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Modelling the impact of deep-water crustacean trawl fishery in the marine ecosystem off Portuguese Southwestern and South Coasts: I) the trophic web and trophic flows

The concentration of the population in coastal regions, in addition to the direct human use, is leading to an accelerated process of change and deterioration of the marine ecosystems. Human activities such as fishing together with environmental drivers (e.g. climate change) are triggering major threats to marine biodiversity, and impact directly the services they provide. In the South and Southwest coasts of Portugal, the deep-water crustacean trawl fishery is not exemption. This fishery is recognized to have large effects on a number of species while generating high rates of unwanted catches. However, taking into account an ecosystem-based perspective, the fishing impacts along the food web accounting for biological interactions between and among species caught remains poorly understood. These impacts are particularly troubling and are a cause of concern given the cascading effects that might arise. Facing the main policies and legislative instruments for the restoration and conservation of the marine environment, times are calling for implementing ecosystem-based approaches to fisheries management. To this end, we use a food web modelling (Ecopath with Ecosim) approach to assess the fishing impacts of this particular fishery on the marine ecosystem of southern and southwestern Portugal. In particular, we describe the food web structure and functioning, identify the main keystone species and/or groups, quantify the major trophic and energy flows, and ultimately assess the impact of fishing on the target species but also on the ecosystem by means of ecological and ecosystem-based indicators. Finally, we examine limitations and weaknesses of the model for potential improvements and future research directions.


Introduction
Bottom trawling is recognized to be one of the most damaging and less environment friendly fisheries in the world (Jackson et al., 2001). This practise poses major risks on the marine ecosystems including for example alteration of habitat structure, damage to bottom communities, loss of biodiversity, structure and productivity of benthic communities (Gray, 1997;Jennings and Kaiser, 1998;Pauly et al., 2002). These changes are ultimately translated into population declines (Lewison et al., 2004) and food web alteration that can affect ecosystem functions (Pauly et al., 1998). The removal of large, slow growth and long-lived predatory fish because of current fishing practices is affecting the reproductive characteristics of stocks even the structure of the whole ecosystem (Myers and Worm, 2003;Worm et al., 2005). In addition, fishing activities generate high rates of unwanted species which represents a wasteful use of living resources (Alverson et al., 1994;Kelleher 2005;Zeller and Pauly, 2005;Ulhmann et al., 2013) with important ecological consequences (Tsagarakis et al., 2014;Heath et al., 2014) and changes at the ecosystem level which may in turn affect fisheries sustainability (Zhou et al., 2010).
In Portuguese waters (ICES subdivision IXa), the deep-sea crustacean trawl fishery is not an exception. This multi-species fishery constitutes a very important part of the fishing fleet particularly in the Algarve region (Costa et al., 2008). The main fishing grounds are located at the shelf edge and continental slope off the southern and southwestern coasts of Portugal at depths of 200-700 m. The main target species rose shrimp (Parapenaeus longirostris), Norway lobster (Nephrops norvegicus) and red shrimp (Aristeus antennatus) inhabit soft habitats composed of a mixture of mud and silt (Monteiro et al., 2001;Campos et al., 2007). Other deep-water crustacean species, namely the purple shrimp (Aristaeomorpha foliacea) and the scarlet shrimp (Aristaeopsis edwardsiana), are only occasionally targeted or incidentally caught at depths greater than 500 m (Silva et al., 2015). This multi-species fishery operates by using codend mesh sizes of 55 mm and 70 mm and targets primarily on the commercial species according to availability, market demand and price (Silva et al., 2015).
Historical landings, economic trends and opportunities for the Portuguese fisheries sector have been fully documented in the area (e.g. Hill and Coelho, 2001;Leitão et al., 2014;Leitão, 2015;Leitão and Baptista, 2017). Trends in landings related to environmental variables have been analysed (Erzini, 2005;Teixeira et al., 2014;Gamito et al., 2015Gamito et al., , 2016) and changes in the trophic level of such landings have been reported (Baeta et al., 2009). Although these studies evidence potential effects in the marine ecosystem structure and functioning, the impact of deep-water crustacean trawl fishery along the food web accounting for biological interactions between and among species caught remains poorly understood. These impacts are particularly troubling and are a cause of concern given the cascading effects that might arise.
The year 2020 is a landmark for the main policies and legislative instruments for the restoration and conservation of the marine environment in Europe (e.g. Common Fisheries Policy (CFP), Marine Strategy Framework Directive (MSFD), EU Biodiversity Strategy, Water Framework Directive (WFD), Marine Protected Areas (MPA), Convention on Biological Diversity-Aichi targets). Hence, it is timely to make operational an ecosystem-based approach to fisheries management (EBFM) (Garcia et al., 2003) that is the most suitable tool to integrate sustainable exploitation activities consistent with the existence of ecological interactions, environmental constraints, protection of natural and cultural heritage, and restoration of marine ecosystems (Link and Browman, 2014). Although the theory behind EBFM is well developed, its full implementation still lags behind (Berkes, 2012).
Within the context of EBFM, the Ecopath with Ecosim (EwE) approach is by far the most commonly used ecosystem modelling platform to create ecological models (Polovina, 1984;Christensen and Pauly, 1992;Christensen andWalters, 2004, 2011). This approach integrates human activities (i.e. fisheries) within an ecosystem context and evaluating their impacts on marine food webs, including trophic interactions between and among species and environmental factors (Coll and Libralato, 2012). Outcomes of this model provide insights into ecosystem structure and functioning, adding to our understanding the control of energy flows transferred through the food web while allowing scientists and managers to modify components of the ecosystem to explore past and future impacts of fishing and environmental disturbances as well as to explore optimal fishing policies. This package also gathers a significant amount of information at an ecosystem level based on network analysis and information theory (Ulanowicz, 1986;Christensen and Pauly, 1993;Heymans and Baird, 2000).
The primary goal of our study is to assess for the first time the impact of deep-water crustacean trawl fishery on the upper trophic levels and explore the effects cascading down the food web to small fish, using a food web modelling approach (Ecopath with Ecosim). In particular, we characterize the structure and resilience of the southern Portuguese marine ecosystem by describing the food web structure and functioning, identifying the main keystone species and/or groups of the ecosystem, quantifying the main trophic flows and energy transferences and, assessing the impact of fishing on the target species but also on the ecosystem by means of ecological and ecosystem-based indicators. Finally, we examine limitations and weaknesses of the model for potential improvements and future research directions.

Study site
The study area, where the deep-water crustacean fishery operates, covers 4,000 km 2 off Portuguese mainland waters (including the EEZ) with depths from 200 to 700 m (Fig. 1). The geographical area is encompassed in ICES Subdivision IXa along the Iberian waters of the Northeast Atlantic Ocean. This region is located in a transition zone between temperate and tropical ecosystems (Briggs, 1974), characterized by high biodiversity associated with the different types of bottoms and relatively low abundance of commercially exploited marine species (Sousa et al., 2005). The zoogeographic importance of this area has been recognized, representing the transition between north-eastern Atlantic warm-temperate and cold-temperate regions, which makes the Portuguese coast an area of great sensitivity to the detection of climate change . Recent studies have shown that ecological responses to global warming are already visible along the Portuguese coast (Vinagre et al., 2011;Teixeira et al., 2014). The continental shelf is very narrow, varying between 7 and 28 km wide, followed by the relatively steep continental slope which is incised by several submarine canyons and trenches (Lopes and Cunha, 2010). The morphology reflects a geological evolution controlled by tectonic and sedimentary dynamics in this region (Hernández-Molina et al., 2016). The only major output source, the Guadiana River, is located on the border between Portugal and Spain exhibiting strong annual and interannual fluctuations (Morais et al., 2007;Cunha et al., 2001;Sonderblohm et al., 2014).
The physical oceanography is dominated mostly by the Gulf of Cadiz system, where the Mediterranean Outflow, a high salinity water mass flowing northward along the margin, plays an important role in controlling circulation dynamics (Relvas, 2002;Peliz et al., 2009;Sánchez-Leal et al., 2017). The area, under the influence of the Iberian system of the North Atlantic upwelling region, is subjected to seasonal hydrodynamic processes (Cunha et al., 2001). The most productive waters are located close to the coast, while significant areas stretch over great depths with relatively levels of productivity (Borges et al., 2001). In this region, the south-west winds are dominant especially during winter, both in frequency and in speed, while the north-west and south-east winds are moderate (Dias, 1988;Sonderblohm et al., 2014).

The food web modelling approach
The food web model to represent the Southern and Southwestern Portuguese coasts (onwards SSWPT) was constructed using the updated Ecopath with Ecosim (EwE) software version 6.6 (Christensen et al., 2008). In our study, we used the static module (i.e. Ecopath) to get a snapshot of the functional groups, according to similar ecological traits, and their interactions in a specific period, namely the year 2000. This year was selected because of the best data were available. The model is parameterized based on two master equations describing the production (Eq. 1) and consumption (Eq. 2) of each functional group: Production = catches + predation mortality + net migration + biomass accumulation + other mortality (Eq. 1) Consumption = production + respiration + unassimilated food (Eq. 2) For each functional group, three of the four basic parameters [biomass (B), production (P/B) rates, consumption (Q/B) rates, and ecotrophic efficiency (EE)] are required and the fourth is left to be estimated by the model. Diet composition and catches are also needed. A detailed explanation of algorithms and equations are described in Christensen and Walters (2004) and Christensen et al. (2008).
Biomass estimates (t km−2 year−1) were estimated from bottom trawl surveys conducted in the study area using stratified and swept-area methods (PTGFS-WIBTS-Q4 targeting mainly fishes and PT-CTS (UWTV (FU 28-29) targeting crustaceans, ICES (2017b)). Both surveys are conducted seasonally on board the R/V Noruega since 1995 (in autumn) and 1997 (in summer) respectively (Azevedo et al., 2014). Estimates of the abundance of vulnerable groups (i.e. seabirds and marine mammals) were taken from the 'MarPro Project' results. Quantitative estimates of phytoplankton, by means of chlorophyll-a values obtained from local sampling procedures were used to calculate the biomass (Table 1S). Detritus biomass was calculated empirically following the equation of Pauly et al. (1993) linking the production of detritus (D) with the average annual primary production (PP) and the average depth of the euphotic zone (E). Finally, biomass estimates for turtles, gelatinous plankton, benthic carnivorous feeders, macrobenthos and macrozooplankton were left to be estimated by the model using realistic EE values (Table 1S).
Under mass-balance conditions, biomass accumulation and other export terms were assumed equal to zero.
According to Allen (1971), the production per unit of biomass (P/B, year -1 ) was assumed as the total mortality (Z), which is the sum of fishing mortality (F, as the ratio between catches (Y) and biomass (B)), and natural mortality (M). Natural mortality for finfish species in absence of catchat-age data was estimated from the empirical equation proposed by Pauly (1980): linking natural mortality (M), the curvature parameter of the von Bertalanffy growth function (K), the asymptotic length (L∞) calculated from Pauly (1984) where L∞ ≈ Lmax / 0.95 and the mean water temperature (T in°C). Maximum length values (Lmax) were obtained from the bottom trawl surveys. Growth parameters were taken from local studies or in absence from literature ( Table 1S). The mean water temperature in the study area in 2000 was 13.5°C (Moreno and Santos, unpublished data). For most invertebrate species, secondary production ratios were taken from other models (Table 1S). Finally, P/Q values were assumed for seabirds, dolphins and benthic carnivorous feeders to let EwE estimate a P/B value (Christensen et al., 2008).
Consumption per unit of biomass (Q/B, year -1 ) for those groups with available information was calculated following the empirical equation of Pauly et al. (1990): where W∞ is the asymptotic body weight (g), T' is an expression for the mean water temperature expressed using T' = 1000 / Kelvin (Kelvin = •C + 273.15), and two variables expressing food type: Pf (1 for predators and zooplankton feeders and 0 for all others) and Hd (1 for herbivores and 0 for carnivores). W∞ was estimated by using length-weight relationships published in the study area or in absence from literature (Table 1S). Information related to the type of food was mostly taken from empirical data or from Fishbase (Froese and Pauly, 2018). For seabirds, consumption was estimated by using the empirical equation proposed by Nilsson and Nilsson (1976): where DR is the daily ration (g) and W is the mean body weight (g) for each species. DR was converted to Q/B considering the biomass of each species and the time spent in the area. In the case of marine mammals, consumption was calculated from the empirical equation of Innes et al. (1987): where R is the daily ration (kg) and W is mean body weight (kg) for each species. R was converted to Q/B considering the biomass of each species (Table 1S). Due to the limited information available for crabs, benthic invertebrate carnivores and polychaetes, consumption was calculated as a function of body size following the equation of Finally, for those groups with no information provided in the area, the values were assumed from other models involving similar species (Table 1S).
The diet composition data were mostly compiled from local studies or from similar areas (Table   1S). Migratory species were taken into account by modeling a proportion of their diet compositions as import to the ecosystem following Torres et al. (2013).
In the SSWPT model we included catches (t km−2 year−1) from two separate fleets: crustacean trawl (OTB_crust) and fish trawl (OTB_fish). The official landings by species and fleet with reference to the year 2000 were provided by the DGRM (Portuguese Direcção-Geral de Recursos Marinhos, Segurança e Serviços Marinhos). Data of discards taken from local studies (Costa et al., 2008) were also included to give a more accurate total catch data. Unfortunately, estimates of illegal, unregulated, or unreported (IUU) landings were lack in the study zone.

Parameterization and mass-balancing
Prior balancing the SSWPT model, we checked that outputs of the model were within expected limits of a model ecologically and thermodynamically balanced (Christensen et al., 2008;Heymans et al., 2016) (Table 2S SM). Unused consumption (U/Q) for those invertebrate groups mainly composed of filters, detritivores, and suspensivores feeders (i.e. FGs 29-34) was assumed to be 0.40 (Christensen et al., 2008). For the rest of the groups, the default value of 0.2 was maintained.
The mass-balancing was performed manually following a top-down strategy. The adjustments were performed according to their uncertainty degree, using the "Pedigree Routine". Initial results showed that the EE > 1 for most fish and invertebrate groups and we assumed that these groups may have had greater uncertainty in the biomass data. We therefore readjusted appropriate inputs parameters to achieve mass-balance following the same strategy applied in other ecosystems (Mackinson and Daskalov, 2008;Torres et al., 2013;Corrales et al., 2015).
We guide our balancing procedure following the PREBAL analysis (Link, 2010) to ensure that the model parameters followed some basic principles of ecosystem ecology. This analysis highlighted that some P/B and Q/B values had to be adjusted since they were too low or too high based on their trophic levels. After testing the model parameters with PREBAL, diets were slightly readjusted where needed, as in other Ecopath models (Coll et al., 2006;Torres et al., 2013). Cannibalism was also decreased in benthopelagic cephalopods, mesopelagic fishes, and crabs. Unfortunately, biomass estimates from the bottom surveys were not sufficient for the mass-balance. In the absence of catchability coefficients to correct biomass estimates, biomass for most groups of fish and invertebrates were left to be estimated by the model using a reasonable value of EE (EE values close but lower than 1) (Heymans et al., 2016).

Pedigree index and quality of the model
The uncertainty associated to the input values in the model was quantified by using the pedigree routine (Christensen et al., 2008). This index varies from 0 for low-quality models (i.e. input data estimated or taken from other models) to 1 for high-quality models (i.e. well-sampled and highprecision local input data). The pedigree index is averaged over all parameters and groups in the model to provide an index of the quality of the input data of the model.

Ecosystem properties and trophic flows
A set of ecological indicators were used to assess the general ecosystem status and its stage of development and maturity sensu Odum (1971) and Christensen (1995). The overall ecosystem size in terms of flows is represented by the Total System Throughput (TST), calculated as the sum of the all the flows of consumption, exports, respiration, import and flows to detritus (Ulanowicz, 1986;Christensen and Pauly, 1993). In addition, the following indicators describing the system maturity are estimated: Total Primary Production/Total System Respiration (TPP/TR), Total Primary Production/Total Biomass (TPP/TB), and Total Biomass/Total System Throughputs (TB/TST) ratios. The Transfer Efficiency (TE) is defined as the fraction of the total flows at each trophic level (TL) that is either exported or transferred to other TLs through consumption. The mean TE is calculated as a geometric mean from the TE in trophic levels (TLs) II-IV (Christensen et al., 2008). Ulanowicz (1986) were calculated. Ascendancy (A) provides information of the degree of development and maturity of an ecosystem as a measure of the average mutual information in a system, scaled by system throughput (Ulanowicz and Norden, 1990). The upper limit of A is called the 'development capacity' (C) and the difference between C and A is called 'system overhead' (O). The latter provides limits on how much the ascendancy can increase and reflect the system's 'strength in reserve' from which it can draw to meet unexpected perturbations (Ulanowicz, 1986). These indices are used as a measure of system stability, information, resilience and ecosystem maturity (Odum, 1969;Ulanowicz and Norden, 1990;Heymans et al., 2007).

Network analysis indices suggested by
The cycling indices include the Finn's Cycling Index (FCI) representing the percentage of flows recycled in the food web (Finn, 1976), the Predatory Cycling Index (PCI) representing the percentage of recycling after the removal of detritus (Christensen et al., 2008), and the Finn's Mean Path Length (MPL), which quantifies the mean number of nodes that energy inflow into the ecosystem passes through before exiting the network. In addition, the System Omnivory Index (SOI) and the Connectance Index (CI) correlated with ecosystem maturity and complexity of the food web (Christensen, 1995) were calculated. SOI is based on the average omnivory index (OI) which is calculated as the variance of the TL of a consumer's prey groups indicating predatory specialization . The CI is the ratio of the number of existing trophic links with respect to the number of possible links (Christensen et al., 2008). The trophic structure was gathered into a Lindeman spine, an analysis of discrete TLs sensu Lindeman (1942) and proposed by Ulanowicz (1995). The system was aggregated into a linear food chain where import (on TL I only), consumption by predators, export, flow to the detritus, respiration, and throughput were calculated for each TL. The detritus box was separated from primary producers to show the amount of energy that is flowing through it. These flows were also represented by means of a flow diagram showing the trophic interactions between all groups within the ecosystem.

Trophic cascading and keystone groups
The Mixed Trophic Impact (MTI) analysis estimates the relative trophic impact (both direct and indirect) that a hypothetical increase in the biomass of a functional group would produce on the others within the ecosystem, including fishing activities (Ulanowicz and Puccia,1990). A positive or negative impact would mean an increase or decrease in the quantity of the impacted group.
Further details or equations are well described in Christensen et al. (2008).
The index of Keystoneness (KS) evaluates the potential ecological roles of each functional group as keystones in the system. This index is a function of a group's trophic impact on other groups in the ecosystem and its biomass. Three methods proposed by Power et al. (1996), Librarato et al. (2006) and Valls et al. (2015) were compared. All methods use the relative overall effect calculated from the MTI against the KS and the contribution of each functional group to the total biomass of the food web.

Exploitation status of the fishery
Fishing impacts on the SSWPT ecosystem were assessed by analysing the mean trophic level of the catch (mTLc), the exploitation rates (F/Z), the relative consumption of total production representing the proportion of total production that is consumed within the system by all the functional groups, fishing mortalities (F), the gross efficiency of the fishery (GE, catch/net primary production), and the percentage of primary production required (PPR) to evaluate the sustainability of fisheries (Pauly and Christensen, 1995).

The food web model analyzed by functional group
The 37 functional groups included in the SSWPT model were combined into four TLs ranged from 1.0 for primary producers and detritus groups to dolphins (TL=4.47) and anglerfish (TL=4.33), representing the top predators in the ecosystem ( Table 1). The remaining fish groups TLs ranged from 3.08 (horse mackerels) to 4.21 (piscivores). Invertebrate groups were estimated to have a TL between 2.00 (benthic filter feeders) and 3.93 for benthic cephalopods. Seabirds which rely heavily on fishery discards probably had their trophic level underestimated (TL=3.04), as Ecopath assigns by default a TL = 1 to the fishery discards.
The highest flows to detritus corresponded to those groups ranked at the base of the food web.
The highest FD values regarding fish groups were provided by small-sized and blue whiting in addition to the benthic groups. R/A ratio ranged from 0.33 to 0.99 t km−2 year−1, with the highest R/A values corresponding to top predators. OI showed that most groups were feeding on multiple TLs ranging from 0 to 1.445. The lower OI values were observed for benthopelagic fishes, horse mackerels and mesopelagic fishes, suggesting high specialization. These groups exert a high predation on zooplanktonic groups (macrozooplankton, micro-y mesozooplankton). By contrast, dolphins, hake, seabirds and minke whales presented the highest OI value with a large number of prey items in its diet. These results suggest more complexity on the upper part of the food web. These groups were followed by crabs, benthopelagic cephalopods, benthic sharks and Norway lobster, presenting a relatively wide trophic spectrum.
Regarding mortalities, most of the functional groups located in the upper part of the food web presented low predation mortalities (M2) as expected. In addition, the target species also showed low M2 values in particular red shrimp, even lower than the followers groups such as piscivores, and deep-sea fishes. Other natural mortality excluding predation (M0) was relatively low in most of the functional groups. However, phytoplankton, gelatinous plankton, and seabirds presented high M0 values. The total biomass supported by the ecosystem (excluding detritus and discard) was calculated as 48.14 t km−2 in 2000. The largest part of the total ecosystem biomass was mainly represented by crustaceans (25.1%), fishes (23.9%), benthic invertebrates (16.9%) and cephalopods (3.2%).
Primary producers occupying the lower trophic levels contributed to 10.5 % of the total biomass (5 t km−2). These results highlight the great importance of the demersal communities and benthic invertebrate producers in the area.
The quantification of the trophic flows among the functional groups are shown in Figure 2. It was observed that many groups from the pelagic compartment were consumed by groups in the benthic compartment. For example, the macrozooplankton (FG33) and benthopelagic and mesopelagic fishes (FG13 and 14) or macrobenthos (FG32) with mackerels (FG10) and horse mackerels (FG11). In addition, demersal groups such as piscivores (FG6), anglerfish (FG7) and hake (FG9) feed on groups from the pelagic compartment. The detritus was shown to be an important compartment in the SSWPT food web supplying most of the biomass and production in the demersal habitat (Fig. 2).

Trophic flows
The Lindeman spine represented in Figure 3 revealed the importance of the TLs II and III in terms of biomass in the ecosystem as well as the later and TL IV in terms of exports and catches.
Most flows were generated at the base of the food web. Primary production and detritus generated over 77% of the TST, underlying the importance of these groups moving the energy to the upper TL groups. TL II also contributed significantly to the major flows (TST = 16.8%). was higher than the general value of 10% estimated by Pauly and Christensen (1995). This means that on average 22% of the production of one TL became a production of its upper TL.
The TE was mainly derived from primary producers (23%) and from detritus (21%). Finally, the exportation flows (i.e. catches) were mainly concentrated in TL III (0.786) followed by TL IV (0.425) and together with the high TE (22%) from TL III underline the strong exploitation by the fisheries on this TL (Fig. 3).

Cascading effects and keystone species
The Matrix Trophic Index analysis (Fig. 4)  In the case of the index proposed by Power et al. (1996), seabirds, dolphins, minke whales, anglerfish, megrim and red shrimp were also highlighted as keystone. Following Valls et al. (2015), the groups identified were benthopelagic cephalopods, rose shrimp, minke whales and benthic carnivorous feeders. Accounting for the relative total impact, some groups with lower TLs were also identified as structuring groups (i.e. groups that have an important role in the ecosystem because have high biomass and high trophic impact) rather that unlikely keystone groups such as phytoplankton, shrimps and macrobenthos. These groups might indicate a possible bottom-up forcing.

Ecosystem status
Ecological indices calculated showed that the total system throughput was represented by consumption (33%), flows to detritus (32%), flows of all exports (23%) and respiration (12%) ( Table 2). TPP/TR ratio (> 1), showed that the energy produced is more than four times higher than respired within the ecosystem (Christensen, 1995).

Fishing impacts
In the SSWPT model, the total catch in 2000 was 1.38 T/km 2 ( Table 2). The mTLc was 3.49, matching the TLs of the dominated groups in landings, which corresponded to rose shrimp, followed by red shrimp, hake, horse mackerels and Norway lobster. By contract, the most discarded group was blue whiting, followed by mesopelagic fishes, benthopelagic fishes, mackerels, benthic sharks and large demersal fishes.
The primary production required to sustain the fisheries (PPR%) when considering the detritus and the primary producers was estimated at 12.61 % in the SSWPT ecosystem in 2000. The major PPR fractions were to sustain mainly the catches of blue whiting, rose shrimp, benthic cephalopods, and mesopelagic fishes. The gross efficiency of the fishery showed a value of 0.001.
The MTI analysis showed the impact of the two trawling fisheries on the ecosystem. Overall, the two fleets presented negative impacts on several groups with the largest negative impacts found on its main target species. In particular, the crustacean trawl fleet had strong negative impacts on the target red shrimp rose shrimp, and Norway lobster, but also on top predators such as benthic sharks, anglerfish, piscivores and rays and skates related to the high rates of discards of these groups by this fishery. Also, it was identified a strong negative impact on seabirds showing indirect impacts on this group. In general, the fish trawl fleet showed the highest negative impact on hake via direct mortality and indirectly on marine mammals possibly due to competition for same resources.