0.02 0.01 0.01 nd 1.1 0.26 C12 Nd 0.03 0.01 0.08 0.02 0.04 Nq 0.01 0.03 0.09 0.01 0.43 0.05 4.0 0.81 C13 Nd 0.02 0.21 0.06 0.01 0.02 Nd 0.01 0.01 0.01 0.01 nd 0.01 0.07 0.38 C14 Nd nq 0.14 0.08 0.01 0.02 Nd nd 0.01 0.01 0.00 0.01 0.01 1.5 0.34 C15 Nq 0.04 0.02 0.04 0.00 nd Nd 0.00 0.00 0.01 0.00 nd 0.00 1.1 0.18 Lacaze-Duthiers canyon L1 Nq 0.04 nq 0.16 0.02 0.06 0.04 0.00 0.02 0.01 nd 0.03 0.01 1.4 0.42 L2 Nq nq nq 0.08 0.01 0.02 0.04 0.00 0.01 0.00 nq 0.01 0.00 0.39 0.24 L3 Nd 0.05 0.01 0.12 0.02 0.09 0.06 0.00 0.02 0.01 nd 0.02 0.03 1.8 0.44 L4 Nd nd nd nq nq nd Nd nd nq nd nd nd nd 0.19 Nd L5 Nd 0.02 nd 0.06 0.01 0.04 Nq 0.00 0.01 0.00 nd nq nq 0.93 0.20 L6 Nd nd nd nd nd nd Nd nd nd nd nd nd nd 0.06 Nd L7 0.05 0.03 nd 0.09 0.02 0.06 0.03 nq 0.02 nd nd 0.01 0.01 2.3 0.32 L8 Nq nd nd nd nd nd Nd nd 0.00 nd nd nd nq 0.17 0.09 L9 Nd nd nd nq nq 0.01 Nd nd 0.00 nd nd nd nd 0.46 0.08 L10 Nd nq nd 0.04 0.01 0.05 Nd nd 0.01 nq nd nq 0.00 1.1 0.17 L11 Nd nd nd nq nq nd Nd nq 0.00 nd nd nd nd 0.29 0.09 L12 Nd nd nd nq nq nd Nd 0.00 0.00 nd nd nd nd 0.30 0.08 L13 Nd nd nd 0.03 nq nq 0.03 0.00 0.00 nd nd nd nd 0.17 0.12 L14 Nd nq nd nq 0.01 0.02 Nd nq 0.01 0.00 nd nq 0.01 0.07 0.13 L15 Nd 0.03 0.01 0.06 0.01 0.02 Nd 0.01 0.00 0.01 0.01 0.01 0.03 2.0 0.20

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Deliverable 2.2 L16 Nd nq nd nd nq nq Nd nd 0.01 nd 0.00 nd nq 0.88 0.08 Open Slope O1 Nd 0.02 nd 0.07 0.01 0.10 Nd 0.00 0.01 0.01 0.00 nd 0.02 0.69 0.27 O2 Nd nq nq 0.06 0.02 0.02 Nd 0.01 0.01 0.01 0.01 nd nd 2.0 0.18 O3 Nd 0.03 0.01 0.07 0.02 0.02 Nd 0.01 0.02 0.05 0.01 nq 0.00 0.15 0.26 S1 Nd nd 0.01 nd 0.01 0.01 Nd 0.00 0.01 0.00 nd nq nd 0.60 0.08 a∑PBDEs refers to the sum of all PBDE congeners except the BDE 209. nd = not detected. nq = not quantified

Radionuclides

Radionuclide concentrations and inventories in sediments are highly variable, being usually the highest on the continental shelf and near the river mouths and the lowest in the deep sea. In fact, the continental margins are areas where fine sediment particles transported by rivers, with associated radionuclides, are deposited.

137 239,240 210 Few publications report inventories of Cs, Pu and Pbex in the Mediterranean deep-sea sediments from the Algero-Balearic Basin, the Ionian Sea, and the Levantine Basin at water depths between 2800 and 4000 m. With the sediment accumulation rate being in the order of a few centimetres per 1000 years, anthropogenic radionuclides reach these sediments to a maximum depth of 2 - 4 cm, and their concentrations generally decrease with depth and are mainly controlled by mixing processes. 137Cs inventories (decay corrected to 2018) range from 0.05 – 0.12 Bq m−2 in the Algero-Balearic Basin to 0.07- 0.1 Bq m−2 in the Ionian Sea, and to 0.07-0.09 Bq m−2 in the Levantine Basin (Figure 9.12). In general, these inventories are only a small fraction, less than 2%, of the water column inventory.

For 239,240Pu, the deep-sea sediment inventories are very similar all over the Mediterranean Sea, with mean values of around 3 Bq m-2 both in western basin and eastern basin (Figure 9.13). In the pelagic areas particle concentration is usually very small, limiting the scavenging of particle-reactive radionuclides. Moreover, a significant part of the biogenic particles present in the surface photic zone, when sinking, is decomposed at intermediate depth, releasing the associated elements to the soluble phase. That is why the 239,240Pu inventories in the open sea are so small.

A particular attention should be paid in the deep sea where geomorphological features like canyon systems connect the continental shelf to the deep sea. Fine sediment particles are preferentially transported down the canyon axes. Hence, in these areas, relatively high sediment accumulation rates are found, associated with radionuclide inventories comparable to those on the shelf and high contaminants concentrations (Tamburrino et al. , 2018).

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Deliverable 2.2

137 -2 Figure 9.12. Inventories of Cs (kBq m ) in sediments (reference time: 2018). Data from Gascò et al. (2002), Noureddine et al. (2008), Garcia-Orellana et al. (2009), Barsanti et al. (2011)

239,240 -2 Figure 9.13: Inventories of Pu (Bq m ) in sediments, Data from Livingston (1979), Delfanti et al. (1995), Delfanti et al. (1996), Gascò et al. (2002.), Garcia-Orellana et al. (2009).

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Deliverable 2.2 9.3 Contaminants in the biota Studies of pollutants in marine biota may involve different species. Within the limited number of available studies on deep-water environments fish have generally been the organisms of choice. Muscle and liver have been investigated at higher extent. Both types of fish samples provide complementary information. The results from one and the other tissue cannot be evaluated under the same criteria.

Studies of organic pollutants in fish muscle have been developed on the most abundant megafaunal species. In 2008-2009, sample collection in the NW Mediterranean encompassed three ﬁsh species belonging to different phylogenetic families, namely Alepocephalus rostratus (Alepocephalidae), Coelorinchus mediterraneus (Macrouridae) and Lepidion lepidion (Moridae), and the red-shrimp Aristeus antennatus (Koenig et al., 2013a). Previous studies on organic pollutants in muscle of deep-sea fish were developed in 1996 and involved Mora moro (Moridae; Sole et al., 2001).

Liver samples studied in deep-sea fish encompassed Chimaera monstrosa, Raja spp (Storelli et al., 2004a), Lophius budegassa (Storelli et al., 2004b), Trachyrincus trachyrincus (Storelli et al., 2009), Coelorhinchus caelorhincus (Storelli et al., 2007; 2009), Nezumia sclerorhynchus (Storelli et al.,2007) and Mora moro (Sole et al., 2001)

Comparisons of the pollutant concentrations in several tissues of the same specimens have also been performed (Sole et al., 2001). Following is a summary of the results of the most extensive studies on deep- water organisms from the Mediterranean Sea.

Organochlorine compounds in muscle

In the specimens collected in 2008-2009, PCB and DDT levels ranged from the highest concentrations in the fish A. rostratus (Σ7PCBs 6.9±0.71 ng/g w.w. and ΣDDTs 8.4±1.10 ng/g w.w.) to the lowest in the crustacean A. antennatus (Σ7PCBs 1.2±0.24 ng/g w.w. and ΣDDTs 2.5±0.26 ng/g w.w.; Table 9.10; Koenig et al., 2013a). The concentrations of ΣHCHs and HCB were more than one order of magnitude lower, ranging from 0.07–0.36 ng/g w.w. and 0.03–0.15 ng/g w.w., respectively, while PeCB was only detected in a few samples above the detection limit. The OC relative abundance followed the sequence PCBs ≈ DDTs ≫ HCHs ≥ HCB and was generally in accordance with previous studies on deep-sea fish conducted in the Mediterranean (Porte et al., 2000; Storelli et al., 2009), the Atlantic (Berg et al., 1997; Mormede and Davies, 2003) and the Pacific Ocean (Ramu et al., 2006; Takahashi et al., 2010) and also with data on deep-sea sediments from the Gulf of Lion (NW Mediterranean) (Salvadó et al., 2012a). These results indicate that the bioaccumulation of HCHs and HCB in deep-sea biota is negligible compared to PCBs and DDTs, which is in agreement with the higher hydrophobicity and bioaccumulation potential of most PCBs and DDTs relative to HCHs and HCB. This effect, together with the enhanced vertical transport of these compounds due to their preferential association to suspended particulate matter (Dachs et al., 2002; Scheringer et al., 2009), explains the dominance of PCBs and DDTs in deep-sea species.

The levels of OC contamination measured in the specimens collected in 2008-2009 (Σ7PCBs 1.2–6.9 ng/g w.w., ΣDDTs, 2.53–8.43 ng/g w.w.; Koenig et al., 2013a) are within the range of values previously measured in deep-sea fish from the same study area (i.e. NW Mediterranean) by Porte et al. (2000)

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Deliverable 2.2 (Σ7PCBs 2.5–10 ng/g w.w.; ΣDDTs 1.9–10 ng/g w.w.) and Solé et al. (2001) (Σ7PCBs 9.0–16 ng/g w.w.; ΣDDTs 7.4–13 ng/g w.w.) in 1996. Accordingly, OC levels in NW Mediterranean deep-sea fish have remained relatively similar over the past decade.

Table 9.10. Organochlorine levels in deep-sea fish and crustacean from the NW Mediterranean collected in 2008-2009 (Koenig et al., 2013a). Values: mean concentrations (ng/g w.w.) ± standard error of mean (min.–max.). n.d.: not detected. A. rostratus C. mediterraneus L. lepidion A. antennatus (n= 30) (n= 25) (n= 20) (n= 3 pools) PeCB 0.02 ± 0.005 0.01 ± 0.003 n.d. 0.01 ± 0.003 (n.d.–0.08) (n.d.–0.04) (0.01–0.02) HCB 0.15 ± 0.02 0.05 ± 0.03 0.05 ± 0.01 0.03 ± 0.01 (n.d.–0.46) (n.d.–0.66) (n.d.–0.09) (0.01–0.04) α-HCH 0.02 ± 0.01 0.04 ± 0.01 0.03 ± 0.01 0.01 ± 0.01 (n.d.–0.11) (n.d.–0.15) (n.d.–0.09) (n.d.–0.03) β-HCH 0.07 ± 0.02 0.11 ± 0.04 n.d. 0.03 ± 0.006 (n.d.–0.51) (n.d.–0.68) (0.02–0.04) γ-HCH 0.17 ± 0.02 0.019 ± 0.06 0.03 ± 0.01 0.03 ± 0.01 (n.d.–0.61) (n.d.–0.99) (n.d.–0.09) (n.d.–0.04) δ-HCH 0.04 ± 0.02 0.02 ± 0.01 0.004 ± 0.004 n.d. (n.d.–0.51) (n.d.–0.14) (n.d.–0.08) ΣHCHs 0.30±0.05 0.36±0.10 0.07±0.02 0.07±0.03 (n.d.-1.2) (n.d.-1.8) (n.d.-0.20) (n.d.-0.10) PCB 28 0.12 ± 0.02 0.03 ± 0.01 n.d. 0.06 ± 0.04 (n.d.–0.35) (n.d.–0.19) (0.02–0.13) PCB 52 0.29 ± 0.05 0.56 ± 0.12 0.65 ± 0.26 0.17 ± 0.02 (n.d.–1.04) (0.04–2.43) (0.03–4.68) (0.14–0.21) PCB 101 0.59 ± 0.05 0.18 ± 0.02 0.23 ± 0.04 0.09 ± 0.02 (0.18–1.23) (n.d.–0.35) (0.10–0.88) (0.06–0.12) PCB 118 0.31 ± 0.05 0.34 ± 0.10 0.37 ± 0.03 0.10 ± 0.04 (n.d.–0.85) (0.10–2.49) (0.18–0.68) (0.06–0.18) PCB 153 2.36 ± 0.28 1.44 ± 0.31 2.53 ± 0.33 0.21 ± 0.04 (0.42–5.69) (0.35–6.44) (0.83–5-65) (0.12–0.26) PCB 138 1.83 ± 0.21 1.17 ± 0.25 1.41 ± 0.18 0.44 ± 0.13 (0.35–4.34) (0.31–5.80) (0.52–3.56) (0.18–0.60) PCB 180 1.44 ± 0.17 0.77 ± 0.16 1.03 ± 0.17 0.12 ± 0.04 (0.26–3.35) (0.17–2.65) (0.12–3.21) (0.05–0.19) ΣPCBs 6.9±0.71 4.5±0.79 6.2±0.64 1.2±0.24 (1.7-15) (1.2-18) (2.0-12) (0.70-1.5) p,p′-DDT 1.83 ± 0.21 0.21 ± 0.03 0.18 ± 0.01 0.39 ± 0.24 (0.35–4.34) (0.09–0.56) (0.10–0.31) (0.08–0.87) p,p′-DDE 6.44 ± 0.91 2.18 ± 0.45 3.38 ± 0.44 0.50 ± 0.24 (0.82.–16.52) (0.65–8.28) (0.37–6.87) (0.24–0.98) p,p′-DDD 0.50 ± 0.07 0.10 ± 0.01 0.08 ± 0.01 n.d. (0.05.–1.48) (0.06–0.21) (n.d.–0.12) o,p′-DDT 0.32 ± 0.03 0.07 ± 0.01 0.01 ± 0.01 0.22 ± 0.06 (0.06–0.70) (n.d.–0.15) (n.d.–0.07) (0.13–0.34) o,p′-DDE 0.08 ± 0.01 0.01 ± 0.003 n.d. 1.33 ± 0.14 (n.d.–0.34) (n.d.–0.05) (1.06–1.47) o,p′-DDD 0.94 ± 0.04 0.27 ± 0.01 0.23 ± 0.02 0.08 ± 0.01 (n.d.–0.94) (0.16–0.43) (0.10–0.46) (0.06–0.09) ΣDDTs 8.4±1.1 2.8±0.4 3.9±0.47 2.5±0.26 (1.3-21) (1.1-9.2) (0.70-7.7) (2.1-3.0)

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Deliverable 2.2 The PCB profiles in deep-sea species were dominated by the high-molecular-weight (HMW) PCBs 153, 138 and 180, which represented 69–80% of Σ7PCBs in fish (Koenig et al., 2013a; Sole et al., 2001) and 60% in the red shrimp A. antennatus. While PCB 153 exhibited the highest abundance in fish, in the shrimp the most abundant congener was PCB 138. The differential PCB accumulation between the fish and the crustacean species has been described in a previous study and is likely related to differences in hepatic cytochrome P450-mediated metabolism of PCBs between fish and crustacea (Koenig et al., 2013a). Overall, the detected PCB profiles are in accordance with the general bioaccumulation patterns of PCBs in deep-sea fish reported in former studies (Porte et al., 2000; Solé et al., 2001; Mormede and Davies, 2003). The predominance of these compounds in biota can be explained by the higher bioaccumulative potential of the more hydrophobic higher chlorinated PCBs (i.e. hexa- to octachloro congeners) (McFarland and Clarke, 1989). As mentioned above, highly chlorinated congeners have higher sediment affinities than low chlorinated compounds and are thus more prone to particle bound transport from surface waters to the deep sea (Dachs et al., 2002; Scheringer et al., 2004).

In fish, the main DDT compound detected is the metabolite p,p′-DDE, which comprised on average 70– 80% of ΣDDTs, while the parent compound p,p′-DDT contribute only 6–10% to ΣDDTs (Table 9.10). This result is a commonly observed distribution in marine organisms (Voorspoels et al., 2004), including deep- sea fish (Mormede and Davies, 2003; Takahashi et al., 2010) and is indicative of old DDT residues, which progressively degrade in aquatic environments into their even more persistent metabolites, primarily DDE (Wolfe et al., 1977). In shrimp however, o,p′-DDE is the main DDT metabolite and represented 49% of ΣDDTs, while the parent compound p,p′-DDT and its metabolite p,p′-DDE exhibit similar proportions of 16% and 21%, respectively (Table 9.10). Hence, the DDT/DDE ratio profile detected in shrimp would indicate a recent input of the parent compound p,p′-DDT within the study area, which is in contrast to the results observed in fish. Thus, it seems that, despite the wide use of the DDT/DDE ratio as a means to discriminate between recent and past use of DDT (Corsolini et al., 2008), these results indicate that it should be applied with caution as it can vary among different organisms.

Polychlorodibenzo-p-dioxins (PCDDs) and polychlorodibenzofurans (PCDFs) accumulation was investigated in A. antennatus collected from depths of 600-2500 m at different points in the Mediterranean Sea, from the western basin off the coast of Barcelona to the central basin off the Peloponnesian Peninsula, with otter trawl gear (Rotllant et al., 2006). Total PCDD/Fs burdens were higher in shrimps caught in the western Mediterranean than in those caught at eastern Mediterranean sites. There was a tendency for higher levels of PCDD/F contamination in samples obtained from deeper (2500 m) than from shallower sites (600 m).

The technical HCH mixtures, containing all four isomers with a α-HCH/γHCH ratio of 4–7, were gradually replaced by lindane, which is the only component exhibiting significant insecticide activity and still being released, although to a limited extent, into the environment. Accordingly, the most abundant HCH isomer detected in the present study is γ-HCH (lindane), contributing approximately 50% to ΣHCH in all species. Moreover, α-HCH/γ-HCH ratios range between 0.2 in A. rostratus and 1.0 in L. lepidion, showing a predominance of lindane over the technical mixture.

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Deliverable 2.2 Table 9.11. Concentrations of organochlorine compounds (ng g-1 ww) in tissues of fish collected in the deep NW Mediterranean Fish species Habitat DDTs HCB PCBs Ref (DDT + DDE) (ICES) Lepidorombus sp. Mesopelagic 0.8±0.2 2.1±0.3 Garcia et al., 2000 Phycis sp. 0.4±0.1 1.0±0.2 Lepidion sp. Deep water 6.0-7.1 0.14±0.17 8.3±9.4 Porte et al., 2000 Coryphaenoides 1.9-4.3 0.25-0.67 2.5-4.6 sp. Bathypterois sp. 5.0-10.2 0.12-0.25 6.0-10 Mora moro 7.4-13 9.0-16 Sole et al., 2000

DDTs were measured in a variety of pelagic and migratory fish. In the Western Mediterranean, mesopelagic and deep-sea species were monitored, the latter showing concentrations similar to those of the coastal species, as shown in Table 9.11.

Sharks (Prionace glauca and Alopias vulpinus) contain 70–4400 ng g–1 ww (mean 980 ng g–1) in fat. The bluefin tuna (Thunnus thynnus) show 170–2200 ng g–1 ww (mean 850 ng g–1) in muscle and 220–1700 ng g–1 ww (mean 930 ng g–1) in liver, in samples collected in 1993, and mean concentrations of 280 ng g–1 in liver and 820 ng g–1 ww in fat, in specimens collected in 1999. Mean PCB values in pooled samples of liver and muscle of swordfish (Xiphias gladius) are 750 and 330 ng g–1 ww, respectively (Kannan et al., 2001). Differences in accumulation encountered in the different species are principally due to different feeding habitats, sex and maturity.

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Table 9.12. Mean concentrations of PCBs, DDTs, HCB and sum of seven congeners (PCBs 28, 52, 101, 118, 138, 153 and 180; ng g-1 lipid weight) in liver of deep-sea ﬁsh from different world regions (from Storelli et al. (2009) and Sole et al (2001)). Species Location PCBs DDTs HCB ΣPCB7 References

Hollowsnout grenadier Mediterranean Sea (Adriatic Sea) 1234 763 6.3 734 Storelli et al. (2009) Roughsnout grenadier Mediterranean Sea (Adriatic Sea) 12,327 5357 1 7814 Storelli et al. (2009) Hollowsnout grenadier Mediterranean Sea (Adriatic Sea) 427 344 3.– 415 Storelli et al. (2007) Roughtip grenadier Mediterranean Sea (Adriatic Sea) 313 239 1 – 300 Storelli et al. (2007) Common mora NW Mediterranean 740-1600 740-1600 Sole et al. (2001) Roughhead grenadier Atlantic Ocean (Davis Strait) 1058 390 59 602 Berg et al. (1997) Round-nose grenadier Atlantic Ocean (North Sea) 2400 5230 15 – Berg et al. (1998) Giant grenadier Pacific Ocean (Tohoku Region) 2200 460 15 – de Brito et al. (2002a) Round-nose grenadier Atlantic Ocean (West of Ireland) 664–876 1010–1500 26–31 460–601 Mormede and Davies (2003) Deep-sea fishes Pacific Ocean (Suruga Bay) 1000 280 19 – Lee et al. (1997) Deep-sea fishes Pacific Ocean (Tosa Bay) 340 290 11 – Takahashi et al. (2001) Deep-sea fishes Pacific Ocean (East China Sea) 230 700 18 – Tanabe et al. (2005) Deep-sea fishes Pacific Ocean (Sulu Sea) 58 150 6.8 – Ramu et al. (2006) Greenland halibut Atlantic Ocean (Davis Strait) 559 473 84 268 Berg et al. (1997) Redfish Atlantic Ocean (Davis Strait) 947 652 29 509 Berg et al. (1997) Black dogfish Atlantic Ocean (Davis Strait) 545 955 37 294 Berg et al. (1997) Blue hake Atlantic Ocean (Davis Strait) 1156 1446 68 672 Berg et al. (1997) Tusk Atlantic Ocean (Davis Strait) 939 992 73 528 Berg et al. (1997) Blue antimora W Greenland 610 610 Berg et al. (1997) Roughhead grenadier W Greenland 110 450 Berg et al. (1997) Rose fish W Greenland 89 160 Berg et al. (1997) Velvet belly Atlantic Ocean (North Sea) 2390 6030 64 – Berg et al. (1998) Ling Atlantic Ocean (North Sea) 2960 4640 31 – Berg et al. (1998) Tusk Atlantic Ocean (North Sea) 11,700 27,000 22 – Berg et al. (1998) Monkfish Atlantic Ocean (Rockall Trough) 275–492 206–376 17–18 – Mormede and Davies (2001) Black scabbard Atlantic Ocean (Rockall Trough) 140–364 385–410 23–24 – Mormede and Davies (2001) Walleye pollock Pacific Ocean (Bering Sea) 780 340 74 – de Brito et al. (2002b, c) Walleye pollock Pacific Ocean (Gulf of Alaska) 1000 420 30 – de Brito et al. (2002b, c) Walleye pollock Pacific Ocean (Japan Sea) 1800 1400 87 – de Brito et al. (2002b, c) Ratfish Pacific Ocean (Tohoku Region) 130 91 13 – de Brito et al. (2002a) Ghostsharks Mediterranean Sea (Adriatic Sea) 387 – – 274 Storelli et al. (2004a) Skates Mediterranean Sea (Adriatic Sea) 889 – – 662 Storelli et al. (2004a) Angler fish Mediterranean Sea (Adriatic Sea) 4372 4423 – 3334 Storelli et al. (2004b) Great northern tilefisha NW Atlantic 210-1100 1000-4700 Steimle et al. (1990) Black scabbardfish E Atlantic 8100 Kramer et al (1984) Black scabbardfish E Atlantic 100-600 70-670 Kramer et al (1984) Abyssal grenadier NW Atlantic 760 360-2700 Stegeman et al. (1986) ain dry weight

Organochlorine compounds in liver

The observed total contaminant concentrations by organochlorine compounds were higher in roughsnout grenadier (PCBs: 12,327 ng g-1; DDTs: 5357 ng g-1; HCB 13.1 ng g-1) than in hollowsnout grenadier (PCBs: 1234 ng g-1; DDTs: 763 ng g-1; HCB: 6.3 ng g-1; Table 9.12) or Common mora (PCBs: 740-1600 ng g-1; DDTs: 740-1600 ng g-1; Table 9.12). This difference could be attributed to the higher trophic level of roughsnout grenadier (Stergiou and Karpouzi, 2002). However, it should be indicated that the feeding habits change depending on age classes in both species. Smaller roughsnout grenadiers eat mainly decapods, while the diet of older animals consists of decapods and ﬁsh. In a similar way, young hollowsnout grenadiers feed on polichaetes, amphipods and copepods, those of medium size prefer polichaetes, decapods and ﬁsh, while larger specimens prey mainly on decapods (Stergiou and Karpouzi, 2002).

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Deliverable 2.2 Another important factor determining the different contaminant burdens between species is habitat depth. In general, higher concentrations of organochlorine compounds have been found in ﬁsh from deeper waters (Mormede and Davies, 2003; Serrano et al., 2000; de Brito et al., 2002a). This agrees well with the present results, roughsnout grenadier being an organism inhabiting deeper marine areas with respect to hollowsnout grenadier.

The organochlorine levels observed in hollowsnout grenadier are generally similar to those found in deep- sea species living in Mediterranean and non-Mediterranean regions. In contrast, roughsnout grenadier shows levels higher than most of the previously published data from deep-sea ﬁsh (Table 9.12).

The total DDT concentrations are similar to those of PCBs, with hollowsnout grenadier having levels of the same order of magnitude as those in literature and roughsnout grenadier having much higher concentrations.

Polybromodiphenyl ethers

The concentrations of PBDEs found in the most abundant megafaunal species in the NW Mediterranean deep sea collected in 2008-2009 (A. rostratus, C. mediterraneus and L. lepidion, and the red-shrimp A. antennatus) are shown in Table 9.13 (Koenig et al., 2013a). They ranged from 0.47±0.20 ng/g w.w. in A. antennatus to 0.92±0.13 ng/g w.w. in A. rostratus and were approximately one order of magnitude lower than PCB and DDT concentrations, which is in agreement with former studies that simultaneously assessed OC and PBDE contamination in deep-sea fish from the Atlantic (Webster et al., 2009, 2011) and the Pacific Ocean (Ramu et al., 2006; Takahashi et al., 2010). Furthermore, this result is also consistent with above reported sediment contamination data from the NW Mediterranean basin, where PCB and DDT contamination clearly exceeded PBDE levels (Salvadó et al., 2012a, 2012b). To our knowledge, only one study has previously measured PBDEs in Mediterranean deep-sea fish (Covaci et al., 2008), however, the reported levels were determined in fish liver and are thus not directly comparable to the results in muscle tissue presented in the study of Koenig et al. (2013a). In comparison to Mediterranean shallow- water species, similar PBDE levels have been detected in the European eel (Anguilla anguilla), with a range of Σ28PBDEs 0.08–1.80 ng/g w.w. (including all 14 congeners analyzed in this work) (Labadie et al., 2010). Similarly, Corsolini et al. (2008) determined the sum of 19 PBDEs in swordfish (Xiphias gladius), and, although it is noteworthy that the more brominated BDEs (i.e. hepta- to decaBDE) were not included, reported values were similar to those found in the study of Koenig et al. (2013a), in the range of 0.04– 1.91 ng/g w.w. Finally, significantly higher concentrations of PBDEs (Σ23PBDEs 15.1 ng/g w.w.) have been observed in tuna (Thunnus thynnus) from the Mediterranean Sea (Borghesi et al., 2009). Σ14PBDE levels expressed on lipid weight basis varied between 62±29 ng/g l.w. in A. antennatus and 190±27 ng/g l.w. in L. lepidion. Webster et al. (2009) reported slightly lower levels in deep-sea fish from North Atlantic Scottish waters, ranging from 12 to 50 ng/g l.w. for Σ17PBDEs, which included all 14 congeners considered in our study except BDE 209. In contrast, the sums of 14 BDE congeners in Pacific deep-sea fish caught in the Sulu Sea (0.9–1.6 ng/g l.w.) (Ramu et al., 2006) and off Tohoku, Japan (1.3–8.5 ng/g l.w.) (Takahashi et al., 2010) were one to two orders of magnitude lower than the Mediterranean results. However, some fish species included in the study by Takahashi et al. (2010) exhibited very high

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Deliverable 2.2 lipid contents in muscle tissue (1.2– 25%). Transforming reported values to wet weight concentrations results in levels in the range of 0.1–0.5 ng/g w.w., which are more similar to the concentrations reported in Table 9.13.

Table 9.13. PBDE levels in deep-sea ﬁsh and crustacean from NW Mediterranean collected in 2008-2009 (Koenig et al., 2013a). Values: mean concentrations (ng/g w.w.) ± standard error of mean (min.–max.). n.d.=not detected. *ng/lipid weight. A. rostratus C. mediterraneus L. lepidion A. antennatus (n= 30) (n= 25) (n= 20) (n= 3 pools) BDE 17 0.06 ± 0.02 0.05 ± 0.02 0.001 ± 0.001 0.003 ± 0.003 (n.d.–0.29) (n.d.–0.35) (n.d.–0.02) (n.d.–0.01) BDE 28 0.10 ± 0.02 0.12 ± 0.01 0.13 ± 0.01 0.06 ± 0.02 (n.d.–0.45) (0.03–0.23) (0.07–0.21) (0.03–0.08) BDE 71 0.01 ± 0.003 0.002 ± 0.001 0.01 ± 0.005 n.d. (n.d.–0.05) (n.d.–0.01) (n.d.–0.09) BDE 47 0.15 ± 0.03 0.14 ± 0.02 0.15 ± 0.03 0.06 ± 0.01 (n.d.–0.58) (n.d.–0.42) (n.d.–0.49) (0.04–0.08) BDE 66 0.004 ± 0.004 n.d. n.d. n.d. (n.d.–0.11) BDE 100 0.14 ± 0.03 0.08 ± 0.01 0.10 ± 0.01 0.02 ± 0.003 (0.03–0.77) (0.03–0.32) (0.03–0.23) (0.01–0.02) BDE 99 0.16 ± 0.05 0.13 ± 0.03 0.13 ± 0.04 0.07 ± 0.03 (n.d.–1.56) (0.05–0.67) (n.d.–0.66) (n.d.–0.11) BDE 85 0.04 ± 0.01 0.03 ± 0.01 0.01 ± 0.002 n.d. (n.d.–0.17) (n.d.–0.15) (n.d.–0.04) BDE 154 0.11 ± 0.03 0.03 ± 0.01 0.02 ± 0.004 0.003 ± 0.003 (0.02–0.79) (n.d.–0.17) (n.d.–0.08) (n.d.–0.01) BDE 153 0.04 ± 0.01 0.01 ± 0.004 0.001 ± 0.004 0.09 ± 0.03 (0.01–0.24) (n.d.–0.07) (n.d.–0.07) (0.03–0.14) BDE 138 0.004 ± 0.001 n.d. n.d. n.d. (n.d.–0.03) BDE 183 0.008 ± 0.002 0.002 ± 0.001 n.d. n.d. (n.d.–0.05) (n.d.–0.02) BDE 190 0.01 ± 0.002 n.d. n.d. n.d. (n.d.–0.03) BDE 209 0.11 ± 0.05 0.02 ± 0.01 0.02 ± 0.01 0.17 ± 0.17 (n.d.–1.66) (n.d.–0.09) (n.d.–0.22) (n.d. ± 0.52) ΣBDEs 0.92 ± 0.13 0.61 ± 0.07 0.58 ± 0.08 0.47 ± 0.20 (0.29–3.02) (0.23–1.97) (0.20–1.63) (0.18–0.84) ΣBDEs 107.9 ± 15.8 172.1 ± 23.4 188.8 ± 26.6 61.9 ± 28.9 l.w.* (16.6–349) (14.5–501) (21.4–448) (23.1–1195)

Σ14PBDE levels expressed on lipid weight basis varied between 62±29 ng/g l.w. in A. antennatus and 190±27 ng/g l.w. in L. lepidion. Webster et al. (2009) reported slightly lower levels in deep-sea fish from North Atlantic Scottish waters, ranging from 12 to 50 ng/g l.w. for Σ17PBDEs, which included all 14 congeners considered in our study except BDE 209. In contrast, the sums of 14 BDE congeners in Pacific deep-sea fish caught in the Sulu Sea (0.9–1.6 ng/g l.w.) (Ramu et al., 2006) and off Tohoku, Japan (1.3–8.5 ng/g l.w.) (Takahashi et al., 2010) were one to two orders of magnitude lower than the Mediterranean results. However, some fish species included in the study by Takahashi et al. (2010) exhibited very high lipid contents in muscle tissue (1.2– 25%). Transforming reported values to wet weight concentrations

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Deliverable 2.2 results in levels in the range of 0.1–0.5 ng/g w.w., which are more similar to the concentrations reported in Table 9.13.

In fish, the most important PBDE congeners detected were BDE 28, 47, 99 and 100, constituting from 68% in A. rostratus to 89% in L. lepidion of ΣPBDEs (Table 9.13), similar to previous results observed in muscle tissue of deep-sea fish (Webster et al., 2009; Takahashi et al., 2010). These congeners are the main components in the commercial penta-BDE formulations (La Guardia et al., 2006). BDE 154 and 209 levels were also significant in all fish species. BDE 154 has been suggested to be a debromination product of BDE 183, the main congener in the technical octa-BDE mixtures (Stapleton et al., 2004; Roberts et al., 2011), while BDE 209 constitutes between 92 and 97% of the total BDE content in the deca-BDE formulations (La Guardia et al., 2006). Therefore, the PBDE composition observed in deep-sea organisms is consistent with the composition of the technical mixtures used in Europe. This PBDE profile differs from that reported in deep-sea sediments, where BDE 209 is the predominant congener (78%; Salvadó et al., 2012b). These differences can be attributed to differences in bioavailability and biotransformation potential between compounds. BDE 209 is thought to have low bioavailability, in agreement with its high molecular size (Eljarrat et al., 2004), and it can be metabolized to less brominated congeners in some fish species (Kierkegaard et al., 1999; Stapleton et al., 2006), thus potentially explaining the relatively low proportion of BDE 209 found in the deep-sea fish.

In contrast to fish, BDEs 153 and 209 were the most abundant congeners in shrimp, although large variability was observed potentially because of the low sample sizes (n=3 pools) (Table 9.13). Previous studies have also observed high concentrations of BDE 209 (Ashizuka et al., 2008; van Leeuwen et al., 2009), as well as BDE 153 (Voorspoels et al., 2003) in shrimp, suggesting a higher uptake or lower biotransformation capacity of these congeners in crustacea. Despite the correspondence between PBDE congeners found in deep-sea organisms and main components in the technical mixtures, the relative abundance of these compounds differs from that found in the commercial formulations. These results can be explained by differences in metabolic transformation rates between congeners (Roberts et al., 2011).

In this sense, BDE 99/100, 153/154 and 47/99 ratios have been used to assess differences in metabolic capacities as well as trophic position among various aquatic organisms (Voorspoels et al., 2003; Xiang et al., 2007; Dickhut et al., 2012). Usually, high BDE 99/100 ratios, similar to those found in the original commercial pentaBDE mixtures such as Bromkal 70 5-DE (approx. 5.3) or DE-71 (approx. 3.7), are found in sediments and lower organisms such as invertebrates, but decrease through the food chain due to higher biotransformation rate of BDE 99 in higher organisms (Christensen and Platz, 2001; Voorspoels et al., 2003; Xiang et al., 2007; Hu et al., 2010). This ratio varied between 0.99 and 1.91 in deep-sea fish species (Table 9.5), indicating a significant degradation of BDE 99. The higher value found in shrimp (3.3) is in accordance with a number of studies reporting higher ratios in crustacea compared to fish (Voorspoels et al., 2003; Xiang et al., 2007; Hu et al., 2010). However, it should be noted that although L. lepidion has the highest trophic level, it does not present the lowest BDE 99/100 ratio, indicating that in the concentrations reported in Table 9.13, the influence of metabolic capacities on this ratio may be more important than trophic position.

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Deliverable 2.2 The ratio between BDE 153 and BDE 154 has been similarly related to the metabolic capacities of different organisms (Xiang et al., 2007), with higher contributions of BDE 154 reflecting the higher biotransformation of more brominated congeners such as BDE 183 (Roberts et al., 2011). The concentrations reported in Table 9.13 show ratios <1 in fish, but shrimp exhibited a very high BDE 153/154 ratio, pointing to a lower metabolic capacity of the crustacean in relation to fish, although it is noteworthy that this result is largely based on very high BDE 153 levels and a lack of BDE 154 in shrimp. However, a significant relationship between the ratios BDE 99/100 and BDE 153/154 in all three fish species (ρ>0.4, pb0.05) indicates the coherent covariation of these two parameters, reinforcing their use as proxies for the BDE metabolization abilities of different species. Furthermore, BDE 99/100 and BDE 153/154 ratios were highest in the shrimp, but also higher in C. mediterraneus compared to the two other fish species. These two species are infaunal feeders, closely associated with the sediment, while A. rostratus and L. lepidion feed on epibenthic and/or pelagic prey. Hence, it is possible that, in addition to differences in metabolic capacities, these BDE congener ratios also reflect differences in feeding strategies among organisms.

BDE 47/99 ratios in deep-sea fish varied between 1.17 and 1.36, with BDE 47 only representing 15–22% of ΣPBDEs. This is in contrast to results found in liver of two Mediterranean deep-sea fish species, where BDE 47 contributed approximately 50% to ΣPBDEs and BDE 99 was clearly depleted (Covaci et al., 2008). BDE 99 has been shown to be metabolized to BDE 47 in carp liver (Stapleton et al., 2004) and the congener ratio BDE 47/99 has been used to assess the level of metabolization of BDE 99 to BDE 47 (Wang et al., 2009). However, a different debromination pathway of BDE 99 has been detected in salmon and trout, suggesting significant differences in efficiency and metabolite formation of BDE 99 debromination among teleost species (Browne et al., 2009; Roberts et al., 2011) and the BDE 47/99 ratio does therefore not necessarily reflect the metabolization rate of BDE 99 in all species. In fact, the similar proportions of BDE 47 and 99 observed in the NW Mediterranean are consistent with BDE 47/99 ratios in the commercial penta-BDE mixtures, which primarily contain these two congeners at equal concentrations, suggesting a lack of debromination of BDE 99 to BDE 47. However, slightly lower BDE 47/99 ratios were observed in the shrimp (0.75) and C. mediterraneus (1.17), compared to the two other fish species (1.36), suggesting again the potential existence of different metabolic capacities and/or differences in BDE uptake related to feeding strategies between the two infaunal feeders and the more pelagic species.

These results indicate that metabolism plays an important role in the PBDE congener distributions in aquatic organisms, resulting in a selective accumulation of the lower brominated congeners. This is relevant to human as well as wildlife health, since lower brominated congeners have higher biomagnification potential and toxicological effects.

Polycyclic aromatic hydrocarbons

Analyses of PAHs in the deep-sea fish, Mora moro, from the NW Mediterranean basin show liver concentrations of 7–16 ng g–1 ww (Sole et al., 2001). This fish species is a gadiform very abundant at about 1000 m water depth. It is a large, demersal species with high-energy requirements that feeds on other

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Deliverable 2.2 fish, crustacea and cephalopods. It actively swims in the upper slope (1000-1200 m) of the western Mediterranean, where it is most frequently caught (Stefanescu et al., 1993).

The distribution of 14 parent PAHs, from 2 to 6 rings, found in the four tissues of this deep-sea fish are shown in Figure 9.14. Alkylated PAHs were also measured, but their presence was not significant, indicating a practical absence of petrogenic contributions, in agreement with the lack of UCM in the aliphatic fraction.

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Deliverable 2.2

Figure 9.14. Distribution of individual PAHs in several tissues of Mora moro. Naphthalene (N), Fluorene (F), phenanthrene (P), anthracene (A), Fluoranthene (FL), pyrene (PY), benz(a)anthracene (BA), chrysene (C), benzo(b)fluoranthene (BbF), benzo(k)fluoranthene (BkF), benzo(a)pyrene (BaP), indeno-(1,2,3-cd) pyrene (IP), dibenz(ah)anthracene (DA) and benzo(ghi)perylene (BP). Groups I, II and III refer to the size of the specimens, 22.0±1.1 cm (<100 g), (II) 34.0±1.4 cm (250-365 g) and (III) 42.0±2.0 cm (>500 g). From Sole et al. (2001).

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Deliverable 2.2 Consistently with the PAH distributions found in the deep-sea sediments of the NW Mediterranean (Tolosa et al., 1996; Lipiatou et al., 1997), all samples are dominated by the heavier PAHs (4- to 6-ring). This is remarkable in the younger specimens (group I) and in tissues reflecting pollutant entries, e.g. gills and digestive tube. The liver and muscle of fish groups II and III are relatively enriched in 3- to 5-ring PAHs, pyrene and benzofluoranthenes being the dominant components. Steimle et al. (1990) also reported relatively higher concentrations of these PAHs in NW Atlantic tilefish tissues. PAH distributions enriched in heavier components are more prone to be incorporated through the diet because they reach more easily the deep water column, as they are preferentially associated with fast sinking particles (Lipiatou et al., 1993; Dachs et al., 1997). Escartin and Porte (1999) have demonstrated that hepatic enzymes (cytochrome P450 and glutathione transferases) in deep-sea fish, e.g. M. moro among them, are able to metabolize PAHs as efficiently as shallow water species and excrete the metabolites through the bile.

Alternatively, fish exposure to PAHs can be assessed by measuring bile PAH metabolites. The method was successfully applied in the NW Mediterranean area, in both coastal species and deep-sea fish (e.g. Mora sp. and Lepidion lepidion) (Escartin and Porte, 1999). Qualitative and quantitative differences were recorded among sampling sites, paralleling the results of the PAH’s patterns. The lowest levels were found in remote areas, but with naphthol and pyrenol among the main metabolites in coastal and deep-sea fish, respectively, which may indicate the higher exposure of the latter to the heavier components that reach the sea bottom.

Tributhyltin

TBT and its degradation products, MBT and DBT, are found in concentrations of 47 ng g–1 ww in muscle of bluefin tuna from the Egadi Islands, close to Sicily (Italy) (Kannan et al., 1996). On the other hand, triphenyltin is the predominant organotin component in deep-sea fish (up to 1430 ng g–1 ww), showing transport capacity far from point sources associated to particulate matter and its higher resistance to degradation (Borghi and Porte, 2002)

Perfluoroacids

Perfluorooctanesulfonate (PFOS) is a widespread contaminant in the Mediterranean wildlife (Kannan et al., 2002). PFOS and other derivatives were detected in 175 samples of liver and blood of bluefin tuna (Thunnus thynnus), swordfish (Xiphias gladius), bottlenose, stripped and common dolphins and whales from the Mediterranean Italian coast. Concentrations in blood decreased in the order of bottlenose dolphins > bluefin tuna > swordfish. Perfluorooctanoate (PFOA) were detected only in certain locations, indicating their sporadic spatial distribution. The highest concentration of FOSA was found in the liver of a common dolphin (880 ng g–1 ww).

Mercury and other heavy metals

Concentrations of Hg and Cd were determined in five deep-sea benthic fauna species: Polycheles typhlops, Acanthephyra eximia, Aristeus antennatus (Crustacea), Bathypterois mediterraneus, and Nezumia sclerorhynchus (Fish) (Table 9.14; Kress et al., 1998).

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Deliverable 2.2 Table 9.14. Concentrations of mercury and cadmium in the species studied in each sampling site (µg g-1 wet weight; Krees et al., 1998).

Length Weight Dry wt. Hg Cd (cm) (g) (%) Polycheles typhlops Atlit Range 2.0-9.5 0.1-12.8 9-34 bdl-0.292 bdl-1.93 Mean (SD) 0.018 (0.044)* 0.44 (0.04) N(bdl) 49 57 52 57 (32) 57 (1) Hadera Range 2.0-7.6 0.1-5.2 7-18 bdl-0.175 0.18-1.07 Mean (SD) (l.016 (0.033)* 0.42 (0.02) N (bdl) 93 93 93 93 (58) 93 Haifa Range 3.0-7.6 0.3-5.1 6-20 bdl-0.271 0.07-1.04 Mean (SD) 0.063 (0.079)* 0.39 (0.03) N(bdl) 81 82 71 82 (24) 82 Acanthephrra eximia Atlit Range 4.2-10.2 0.7-14.0 15-30 bdl-1.896 0.37-11.9 Mean (SD) 0.098 (0.036) 1.28 (0.17) N (bdl) 33 39 37 39 (4) 39 Hadera Range 4.1-11.3 0.6-17.5 12-24 bdl-1.443 0.26-10.8 Mean (SD) 0.114 (0.029) 1.37 (0.15) N (bdl) 68 67 66 67 (3) 68 Haifa Range 5.4-12.0 1.1-20.4 11-23 bdl-1.651 0.03-9.08 Mean (SD) 0.059 (0.017) 1.25 (0.16) N (bdl) 57 57 57 57 (8) 57 Aristeus antenna/us Atlit Range 5.0-12.5 1.2-16.2 18-26 bdl-0.038 bdl-0.35 Mean (SD) 0.003 (0.003)* 0.06 (0.01) N (bdl) 16 23 23 23 (12) 23 (1) Hadera Range 6.8-15.3 3.1-27.8 17-26 bdl-0.294 bdl-0.19 Mean (SD) 0.044 (0.083)* 0.06 (0.01) N (bdl) 27 27 27 27 (11) 26 (1) Haifa Range 7.4-16.5 2.96-32.0 17-23 bdl-0.171 bdl-0.33 Mean (SD) 0.015 (0.032) 0.06 (0.01) N(bdl) 28 28 26 28 (5) 28 (I) Range 6.8-11.2 1.6-6.4 15-20 bdl-0.357 bdl-0.19 Mean (SD) 0.111 (0.108) 0.06 (0.03) N (bdl) 5 6 6 6 (1) 6 (1) Hadera Range 5.5-14.0 0.8-17.6 12-21 bdl-0.934 0.03-0.19 Mean (SD) 0.129 (0.035) 0.07 (0.01) N(bdl) 26 26 26 26 (1) 26 Haifa Range 7.0-13.5 1.8-12.1 18-24 bdl-0.239 bdl-0.15 Mean (SD) 0.018 (0.006) 0.04 (0.01) N(bdl) 27 27 27 27 (6) 27 (2) Range 14.5-23.8 8.7-22.5 19-22 bdl-0.150 bdl-0.07 Mean (SD) <0.036 (0.045)* N (bdl) 16 16 16 16 (1) 16 (2) *aritmethic mean

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Deliverable 2.2 Specimens (n = 548) were collected in the southeastern Levantine Basin between 32°30'-33°00'N, 34°00'- 34°40'E, at two deep-sea dump sites and at a deep-sea control area in the southeastern Mediterranean Sea between 1988-1995. One dumpling site, Haifa, encompassed an area of 50 km2 and was located 40 km offshore from Haifa. This site receives gypsum acidic wastes from a single source (36000 t yr-1). The other dumping site, Hadera, has an area of 210 km2 and is located 70 km offshore from Hadera. It receives coal fly ash (approximately 1.25 million tons dumped between 1989 and 1995). Atlit, located 50 km offshore from Atlit, was a control area located between the two dump sites, 30 km from Hadera and 23 km from Haifa, sufficiently removed from them to prevent inaccurate dumps or bottom transport of the waste material to the control area.

Specimens collected at the dump sites were used as biomonitors to assess the impact of waste dumping operations. Comparison of results to the natural metal levels showed some significant differences among the sites, but no systematic trends. The high levels of Hg and Cd in the reference site found in some of the species sampled are intriguing, in particular Hg in B. mediterraneus and A. eximia, and Cd in A. eximia and P. typhlops (Table 9.14). Similar elevated concentrations were reported for various species from the eastern Mediterranean and from other deep-sea areas. High Cd levels were found in pelagic species of the eastern Mediterranean (in particular the carnivore amphipod Phrosina semilunatica) (Fowler, 1986) and high Hg levels were found in some pelagic species of the eastern Mediterranean (Fowler, 1986). The origin of these metals, and determinants for the differences in accumulation between crustacean species, as well as between crustaceans and fish are not well known.

The order of decreasing Cd and Hg concentrations in the crustaceans studied was: A. eximia > P. typhlops > A. antennatus. A. eximia feeds mainly on plankton and micronekton (Christiansen, 1989), P. typhlops is a scavenger feeder, feeding mainly on the remains of fish and decapod crustaceans (Cartes and Abello, 1992), while A. antennatus feeds entirely on benthic organisms that live completely or partially buried in the substratum (Cartes and Sarda, 1989). High concentrations of Cd were found in plankton and nekton in the eastern Mediterranean (Fowler, 1986) and in crustaceans (White and Rainbow, 1987; Ridout et al., 1989). Thus, A. eximia and P. typhlops that feed on a Cd rich diet, together with the absence of Cd metabolic regulation in crustaceans (Rainbow and White, 1989; Rainbow et al., 1990), accumulated Cd while A. antennatus did not.

Mercury is the only metal that shows biomagnification (Eisler, 1981). Fowler (1986) found that Hg concentration increased in the deep-sea pelagic food chain in the eastern Mediterranean: microplankton (autotrophs, herbivores), euphasiids, myctophids (carnivorous). Therefore, it seems that Hg accumulated in crustaceans in a way similar to Cd, through the diet. Moreover, it is reasonable that accumulation of Hg in B. mediterraneus might he similar to that in the crustaceans because, like A. eximia, B. mediterraneus feeds almost exclusively on benthopelagic and pelagic small sized prey (copepoda, gammaridea, mysidacea; Carrasson and Matallanas, 1990). Pellegrini and Barghigiani (1989) found that flat fish feeding on fish and crustacea have higher Hg levels than flat fish living in the same environment, but feeding on

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Deliverable 2.2 polychaeta and molluscs. Therefore, based on its diet, it is reasonable to expect that N. sclerorhynchus will have lower Hg level than B. mediterraneus.

Mercury and cadmium were analyzed in the muscle and other organs from several species of deep-water sharks: Centrophorus granulosus, Galeus melastomus, Somniousus rostratus, Etmopterus spinax and Hexanchus griseus (Tables 9.13 and 9.14; Hornung et al., 1993). They were collected from 1280 to 1500 m depth in the eastern Mediterranean between 1985 and 1991. Most of the sharks examined were collected in a rectangular area between 32°31'N-33°02'N and 34°02'E-34°37'E, approximately 45 km west of Haifa (Israel). As described above, since 1985 the area was used as a deep water disposal site for gypsum wastes (total amount dumped at this site in 1990: 144 000 t). Other sharks were collected in Hadera and one station opposite Tel-Aviv (32°30'N and 33°38 'E) used as a control area.

The mercury content in the muscle, liver and kidneys increased with length and weight. For G. melastomus a quasi-exponential relationship between mercury content in the muscle and body weight was observed (Figure 9.15). The mercury content in the liver and kidneys of G. melastomus covaried with the concentration in the muscle throughout the size range sampled (Figure 9.16) with linear regression coefficients of 0.72 and 0.69 for liver and kidney, respectively. Individuals with dark livers followed the general trend.

Table 9.15. Average concentrations of Hg (µg g-1 wet weight) in various tissues of five species of sharks from the Eastern Mediterranean caught between 1300 m and 1500 m and at 700 m depth (Hornung et al., 1993). Centroporus Galeus Somniosus Etmopterus Hexanchus granulosus melastomus rostratus spinax griseus Muscle 0.48-8.4 0.99-8.8 0.86-5.6 1.8-4.6 0.41-4.5 Skin 0.42-2.0 1.2-1.3 Brain 0.92-2.7 2.0 Heart 2.5-3.6 0.56-2.1 14 Kidney 0.77-11 1.1-25 2.4-24 2.0-13 0.39-11 Livera 0.60-10 0.28-5.9 14 2.0-3.0 0.42-13 Liverb 4.5-23 3.1-17 2.4-24 4.6-6.3 Gonads bdl-3.7 0.24-2.1 0.58-3.0 1.2-2.7 Intestines 0.20-4.8 0.18-3.6 0.04-11 1.3-3.7 0.05-3.3 Stomach 0.11-1.2 0.48-1.7 0.46-1.8 0.60-0.70 contets Egg sac 0.46-1.0 anormal healthy liver. bvery dark liver or dark patches on liver.

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Deliverable 2.2 Table 9.16. Average concentrations of Hg (µg g-1 wet weight) in various tissues of five species of shaks from the Eastern Mediterranean caught between 1300 m and 1500 m and at 700 m depth (Hornung et al., 1993). Centroporus Galeus Somniosus Etmopterus Hexanchus granulosus melastomus rostratus spinax griseus Muscle 0.06 0.07 0.07 0.08 0.04 Skin 0.13 1.1 Brain 0.11 0.14 Heart 1.2 0.25 0.38 Kidney 1.2 0.99 1.9 1.4 0.71 Livera 1.8 1.6 4.0 0.79 1.1 Liverb 7.7 5.1 12 3.1 Gonads 0.13 0.16 0.10 0.17 Intestines 1.7 1.8 5.5 2.3 1.3 Stomach 0.11 1.9 0.28 0.91 contets Egg sac 0.16 anormal healthy liver. bvery dark liver or dark patches on liver.

The mercury content in the muscle of C. granulosus increased with increasing body weight until it reached a constant mercury content at higher (increasing) body weight (Figure 9.17). Dark spots were found on the livers of several of the largest specimens of this species. The concentration of mercury in the muscle of these individuals was not unusually high, though the mercury in liver and kidney vs mercury in muscle appeared to be higher for several of those individuals than for the rest of the population (Figure 9.18).

For all the shark species sampled, mercury levels were generally lowest in the gonads and intestines and highest in the muscle, liver and kidneys (in ascending order; Table 9.15). There was no apparent relationship between mercury content and the sex of the individual for any of the species considered.

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Deliverable 2.2

Figure 9.15. Distribution of mercury in the muscle of G. melastomus vs body weight. FW: fresh weight. Data from Renzoni and Baldi (1975) and Hornung et al. (1993)

For Cd no significant correlations were found between the metal content in the muscle and the body weight of G. melastomus and C. granulosus. In all species the levels in the muscle was low compared to the other organs (Table 9.16). Unusually high values of Cd (similar to Hg) were observed in livers with dark patches or in livers which were small and very dark.

Although most of the sharks were collected near a deep-water dumpsite, there is no evidence that the levels of trace metals were affected by the disposal operation. There were no observed differences be• tween the levels of metals in the sharks sampled adjacent to the waste disposal site and those sampled at the Hadera and Tel-Aviv areas. The gypsum wastes which are disposed of do not contain significant amounts of mercury (2.4 kg yr-1) and thus it is considered unlikely that mercury levels in sharks were elevated by local pollution. It is suggested that the results reflect the natural contents of trace metals of these species in the eastern Mediterranean area.

Mercury concentration increase with the age of marine organisms as well as with their size (length and weight) if a positive correlation exists between length/weight and age. This has been observed in many fish species, molluscs, crustaceans and marine mammals (e.g. Peterson et al., 1973; Shultz et al., 1976; Stoeppler et al., 1979). Exceptions have been found only in some bivalves, where total mercury decreased with weight (Hornung and Oren, 1980/1981). Position in the food chain is another important variable because mercury is preconcentrated in the prey and therefore organisms located higher in the food chain have at comparable age higher HgT concentrations than those with shorter food chains.

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Figure 9.16. Distribution of mercury in liver and kidney of G. melastomus vs mercury concentrations in liver. FW: fresh weight. Data from Hornung et al. (1993).

Figure 9.17. Distribution of mercury in the muscle of C. granulosus vs body weight. FW: fresh weight. Data from Renzoni and Baldi (1975) and Hornung et al. (1993).

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Deliverable 2.2 The increase of HgT with size has also been observed in shark species (Lyle, 1986), and the data on deep- water sharks from the eastern Mediterranean (Hornung et al., 1993).

Figure 9.18. Distribution of mercury in liver and kidney of C. granulosus vs mercury concentrations in liver. FW: fresh weight. Data from Hornung et al. (1993).

In order to make a detailed comparison of the findings of the present study and those from other areas, it is necessary to compare mercury vs weight for the same species. Comparison of the results of G. melastomus (Figure 9.16) and for C. granulosus (Figure 9.17) from the study of Hornung et al. (1993) with those available from the west coast of Italy (Renzoni and Baldi, 1975) showed higher mercury levels in the first case.

Radionuclides

The low radioactivity levels explain why very few papers have been published on radionuclide concentration in organisms in Mediterranean Sea. 137Cs (the main anthropogenic source of radioactive dose in population from the marine pathway) is the only radionuclide measured by the different countries through their own national networks for monitoring environmental radioactivity. In the deep environment, concentrations of 137Cs were determined in Merluccius merluccius in Gulf of Lions down to 600 m depth ranging between 0.47 to 0.96 (Bq kg-1 dw) (Harmelin-Vivien et al., 2012).

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Deep-sea organisms have adapted to particular environmental conditions (high pressure, low temperature and absence of light), and they may respond differently than coastal species to pollution stress (Gross and Jaenicke, 1994). Xenobiotics within the organism undergo a suite of reactions to facilitate their excretion. In all eukaryotes, the cytochrome P450 system and glutathione 5-transferases (GSTs) play a key role in the biotransformation (monooxygenation and conjugation) of lipophilic foreign chemicals, such as PAHs and polychlorinated biphenyls (PCBs). Substantial differences in monooxygenase or transferase activities and number of isoenzymes have been reported in marine organisms, depending on habitat, pollutants load, etc. (Fortin et al., 1995; Beyer et al., 1996; Stegeman et al., 1997). Metabolic rates of deep-sea fish are known to decrease with depth as a result of several interacting factors, such as low temperatures, low food availability and poor locomotor capabilities (Gibbs, 1997); thus, strong differences from coastal species in terms of xenobiotic metabolising enzymes could be anticipated.

Many pollutants in aquatic systems exert their toxic effects due to oxidative stress (Winston and Di Giulio, 1991; Thomas and Wofford, 1993; van der Oost et al., 1996). Oxyradicals are continually produced in eukaryotes as unwanted byproducts of normal oxidative metabolism, and their production can be increased by condition s such as hypoxia/hyperoxia, redox cycling xenobiotics (e.g., metals, quinones, nitroaromatic compounds) and induction of enzymes, such as cytochrome P450 and P450 reductase (Premereur et al., 1986). Consequently, aerobic organisms have developed defence systems against oxidative damage (Di Giulio et al., 1989), consisting of antioxidant scavengers (glutathione, vitamin C, vita min E, carotenoid pigments).

The biotransformation of xenobiotics mainly takes place in the liver/hepatopancreas and usually consists of an alteration (oxidation, reduction or hydrolysis) of the original foreign molecule (Phase I), followed by a conjugation reaction (Phase II) where a large endogenous molecule (e.g. glucoronid, glutathione, sulfate) is added to increase its water-solubility and facilitate excretion. Enzymes involved in the Phase I metabolism of xenobiotics include cytochrome P450 (CYP) enzymes (oxidation) and carboxylesterases (hydrolysis), while glutathione-S-transferase (GST) plays an important role in the Phase II metabolism. Moreover, the exposure of aquatic organisms to anthropogenic contaminants can result in the production of reactive oxygen species (ROS), potentially causing adverse effects related to oxidative stress (Winston and Di Giulio, 1991). An increase in oxidative stress can result in the induction of antioxidant enzymes such as catalase (CAT), glutathione peroxidase (GPX) and superoxide-dismutase (SOD) as a protection mechanism against the generation of oxidative radicals (Winston and Di Giulio, 1991; Valavanidis et al., 2006). However, when this protective mechanism is overwhelmed, lipid peroxidation (LP) (i.e. the oxidation of polyunsaturated fatty acids) can occur (Winston and Di Giulio, 1991; Valavanidis et al., 2006).

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Deliverable 2.2 Enzymes involved in the detoxiﬁcation of xenobiotics and their metabolites, including biotransformation and antioxidant enzymes, are the most extensively studied biomarkers (van der Oost et al., 2003). Strong differences among species were seen in EROD activity of L. lepidion, C. guentheri and B. mediterraneus from sites located at 1500-1800 m below sea level in the NW Mediterranean. EROD activities were significantly elevated in L. lepidion (111 to 192 pmol min-1 mg-1 protein) compared with C. guentheri (23 to 29 pmol min-1 mg-1 protein) and B. mediterraneus (2.6 to 7.2 pmol min-1 mg-1 protein) (Table 9.17; Porte et al., 2000). In contrast, the strong differences among species concerning GST involved the highest specific activity detected in C. guentheri, 1.4 to 1.6 nmol min-1 mg-1 cytosolic protein, and lower in L. lepidion, 0.9 nmol min-1 mg-1 cytosolic protein, and B. mediterraneus, 0.2 nmol min-1 mg-1 cytosolic protein.

Hepatic microsomal preparations of the species studied showed active electron transport components and native cytochrome P450. Total cytochrome P450 content (P450+P420) in C. guentheri and L. lepidion were in the higher range (0.9 to 2.0 mol mg-1 protein) reported for teleosts. The specific activity of NADPH- cytochrome c (P450) reductase, the microsomal flavoprotein that transfers electrons from NADPH to cytochrome P450, was also similar to that reported for other species (Mathieu et al., 1991; Haasch et al., 1993; Sleiderink et al. 1995). However, only L. lepidion exhibited an EROD activity comparable to coastal fish (Collier et al., 1995; Forlin et al., 1995). C. armatus sampled in the western North Atlantic and C. rupestris caught in the North Sea (Stegeman et al. 1986; Forlin et al., 1995), showed similar EROD activities to those found in the NW Mediterranean. High pressure and cold temperatures may influence membrane functioning by reducing fluidity and affecting protein-protein interactions (Gibbs, 1997). Thus, the interaction between cytochrome P450 and reductase in the membrane will be less effective in deep-sea fish than in shallow-water fish, which may lead to low cytochrome P450 catalytic activities L. lepidion, a typical middle slope species (1 000 to 1 400 m), exhibited higher EROD activity (4-74 fold) than B. mediterraneus and C. guentheri; both of which are adapted to live at greater depths, with maximum abundance at 1600 to 2200 m (Stefanescu et al. 1992).

Cytosolic GST activity was in the range or even higher than the activities reported for coastal fish (Forlin et al., 1995). The high CST activity detected in C. guentheri is related to high P450 specific content, which is in accordance with the important role of GSTs in conjugation of electrophiles produced by P450 monooxygenation. Elevated GST activity may have been selected in these organisms as protection against toxic dietary chemicals. C. guentheri lives in direct contact with the sediment, it feeds actively on small epibenthic and endobenthic in vertebrates (Stefanescu et al., 1992), and it might the re fore be exposed to a higher amount of sediment-trapped pollutants than the other two species. Although levels of PCBs in muscle tissue of C. guentheri were relatively low, previous work by Escartin and Porte (1999) found PAH metabolites in bile of this species up to one order of magnitude higher than in B. mediterraneus or L. lepidion. Alternatively, low GST activities were observed in B. mediterraneus. This species also presented the lowest P450 content and has a diet restricted to zooplankton (Carrason and Matallanas, 1990). lnterspecies variations were also observed in antioxidant enzyme activities. L. lepidion showed higher catalase and SOD activities, higher catalytic efficiency (EROD/P450) and elevated cytochrome P450 content. A higher production of oxyradicals via reactions catalysed by cytochrome P450 enzymes can be anticipated, requiring mechanisms to remove those oxyradicals

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Deliverable 2.2

In the NW Mediterranean, at 900 m depth, the fish L. lepidion and the crustacean A. antennatus exhibit higher organic contaminant accumulation as well as increased biomarker responses (Table 9.17; Koenig et al., 2013b). The differences appear to be more pronounced in L. lepidion than in A. antennatus, suggesting differential bioaccumulation dynamics (uptake and excretion) between the two species. This is in accordance with previous results showing that L. lepidion has lower CYP-mediated PCB metabolizing capacities than A. antennatus and thus accumulates PCBs more readily (Koenig et al., 2012). In accordance with chemical results exhibit significantly higher activities of cytochrome P450-related EROD and PROD enzymes, respectively, as well as antioxidant enzymes CAT and GPX, respectively (Table 9.17). A multitude of chemicals have been shown to be able to induce CYP enzymes and numerous studies have linked increases in EROD/ PROD activity to exposure to environmental stressors (Goksøyr and Förlin, 1992; Whyte et al., 2000; van der Oost et al., 2003). L. lepidion exhibit a 600% increase in CYP1A-related EROD activity, while A. antennatus CYP2B-related PROD activity increase by approximately 100%. As reviewed in van der Oost et al. (2003), an increase in EROD activity is considered as strong if the activity is 500% of the control value. In the study of Koenig et al (2013b), the strong increase in EROD activity in significantly smaller juvenile fish is particularly remarkable considering the fact that former studies revealed that the inducibility of EROD, as a result of contaminant exposure, is lower in smaller fish (Whyte et al., 2000).

Concomitant with the above-mentioned increase in CYP activity, both species experience an increase in antioxidant enzyme activities (Table 9.17). This finding is consistent with the fact that the CYP-mediated biotransformation of organic contaminants may enhance the production of reactive oxygen species (ROS), potentially causing an increase in oxidative stress and subsequent antioxidant responses (i.e. CAT and GPX in this case) (Winston and Di Giulio, 1991). The lack of increase in LP levels for L. lepidion samples suggests that oxidative stress does not result in the peroxidation of lipids. The fact that different antioxidant enzymes respond in the fish and the crustacean species might reflect a potentially different role of CAT and GPX between these organisms. As seen in other deep-sea fish (Janssens et al., 2000), A. rostratus and L. lepidion, exhibit low GPX activities, whereas CAT activities are almost one order of magnitude higher than those of fish dwelling at shallower depths (Solé et al., 2010). In this context, it has been postulated that GPX is mainly used as defense mechanism against metabolically produced ROS, while CAT could be essentially responsible for eliminating exogenously-generated H2O2 in deep-sea fish (Janssens et al., 2000). Thus, in deep-sea fish, CAT might exhibit a more marked response than GPX as a result of xenobiotic exposure, which is also in accordance with other studies (Lemaire et al., 2010). In contrast to fish, A. antennatus exhibit low CAT and high GPX activity, indicating that crustaceans potentially rely on different antioxidant mechanisms than fish to respond to oxidative stress. This, in turn, would explain why in A. antennatus GPX, and not CAT, enzymes respond to the putative increase in pollutant-induced ROS production.

In addition, CbE activity is signiﬁcantly lower in L. lepidion from the Blanes Canyon (NW Mediterranean). In general, CbE activity is believed to increase in the presence of xenobiotics due to its role in the metabolism of ester-containing compounds (Wheelock et al., 2005). However, certain chemicals such as tributyltins (TBTs) and organophosphate pesticides are known to inhibit CbE activity (Al-Ghais et al., 2000;

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Deliverable 2.2 Wheelock et al., 2005). In particular, elevated levels of TBTs have been found in deep-sea fish within the same region, indicating the presence of this contaminant at great depths (Borghi and Porte, 2002). The significant inhibition of CbE could thus reflect higher exposure to the above-mentioned pollutant classes. This trend is however not reflected in A. antennatus, for which CbE activities do not differ between sites at 900 m depth. The fact that other biomarker responses applied in the present work (i.e. GST, GR, SOD and LP) do not reveal any significant differences between sites in autumn at 900 m depth can be due to their lower sensitivity as pollution biomarkers compared to, for instance, the EROD assay (van der Oost et al., 2003).

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Deliverable 2.2 Table 9.17. Biomarker responses (mean ± S.E.) for Alepocephalus rostratus, Lepidion lepidion and Aristeus antennatus in deep-sea samples from the NW Mediterranean (Koenig et al., 2013b) and L.lepidion and Coryphaenoides guentheri and Bathypterois mediterraneus (Porte et al., 2000). Values are shown for ethoxyresoruﬁn-O-deethylase (EROD), pentoxyresoruﬁn-O-deethylase (PROD), carboxylesterase (CbE) , glutathione S-transferase (GST), total glutathione peroxidase (GPX), glutathione reductase (GR), catalase (CAT), superoxide-dismutase (SOD) activities and lipid peroxidation levels (LP). Species Site Depth (m) Length (mm) Sex ratio EROD/PRODa CbEb GSTb GPXb GRb CATc SODd LPe M:F:Im A. rostratus OS 1500 320 ± 13 3:7:0 260 ± 85 140 ± 11 180 ± 10 51 ± 1.7 2.9 ± 0.4 2700 ± 180 13 ± 1.2 n.a. BC 1500 310 ± 9.1 8:2:0 390 ± 95 170 ± 15 210 ± 10 50 ± 1.8 1.7 ± 0.3 2800 ± 150 12 ± 0.7 n.a.

OS 1500 320 ± 13 1:9:0 280 ± 120 230 ± 37 290 ± 12 53 ± 2.1 3.6 ± 0.2 2800 ± 150 22 ± 1.5 63 ± 8.7 BC 1500 290 ± 10 7:3:0 450 ± 110 230 ± 33 290 ± 34 50.7 ± 4.1 3.1 ± 0.5 2600 ± 110 15 ± 1.0 96 ± 24.6 L. lepidion OS 900 190 ± 5.2 3:6:1 7900 ± 2400 53 ± 3.3 320 ± 23 52 ± 2.9 3.6 ± 0.6 920 ± 89 n.a. 22 ± 3.9 BC 900 110 ± 9.7 1:1:8 44000 ± 13000 30 ± 6.6 280 ± 24 43 ± 1.8 2.1 ± 0.8 1300 ± 78 n.a. 20 ± 2.4

OS 1500 220 ± 14.4 4:6:0 4100 ± 2200 32 ± 3.2 350 ± 15 35 ± 1.7 3.8 ± 0.4 1000 ± 72 11 ± 1.9 23 ± 2.8

BC 1500 220 ± 7.7 1:7:2 2100 ± 430 29 ± 2.4 380 ± 22 36 ± 1.7 3.1 ± 0.4 630 ± 42 20 ± 5.9 19 ± 2.5

1500-1800 213 ± 7 130000 ± 56000 960 ±120 72.7 ± 8.8 92 ± 9.8 13 ± 1.8

A. antennatus OS 900 44 ± 3.7 1:9:0 190 ± 16 340 ± 45 210 ± 62 180 ± 16 0.9 ± 0.3 9.4 ± 1.4 n.a. 35 ± 4.9 BC 900 32 ± 1.5 0:10:0 160 ± 11 690 ± 50 190 ± 37 220 ± 14 1.0 ± 0.3 7.1 ± 1.1 n.a. 27 ± 1.4 OS 900 43 ±3.1 5:5:0 160 ± 17 360 ± 24 67 ± 11 140 ± 8.0 0.9 ± 0.3 7.7 ± 1.2 89 ± 23 n.a. BC 900 38 ±4.3 5:5:0 270 ± 21 320 ± 32 71 ± 8.9 200 ± 16 1.3 ± 0.4 4.4 ± 0.9 54 ± 19 n.a. OS 1500 24 ± 1.6 5:5:0 200 ± 62 800 ± 81 240 ± 71 310 ± 48 1.8 ± 0.4 3.1 ± 1.4 47 ± 13 n.a.

BC 1500 34 ± 4.0 8:2:0 240 ± 26 590 ± 50 83 ± 9.9 220 ± 19 1.0 ± 0.2 2.4 ± 0.6 81 ± 26 n.a.

C. guentheri 1500-1800 149 ± 3.0 27000 ± 7200 1500 ± 210 99 ± 17 99 ± 17 78 ± 9.5 10 ± 1.3

B. mediterraneous 1500-1800 157 ± 4.0 7200 ± 1100 210 ± 130 19 ± 0.5 20 ± 0.5 33 ± 6.4 8.9 ± 3.1 n.a., not analyzed. afmol min-1 mg protein-1. bnmol min-1 mg protein-1. cµmol min-1 mg protein-1. da.U mg protein-1. e nmol MDA g-1 wet weight.

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Sanchez-Vidal A, Higueras M, Marti E, Liquete C, Calafat A, Kerherve P, Canals M (2013) Riverine transport of terrestrial organic matter to the North Catalan margin, NW Mediterranean Sea. Progr. Oceanogr., 118, 71-80. Scheringer M, Stroebe M, Wania F, Wegmann F, Hungerbühler K (2004) The effect of export to the deep sea on the long-range transport potential of persistent organic pollutants. Environ Sci Pollut Res Int 11:41–48. Schoeib M, Harner T (2002) Using measured octanol-air partition coefficients to explain environmental partitioning of organochlorine pesticides. Environ Toxicol Chem 21, 984-990. Schroeder W, Munthe J (1998) Atmospheric mercury—an overview. Atmos Environ 32:809–822 Schulz DE, Petrick G, Duinker JC (1988) Chlorinated biphenyls in North Atlantic surface and deep wáter. Mar Pollut Bull 19:526-531 Schulz DE, Petrick G, Duinker JC (1991) Polychlorinated biphenyls in North Sea water. Mar Chem 36:365-384. Sericano JL, Brooks JM, Champ MA, Kennicutt Ii MC, Makeyev VV (2001). Trace contaminant concentrations in the Kara Sea and its adjacent rivers, Russia. Mar Pollut Bull 42, 1017-1030. Serrano R, Fernandez M, Rabanal R, Hernandez M, Gonzales MJ (2000) Congener-specific determination of polychlorinated biphenyls I shark and grouper livers from the Northwest African Atlantic Ocean. Arch Environ Contam Toxicol 38, 217–224. Shultz CD, Crear D, Pearson JE, Rivers JB, Hylin JW (1976) Total and organic mercury in the Pacific blue marlin. Bull Envir Contam Toxic 15, 230-234. Sicre MA, Marty JC, Saliot A, Aparicio X, Grimalt JO, Albaigés J (1987) Aliphatic and aromatic hydrocarbons in different sized aerosols over the Mediterranean Sea: Occurrence and Origin. Atmos Environ 21, 2247- 2259. Simoneit BRT (1977) Organic matter in eolian dusts over the Atlantic Ocean. Mar Chem 5, 443-464. Simoneit BRT, Grimalt JO, Wang T, Cox RE, Hatcher PG, Nissenbaum A (1986) Cyclic terpenoids of contemporary resinous plant detritus and of fossil woods, ambers and coals. Org Geochem 10, 877-889. Sleath JFA (1987) Turbulent oscillatory flow over rough beds. J. Fluid Mech. 182, 369-409. Sleiderink HM, Oostingh I, Goksoyr A, Boon JP (1995) Sensitivity of cytochrome P450 IA induction in dab (Limanda limanda) of different age and sex as a biomarker for envi• ronmental contaminants in the Southern North Sea. Arch Environ Contam Toxicol 28:423- 430 Soderstrom G, Sellstrom U, De Wit CA, Tysklind M (2004) Photolytic debromination of decabromodiphenyl ether (BDE 209). Environ Sci Technol 38, 127-132. Solé M, Antó M, Baena M, Carrasson M, Cartes JE, Maynou F (2010) Hepatic biomarkers of xenobiotic metabolism in eighteen marine fish from NW Mediterranean shelf and slope waters in relation to some of their biological and ecological variables. Mar Environ Res 70, 181–188. Solé M, Porte C, Albaiges J (2001) Hydrocarbons, PCBs and DDT in the NW Mediterranean deep-sea fish Mora moro. Deep Sea Res. Part I 48:495–513. Sprovieri F, Pirrone N, Gardfeldt K, Sommar J (2003) Atmospheric mercury speciation in the marine boundary layer along 6000 km cruise path over the Mediterranean Sea. Atmos Environ 37(S1):63–72 Stapleton HM, Brazil B, Holbrook RD, Mitchelmore CL, Benedict R, Konstantinov A, et al. (2006) In vivo and in vitro debromination of decabromodiphenyl ether (BDE 209) by juvenile rainbow trout and common carp. Environ Sci Technol 40:4653–4658. Stapleton HM, Letcher RJ, Baker JE (2004) Debromination of polybrominated diphenyl ether congeners BDE 99 and BDE 183 in the intestinal tract of the common carp (Cyprinus carpio). Environ Sci Technol 38:1054– 61. Stastna M, Lamb KG (2008) Sediment resuspension mechanisms associated with internal waves in coastal waters. J. Geophys. Res-Oceans 113, C10016, doi: 10.1029/2007JC004711. Stefanescu C, Rucabado J, Lloris D (1992) Depth-size trends in western Mediterranean demersal deep-sea fishes. Mar Ecol Prog Ser 81:205-213.

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Stefanescu, C, Lloris, D, Rucabado, J (1993) Deep-sea fish assemblages in the Catalan Sea (western Mediterranean) below a depth of 1000 m. Deep-Sea Research I 40, 695-707. Stegeman JJ, Kloepper-Sams PJ, Farrington JW (1986) Monooxygenase induction and chlorobiphenyls in the deep-sea fish Coryphaenoides armatus. Science 231, 1287-1289. Stegeman JJ, Woodin BR. Singh H, Oleksiak MF, Celander M (1997) Cytochromes P450 (CYP) in tropical fishes: catalytic activitities, expression of multiple CYP proteins and high levels of microsomal P450 in liver of fis hes from Be rmuda . Comp Biochem Physiol 116C61-75 Steimle FW, Zdanowicz VS, Gadbois DF (1990) Metals and organic contaminants in Northwest Atlantic deep- sea tile"sh tissues. Mar Pollut Bull 21, 530-535. Stergiou KI, Karpouzi VS (2002) Feeding habits and trophic levels of Mediterranean fish. Rev Fish Biol Fish 11, 217–254. Stoeppler, M., Bernhard, M., Backhaus, F., Schulte, E. (1979) Comparative studies on trace metal levels in marine biota. I. Mercury in marine organisms from western Italian coast, the Strait of Gibraltar and the North Sea. Sci. total Envir. 3: 209- 223 Storelli MM, Losada S, Marcotrigiano GO, Roosens L, Barone G, Neels H, et al. (2009) Polychlorinated biphenyl and organochlorine pesticide contamination signatures in deep-sea fish from the Mediterranean Sea. Environ Res 109:851–856. Storelli MM, Perrone VG, Marcotrigiano GO (2007) Organochlorine contamination (PCBs and DDTs) in deep- sea fish from Mediterranean Sea. Mar Pollut Bull 54, 1968–1971. Storelli MM, Storelli A, Barone G, Marcotrigiano GO (2004b) Polychlorinated biphenyl and organochlorine pesticide residues in Lophius budegassa from the Mediterranean Sea (Italy). Mar Pollut Bull 48, 743–748. Storelli MM, Storelli A, D’Addabbo R, Barone G, Marcotrigiano GO (2004a) Polychlorinated biphenyl residues in deep-sea fish from Mediterranean Sea. Environ Int 30, 343–349. Strandberg B, Van Bavel B, Bergqvist PA, Broman D, Ishaq R, Naf C et al. (1998) Occurrence, sedimentation, and spatial variations of organochlorine contaminants in settling particulate matter and sediments in the northern part of the Baltic Sea. Environ Sci Technol 32, 1754-1759. Takahashi S, Hayashi S, Kasai R, Tanabe S, Kubodera T (2001) Contamination of deep-sea organisms from Tosa bay, Japan by organochlorine and butyltin compounds. Nat Sci Mus Monogr 20, 364–380. Takahashi S, Oshihoi T, Ramu K, Isobe T, Ohmori K, Kubodera T, et al. (2010) Organohalogen compounds in deep-sea fishes from the western North Pacific, off-Tohoku, Japan: contamination status and bioaccumulation profiles. Mar Pollut Bull 60: 187–196. Tamburrino S, Passaro S, Barsanti M, Schirone A, Delbono I, Conte F, Delfanti R, Bonsignore M, Del Core M, Gherardi S, Sprovieri M (2019) Pathways of inorganic and organic contaminants from land to deep sea: The case study of the Gulf of Cagliari (W Tyrrhenian Sea). Sci. Total Environ. 647, 334–341. Tanabe S (2002) Contamination and toxic effects of persistent endocrine disrupters in marine mammals and birds. Mar Pollut Bull 45, 69-77. Tanabe S, Ramu K, Isobe T, Takahashi S. (2008) Brominated flame retardants in the environment of Asia- Pacific: an overview of spatial and temporal trends. J Environ Monit 10:188–197. Tanabe S, Ramu K, Mochizuki H, Miyasaka H, Okuda N, Muraoka M, Kajiwara N, Takahashi S, Kubodera T (2005) Contamination and distribution of persistent organochlorine and organotin compounds in deep- sea organisms from East China Sea. Nat Sci Mus Monogr 29, 453–476. Telli-Karakoç F, Tolun L, Henkelmann B, Klimm C, Okay O, Schramm KW (2002) Polycyclic aromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs) distributions in the Bay of Marmara sea: İzmit Bay. Environ Pollut 119:383-397. Thomas P, Wofford HW (1993) Effects of cadmium and Aroclor 1254 on lipid peroxidation, glutathione peroxidase activity, and selected antioxidants in Atlantic croaker tissues. Aquat Toxicol 27 159-178

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Tokarz JA, Ahn M-Y, Leng J, Filley TR, Nies L (2008) Reductive debromination of polybrominated diphenyl ethers in anaerobic sediment and a biomimetic system. Environ Sci Technol 42, 1157-1164. Tolosa I, Bayona JM, Albaiges J (1995) Spatial and temporal distribution, fluxes, and budgets or organochlorinated compounds in Northwest Mediterranean sediments. Environ Sci Technol 29, 2519- 2527. Tolosa I, Bayona JM, Albaigés J (1996) Aliphatic and polycyclic aromatic hydrocarbons and sulfur/oxygen derivatives in Northwestern Mediterranean sediments: Spatial and temporal variability, fluxes, and budgets. Environ Sci Technol 30:2495-2503. Tolosa I, Readman JW, Fowler SW, Villeneuve JP, Dachs J, Bayona JM, Albaigés J (1997a) Zooplankton ecology in the eastern tropical Pacific oxygen minimum zone above a seamount: 1. General trends. Deep-Sea Res 44:907-930. Tolosa I, Readman JW, Fowler SW, Villeneuve JP, Dachs J, Bayona JM, et al. (1997b) PCBs in the western Mediterranean. Temporal trends and mass balance assessment. Deep-Sea Res. Pt. II, 44, 907-928. Tsabaris C, Zervakis V, Kaberi H, Delfanti R, Georgopoulos D, Lampropoulou M, Kalfas CA (2014) 137Cs vertical distribution at the deep basins of the North and Central Aegean Sea, Greece. Journal of Environmental Radioactivity, 132, 47-56. Tsapakis M, Apostolaki M, Eisenreich S, Stephanou EG (2006) Atmospheric deposition and marine sedimentation fluxes of polycyclic aromatic hydrocarbons in the eastern Mediterranean basin. Environ Sci Technol 40, 4922-4927. UNEP (1992) Mediterranean Action Plan, ‘Assessment of the State of Pollution of the Mediterranean Sea by Radioactive Substances’, UNEP, Athens, 1992. UNEP (2002) Regionally based assessment of persistent toxic substances. Mediterranean Regional Report, UNEP, Switzerland, 2002. UNEP (2003) Regionally based assessment of persistent toxic substances. Global Report, UNEP, Switzerland, 2003. Valavanidis A, Vlahogianni T, Dassenakis M, Scoullos M (2006) Molecular biomarkers of oxidative stress in aquatic organisms in relation to toxic environmental pollutants. Ecotoxicol Environ Safety 64, 178–189. van der Oost R, Beyer J, Vermeulen NPE (2003) Fish bioaccumulation and biomarkers in environmental risk assessment: a review. Environ Toxicol Pharmacol 13, 57–149. van der Oost R, Goksoyr A, Celander M, Heida H, Vermeulen PE (1996) Biomonitoring of aquatic pollution with feral eel (Anguilla anguilla. ll. Biomarkers: pollution-induced biochemical responses. Aquat Toxicol 36 189-222. van Drooge BL, Fontal M, Bravo N, Fernández P, Fernández MA, Muñoz-Arnanz J, Jiménez B, Grimalt JO (2014) Seasonal and spatial variation of organic tracers for biomass burning in PM1 aerosols from highly insolated urban areas. Environ Sci Pollut Res 21, 11661-11670. van Drooge BL, Grimalt JO, Camarero L, Catalan J, Stuchlik E, Torres Garcia CJ (2004a) Atmospheric semivolatile organochlorine compounds in European high-mountain areas (Central Pyrenees and High Tatras). Environ Sci Technol 38, 3525-3532. van Drooge BL, Grimalt JO, Camarero L, Catalan J, Stuchlik E, Torres Garcia CJ (2004b) Atmospheric semivolatile organochlorine compounds in European high-mountain areas (Central Pyrenees and High Tatras). Environ Sci Technol 38, 3535-3532 (2004) van Drooge BL, Grimalt JO, Torres García CJ, Cuevas E (2002) Semivolatile Organochlorine Compounds in the Free Troposphere of the Northeastern Atlantic. Environ Sci Technol 36, 1155-1161. van Drooge BL, Grimalt JO, Torres-García CJ, Cuevas, E (2001) Deposition of semi-volatile organochlorine compounds in the free troposphere of the Eastern North Atlantic Ocean. Mar Pollut Bull 42, 628-634. van Leeuwen SPJ, van Velzen MJM, Swart CP, van der Veen I, Traag WA, de Boer J (2009) Halogenated contaminants in farmed salmon, trout, tilapia, pangasius, and shrimp. Environ Sci Technol 43:4009–4015.

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Vaupotic J, Gregoric A, Kotnik J, Horvat M, Pirrone N (2008) Dissolved radon and gaseous mercury in the Mediterranean seawater. J. Environ. Radioact. 99, 1068-1074. Voorspoels S, Covaci A, Maervoet J, De Meester I, Schepens P (2004) Levels and profiles of PCBs and OCPs in marine benthic species from the Belgian North Sea and the Western Scheldt Estuary. Mar Pollut Bull 49:393–404. Voorspoels S, Covaci A, Schepens P. (2003) Polybrominated diphenyl ethers in marine species from the Belgian North Sea and the Western Scheldt Estuary: levels, profiles, and distribution. Environ Sci Technol 37:4348–4357. Wakeham SG, Schaffner C, Giger W (1980) Polycyclic aromatic hydrocarbons in recent lake sediments. II. Compounds derived from biogenic precursors during early diagenesis. Geochim Cosmochim Acta 44, 415- 429. Walker K, Vallero DA, Lewis RG (1999) Factors influencing the distribution of lindane and other hexachlorocyclohexanes in the environment. Environ Sci Technol 33, 4373-4378. Wang Z, Ma X, Lin Z, Na G, Yao Z. (2009) Congener specific distributions of polybrominated diphenyl ethers (PBDEs) in sediment and mussel (Mytilus edulis) of the Bo Sea, China. Chemosphere 74:896–901. Wängberg I, Munthe J, Pirrone N, Iverfeldt Å, Bahlman E, Costa P, Ebinghaus R, Feng X, Ferrara R, Gårdfeldt K, Kock H, Lanzillotta E, Mamane Y, Mas F, Melamed E, Osnat Y, Prestbo E, Sommar J, Schmolke S, Spain G, Sprovieri F, Tuncel G (2001) Atmospheric mercury distributions in North Europe and in the Mediterranean Region. Atmos Environ 35:3019–3025 Webster L, Walsham P, Russell M, Hussy I, Neat F, Dalgarno E, et al. (2011) Halogenated persistent organic pollutants in deep water fish from waters to the west of Scotland. Chemosphere 83:839–850. Webster L, Walsham P, Russell M, Neat F, Phillips L, Dalgarno E, et al. (2009) Halogenated persistent organic pollutants in Scottish deep water fish. J Environ Monit 11:406–417. Wheelock CE, Shan G, Ottea J (2005) Overview of carboxylesterases and their role in the metabolism of insecticides. J Pest Sci 30, 75–83. White SL, Rainbow PS (1987) Heavy metal concentrations and size effects in the mesopelagic decapod crustacean Systel/aspis dehilis. Mar Ecol Prog Ser 37, 147-151. Whyte JJ, Jung RE, Schmitt CJ, Tillitt DE (2000) Ethoxyresorufin-O-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit Rev Toxicol 30, 347–570. Willett KL, Ulrich EM, Hites RA (1998) Differential toxicity and environmental fates of hexachlorocyclohexane isomers. Environ Sci Technol 32, 2197-2207. Winston GW, Di Giulio RT (1991) Prooxidant and antioxidant mechanisms in aquatic organisms. Aquat Toxicol 19, 137–161. Wolfe NL, Zepp RG, Paris DF, Baughman GL, Hollis RC (1977) Methoxychlor and DDT degradation in water - rates and products. Environ Sci Technol 11, 1077-1081. Wollast R. (1991) The coastal organic carbon cycle: Fluxes, sources and sinks. In Ocean Margin Processes in Global Change (eds. R. F. C. Mantoura, J. M. Martin and R. Wollast). Dahlem Workshop Reports Wiley, Chichester. pp. 365–381. Xiang C-H, Luo X-J, Chen S-J, Yu M, Mai B-X, Zeng EY (2007) Polybrominated diphenyl ethers in biota and sediments of the Pearl River Estuary, South China. Environ Toxicol Chem 26:616–623. Yang RQ, Jiang GB, Zhou QF, Yuan CG, Shi JB (2005) Occurrence and distribution of organochlorine pesticides (HCH and DDT) in sediments collected from East China Sea. Environ Int 31, 799-804. Yilmaz K, Yilmaz A, Yemenicioglu, Sur M, Salihoglu I, Karabulut Z, Telli Karacoç F, Hatipoglu E, Gaines AF, Phillips D, Hewer A (1998) Polynuclear aromatic hydrocarbons (PAHs) in the Eastern Mediterranean Sea Mar Pollut Bull 36:922-925.

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10. DESCRIPTOR 10: MARINE LITTER

Descriptor 10 refers to marine litter. The assessment framework considers four criteria, two focused on the composition, amount and distribution of litter and micro-litter (D10C1 and D10C2) and two targeting the effects on organisms by ingestion or organisms adversely affected in general terms (D10C3 and D10C4). For further details see COMMISSION DECISION (EU) 2017/848.

The meta-analysis planned for Task 2.2 requires a minimum of 10-15 papers regarding the same measure. Additionally the measure should refer to ecosystem impacts caused by the descriptor-specific pressure. The same data was required from control cases, where the ecosystem remained unaffected. The dataset compiled for descriptor 10 did not fulfil these conditions, preventing a meta-analysis equivalent to the ones performed for other descriptors. In return, a semi-quantitative analysis was performed and summarized in the following four figures targeting the four D10 criteria.

Figure 10.1 summarizes the distribution of datasets referring to marine litter in the deep Mediterranean Sea. The three parts of the figure differentiate between geographical (Box 1) and temporal distribution (Box 2), taking into account the distinction of macro- and micro-litter (Boxes 1 and 2a), the different methodologies applied (Box 2b) and the depth-range investigated (Box 3). Geographical distribution of the references illustrates already a clear north-south gradient at basin scale. A smoother gradient is also observed in the east-west direction. Research on deep-sea litter distribution in the Mediterranean Basin has focused on four main areas: the Northwestern Mediterranean (i.e. Gulf of Lion and Catalan-Balearic Sea), the Central-western Mediterranean (i.e. Sardinian coast), the Central Mediterranean (i.e. Strait of Sicily) and the Eastern Mediterranean (i.e. Southwestern Aegean Sea). The map in the upper box also evidences the disproportion between articles on macro- and micro-litter in the deep Mediterranean Sea. This big difference is quantitatively presented in the histogram of Box 2a. The current high concern and interest in marine litter is showed with the temporal distribution of publications. Box 2b illustrates the cumulative evolution of the published bibliography referring to marine litter, drawing an exponential trend. Most of the publications have to some extent inspected the deep-sea (61.4%), whereas the rest (38.6%) only examined the continental shelf. Within deep-sea publications, trawling appears as the main methodology applied. However, the usage of ROVs is increasing in the recent years. Other techniques such as multi-corers and Van Veen grabs have been mainly used for microplastic and microfiber sampling (e.g. Van Cauweberghe et al., 2013; Woodall et al., 2014).

Regarding the depth-range investigated, only 11% of all deep-sea publications have totally or partially examined the compartment found below 2500 m depth (e.g. Galil et al., 1995; Ramirez-Llodra et al., 2013), whereas the majority of them (55%) had to some extent inspected the 200 – 2500 m depth range. However, some publications categorized in the 200 – 2500 depth range had only few single data points located deeper than 200 m. Therefore, the actual proportion of deep-sea litter information is suspected to be slightly lower than presented. In addition, only a reduced number of publications provide accurate data on depth, coordinates and litter concentrations at each sampling site which in turn hampers direct comparisons between investigated areas and depths.

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An accurate analysis, illustrated in Box 3, reveals that the majority of the publications do not cover a wide depth range. The outcome results in fragmented knowledge, hindering the obtaining of a rigorous picture of marine litter contamination in the whole deep Mediterranean Sea. In addition, sampling is usually performed punctually at specific locations, stretching only to a few weeks or months. This is well represented in Figure 10.2, which shows the sampling date intervals in each publication during the last two decades. Apparently, monitoring of marine litter in the deep-sea has been almost absent in the Mediterranean Sea, thus hampering qualitative and quantitative assessments of litter impacts and distribution through time.

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Figure 10.1. Distribution (1) and trends (2 and 3) of publications that report marine litter in the deep Mediterranean Sea. Box 1 shows the approximate location of geographic regions where research has been 182

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Deliverable 2.2 conducted on macro-litter (yellow bubbles) and micro-litter (red bubbles). References: (1) Alvito et al. (2018), (2) Angiolillo et al. (2015), (4) Bo et al. (2014), (5) Van Cauwenberghe et al. (2013), (6) Cau et al. (2017), (7) Consoli et al. (2018), (8) D'Onghia et al. (2017), (9) Fabri et al. (2014), (10) Fiorentino et al. (2015), (11) Galgani et al. (1995), (12) Galgani et al. (2000), (13) Galgani et al. (1996), (14) Galil et al. (1995), (15) Garcia-Rivera et al. (2017), (16) Gerigny et al. (n/a), (17) Güven et al. (2013), (18) Ioakeimidis et al. (2014), (19) Koutsodendris et al. (2008), (20) Lastras et al. (2016), (21) Mecho et al. (2017), (22) Mifsud et al. (2013), (23) Orejas et al. (2009), (24) Pham et al. (2014), (25) Ramirez-Llodra et al. (2013), (26) Sanchez-Vidal et al. (n.d.), (27) Stefatos et al. (1999), (28) Tubau et al. (2015), (29) Woodall et al. (2014), (30) Alomar et al. (2017), (31) Cannizzaro et al. (1995), (32) Lefkaditou et al. (2013), (33) Pace et al. (2007a), (34) Pace et al. (2007b), (35) Papadopoulou (2015a), (36) Papadopoulou (2015b), (37) Ragonese et al. (1994), (38) Serena et al. (2011), (39) Topcu et al. (2010), (42) IEO (2012a), (43) IEO (2012b), (47) Valchogianni et al. (2017). Box 2a shows the percentage of publications (of a total of 42) that have analyzed macro-litter and micro-litter and by some means inspected different depth compartments. Box 2b shows the cumulative number of publications from 1987 to 2018 and the techniques used to sample litter in the deep-sea. Five reviews have been excluded from the analysis. Box 3 indicates the inspected depth range of each publication (bar) from 1987 to 2018 and differentiates between litter size and method used. ROV = Remote Operated Vehicle. Figure from de Haan et al. (2018). Disclaimer: Not to be used for purposes other than the IDEM project analyses, deliverables and reporting without the written permission of GRC Geociències Marines, Universitat de Barcelona.

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Figure 10.2. Timeline of sampling periods of publications. Horizontal lines indicate the start and ending date of sampling of each publication. Sampling is not continuous during each time period as it can extend from punctual days to several months.

Figure 10.3 illustrates the mean or range of concentrations of litter among publications. The upper Box 1 shows the mean concentrations stated in 25 analyzed publications, differentiating between the two main reporting units used in the bibliography. The plastic amounts reported in these articles range from 5 to 96% of the total marine litter, yet the second largest litter typology found is fishing gear (e.g. longlines and trawl nets). However, the average plastic proportion falls at around 60%. The gradient observed in the number of publications results in the geographical data heterogeneity illustrated in the map of the lower Box 2, which also shows a general lack of single data points. Many of the publications shown in Figure 10.1 have not submitted all data underlying their findings. For instance, geographical coordinates or litter concentrations

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Deliverable 2.2 are not acknowledged or are not crosschecked, thus hindering direct comparison of litter concentrations among the investigated locations.

The highest mean litter concentrations in the deep Mediterranean Sea amount up to 21300 ± 8400 items km- 2 found in the Strait of Sicily (Consoli et al., 2018) and 155 ± 57 kg km-2 in abyssal plains across the Mediterranean Sea (Pham et al., 2014). Interestingly, the first five highest concentrations reported in items km-2 have been obtained using surveys with towed cameras or Remote Operated Vehicles (ROV’s) (Angiolillo et al., 2015; Tubau et al., 2015; D’Onghia et al., 2017; Cau et al., 2017; Consoli et al., 2018) and yet these methods may underestimate actual litter concentrations as they omit very small debris (e.g. micro-litter) and litter items buried by either natural sedimentation or bottom trawling-induced sediment flows (Tubau et al., 2015).

Box 2 also shows the utilization of different units for reporting marine litter abundance and density. Whereas the majority of data is expressed in items per area (km2 or m2), five other measuring units were identified. Certainly, this heterogeneity hampers comparisons and consistent assessments that are required for providing an accurate evaluation of this pressure and its impacts.

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Figure 10.3. Semi-quantitative data referring to marine litter for the deep Mediterranean Sea. Marine litter concentrations, distribution, plastic proportions and units are described. The upper box shows mean litter concentrations reported in items km-2 and kg km-2 in each publication. Mean values of litter concentration were calculated for those publications with accurate data deeper than 200 m and concentrations at each sampling station. Litter concentrations found below 200 m depth were therefore excluded in Box 1 and clipped to the deep-sea area in Box 2. In addition, concentrations of litter outside the Mediterranean geographical area reported in a number of publications were also excluded. Figure from de Haan et al. (2018). Disclaimer: Not to be used for purposes other than the IDEM project analyses, deliverables and reporting without the written permission of GRC Geociències Marines, Universitat de Barcelona.

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At least 49 species and 21 groups have been reported to interact with marine litter, as shown in Figure 10.4. Over 50% of the publications indicated that marine litter had somehow impacted the deep-sea coral Dendrophylia cornigera although this species is not typically the most impacted in each publication. Cold Water Corals (CWC) are usually the most reported impacted species. For instance, Cau et al. (2017) found the yellow gorgonian (E. cavolini) and the red coral (C. rubrum) as the most impacted species in deep Sardinian fishing grounds at depths between 100 and 480 m, with 23% and 18% of them impacted with regard to the total number of species found, respectively. Bo et al. (2014) found the most of the impacted frame shots overlapping the ones with the presence of L. glaberrima at the Marco Bank in western Sicily at depths between 250 and 280 m. Angiolillo et al. (2015) conducted observations in Campania, Sardinia and Sicily and found E. cavolini, C. rubrum and A. subpinnata as the main impacted species at each location, respectively. However, only 27% of the inspected sites encompassed depths larger than 200 m. Likewise, Consoli et al. (2018) also found E. cavolini as the most impacted species, yet with E. verrucosa and P. clavata as second and third most impacted species in the Strait of Sicily, yet they mostly inspected depths above 200 m (20 – 220 m). Many facies with some of these species represent priority habitats according to the SPA/BD Protocol of the Barcelona Convention and are in critical conservation status according to the IUCN assessment (Consoli et al., 2018). However, a general lack of accurate data on litter impacts on deep-sea benthic species exists as the majority of publications usually omit reporting data on observations of individual species and depths. Lastras et al. (2016) also found CWC such as M. oculata, D. cornigera, C. rubrum and M. tenuimana overlapping in the same field of view with litter in 23% of the cases in a ROV study of La Fonera Canyon, Northwestern Mediterranean Sea, and also found a negative correlation between litter and CWC presence. However, Orejas et al. (2009) found a positive correlation between variables in Cap de Creus Canyon, a few kilometres north of La Fonera Canyon, which was suggested to occur because of the larger extension of coral- covered hard substrate regions where longlines got trapped more effectively than on soft sediment or coral- free seabed areas. Eventually, these authors found M. oculata, L. pertusa and D. cornigera highly impacted by longlines.

The presence of litter was in most cases dominated by fishing gear such as longlines, trawl nets and ropes (usually found in higher proportions than 70% of all litter items) (Box 2). Only two publications found plastics as the most abundant litter type (Tubau et al., 2015; Mecho et al., 2017). Covering, abrasion or entrapment by fishing gear are common impacts on local fauna and habitats in some locations (up to 54%), whereas the rest may be found lying on the seabed or hanging from rocks (up to 55%) (Angiolillo et al., 2015; Cau et al., 2017; Consoli et al., 2018). Marine litter is also often used as habitat. Epibionts, such as by Hydroids, Anthozoans, encrusting sponges and colonial tunicates have been recurrently observed to use plastic objects and fishing gear as hard substrate, sometimes reaching up to 65% of observed litter colonized (Consoli et al., 2018). Most of the research has been conducted mainly on hard-bottom habitats such as rocky or coralligenous outcrops, hardground crusts and boulders and buried rubble (Box 3) (e.g. Orejas et al., 2009; Bo et al., 2014; D’Onghia et al., 2017). Litter on soft-bottom habitats has been scarcely investigated. Consoli et al. (2018) recently found a significant correlation of litter with increasing habitat complexity, with rocky habitats showing larger quantities of litter (70%) rather than soft bottom habitats (30%), also with 42% on habitats with high complexity. They suggested that large boulders and outcrops usually representing complex reliefs increase the chances for fishing gear to get entangled there. 187

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Figure 10.4. Summary of species, groups and habitats affected by marine litter (Boxes 1, 2 and 3) and their impact and interactions to benthic species (Box 4). Box 1 summarizes all reported species among 11 published articles that have directly or indirectly interacted with marine litter (e.g. entangled, abraded or present in the same field of view marine litter). Similarly, classification was also performed using groups. Box 2 indicates the dominant litter typology in each publication along with average or maximum proportions of each. A qualitative analysis performed in Box 3 shows the main habitat inspected and often impacted. Box 4 summarizes the main types of impacts and interactions of marine litter with benthic species. Figure from de Haan et al. (2018). Disclaimer: Not to be used for purposes other than the IDEM project analyses, deliverables and reporting without the written permission of GRC Geociències Marines, Universitat de Barcelona.

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References: Alomar C, Guijarro B, Deudero S. Macro- and Microplastic in Seafloor Habitats Around Mallorca. In: Baztan J, Jorgensen B, Pahl S, Thompson RC, Vanderlinden J-PBT-F and I of M in ME, editors. Elsevier; 2017. pp. 100–101. Alvito A, Bellodi A, Cau A, Moccia D, Mulas A, Palmas F, et al. Amount and distribution of benthic marine litter along Sardinian fishing grounds (CW Mediterranean Sea). Waste Manag. Elsevier Ltd; 2018;75: 131–140. Angiolillo M, Lorenzo B di, Farcomeni A, Bo M, Bavestrello G, Santangelo G, et al. Distribution and assessment of marine debris in the deep Tyrrhenian Sea (NW Mediterranean Sea, Italy). Mar Pollut Bull. Elsevier Ltd; 2015; 92: 149–159. Bo M, Bava S, Canese S, Angiolillo M, Cattaneo-Vietti R, Bavestrello G. Fishing impact on deep Mediterranean rocky habitats as revealed by ROV investigation. Biol Conserv. Elsevier Ltd; 2014;171: 167–176. Cannizaro L, Garofalo G, Giusto G, Rizzo P, Levi D. Qualitative and quantitative estimate of solid waste in the channel of Sicily. Proc Second Int Conf Mediterr Coast Environ MED-COAST. 1995; 95. Cau A, Alvito A, Moccia D, Canese S, Pusceddu A, Rita C, et al. Submarine canyons along the upper Sardinian slope (Central Western Mediterranean) as repositories for derelict fishing gears. Mar Pollut Bull. Elsevier; 2017; 123: 357–364. Consoli P, Andaloro F, Altobelli C, Battaglia P, Campagnuolo S, Canese S, et al. Marine litter in an EBSA (Ecologically or Biologically Significant Area) of the central Mediterranean Sea: Abundance, composition, impact on benthic species and basis for monitoring entanglement. Environ Pollut. 2018; 236: 405–415. de Haan W, Sanchez-Vidal A, Canals M. Marine litter in the deep Mediterranean Sea; Progress in Oceanography (in prep.), 2018. D’Onghia G, Calculli C, Capezzuto F, Carlucci R, Carluccio A, Grehan A, et al. Anthropogenic impact in the Santa Maria di Leuca cold-water coral province (Mediterranean Sea): Observations and conservation straits. Deep Res Part II Top Stud Oceanogr. Elsevier Ltd; 2017; 145: 87–101. Fabri MC, Pedel L, Beuck L, Galgani F, Hebbeln D, Freiwald A. Megafauna of vulnerable marine ecosystems in French mediterranean submarine canyons: Spatial distribution and anthropogenic impacts. Deep Res Part II Top Stud Oceanogr. Elsevier; 2014; 104: 184–207. Fiorentino F, Gancitano V, Giusto GB, Massi D, Sinacori G, Titone A, et al. Marine litter on trawlable bottoms of the Strait of Sicily. Biol Mar Medit. 2015; 22: 225–228. Galgani F, Jaunet S, Campillo A, Guenegen X, His E. Distribution and abundance of debris on the continental shelf of the north-western Mediterranean Sea. Mar Pollut Bull. 1995; 30: 713–717. Galgani F, Leaute JP, Moguedet P, Souplet A, Verin Y, Carpentier A, et al. Litter on the sea floor along European coasts. Mar Pollut Bull. 2000; 40: 516–527. Galgani F, Souplet A, Cadiou Y. Accumulation of debris on the deep sea floor off the French Mediterranean coast. Mar Ecol Ser. 1996; 142: 225–234. Galil BS, Golik A, Turkay M. Litter At the Bottom of the Sea: a Sea-Bed Survey in the Eastern Mediterranean. Mar Pollut Bull. 1995; 30: 22–24. Garcia-Rivera S, Lizaso JLS, Millán JMB. Composition, spatial distribution and sources of macro-marine litter on the Gulf of Alicante seafloor (Spanish Mediterranean). Mar Pollut Bull. Elsevier; 2017; 121: 249–259. Gerigny, M. Brun, M.C Fabri, C. Tomasino, A. Jadaud and F. Galgani (2019). Seafloor litter in the continental shelf and canyons in French water of Mediterranean Sea: quantities, distribution and typology. In preparation Gobierno de España y Instituto Español de Oceanografia. 2012a. Parte IV. Descriptores del buen estado ambiental. Descriptor 10: basuras marinas. Evaluación inicial y buen estado ambiental. Demarcación Balear-Levantine. Madrid, 47p. Accessed March 7, 2018. Available from: http://cdr.eionet.europa.eu/es/eu/msfd8910/msfd4text/envuhupiw/IV_D10_Levantino- Balear.pdf/manage_document 190

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Gobierno de España y Instituto Español de Oceanografia. 2012b. Parte IV. Descriptores del buen estado ambiental. Descriptor 10: basuras marinas. Evaluación inicial y buen estado ambiental. Demarcación del Estrecho y Alboran. Madrid, 36p. Accessed March 12, 2018. Available from: http://cdr.eionet.europa.eu/es/eu/msfd8910/msfd4text/envuhuorw/IV_D10_Estrecho_y_Alboran.pdf/ manage_document Guven O, Gulyavuz H, Mehmet Cengiz D. Benthic Debris Accumulation in Bathyal Grounds in the Antalya Bay, Eastern Mediterranean. Turkish J Fish Aquat Sci. 2013; 13: 881–896. Ioakeimidis C, Zeri C, Kaberi H, Galatchi M, Antoniadis K, Streftaris N, et al. A comparative study of marine litter on the seafloor of coastal areas in the Eastern Mediterranean and Black Seas. Mar Pollut Bull. Elsevier Ltd; 2014; 89: 296–304. Katsanevakis S, Katsarou A. Influences on the distribution of marine debris on the seafloor of shallow coastal areas in Greece (eastern Mediterranean). Water Air Soil Pollut. 2004; 159: 325–337. Koutsodendris A, Papatheodorou G, Kougiourouki O, Georgiadis M. Benthic marine litter in four Gulfs in Greece, Eastern Mediterranean; abundance, composition and source identification. Estuar Coast Shelf Sci. 2008; 77: 501–512. Lastras G, Canals M, Ballesteros E, Gili JM, Sanchez-Vidal A. Cold-water corals and anthropogenic impacts in la Fonera submarine canyon head, Northwestern Mediterranean Sea. PLoS One. 2016; 11: 1–36. Lefkaditou E, Karkani M, Kavadas S, Aikaterini A, Christidis G, Mytilineou C. Litter composition on the shelf and upper slope of the Argosaronikos region and the eastern Ionian Sea, as evidenced by MEDITS surveys 1995‐2008. Mediterr Mar Sci. 2013;ICES CM 20. Mecho A, Aguzzi J, De Mol B, Lastras G, Ramirez-Llodra E, Bahamon N, et al. Visual faunistic exploration of geomorphological human-impacted deep-sea areas of the north-western Mediterranean Sea. J Mar Biol Assoc United Kingdom. 2017; 1–12. Mifsud R, Dimech M, Schembri PJ. Marine litter from circalittoral and deeper bottoms off the Maltese islands (Central Mediterranean). 2013; 2: 95–118. Orejas C, Gori A, Lo Iacono C, Puig P, Gili JM, Dale MRT. Cold-water corals in the Cap de Creus canyon, northwestern Mediterranean: Spatial distribution, density and anthropogenic impact. Mar Ecol Prog Ser. 2009; 397: 37–51. Pace R, Dimech M. Distribution and density of discarded limestone slabs used in the traditional Maltese lampuki fishery. CIESM Congr. 2007; 38: 568. Pace R, Dimech M. Litter as a source of habitat islands on deep water muddy bottoms. CIESM Congr. 2007; 38: 567 Papadopoulou N. Seabed marine litter, comparison of 4 Aegean trawling grounds. 2015. Papadopoulou N. Trawled up marine litter , first observations from Heraklion Bay. 2015. Pham CK, Ramirez-Llodra E, Alt CHS, Amaro T, Bergmann M, Canals M, et al. Marine litter distribution and density in European seas, from the shelves to deep basins. PLoS One. 2014; 9. Ragonese S, Rizzo P, Giusto GB. Rifiuti antropici e pesca dello scampo Nephrops norvegicus (L., 1758) (Crustacea, Nephropidae) nello Stretto di Sicilia. Biol Mar Medit. 1994; 1: 309–310. Ramirez-Llodra E, De Mol B, Company JB, Coll M, Sardà F. Effects of natural and anthropogenic processes in the distribution of marine litter in the deep Mediterranean Sea. Prog Oceanogr. Elsevier Ltd; 2013; 118: 273–287. Sanchez-Vidal A, Thompson RC, Canals M, and de Haan WP, Submitted. The Imprint of Microplastics from Textiles in Southern European Deep Seas. Serena F, Abella AJ, Baino RT, Cecchi E, Ria M, Silvestri R, et al. Anthropogenic waste in the marine ecosystem. Biol Mar Mediterr. 2011; 18: 161–164. Stefatos A, Charalampakis M, Papatheodorou G, Ferentinos G. Marine Debris on the Seafloor of the Mediterranean Sea: Examples from Two Enclosed Gulfs in Western Greece. 1999;36: 389–393.

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Topcu EN, Tonay AM, Öztürk B. A Preliminary Study on Marine Litter in the Aegean Sea. Rapp Comm Int Mer Médit. 2010; 804. Tubau X, Canals M, Lastras G, Rayo X, Rivera J, Amblas D. Marine litter on the floor of deep submarine canyons of the Northwestern Mediterranean Sea: The role of hydrodynamic processes. Prog Oceanogr. Elsevier Ltd; 2015; 134: 379–403. Van Cauwenberghe L, Vanreusel A, Mees J, Janssen CR. Microplastic pollution in deep-sea sediments. Environ Pollut. Elsevier Ltd; 2013; 182: 495–499. Vlachogianni Th, Anastasopoulou A, Fortibuoni T, Ronchi F, Zeri Ch, 2017. Marine Litter Assessment in the Adriatic and Ionian Seas. IPA-Adriatic DeFishGear Project, MIO-ECSDE, HCMR and ISPRA. pp. 168 (ISBN: 978-960-6793-25-7). Woodall LC, Sanchez-Vidal A, Paterson GLJ, Coppock R, Sleight V, Calafat A, et al. The deep sea is a major sink for microplastic debris. R Soc open Sci. 2014; 1.

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11. DESCRIPTOR 11: INTRODUCTION OF ENERGY

Task 2.1 and the related deliverable (report 2.1) evidenced the existence of poor datasets regarding the introduction of energy in the deep Mediterranean Sea (5 published papers in total and any open access repositories at the moment). Spatial but also temporal gaps were pointed out while reviewing the available datasets. Only two studies were conducted in the western Mediterranean Sea and none in the Aegean- Levantine basin. In addition long-term data series are still lacking. Thus, a proper meta-analysis has not been carried out for this Descriptor.

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12. CONCLUSIONS AND FUTURE WORK

The present report on the first assessment of deep Mediterranean status described the results of analyses (meta-analysis and semi-quantitative analyses) performed on Descriptors 1, 3, 4, 6, 7, 8, 9 and 10 across the different Mediterranean sub-basins.

During the last decades, exploration in the deep Mediterranean Sea has increased and led to the description of a wide variety of habitats which are responsible of the high spatial heterogeneity at different spatial scale. Despite this, most of the data published in scientific papers cover few habitat types, i.e., open slopes, canyons and cold-water coral ecosystems. Similarly, there are several publications including data on fishes and various macro-invertebrates on both soft and hard bottoms, but studies related to plankton, bacteria and foraminifera are instead limited. Meta-analysis on Descriptor 1 has been carried out on data regarding meiofauna, megafauna and hard-substrate communities, spanning from western to eastern Mediterranean Sea. Our results confirmed the importance of topographic complexity contributing to the high biodiversity of meio- macro- and megafaunal assemblages which are in turn also structured by food availability. We pointed out that some of the current sampling instruments such as gears can damage the vulnerable deep-sea ecosystems, thus less invasive methodologies (i.e., video surveys) should be used in future studies. Video surveys can be used to cover indicators on sediment conditions but also presence and abundance of species, their coverage and biomass. Recently developed technologies (i.e., molecular tools) should help to determine not only the biodiversity and taxonomic composition but also population genetic structure (indicator 1.3.2), feeding the needs of the Marine Strategy Framework Directive.

The analysis performed on Descriptor 3 evidenced that a clear overfishing for most of the stocks assessed is ongoing in the Western and Central Mediterranean Sea and in the Ionian Sea. Many knowledge gaps and challenges need to be addressed to improve the accuracy of future Descriptor 3 assessments of the deep Mediterranean Sea. Particularly, more stock assessments are urgently needed for deep-water Mediterranean species and a list of exploited deep-sea stocks should be compiled at sub-region level, and assessments covering a significant proportion of landings (e.g. >70%) should be carried out. Still, the timeseries coverage of SSB trend data should be improved, and suitable proxies to qualitatively estimate biomass against reference point proxies available in the Mediterranean Sea should be investigated (i.e. where SSB estimates based on stock assessments are not available, biomass indices based on MEDITS data can be used instead). Finally, suitable thresholds for the population and age size distribution indicator should be identified and specifically tested for the deep-sea.

The meta-analysis carried out on D4 evidenced the organization of Mediterranean deep-sea food webs in essentially four trophic levels, encompassing different trophic guilds each. This analysis fully complied with 13 15 the first D4-MSFD criterion, i.e. D4C1 - diversity of trophic guilds, based on datasets where  C and  N values of both potential food sources for deep-sea food webs and consumers were analysed. Stable isotope analysis seemed to be a valid tool to depict deep-sea food webs as allows to discriminate also among different trophic guilds within the same trophic level, based also on the 13C values (i.e. POM feeders vs. surface deposit feeders). However, as previously observed while the food web of the western basin can be described in details, also concerning the potential food sources, some gaps, related to lower trophic levels (generally analysed as a whole and not at species level, i.e. suprabenthos, meso- and macrozooplankton) exist for the central and the eastern sub-basins and further data would be needed.

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The semi-quantitative analysis performed for descriptor 6 provided an overview of the existing research and associated data regarding sea-floor integrity in the deep Mediterranean Sea. Bottom trawling and waste disposal are the two mostly investigated pressures in the deep Mediterranean Sea. Bottom trawling leads to a decrease of seabed integrity resulting in physical disturbance, alteration of seafloor properties, lower abundance, biomass, and diversity indices in the trawled area. Conversely, the impact of waste disposal in the deep Mediterranean is partly described in the literature, and most studies are out-dated (more than 20 years ago). The review highlighted the actual fragmented knowledge, concerning MSFD criteria, anthropogenic pressures and geographical scales. The majority of the investigations targeted physical loss and disturbance of the seafloor caused by bottom trawling activities. However, bottom trawling bibliography is quite heterogeneous in the reported physical effects and lacks the evaluation of ecosystem impacts in most of the articles compiled. The studies taking into account deep-sea habitats focused mainly on specific biological groups, hindering the evaluation of systemic effects. Finally, not all potential pressures, either existing or future, are identified and revised in the literature.

Descriptor 7 analysis was focused in reviewing the most prominent events, processes and water masses properties identified and described with data and literature collected during the last 50 years. Knowledge and data gaps were identified in water depths below 2000 m, in southern Mediterranean countries and in long-term datasets of the deep-sea. Despite knowledge fragmentation, the data-sets analyzed highlighted changes in thermohaline properties such as increased temperature and salinity levels which can be associated to intensive events of intermediate and deep-water formation. We reported variations of dissolved oxygen and pH of seawater, as the main pressures impacting the deep Mediterranean Sea. Our results pointed out that all Mediterranean waters are acidified, particularly those in the Western Basin, where deep-water renewal is faster. Most of the articles did not identify specific anthropogenic actions as pressures for benthic habitats. Climate change importance and influence was generally addressed, together with the significance of dense shelf water cascading processes, the EMT and the WMT. Finally, alteration of biogeochemical properties and investigation of habitat and ecosystem impacts were poorly revised because of the low information available in the bibliography. In addition most of the studies cover the Western Mediterranean and are snapshot, not considering the temporal evolution of the habitat.

Coastal discharges of anthropogenic compounds from rivers, urban and industrial effluents, dumped mining waste, and others, are very significant for the Mediterranean Sea due to geography and the intensive use of littoral land. These inputs add to the general atmospheric pollutant fallout in marine systems. The intense ship activity constitutes an additional pollution source. Conversely, the anti-estuarine Mediterranean water circulation exports pollutants to the Atlantic and in some cases to the atmosphere, e.g. mercury, providing detoxification mechanisms. Assessment of the toxicity impact resulting from these counter-acting processes requires much more data than presently available. Different budgets depending on the physical-chemical characteristics of the contaminants may be observed. The existing information is, in addition, unevenly distributed, being restricted to the North-Western area for many pollutants and processes. The lack of information is particularly relevant when trying to elucidate the temporal evolution of the pollution impacts.

Dense shelf water cascading (DSWC) through submarine canyons has been identified as one specific North- Western Mediterranean mechanism for pollutant transfers to deep environments. These processes remove compounds retained in the continental margin over several tens of kilometers offshore in a few days. The highest marine deposition fluxes ever described for PCBs, DDTs, HCB, PeCB and PAHs have been observed in association to these events, which are also relevant for the transport of radionuclides. The significance of

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DSWC in other Mediterranean areas should be investigated given the high accumulation of pollutants in the continental margins.The DSWC processes transfer pollutants associated to particles. Obviously, food web vertical sedimentation also transports these compounds to deep environments. The pollutants predominantly found in deep waters are those associated to particles, e.g. pyrolytic PAHs, highly chlorinated or brominated compounds, which are those also present in organisms living in these environments, e.g. fish and crustaceans.

Fish from deep Mediterranean waters have organic pollutant concentrations that are comparable to those living in marine surface environments. There is a need for assessment of the sources and processes leading to the high mercury accumulation in some fish of common human consumption as the concentrations are often higher than the maximum levels set forth by the European Union. Fish and species of smaller size have lower mercury levels. The Mediterranean Sea can be considered a clean environment for what concerns radioactivity distribution. Anthropogenic radionuclides enter the Mediterranean Sea mainly through the atmosphere and rivers. There is a general westward 137Cs transport from the Aegean Sea, where the Chernobyl contribution was higher.

The analysis performed on marine litter data revealed that fishing gear such as longlines, trawl nets and ropes dominate litter found in the deep Med, followed by plastics. Marine organisms can be differently affected by marine litter through a wide variety of interactions. The most impacted area seems to be the strait of Sicily, followed by abyssal plains across the Med. Most of the publications cover the 200-2500 m depth range, and very few data have been reported below the 2500 m depth. Visual surveys with ROVs are useful to obtain data on marine litter even if they can lead to an underestimation of actual litter concentrations. Also for marine litter, we found both spatial and temporal gaps, due to local studies and data covering few weeks or months. In addition, most of the surveys have been conducted on hard-bottom habitats, leaving soft-bottoms ones less studied.

Concluding, the heterogeneity of available data on habitats and taxonomic groups for the different geographic areas hampers a consistent assessment of deep Mediterranean status for the four sub-basins. Also, the different unit of measure used in published papers for some indicators (e.g., abundance, coverage by the species, concentration of marine litter) do not allow comparisons that are required for providing an accurate evaluation of the different pressures and impacts on habitat and ecosystems. The cumulative effects of multiple activities and stressors are rarely investigated and the scant knowledge on spatial and temporal trends hinders the establishment of thresholds that in turn complicates the implementation of measures.

In order to implement Marine Strategy Framework Directive in the deep Mediterranean Sea, we need to feed the open access databases, define common procedures for data sharing and rigorous and standard protocols for data acquisition. The development of innovative tools, the implementation of new technologies, together with the optimization of the monitoring strategies (spatial resolution, long-term monitoring positions, data sharing…) are crucial to provide a holistic assessment of deep Mediterranean status.

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