Malcolm R. Clark*,
Franziska Althaus
Thomas A. Schlacher
Alan Williams
David A. Bowden
Ashley A. Rowden
1National Institute of Water and Atmospheric Research (NIWA),
2Marine Laboratories,
3University of the Sunshine Coast,
↵*Corresponding author: e-mail:
Handling editor: Michel Kaiser
Received December 19, 2014.
Revision received June 19, 2015.
Accepted June 25, 2015.
Deep-sea fisheries operate globally throughout the world's oceans, chiefly targeting stocks on the upper and mid-continental slope and offshore seamounts. Major commercial fisheries occur, or have occurred, for species such as orange roughy, oreos, cardinalfish, grenadiers and alfonsino. Few deep fisheries have, however, been sustainable, with most deep-sea stocks having undergone rapid and substantial declines. Fishing in the deep sea not only harvests target species but can also cause unintended environmental harm, mostly from operating heavy bottom trawls and, to a lesser extent, bottom longlines. Bottom trawling over hard seabed (common on seamounts) routinely removes most of the benthic fauna, resulting in declines in faunal biodiversity, cover and abundance. Functionally, these impacts translate into loss of biogenic habitat from potentially large areas. Recent studies on longline fisheries show that their impact is much less than from trawl gear, but can still be significant. Benthic taxa, especially the dominant mega-faunal components of deep-sea systems such as corals and sponges, can be highly vulnerable to fishing impacts. Some taxa have natural resilience due to their size, shape, and structure, and some can survive in natural refuges inaccessible to trawls. However, many deep-sea invertebrates are exceptionally long-lived and grow extremely slowly: these biological attributes mean that the recovery capacity of the benthos is highly limited and prolonged, predicted to take decades to centuries after fishing has ceased. The low tolerance and protracted recovery of many deep-sea benthic communities has implications for managing environmental performance of deep-sea fisheries, including that (i) expectations for recovery and restoration of impacted areas may be unrealistic in acceptable time frames, (ii) the high vulnerability of deep-sea fauna makes spatial management—that includes strong and consistent conservation closures—an important priority, and (iii) biodiversity conservation should be > balanced with options for open areas that support sustainable fisheries.
deep sea
fisheries
fisheries management
fishing impacts
recovery
sensitivity
Fishing operations that contact the seabed can have unwanted, and often severe, environmental effects. Impacts most commonly documented include the scraping and ploughing of the seabed, resuspension of sediments smothering the fauna, killing of non-target benthic animals, and the dumping of processing wastes (Jones, 1992; Dayton et al., 1995; Jennings and Kaiser, 1998; Hall, 1999; Clark and Koslow, 2007). There is also growing evidence that environmental changes attributable to fisheries practices can have negative impacts on habitat quality, biodiversity, and the structural and functional integrity of ecological assemblages (Hutchings, 1990; Auster et al., 1996; Collie et al., 1997; Auster and Langton, 1999; Koslow et al., 2001). The majority of studies reporting on fishing impacts come from coastal areas or the continental shelf (Collie et al., 2000; Kaiser et al., 2002; Kaiser et al. 2006), compared with more limited work in the deep ocean.
Conventionally, the deep sea is regarded to be >200 m depth and beyond the shelf break (Thistle, 2003). In this zone, a number of finfish species characterized by low productivity and high vulnerability are the target of commercial fishing (FAO, 2009), including species that can be abundant on offshore seamounts and ridge systems, such as alfonsino (Beryx splendens), orange roughy (Hoplostethus atlanticus), pelagic armourhead (Pseudopentaceros wheeleri), macrourid rattails (roundnose grenadier Coryphaenoides rupestris), and oreos (several species of the family Oreosomatidae) (Koslow et al., 2000; Clark et al., 2007).
Deep-sea fisheries have become economically important in recent decades. As many shelf stocks became overexploited, the search for commercial fisheries moved into deeper offshore waters (Koslow et al., 2000; Clark et al., 2007; Pitcher et al., 2010). Expansion of fishing into upper (200–700 m) and mid (700–1500 m) continental slope environments extended harvests to new target species and it broadened the depth range over which previously fished species are caught (Morato et al., 2006; Watson and Morato, 2013). Many of these fisheries were not sustainable (Clark, 2009; Pitcher et al., 2010; Norse et al., 2012). They were also recognized as causing substantial ecological impacts in several areas of operation, and probably globally (Koslow et al., 2001; Hall-Spencer et al., 2002; Waller et al., 2007; Althaus et al., 2009; Clark and Rowden, 2009). Halpern et al. (2007) identified ‘demersal, destructive fishing’ (e.g. demersal trawl) as the most consistently high-scoring threat to oceanic deep-sea ecosystems.
However, are the effects of fishing in the deep sea any different from in shallower waters? In this review, we summarize the state of knowledge of fisheries impacts on benthic fauna and communities in the deep sea, bringing together published studies and grey literature reports, as well as drawing inferences from appropriate shallow-water studies. We focus primarily on hard-substrate invertebrate communities, although also consider soft sediment in less detail. The review has three main sections which move logically through a summary and review of fishing impacts, to an assessment of the sensitivity of deep-sea fauna to fishing, then an examination of their recovery potential. We conclude by considering the implications of benthic impacts for deep-sea fisheries management.
There is no universally accepted or applicable definition of what constitutes a “deep-sea” species for commercial fisheries, but generally include species being fished mainly deeper than 200–500 m (Clark, 2001; FAO, 2005). Species lists of deep-sea fisheries typically include species with lower productivity (based on characteristics such as slower growth rates, higher longevity, and lower fecundity) than shallow shelf species, and those which often occur on offshore topographic features such as ridges and seamounts (FAO, 2004; Sissenwine and Mace, 2007; Clark, 2009; European Parliament, 2014). It is not our intention here to give a detailed list of deep-sea fish and fisheries (e.g. Sissenwine and Mace, 2007; European Parliament, 2014) but it is useful to illustrate the types of fisheries (Table
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A key consideration when estimating the ecological impacts of deep-sea fisheries is the geographic extent over which they operate (i.e. the likely area of impact). There are, however, few published accounts to make these estimates. Benn et al. (2010) estimated that in 2005 ∼30 000 km2 at depths >200 m in the Hatton Bank and Rockall areas of the Northeast Atlantic Ocean were trawled with bottom gear. Within the New Zealand EEZ, the total trawled area deeper than 200 m amounts to 180 000 km2 (Black et al., 2013) out of a fishable area of 1.4 million km2-hence deep-sea fishery impacts can be widespread. In addition, deep-sea fishing can be concentrated on particular habitat types, especially oceanic ridges and seamounts. For example, in New Zealand the seabed trawled between 800 and 1200 m accounts for 12% of the total swept-area deeper than 200 m (Black et al., 2013) but 80% of known seamount features in this depth range have been fished; in some years, these seamount fisheries comprise up to 50% of total orange roughy trawling effort and catch (Clark and O'Driscoll, 2003; O'Driscoll and Clark, 2005).
Deep-sea fisheries use several types of gear that can damage seabed habitats and their fauna: bottom otter trawls, bottom longlines, deep midwater trawls, sink/anchor gillnets, pots and traps, and tanglenets (Clark and Koslow, 2007). None is exclusive to deep-sea fisheries, and hence the general types of impacts would be expected, a priori, to be similar to shelf fisheries. The most common techniques used in many deep-sea fisheries are bottom trawling and bottom longlines. The main difference to shallow-water trawl rigs is the size and weight of the groundgear: trawl doors can weigh up to 2000 kg, and nets designed for fishing on rough seabed at mid-ocean ridges or seamounts are frequently fitted with many bobbins or rock-hopper discs of 60 cm diameter or greater, weighing several tonnes.
Hence, although the amount of deep-sea fishing is much less than on the shelf, effects are, nevertheless, likely to be appreciable due to larger and heavier trawl gears used, and the very high intensity of fishing in localized areas on certain features, such as seamounts and ridges, where deep-sea fish aggregate (Clark et al., 2007).
Ecological impacts on seabed communities attributable to fishing in the deep sea are—in general terms—of the same type as documented in shallow systems (Gage et al., 2005; Clark and Koslow, 2007). Ploughing and scraping of the seabed and resuspension of sediment are likely to occur also in deep-sea habitats (and see other papers in this volume). In the following sections, we summarize some of the key impacts of demersal fishing in the deep sea on the physical environment and benthic fauna. General lessons from shelf work are sometimes included, but whenever possible we focus on results from deep-sea studies.
Few deep-sea studies have explicitly examined direct changes to substrate and topography of the seabed caused by fishing gear. Clark and Koslow (2007) summarized general impacts, which they noted depend on the gear type, its weight and rigging, as well as the nature of the substrate and the frequency of disturbance. In shelf habitats heavy trawl doors and sleds gouge, scrape and plough the seabed and homogenize unconsolidated sediments (Handley et al., 2014; Palanques et al., 2014). In the deep sea, trawling can alter the physical properties of surface sediments, either by thorough mixing of soft sediments, or by causing the erosion of upper layers, exposing denser, older sediments in the trawl path (Martin et al., 2014). Trawls can also uproot semi-buried glacial drop stones or boulders (Gage et al., 2005; Hall-Spencer et al., 2007). Line gears alter the seabed to a lesser extent due to their much narrower footprint; lines can, however, drag on the seabed stirring up sediments (Ewing and Kilpatrick, 2014). Trawl gear mobilizes sediments creating plumes of particles in their wake (O'Neill et al., 2013) which are typically 2–4 m high (Palanques et al., 2001; Durrieu de Madron et al., 2005), and 120–150 m in width depending on the size of trawl gear (Bradshaw et al., 2012). In low-current deep-sea environments, these can disperse very slowly over large distances (Bluhm, 2001; Rolinski et al., 2001), and potentially affect areas well beyond, and deeper than the area of the fishery (Black and Parry, 1999; Martin et al., 2014). O'Neill and Summerbell (2011) estimated that a typical Scottish demersal trawl would suspend up to 3 kg m−2 of sediment between the trawl doors, and trawling-induced sediment gravity flows can remove large volumes of sediment from the shelf (Puig et al., 2012).
Compared with inshore fisheries, it is likely that deep-sea trawling and lining have very similar effects on the seabed. The main difference would be the heavier groundgear often used in deep-sea fisheries on rough-bottom habitat such as seamounts, which can increase the depth of gouging in areas of soft sediment. The physical effects can also remain longer than in shallow shelf waters. Whereas trawl door gouges and tracks can often disappear from shallower sandy substrate after just a few months (Lokkeborg and Fossa, 2011) or 1–2 years in mud substrate (Ball et al., 2000), in the deep sea the physical scars can remain much longer. Clear marks from orange roughy trawling on soft-sediment areas were visible 5 years after fishing ceased on several seamount features off New Zealand (Clark et al., 2010a).
Direct interactions of fishing gear with epibenthic animals that results in physical damage can be classified into three basic types (Ewing and Kilpatrick, 2014): (i) blunt impacts—the motion of a broad object through the benthos (e.g. groundrope, trawl doors, mesh, codend, or chafe mat); (ii) line shear—the motion of a narrow object across or through the benthos (e.g. trawl sweeps and lower bridles, longlines when dragging across the seabed); (iii) hooking—direct interaction of hooks with the benthos (e.g. snagging animals). Blunt interactions generally result in the dislodgement or crushing of individuals, particularly larger, erect forms that are anchored to the seabed such as corals, sponges, and crinoids (Koslow et al., 2001; Hall-Spencer et al., 2002; Denisenko, 2007; Althaus et al., 2009; Clark and Rowden, 2009; Rooper et al., 2011; Munoz et al., 2012). These organisms can also be sheared off, hooked, or tangled in longlines (Orejas et al., 2009; Munoz et al., 2011; Bo et al., 2014; Sampaio et al., 2012).
Table Hall-Spencer et al. (2002) and Williams et al. (2009); Fossa et al. (2002) extrapolated such observations into an estimate of spatial damage of the total Lophelia reef in their Norwegian study area (Table
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The design of studies on seamounts and slope environments off Australia and New Zealand (Koslow et al., 2001; Cryer et al., 2002; Althaus et al., 2009; Clark and Rowden, 2009) is the ‘compare-and contrast’ type, examining biological differences between areas that are known to have different fishing histories. Similarly, Atkinson et al. (2011) described differences in epifaunal abundance and diversity off West Africa with different levels of trawl intensity, and Munoz et al. (2012) also observed differences in bycatch levels between areas of high and low fishing effort in the North Atlantic. Denisenko (2007) reported results of surveys in the Barents Sea where there was a widespread reduction in biomass and distribution of 11 of 13 epibenthic species exposed to the demersal trawl fishery. Other bycatch studies have shown, in association with trawling on seamounts, a decline in benthic invertebrate abundance, biomass, or richness over a period of a fishery (Anderson and Clark, 2003; Niklitschek et al., 2010). Most of these studies have indicated strong differences in the biodiversity of benthic fauna, especially coral-associated communities (Table Clark et al., 2010a) (Figure
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The trawl gear varies between the fisheries covered in Table Clark and Koslow, 2007). The more mixed trawl fisheries in the North Atlantic and upper continental slope off Australia and South Africa (“trawl” in Table
Experimental studies of fishing impacts in the deep sea are rare, but Freese et al. (1999) found in the Gulf of Alaska density differences of 30% for finger sponges and similar reductions in anemone populations between experimental trawl tracks and adjacent un-trawled areas. Off Norway, marked changes in the height of Paragorgia and Lophelia pertusa colonies have been attributed to trawling, whereby coral colonies were on average about half as high in fished areas relative to intact colonies (Buhl-Mortensen et al., 2013).
Static gears, such as longlines and traps are considered to have lower impacts than mobile gear types (Pham et al., 2014). However, in certain conditions, for example during retrieval, static gear may move laterally across the seabed, resulting in impacts to the habitat and biota (Sampaio et al., 2012; Ewing and Kilpatrick, 2014). Longline impacts on sessile fauna such as sponges and corals have been observed (Fossa et al., 2002; Mortensen et al., 2008), where the animals have been broken by longline weights or by the mainline cutting through them while moving laterally during fishing or hauling (Welsford and Kilpatrick, 2008).
The amount of trawling effort required to cause a certain level of impact has not been well researched in the deep sea. However, several studies using fish and prawn trawls have occurred at shelf depths that indicate what impacts might be expected with similar taxa in deeper waters. Table Pitcher et al., 2000; Burridge et al., 2003) showed that 10–20% of gorgonian corals and large sponges were removed during each trawl pass. This work used a prawn trawl which swept a width of ∼18 m, and a ground chain was used for close bottom contact. This gear may be more efficient than a deep-sea trawl, as the latter will have heavier groundgear and potentially less-continuous contact with the seabed (and perhaps less direct contact than the other gears used in Table Clark et al., 2010a). The extra weight of such trawl gear over that often used at shelf depths means that the impacts of crushing and gouging can be more severe in deep-sea fisheries.
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Indirect impacts on epifauna can arise from the sediment plumes caused by the trawlnet or longline contact with the seabed. Small amounts of sediment settling on the bottom, of the order of only several mm, can smother small cold-water corals such as Lophelia, and prevent expansion or recovery of the colony (Rogers, 1999). Impacts on coral feeding and metabolic function are uncertain, although stony corals can actively shed sediment, both in shallow-water species (Riegl, 1995) and on the slope. Larsson and Purser (2011) observed that Lophelia pertusa in an aquarium setting was able to survive repeated light smothering by sediment, but polyps died when wholly covered by particles. Hence some taxa can potentially cope with a degree of sediment increase caused by trawling. However, deep-sea sponge respiration has been reported as largely shutting down when subjected to heavy sedimentation loads (Tjensvoll et al., 2013). A number of oil and gas related studies have examined the effects of drilling muds on benthic fauna, but most have focused on aspects of ecotoxicity with chemical contaminants contained in the discharges, which confounds the effects of sedimentation.
Direct effects of fishing disturbance on infauna are relatively well studied in shallow waters (Jennings and Kaiser, 1998; Collie et al., 2000; Kaiser et al., 2000Sanchez et al. (2000) found no short-term impact on muddy bottoms; and similarly O'Neill et al. (2013) describe no change in infauna after a single passage of a scallop dredge, despite the animals being swept up in the sediment plume. However, Handley et al. (2014) reported that soft-sediment shelf habitats impacted by fishing over longer terms were devoid of large bodied species, as they are more likely to be crushed or removed. Given the more stable nature of deep-sea environments, there would be an expectation that more species could be affected by disturbance, with declines in abundance and species richness (Grassle and Sanders, 1973). Several studies at abyssal plain depths have noted reductions in small infauna (e.g. nematodes, polychaetes) and larger mobile burrowing forms (e.g. urchins, asteroids) following experimental disturbance that ploughed the seabed (Ahnert and Schriever, 2001; Miljutin et al., 2011). More relevant to fishing depths, Leduc and Pilditch (2013) conducted a small-scale experimental disturbance in the laboratory with sediment cores from 345 m depth, and after 9 days found nematode species richness remained similar, but there were changes in the vertical distribution of nematode species, and community structure. Mangano et al. (2013) found significantly lower numbers of individuals and species and a shift in the community composition (i.e. more worms, bivalves, and scavengers) in areas of higher trawl frequency on the shelf, but not on the slope (although the latter result may have been confounded by illegal fishing activity).
The vertical penetration of various parts of trawl gear into the seabed can be significant, at least 30 cm for doors, and several centimetres for the groundgear (Buhl-Mortensen et al., 2013) (and see other papers in this issue). This can affect infaunal composition and distribution (Leduc et al., 2012) but also potentially epifauna that are adapted to a certain substrate mix of bedrock, boulders, pebbles, or gravel. Mixing of the upper sediment layers can also alter the chemical composition, especially in the more stable waters of the deep sea (Rumohr, 1998). Chemical release from the sediment can also be enhanced, especially if enriched nodules or sediments containing elements such as phosphorites are broken up or disturbed by trawling (ICES, 1992). In a detailed study of the sedimentary environment of a canyon in the western Mediterranean Sea, affected by intensive and regular trawling at depths of 200–800 m, Pusceddu et al. (2014) found substantial decreases in organic matter content of the sediments, slower organic carbon turnover, and reduced meiofauna diversity and abundance. They concluded that the majority of the daily organic carbon input could be removed by trawling, causing a general degradation of sedimentary habitats, and infaunal depauperation.
Offal discards from fisheries may result in localized organic enrichment of the sediment, and provide a trophic subsidy to deep-sea consumers. Discarded catch and processing waste that is not taken at the surface by seabirds or scavenged in the water column, can result in localized, and relatively large, food falls (Connolly and Kelly, 1996). This can lead to an influx of scavengers and predators (Britton and Morton, 1994; Clark and Koslow, 2007; Williams et al., 2009; Dannheim et al., 2014); but where deep-sea communities are already well adapted to a scavenging role for natural flux of dead animals, it is unclear how significant the supply of fishing discards might be (Gage et al., 2005). A single study on discards from the hoki (Macruronus novaezelandaie) fishery off New Zealand suggested that there might be oxygen depletion (Livingston and Rutherford, 1988), although this has not been confirmed. Dannheim et al. (2014) observed that the trophic level of soft bottom communities in fished areas was higher than in unfished areas, and Shephard et al. (2014) noted increased scavenging by some fish in trawled areas affecting size structure of communities.
Smothering impacts on infauna by a sediment cloud are likely to be less severe than for epifauna. There are few deep-sea studies, but Trannum et al. (2010) experimented with sediment obtained from 40 m depth off Norway. Sediment thicknesses between 3 and 24 mm were applied, and no changes were found in number of taxa, abundance, biomass, or diversity of macrofauna. However, in situ studies at depth are required to improve our understanding of sedimentation processes and biological impacts.
The direct impacts on fauna through dislodgement or damage of individuals are the most obvious effects caused by fishing gear, but the range of biological changes extends well beyond these physical impacts and can significantly alter the community composition and foodweb architecture in the ecosystems subjected to fishing disturbance. On the shelf high levels of trawling results in changes to overall community composition through substantial habitat alterations, removal of non-target species, and through attraction of scavengers and predators to trawled areas (Tillin et al., 2006; Hinz et al., 2009).
Removal of structural engineers and homogenizing of sediments alters the benthic habitat in ways that may not be suitable for settlement of recruits from the original community, leading to long-term or potentially permanent changes in community composition and structure. Such shifts are well documented in shelf communities as a result of fishing (Kaiser et al., 2000). In deeper water, trawling with heavy bottom gear has removed habitat forming stony coral from seamounts and offshore reef areas (Koslow et al., 2001; Fossa et al., 2002; Hall-Spencer et al., 2002; Althaus et al., 2009; Clark and Rowden, 2009; Buhl-Mortensen and Buhl-Mortensen, 2014) which is implicated in changes in community structure (Koslow et al., 2001; Althaus et al., 2009; Clark and Rowden, 2009) whereby the species composition and relative abundances have changed with removal of the coral habitat. Such shifts have yet to, or may never, recover to a pre-impact state (Williams et al., 2010). Other deep-sea environments dominated by vulnerable structural species include bryozoans that consolidate soft substrates to provide habitats and attachment points for other sessile fauna, forming species rich communities on the continental slope (Schlacher et al., 2010). Williams et al. (2009) observed that trawling has impacted such bryozoan habitat off Australia, where there is a high overlap with bottom trawling, and bryozoan turf was observed to be vulnerable to damage by a relatively light research sled.
In soft-sediment slope environments without significant habitat structure, epifauna community structure has also been demonstrated to be markedly different between lightly and heavily trawled areas off southwest Africa (Atkinson et al., 2011). In the same study, infaunal community structure was reported as very different between two of the four pairs of lightly and heavily trawled sites (Atkinson et al., 2011). An extensive study of the effects of trawling on deep-sea infaunal communities was conducted using data from research trawls from a 2400 km2 area of slope off New Zealand (Cryer et al., 2002). This study demonstrated that 11–40% of variation in infaunal community structure was attributable to fishing (over many years for both finfish and scampi), and inferred that trawling probably changes benthic community structure over broad spatial scales on the continental slope as well as in coastal systems (Cryer et al., 2002).
Changes in community structure include alterations in the proportions of ecological or trophic “types” of fauna. Marked differences in both epi- and infaunal communities away from larger slow-growing species, such as echinoderms, towards smaller fast-growing species such as worms and scavengers, have been observed in deep-sea environments subject to regular trawling (Atkinson et al., 2011; Mangano et al., 2013). Denisenko (2007) (cited in Lyubin et al. (2011) observed changes in Barents Sea communities down to 300 m due to demersal trawling, and a shift from abundance of large and long-lived suspension-feeders to smaller deposit feeders.
Changes may also occur through an altered balance in the composition of fish species associated with benthic habitat. Cold-water coral reef structures often have high diversity or abundance of fish species (Costello et al., 2005; Auster, 2007) and may provide nursery ground, spawning, and protective habitat (Husebo et al., 2002; D'Onghia et al., 2010; Clark and Dunn, 2012). Hence, there could be potential flow-on effects into deep-sea fish communities, and subsequently predator–prey interactions with benthic invertebrates.
The impacts of fishing on benthic communities are determined by interactions between the physical, behavioural, and life history characteristics of individual taxa and the nature of the disturbance itself (Thrush and Dayton, 2002; Gray et al., 2006; Hewitt et al., 2011). Sensitivity can be thought of as the balance between intolerance—impairment or death of individuals, populations, or communities in response to disturbance, and recoverability—the re-colonization or re-growth following disturbance (Hiscock and Tyler-Walters, 2006). Alternative but equivalent terms, such as ecological resistance and resilience, have also been applied to these concepts (Bax and Williams, 2001; Halpern et al., 2007; Williams et al., 2010); in these cases, vulnerability accounts for the exposure of the ecological unit to fishing disturbance. Sensitivity is typically applied at the level of individual taxa by reference to a suite of relevant biological attributes or traits (MacDonald et al., 1996; Bremner, 2008; Tyler-Walters et al., 2009; de Juan and Demestre, 2012), but the concept can also be applied at the level of populations, habitats, biotopes, or ecosystem functions (e.g. Hiddink et al., 2007; Tyler-Walters et al., 2009; Bolam et al., 2014; Lambert et al., 2014). Estimating the relative sensitivity of fauna is important because it provides a basis for identifying the potential vulnerability of ecological units, and for assessing the risk stemming from impacts. But for management uptake of this information, for example to prioritize areas for protection or designing monitoring programmes, metrics of sensitivity are needed.
Sensitivity is necessarily defined in relation to the characteristics of a specific fishing method because the intensity, spatial scale, and frequency of disturbance can vary greatly between methods (e.g. bottom trawling vs. longlining) (MacDonald et al., 1996). The principal characteristics of fishing disturbance relevant to direct impacts on benthic organisms are the gear's spatial extent, speed, degree of penetration into the substratum, and the frequency of the disturbance (Thrush and Dayton, 2002; Hewitt et al., 2011). Spatial extent and speed influence whether or not mobile organisms will be able to avoid the disturbance. Spatial scale will also influence which taxa are able to benefit from exploiting food resources, such as carrion and exposed infauna, caused by the disturbance. The frequency of disturbance will influence recoverability (re-growth, migration, or re-colonisation), with higher frequencies having greater impact on taxa that are less tolerant and/or with lower recoverability. In soft sediments, the degree of penetration of the fishing gear will strongly influence the range of taxa affected, with fewer taxa being tolerant of deeper penetration (Thrush et al. 1998).
Attributes useful for defining the sensitivities of individual benthic taxa to fishing disturbances can be separated broadly into two categories: (i) physical and behavioural attributes including feeding mode (e.g. deposit-feeding vs. suspension-feeding), living position (e.g. infaunal vs. epifaunal), growth form (e.g. encrusting vs. erect), and mobility (e.g. sessile vs. mobile) and (ii) life-history attributes such as growth rate, capacity to regenerate, reproductive mode, and dispersal potential (Bremner et al., 2006; de Juan et al., 2009). For deep-sea studies, an important practical issue is the availability of reliable data to inform these categories. Thus, while knowledge of life history characteristics of shallow-water benthic taxa may be derived through observation and experimentation (e.g. MacDonald et al., 1996), the difficulty of applying an experimental framework in deep-sea studies leads to relatively sparse ecological knowledge for benthic fauna, particularly their reproductive and growth characteristics. Estimates of relative sensitivity or vulnerability incorporating life history attributes have been generally undertaken by expert consensus where such data are sparse (CCAMLR, 2009; Halpern et al., 2007; Williams et al., 2010). However, these assessments for deep-sea taxa rely heavily on the more obvious physical and behavioural characteristics, particularly living position, growth form, mobility, and fragility, which are simpler in concept and more robust in practice because the data underlying them are readily accessible. Because estimates of recoverability depend on often unknown life-history attributes, it is more conservative to assess impacts and risk of disturbance to benthic fauna and habitats in the deep-sea context by considering intolerance separately (Thrush et al., 2009; Tyler-Walters et al., 2009; Lambert et al., 2014).
In a simple biological traits scheme developed for categorizing sensitivities of deep-sea benthic fauna Hewitt et al. (2011) assigned taxa to one or more ‘traits’ in each of five physical attributes. This was based on their expected responses to an along-surface disturbance, such as that resulting from bottom trawling (Table Chevenet et al., 1994). By combining scores across all attributes, individual taxa can be ranked in sensitivity categories (Table Hewitt et al. (2011), taxa were ranked based on the degree of mortality resulting from exposure to disturbance. The highest ranked were sedentary, erect, and fragile forms, with sensitivity decreasing either as the degree of fragility decreased or there was greater mobility or a living position deeper in sediment. Thus, taxa living mainly subsurface with high burrowing capacity were considered tolerant, whereas mobile scavengers with potential to benefit from moving into a disturbed area with increased food availability were considered “favoured”. Highly sensitive taxa in the deep sea include those with erect and fragile forms such as arborescent octocorals and thicket-forming stony corals (Figure
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In management applications, for example to prioritize areas for protection, sensitivity estimates are needed for broader ecological units—populations, communities, or ecosystems. Aggregate metrics can be developed from the traits of their constituent species using a number of methods: (i) basing higher-level sensitivity on the taxon with the highest sensitivity rank (Tyler-Walters et al., 2009); (ii) rank-weighted average sensitivity, based on the product of sensitivity rank and abundance for all taxa at a site (de Juan et al., 2009); and (iii) the number or proportion of taxa present in each sensitivity category defined from biological traits (Hewitt et al., 2011). Ideally, knowledge of population dynamics and density-dependent effects would be incorporated into metrics scaled to population levels, but this is not yet possible for deep-sea benthos. Hewitt et al. (2011) applied the above three methods to data from a deep-sea fishery area in New Zealand and concluded that community sensitivity was best assessed by a combination of (ii) and (iii), both of which yielded a graduated negative relationship between sensitivity and trawl intensity (Figure
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Time-series observations of seamount benthic communities exposed to bottom trawl impacts (Williams et al., 2010) indicated that fine-level grouping of taxa may be needed to classify their sensitivity. Relatively high abundances of sessile forms (unstalked crinoids, chrysogorid octocorals, solitary scleractinians, and gorgonians), and mobile forms (the urchin Dermichinus horridus and species of echinoids and prawns) occurred where trawl intensity had been relatively high. While these patterns might have contained some early signals of recolonization (including immigration by mobile species), a more detailed examination of the size and distribution of some of these taxa in still images of the seabed indicated that they were more likely to have resulted from individuals that were tolerant of the direct impacts of trawling and those that existed in natural refuges inaccessible to trawls (authors unpublished data). Tolerance of erect taxa to bottom trawling was apparent in chrysogorgid corals on account of their small size and high structural flexibility, and in solitary scleractinians and stylasterids with their small size and robust hard exoskeleton. Thus, analysis of sensitivity may need to treat hard and soft bottom separately because the expected responses of fauna to disturbance may be partially mitigated by natural refuges which are more common on complex rocky seabed.
A critical element in managing the environmental performance of deep-sea fisheries is to identify the capacity for impacted populations, assemblages and ecosystems to re-establish biological structures and functions after the impacts have ceased or diminished. Recovery—or the return to conditions that resemble background values in systems not damaged by fishing activities—is not unique to managing impacts in the deep sea (Paine et al., 1998). It is a feature common to all ecosystems and encompasses disturbance regimes that can be natural (e.g. turbidity flows, benthic storms, and volcanic eruptions) or anthropogenic (i.e. mining, trawling, and longlining). Many studies have examined this aspect in shallow and shelf waters, where it has become evident that responses are based on a complex set of site-specific factors that are often poorly understood and difficult to estimate. Collie et al. (2000) documented a number of studies where results from carefully designed studies were contrary to expectations, or changes could not be detected. Severe storm events can also influence benthic communities to a depth of ∼100 m (Sharma, 1974). This is much shallower than the depths considered in this paper, but it is nevertheless useful to bear in mind that such natural influences can have as large, or a greater, influence on changes in species abundance than bottom trawling (McConnaughey and Syrjala, 2014). What is, however, unique in the deep sea is (i) the rates of recovery may be much slower than in shallower systems and (ii) the almost complete lack of empirical data on faunal recovery in the deep sea means that inferences about recovery have to be made using proxies based on the longevity and growth rates of the organisms that have been damaged.
There is a long-standing and widely held belief that recovery is extraordinarily sluggish in the deep sea (Grassle, 1977). The expectation of slow recovery arises primarily from low biological rates in deep-sea species (Smith, 1994), life history traits that are predicted to delay recovery (Young, 1983), and variable larval dispersal and intermittent recruitment and settlement (Lacharité and Metaxas, 2013). While it is true that organisms in colder, deeper waters have slower turnovers, this is primarily a temperature effect: when body size and temperature are accounted for, deep-sea benthic species have similar metabolic rates (McClain et al., 2012). Some deep-sea benthos also have comparatively longer lifespans and tend to grow slower as a consequence of living in food-poor and cold environments (McClain et al., 2012). A further critical factor in determining recovery is the supply and fitness of colonizers: dispersal in the deep sea can thus be a limiting factor if disturbed areas are widely separated from colonizer pools, resulting in potentially low larval supply to impacted areas (Lacharité and Metaxas, 2013).
The question whether deep-sea systems recover as slowly as expected cannot at present be answered with empirical data. While there exist a number of studies that have measured post-disturbance processes in the deep-sea, these are, with one exception, limited to soft-sediment habitats (Smith and Hessler, 1987; Borowski and Thiel, 1998; Bluhm, 2001; Thiel et al., 2001; Khripounoff et al., 2006; Miljutin et al., 2011; Gates and Jones, 2012). In contrast, trawling most commonly targets hard grounds in the deep sea (e.g. seamounts) where ecological impacts are often most severe (Clark et al., 2010b). On seamounts that have been the target of fisheries for several decades in New Zealand and Australia, Williams et al. (2010) attempted to measure actual recovery rates of the fauna: they found no consistent and clear signal of recovery in the megabenthos 5–10 years after fishing had ceased, suggesting that any recovery is likely to be very prolonged.
A complementary line of evidence that suggest limited recovery potential for the megabenthos (e.g. sponges, corals, and crinoids) impacted by trawling gear on hard bottoms in the deep sea comes from data on growth, age, and lifespans of the fauna (Table Schlacher et al., 2014). Formation of new habitat could operate at geological time-scales (centuries or longer).
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Limited recovery potential is a key difference between shallow-water and deep-sea benthic communities (and many fish species). Even if the sensitivity of the benthos is similar, the recovery time from any given disturbance will be much greater in the deep sea.
This review has identified a number of studies demonstrating that direct and indirect fishing disturbances can severely impact deep-sea benthos by reducing diversity and abundance. Impacts from bottom trawling are better understood than those from other gears such as bottom-set longlines, and show that changes to benthic communities can be rapid, persistent, and occur with low levels of fishing effort. This is because many individual taxa are sessile with erect and fragile forms, can be relatively long-lived and slow-growing (especially as depths increase beyond ∼500 m), and may attain large body size. Communities associated with biogenic habitats formed by deep-sea corals and sponges are among the most susceptible to fishing impacts because their tolerance (ecological resistance) and recoverability (ecological resilience) is low. The few post-impact time-series data available from the deep sea show, unsurprisingly, that recovery times of benthic communities may be very long. Deep-sea benthic communities have the collective properties of high susceptibility and low recoverability, and hence it is unrealistic to expect them to recover from ongoing fishing impacts, or in the time-spans (years) typically applied to management planning. Restoration concepts are unachievable in the short term, and will be prohibitively expensive (see Van Dover et al., 2014). What then are the options for fisheries managers tasked with balancing sustainable fisheries exploitation and environmental conservation?
The variety of management actions taken to date include regulating fishing methods and gear types, specifying the depths fished, limiting the volumes of bycatch or limiting catch, move-on rules, and closing areas of particular habitat and individual seamounts (Probert et al., 2007; Morato et al., 2010). In terms of fishing methods and gears, various technical modifications to trawl gear such as a lighter groundrope, reduced trawl door weight, shortening the sweep wires that connect the doors to the net, using fly-wires to reduce ground contact, as well as elevating the sweep wires are possible (Mounsey and Prado, 1997; Valdemarsen et al., 2007; Rose et al., 2010; Skaar and Vold, 2010). However, while these can reduce fisheries bycatch and small invertebrates, they are unlikely to substantially reduce the impact on benthic communities, particularly sessile invertebrates with fragile and erect body forms. Use of midwater trawling gear close to the seabed has potential to reduce impact, and longline fishing may be appropriate in some environments—but the practicality of using these methods will vary with target species and location, and operationally will almost certainly involve a trade-off between bottom impact and catch rate of fish species. Many of the main target commercial species have a diving behaviour when disturbed (e.g. orange roughy, oroes, and alfonsino) and hence fishing clear of the seabed can leave an escape channel open that will reduce catch rates. Nevertheless, environmental management is as important in many nation's fisheries policies as target fish species catches. The “ecosystem approach” to fisheries management is now widely advocated and applied in deep-sea fisheries (Garcia et al., 2003). In the deep sea, however, the inherent restrictions on obtaining sufficient stock assessment or benthic habitat data (compared with nearshore shelf/slope fisheries) mean that management regimes typically operate at a low level of knowledge, and management action must occur in a highly precautionary manner. Move-on rules have recently become a common management tool, promoted by United Nations General Assembly resolutions for high seas fisheries that force vessels to move a certain distance if a threshold catch of vulnerable marine ecosystem (VME) indicator species is exceeded (Rogers and Gianni, 2010; Auster et al., 2011). However, the impacts from a single deep-sea trawl will potentially affect the benthos over a large distance (up to 150 m width along the length of a tow). The cumulative area swept by bottom trawl fisheries is typically the most extensive human impact on the seabed (Benn et al., 2010; Ramirez-Llodra et al., 2011), and there are further issues with move-on rules, such as threshold criteria and forcing fishing effort to spread further (Auster et al., 2011; Clark and Dunn, 2012).
Spatial management is likely to be the most effective strategy, and perhaps the only approach that can be successful for protection of vulnerable benthic fauna in the deep sea (Clark and Dunn, 2012; Schlacher et al., 2014). This approach is best achieved by restricting the distribution of fishing effort, and putting in place a system of zones which can allow exploitation in productive fishing areas, but protect vulnerable or sensitive species and habitats. Typically, this involves a network of open and closed areas, with closure of unfished areas where benthic communities occur in their natural state.
Management of the deep-sea lags behind that of the continental shelf, but there is a growing array of protection measures. Fishery closures are becoming common, with large areas within EEZs being closed zones for bottom trawling (e.g. New Zealand, North Atlantic, Gulf of Alaska, Bering Sea, USA waters, Azores) (Hourigan, 2009; Morato et al., 2010), and there are even some closures implemented on the high seas under international fishery management agreements (e.g. South Pacific, Penney et al., 2009). The effectiveness of such deep-sea fishing closures is, usually, yet to be formally established. One of the most thorough evaluations conducted to date suggests that the spatial closures instigated by New Zealand on the Louisville Seamount Chain, Lord Howe Rise, Challenger Plateau, and West Norfolk Ridge are suboptimal for the protection of VMEs and alternative closures would better balance protection against economic loss to fishers from closure of historically fished areas (Penney and Guinotte, 2013). There are increasing efforts to identify areas of importance for deep-sea benthic biodiversity, such as Ecologically or Biologically Significant Areas (CBD, 2009) and VMEs (FAO, 2009), and systematic methods are being developed (Taranto et al., 2012; Ardron et al., 2014; Clark et al., 2014) to help incorporate them in spatial management measures. Even without extensive biological data on deep-sea communities, it is possible to use habitat suitability models to predict the likelihood of regions hosting particularly vulnerable taxa (Davies and Guinotte, 2011; Vierod et al., 2014), derive risk indices to rank the threat of fishing (Clark and Tittensor 2010), and use biophysical variables as surrogates for biological assemblages (Anderson et al., 2011). Such methods and techniques will always have their limitations given the paucity of hard data in the deep sea, but together with the application of planning software tools (such as Marxan (Ball and Possingham, 2000), Zonation (Moilanen, 2007)) these methods can give managers a potentially powerful array of information and scientific approaches on which to base improved management of the impacts of fishing in the deep sea.
We acknowledge participants at the ICES Symposium on effects of fishing on benthic fauna, habitat, and ecosystem function, for their questions and discussion about aspects of fishing impacts in the deep sea. This work builds on international collaboration started in 2005 under the Census of Marine Life field project on Seamounts (“CenSeam”). MRC, DAB, and AAR were supported by funding from the New Zealand Ministry of Business, Innovation and Employment (Contract CO1X0906: “Vulnerable deep-sea communities” research project). AW and FA were supported through the Marine Resources and Industries Theme within the CSIRO Oceans and Atmosphere Flagship.
© International Council for the Exploration of the Sea 2015.