The Marine Life Information Network

Temperature increase (local)

Benchmark. A 5°C increase in temperature for one month, or 2°C for one year. Further detail

Evidence

The species that occur in this biotope are widely distributed and found both north and south of the British Isles (Hayward & Ryland, 1995b).

Brissopsis spp. exhibit a wide climatic and geographical distribution, occurring in tropical, subtropical and temperate zones, and shallow waters of cold regions (Ghiold, 1989; Borghi & Garilli, 2022). Brissopsis lyrifera is generally distributed in the northeast and southeast Atlantic and the western Indian Ocean (Ghiold, 1989). Borghi & Garilli (2022) found that Brissopsis lyrifera and Brissopsis atlantica have a long historical presence in the Mediterranean and still occur there today, and their biogeography is not remarkably affected by sea surface temperature. OBIS (2025) recorded Brissopsis lyrifera and Amphiura chiajei from 0 to 30°C (SST), with the majority of records occurring between 10 to 15°C.

In shallower locations e.g. sea lochs, sedimentary biotopes typically experience seasonal changes in temperature of about 10°C (5 to 15°C) (Hughes, 1998b) and it is likely that the SS.SMu.CFiMu.BlyrAchi community would be resistant of an acute temperature increase. For most offshore burrowing species, temperature changes in the water column are likely to be buffered by the insulation offered by the substratum and the depth of overlying water. A temperature increase may enhance growth and fecundity in brittlestars. Muus (1981) showed that juvenile Amphiura filiformis are capable of much higher growth rates in experiments with temperatures between 12 and 17°C (and unlimited food supply). Juvenile disc diameter increased from 0.5 to 3.0 mm in 28 weeks under these conditions, compared to over two years in the North Sea. Mean summer temperatures of 14°C and an apparent abundant food supply may also have accounted for the early rapid growth of Amphiura chiajei in Killary Harbour (Munday & Keegan, 1992).

In Brissopsis lyrifera, processes such as mobility, sediment turnover and re-mineralization may increase (K. Hollertz, pers. comm., Hollertz & Duchêne, 2001). Hollertz & Duchêne (2001) found that in Brissopsis lyrifera, the amount of re-worked sediment due to burrowing almost doubled from 14 to 22 ml/l sediment per hour when the temperature increased from 7 to 13°C. This temperature increase also saw the amount of ingested sediment increase from 0.02 to 0.08 g dry sediment per hour. However, increased water temperature may enhance microbial decomposition within the substratum and promote de-oxygenation, to which Brissopsis lyrifera is not resistant.

Sensitivity assessment. The biotope is subtidal and a low-energy environment where wide and rapid variations in temperature are less common. The community is therefore less likely to be resistant to an acute increase in temperature. However, the distribution data suggest that species in this biotope are unlikely to be adversely affected by an increase in temperature at the pressure benchmark level. Furthermore, the evidence presented suggests that growth and recruitment of the characterizing species of the biotope, Brissopsis lyrifera and Amphiura chiajei may benefit from an increase in temperature at the pressure benchmark level. Resilience and resistance are therefore assessed as 'High', and the biotope is considered 'Not Sensitive' at the pressure benchmark.

Temperature decrease (local)

Benchmark. A 5°C decrease in temperature for one month, or 2°C for one year. Further detail

Evidence

The species occurring in this biotope are widely distributed and found both north and south of the British Isles (Hayward & Ryland, 1995b).

Brissopsis spp. exhibit a wide climatic and geographical distribution, occurring in tropical, subtropical and temperate zones, and shallow waters of cold regions (Ghiold, 1989; Borghi & Garilli, 2022). Brissopsis lyrifera is generally distributed in the northeast and southeast Atlantic and the western Indian Ocean (Ghiold, 1989). Borghi & Garilli (2022) found that Brissopsis lyrifera and Brissopsis atlantica have a long historical presence in the Mediterranean and still occur there today, and their biogeography is not remarkably affected by sea surface temperature. OBIS (2025) recorded Brissopsis lyrifera and Amphiura chiajei from 0 to 30°C (SST), with the majority of records occurring between 10 to 15°C.

In shallower locations e.g. sea lochs, sedimentary biotopes typically experience seasonal changes in temperature of about 10°C (5 to 15°C) (Hughes, 1998b) and it is likely that the SS.SMu.CFiMu.BlyrAchi community would be resistant to a long-term temperature decrease. For most offshore burrowing species temperature changes in the water column are likely to be buffered to some extent by the insulation offered by the substratum and the depth of overlying water. However, burrowing itself has been found to be significantly affected by temperature in Brissopsis lyrifera. Hollertz & Duchêne (2001) found that Brissopsis lyrifera re-worked almost half the amount of sediment per hour at 7°C compared to activity at 14°C. Furthermore, Brissopsis lyrifera maintains a continuous contact with the overlying water column through the funnel (Hollertz, 2002). Also, the biotope community seems to be periodically affected by severe winters. Mean densities of Amphiura chiajei in Killary Harbour, west coast of Ireland, decreased following months with the lowest recorded bottom temperatures, 4°C and 6°C, for February 1986 and January 1987, respectively. Lack of resistance of the acute change and depressed temperatures on the part of some of the older individuals probably led to their demise (Munday & Keegan, 1992). Low temperatures are also a limiting factor for breeding, which occurs in the warmest months in the UK. Temperature tolerances of Brissopsis lyrifera are unknown but low water temperatures have caused mass mortalities of other similar echinoderms, such as Echinocardium cordatum. In the severe winter of 1962 to 63 masses of dead Echinocardium cordatum were observed in regions of the North Sea and English Channel, although it was reported that living specimens were obtained easily enough by digging (Crisp, 1964).

Sensitivity assessment. Although the evidence suggests possible effects on metabolism and recruitment for the characterizing species of this biotope caused by a decrease in temperature, their cosmopolitan distribution suggests that Brissopsis lyrifera and Amphiura chiajei are likely to be able to resist a decrease in temperature at the benchmark level. Resistance and resilience are therefore assessed as 'High', and the biotope is considered 'Not Sensitive' at the pressure benchmark.

Salinity increase (local)

Benchmark. A increase in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

The biotope SS.SMu.CFiMu.BlyrAchi is found within fully marine subtidal locations and it is highly unlikely that the biotope would experience conditions of hypersalinity. However, it is likely that key components of the biotope community would not be resistant of an increase in salinity. For instance, echinoderms such as Brissopsis lyrifera and Amphiura chiajei are stenohaline owing to the lack of an excretory organ and a poor ability to osmo- and ion-regulate (Stickle & Diehl, 1987). Echinoderm larvae are particularly sensitive to reduced or increased salinity (Stickle & Diehl, 1987).

Sensitivity assessment. There is little direct evidence of the effects of hypersaline conditions on the characterizing species of this biotope, Brissopsis lyrifera and Amphiura chiajei. However, echinoderms are generally considered to be stenohaline (Stickle & Diehl, 1987; Russell, 2013). Therefore, an increase in salinity to >40 psu is likely to result in mortality and resistance is assessed as 'Low' but with low confidence. Resilience is probably 'Medium' so sensitivity is therefore assessed as 'Medium'.

Salinity decrease (local)

Benchmark. A decrease in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

The biotope SS.SMu.CFiMu.BlyrAchi is found within fully marine subtidal locations and it is highly unlikely that it would experience conditions of hyposalinity, except from effluent. However, it is likely that key members of the biotope community would not be resistant of a decrease in salinity. For instance, echinoderms such as Brissopsis lyrifera and Amphiura chiajei are stenohaline owing to the lack of an excretory organ and a poor ability to osmo- and ion-regulate (Stickle & Diehl, 1987). Pagett (1981) examined the resistance of Amphiura chiajei to brackish water (0.5-30 psu) in specimens taken from Loch Etive, Scotland. Loch Etive is a sea loch subject to periods of reduced salinities owing to heavy rain and freshwater runoff. The author suggested that localised physiological adaption to reduced or variable salinities may occur in nearshore areas subject to freshwater runoffs. Amphiura chiajei taken from an area of 24 psu had an LD50 of > 21 days for a 70% dilution (17 psu) and an LD50 of 8.5 days for a 50% dilution (12 psu). In comparison, specimens taken from an area with salinity 28.9 psu, had an LD50 of >12.5 days for a 70% dilution (20 psu) and an LD50 of 6 days for a 50% dilution (14 psu). As Amphiura chiajei is mobile and burrows it may be able to avoid changes in salinity outside its preference, e.g. burrowing may help Amphiura chiajei to withstand depressed salinities owing to the 'buffering' effect of the substratum. A review by Russell (2013) reported that Amphiura chiajei tolerated salinities of 14.8‰, 20.7‰ and 18‰ in Portugal, Scotland and the Black Sea, respectively. However, no information concerning the specific tolerance of Brissopsis lyrifera to a decrease in salinity was found but burrowing in the muddy sediment may offer some protection to this species.

Sensitivity assessment. Populations that occur in sheltered nearshore situations, such as sea lochs, which periodically receive inputs of freshwater are unlikely to experience the reduced salinities recorded at the surface. However, this circalittoral biotope is less likely to experience variable salinities, and resident species, therefore, less likely to adapt to variation in salinity, as suggested by the results given by Pagett (1981). Therefore, resistance to a decrease in salinity from full to reduced (18 to 30 psu) is assessed as Low (loss of 25 to 75% of individuals). Resilience is assessed as 'Medium' (2 to 10 years), and the biotope is considered to have 'Medium' sensitivity at the benchmark level.

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s to 0.2 m/s for more than one year. Further detail

Evidence

The presence of the biotope is determined by a low energy hydrodynamic regime facilitating the deposition of cohesive fine silts and clays. The substratum is likely to remain unchanged because the smooth compacted fine sediments require very high shear stresses to be eroded. However, the settlement of the planktonic larvae of these key species may be inhibited with an increase in water flow owing to re-suspension along with particulate matter. Consequently, the viability of the population may be reduced. Furthermore, the deposit feeding community may experience a reduction in food availability owing to reduced deposition of organic matter. The community is not likely to be directly vulnerable to a decrease in water flow rate but sediments may become muddier owing to increased settlement of particulate matter. However, as deposit feeders are the dominant trophic group such additional material may be utilized as a food resource and the community may benefit indirectly.

Brissopsis lyrifera is characteristic of offshore muddy sedimentary habitats exposed to only weak or very weak currents (Budd, 2004). Amphiura chiajei shows no clear response to directional bottom currents or an increase in water current rate (Buchanan, 1964). In laboratory conditions, Amphiura chiajei maintained a position within the sediment with its arms stretched out across the sediment until 0.3 m/s, when the arms streamed out in the direction of the water current (Buchanan, 1964). Increases in water flow rates could mean the brittlestars would withdraw arms from the current and cease feeding.

Sensitivity assessment. The assessment is based largely on the Hjulström-Sundborg diagram (Sundborg, 1956). This relates current velocity to deposition, erosion and transport. While this model has largely been superseded in by more recent models that take into account other factors such as shear stress and water depth, these newer models are more complex, site specific and do not relate sediment transport to water velocity. The curve is therefore used to assess generally the potential effects of changes in water velocity but it should be recognized that a number of other factors will mediate effects. Both characterizing species of this biotope are normally associated with weak and very weak tidal streams (negligible - <0.5 m/s). A decrease in water flow rate would likely result in increased siltation, potentially associated with increased deposition organic matter, which would likely indirectly benefit the characterizing species of the biotope. An increase in water flow may alter the character of the sediment of the biotope by washing away finer particle, which could potentially exclude colonization of the characterizing species of the biotope, such as Brissopsis lyrifera, which is associated to muddy/silty substrata. However, records indicate the biotope occurs in weak and very weak tidal streams (Connor et al., 2004). This range is likely to exceed the pressure benchmark. Resistance and resilience are therefore assessed as 'High' and the biotope is considered 'Not Sensitive' at the pressure benchmark.

Emergence regime changes

Benchmark.  1) A change in the time covered or not covered by the sea for a period of ≥1 year or 2) an increase in relative sea level or decrease in high water level for ≥1 year. Further detail

Evidence

Changes in emergence are Not Relevant to biotopes which are restricted to fully subtidal/circalittoral conditions. The pressure benchmark is relevant only to littoral and shallow sublittoral fringe biotopes.

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year. Further detail

Evidence

The SS.SMu.CFiMu.BlyrAchi biotope occurs offshore and in sheltered nearshore habitats where wave exposure is negligible (Connor et al., 2004), so the biotope is probably not resistant to changes in wave exposure. However, as the effects of wave action are attenuated with depth, the factor is only likely to affect the biotope where it occurs at depths of less than 60 m in a strong swell or force 8 gale (Hiscock, 1983). Wave action resulting from storms may disturb the surface sediment. McIntosh (1975) reported specimens of Amphiura chiajei thrown on to West Sands, St. Andrews Bay after storms. Long-term increases in wave exposure is likely to cause the substratum character to be altered, as wave action would penetrate the substratum to a greater depth, and become outside the habitat preference of the species. The community would no longer occur at that location. However, Amphiura chiajei is a burrower and may withdrawal its arms into the burrow for additional protection, but because wave action may cause displacement and stranding, it is likely to cause some mortality in shallower locations.

Brissopsis lyrifera is characteristic of offshore and stable muddy nearshore habitats, where wave exposure is negligible. Populations situated in normally sheltered stable habitats at shallower depths may experience some disturbance to the sediment surface. However, as Brissopsis lyrifera burrows in the sediment to a depth of 10 cm, it is unlikely that turnover and displacement would occur to an extent where the population is significantly reduced.

Sensitivity assessment. A decrease in wave exposure is not relevant because this biotope is characterized as a low energy environment. An increase in wave exposure is likely to adversely affect the characterizing species in this biotope, limiting or removing the shallower proportion of the community, and potentially modifying sediment and therefore habitat preferences in the longer-term. However, records indicate the biotope occurs in a range of wave exposures (Connor et al., 2004). This range is likely to exceed the pressure benchmark. Additionally the depth at which the biotope occurs is likely to provide some protection to changes in wave height at the pressure benchmark level. Resistance and resilience are therefore assessed as 'High' and the biotope considered 'Not Sensitive' to a change in nearshore significant wave height >3% but <5%.

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed but evidence is presented where available.

Information concerning the effects of heavy metals on echinoderms is limited and no information specific to Brissopsis lyrifera and Amphiura chiajei was found. In Norwegian fjords, Rygg (1985) found a relationship between species diversity in benthic fauna communities and sediment concentrations of heavy metals Cu, Pb and Zn. Cu in particular showed a strong negative correlation and the author suggested a cause-effect relationship. Those species not present at sites where Cu concentrations were greater than ten times the background level, such as Amphiura filiformis and the bivalve Nucula sulcata (also found in SS.SMu.CFiMu.BlyrAchi), were assessed as non-resistant species. Resistant species were all polychaete worms. Polychaete worms are the dominant component of the biomass in SS.SMu.CFiMu.BlyrAchi and thus may not be as sensitive as the characterizing species. Crompton (1997) reports that the concentrations above which mortality of crustaceans can occur is 0.01-0.1 mg/l for mercury, copper and cadmium, 0.1-1.0 mg/l for zinc, arsenic and nickel and 10 mg/l for lead and chromium. Some burrowing crustaceans, brittlestars and bivalves may disappear from the biotope and lead to an increasing dominance of polychaetes. Bryan (1984) suggested that metal-contaminated sediments can exert a toxic effect on burrowing bivalves and echinoderms, especially at larval stages, and that polychaetes were fairly resistant. Adult echinoderms, such as Ophiothrix fragilis are known to be efficient concentrators of heavy metals including those that are biologically active and toxic (Hutchins et al., 1996). However, there is no information available regarding the effects of this bioaccumulation. More recent studies by Deheyn & Latz (2006) at the Bay of San Diego found that heavy metal accumulation in brittlestars occurs both through dissolved metals as well as through diet, to the arms and disc, respectively. Similarly, Sbaihat et al. (2013) measured concentrations of heavy metals (Cu, Ni, Cd, Co, Cr and Pb) in the body of Ophiocoma scolopendrina collected from the Gulf of Aqaba, and found that most concentration was found in the central disc rather than arms and no simple correlations could be found between contaminant and body length. It is logical to suppose that brittlestar beds would be adversely affected by major pollution incidents such as oil spills, or by continuous exposure to toxic metals, pesticides, or the antiparasite chemicals used in cage aquaculture. So far, however, there are no field observations of epifaunal brittlestar beds being damaged by any of these forms of pollution, and there seems to be no evidence of the toxicity effects of heavy metal accumulation on brittlestars.

The evidence presented suggests that the characterizing species in this biotope are likely to bio-accumulate heavy metals which could result in changes in the faunal composition of the community and decrease overall species diversity.

Hydrocarbon & PAH contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed but evidence is presented where available.

Chronic sub-lethal effects were detected around the Beryl oil platform in the North Sea where the hydrocarbon content of the sediment was very low (<3 ppm total hydrocarbons in sediment), and Amphiura chiajei was excluded from areas nearer the platform with higher sediment hydrocarbon content (> 10 ppm) (Newton & McKenzie, 1998). Amphiura chiajei is also host to symbiotic sub-cuticular bacteria (Kelly & McKenzie, 1995). After exposure to hydrocarbons, loadings of such bacteria were reduced indicating a possible sub-lethal stress to the host (Newton & McKenzie, 1995). Furthermore, Brissopsis lyrifera has a continuous water flow over the test so the exposure route through the epidermis may also be important (K. Hollertz, pers. comm., Hollertz, 2002). However, as a burrower and deposit feeder, ingestion of contaminated sediments is likely to be a more important route of exposure. A range of effects (mortalities, feeding/growth inhibition and embryological abnormalities) have been reported for other echinoderms following hydrocarbon exposure (reviewed by Suchanek, 1993).

Untreated oil (e.g. from oil spills) is not a risk, since it is concentrated mainly at the surface, and circalittoral biotopes are likely to be protected by their depth. If oil is treated by dispersant, the resulting emulsion will penetrate down the water column, especially under the influence of turbulence (Hartnoll, 1998).

Synthetic compound contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed but evidence is presented where available.

Effects caused by synthetic chemicals have been reported for some of the individual species in the SS.SMu.CFiMu.BlyrAchi biotope. Dahllöf et al. (1999) studied the long-term effects of tri-n-butyl-tin (TBT) on the function of a marine sediment system. TBT spiked sediment was added to a sediment that already had a TBT background level of approximately 27 ng/g (83 pmol TBT per g) and contained the following fauna: Amphiura spp., Brissopsis lyrifera and several species of polychaete. Within two days of treatment with a TBT concentration above 13.7 µmol/m² all species except the polychaetes had crept up to the surface and after six weeks these fauna had started to decay. Thus, contamination from TBT is likely to result in the death of some not-resistant species such as brittlestars and heart urchins. Amphiura chiajei is also known to bioaccumulate PCBs, although direct effects of synthetic chemicals on this species are unknown (Gunnarsson & Skold, 1999). However, Walsh et al. (1986) observed inhibition of arm regeneration in another brittlestar, Ophioderma brevispina, following exposure to TBT at levels between 10 ng/l and 100 ng/l. Loizeau & Menesguen (1993), found that 8-15% of the PCB burden in dab, Limanda limanda, from the Bay of Seine could be explained by ophiuroid consumption. Thus, Amphiura communities may play an important role in the accumulation, re-mobilization and transfer of PCBs and other sediment associated contamination to higher trophic levels.

Detergents used to disperse oil from the Torrey Canyon oil spill caused mass mortalities of a similar species, Echinocardium cordatum (Smith, 1968). Sea-urchins, especially the eggs and larvae are used for toxicity testing and environmental monitoring (reviewed by Dinnel et al., 1988). It is likely therefore, that Brissopsis lyrifera and its larvae are not resistant to synthetic compound contamination.

Radionuclide contamination

Benchmark. An increase in 10µGy/h above background levels. Further detail

Evidence

There is insufficient information on the resistance of Amphiura chiajei to radionuclides, although adult echinoderms, such as Ophiothrix fragilis, are known to be efficient concentrators of radionuclides (Hutchins et al., 1996). However, no information concerning the effects of such bioaccumulation was found.

Carvalho (2011) determined the concentrations of ²¹⁰Po and ²¹⁰Pb in marine organisms from the seashore to abyssal depths, as these two radioactive elements tends to be higher in the marine environment. The author’s results showed that concentrations varied greatly, even between organisms of the same biota, mainly related with the trophic levels occupied by the species, suggesting that the more levels between a species and the bottom of the food chain, the more likely that the concentrations of radioactive elements were likely to be diluted. This may have great implications for the detritus feeders that characterize this biotope. There was no information available about the effect of this bioaccumulation.

Sensitivity assessment: Although species in this biotope are likely to bio-accumulate radionuclides with potential impacts on the biological community, no information concerning the effects of such bioaccumulation was found. Therefore, there is insufficient evidence to assess this pressure against the benchmark.

Introduction of other substances

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed.

De-oxygenation

Benchmark. Exposure to dissolved oxygen concentration of less than or equal to 2 mg/l for one week (a change from WFD poor status to bad status). Further detail

Evidence

Infaunal burrowers in the community live in close association with hypoxic and even anoxic muddy substrata. In experiments, Amphiura chiajei exposed to decreasing oxygen levels only left its protected position in the sediment when oxygen levels fell below 0.54 mg/l (Rosenberg et al., 1991). This escape response increases its risk to predators. Mass mortality in a superficially similar species of ophiuroid, Amphiura filiformis from the south-east Kattegat was observed during severe hypoxic events (< 0.7 mg/l), while the abundance of Amphiura chiajei remained unchanged at the same site and time (Rosenberg & Loo, 1988). Nilsson (1999) maintained specimens of Amphiura chiajei in hypoxic conditions (1.8 to 2.2 mg/O₂/l for eight weeks and recorded no deaths or witnessed specimens escaping to the surface. Infaunal burrowers in the community live in close association with hypoxic and even anoxic muddy substrata. In experiments, Amphiura chiajei exposed to decreasing oxygen levels only left its protected position in the sediment when oxygen levels fell below 0.54 mg/l (Rosenberg et al., 1991). This escape response increases its risk to predators. Mass mortality in a superficially similar species of ophiuroid, Amphiura filiformis, from the south-east Kattegat was observed during severe hypoxic events (< 0.7 mg/l), while the abundance of Amphiura chiajei remained unchanged at the same site and time (Rosenberg & Loo, 1988). Nilsson (1999) maintained specimens of Amphiura chiajei in hypoxic conditions (1.8 to 2.2 mg/O₂/l for eight weeks and recorded no deaths or witnessed specimens escaping to the surface.

Brissopsis lyrifera was reported to be not resistant of hypoxia (Diaz & Rosenberg, 1995). It was reported to leave its position within the substratum and lie exposed on the sediment surface in bottom waters with an oxygen concentration of 1 ml/O₂/l (ca 1.4 mg/l O₂) (Baden et al., 1990). At a bottom water oxygen concentration of ca. 1 ml/l (15% saturation) in the Kattegat, Baden et al. (1990) caught no fish, but 200 to 400 kg per hour of benthic invertebrates that included the echinoderms Brissopsis lyrifera. Similar mass migration of benthic infauna (including Brissopsis lyrifera) to the sediment surface was recorded during trawling in the North Sea with low values of oxygen (ca 2 ml/l, ca 2.8 mg/l) (Dyer et al., 1983). Hollertz (2002) reported that Brissopsis lyrifera could tolerate ca 4 ml/l (ca 5.6 mg/l) for at least 15 hours in the laboratory and that the animals recovered quickly. In the Gullmarsfjord (where Brissopsis lyrifera is recorded; Brattström, 1946; Vasseur & Carlsen, 1949), a 1980/1981 hypoxia event, ca 0.2 ml/l, eliminated all the macrobenthic fauna below 115 m depth. The recovery sequence was slow, and communities were not re-established eighteen months after the collapse (Josefson & Widbom, 1988).

Evidence from the burrowing urchin Brissopsis pacifica showed the species could tolerate deoxygenation along the continental shelf and slope of the eastern Pacific (Sato et al., 2017). Although no significant shifts in density or depth distribution were found during the study period (2003 to 2013), Brissopsis pacifica was observed occupying shallower depths during high oxygen El Niño periods. This indicates that the species can tolerate low-oxygen conditions and may expand upslope as deoxygenation intensifies (Sato et al., 2017).

Sensitivity assessment: The evidence presented suggests that exposure to dissolved oxygen concentrations of less than or equal to 2 mg/l may result in some mortality of the key species in this biotope. Resistance is therefore assessed as ‘Low’, resilience as ‘Medium’ and as  ‘Medium’ sensitivity to de-oxygenation at the pressure benchmark level.

Nutrient enrichment

Benchmark. Compliance with WFD criteria for good status. Further detail

Evidence

Increased nutrients are most likely to affect the abundance of phytoplankton, which may include toxic algae (OSPAR, 2009). This primary effect resulting from elevated nutrients will impact other biological elements or features (e.g. toxins produced by phytoplankton blooms or de-oxygenation of sediments) and may lead to ‘undesirable disturbance’ to the structure and functioning of the ecosystem. With enhanced primary productivity in the water column, organic detritus that falls to the seabed may also be enhanced. 

Bioturbators such as the characterizing species of this biotope are likely to be able to utilize extra nutrients that may end up in the biotope due to nutrient enrichment. Olsgard et al. (2008) highlighted the importance of macrofauna as facilitators of nutrient flux in soft-sediments. The authors suggested that Brissopsis lyrifera plays an important role in nutrient cycling and ecosystem functioning. Similarly, under laboratory conditions, Cassidy et al. (2020) showed that Amphiura chiajei can influence nutrient levels in the sediment through bioturbation, indicating that Amphiura chiajei also plays an important role in nutrient cycling and ecosystem functioning. Equally, Widdicombe et al. (2013) suggested that the presence of urchins preferentially stimulates nitrification (a process by which nutrients are recycled in the sediment), increasing the production of nitrite and nitrate within the sediment and therefore reducing the gradient in nutrient concentrations that exists between the overlying water and sediment porewaters.

Sensitivity assessment. While organic enrichment (discussed below) is likely to result in nutrient enrichment, no evidence of the effects of nutrient enrichment (nitrates, phosphates, silicates or iron) as defined by the pressure was found. Hence, there is ‘Insufficient evidence’ on which to base an assessment, at present.

Organic enrichment

Benchmark. A deposit of 100 gC/m²/yr. Further detail

Evidence

Amphiura chiajei has been described as a bioindicator species for organic enrichment, often associated with organically enriched mud and muddy sands (Aslan, 2024). Nilsson (1999) reported a positive response by Amphiura chiajei to increased organic enrichment (27 and 55 gC/m², applied four times over eight weeks) demonstrated by an increase in arm tip regeneration rate. Nilsson (1999) also found that Amphiura chiajei was able to utilize an increased input of organic matter for growth in conjunction with moderate hypoxia. In the Skagerrak in the North Sea, Josefson (1990) reported a massive increase in abundance and biomass of Amphiura species between 1972 and 1988 attributable to organic enrichment. Sköld & Gunnarsson (1996) reported enhanced growth and gonad development in response to short-term enrichment of sediment cores containing Amphiura chiajei maintained in laboratory mesocosms. For benthic deposit feeders, food is suggested to be a limiting factor for body and gonad growth, at least between events of sedimentation of fresh organic matter (Hargrave, 1980; Tenore, 1988).

Pearson (1975) reported that Amphiura chiajei and Amphiura filiformis can tolerate “moderately increased” organic-enriched sediments exposed to effluent from a pulp and paper mill in Lochs Linnhe and Eil on the West coast of Scotland between 1966 and 1973. However, both species were excluded from areas of ‘high’ organic enrichment (dominated by polychaetes) and ‘excessive’ enrichment (dominated by Capitellids and bacterial mats).  Similarly, Amphiura filiformis recolonized a benthic community in a Swedish fjord, after the discontinued use of a sulphite mill, which reduced the amount of wastewater discharge (Rosenberg, 1972). Rosenberg (1972) described Amphiura filiformis as part of a mature community able to tolerate slight pollution only. Pearson et al. (1985) reported that Amphiura filiformis, previously absent or of minor importance, increased in biomass and dominance at 70% of stations studied in the Kattegatt following exposure to eutrophication. However, as organic enrichment increased, Amphiura chiajei and Amphiura filiformis have been reported to be lost from affected communities and may have been outcompeted by more tolerant macrofauna in “high and excessive” organic-enriched sediments (Pearson, 1975). Mattsson & Linden (1983) found Amphiura chiajei was amongst the first species to disappear from the faunal community under mussel cultures along the Swedish coast, six months after the start of mussel cultivation. The organic content in the sediment was between 13 and 22% (Mattsson & Linden, 1983). Bustos-Baez & Frid (2003) investigated pollution indicator species and found a significantly negative relationship between Amphiura filiformis abundance and macro-litter density found at the Tyne sewage sludge dumping ground. Pearson & Black (2001) describe Amphiura filiformis in a normal ‘unpolluted’ community, but absent in “lightly to highly enriched” sediment from fish farm waste.

Brissopsis lyrifera is a non-selective deposit feeder, characteristic of muddy sediments with significant organic matter content, so an increase in the suspended matter settling out from the water column to the substratum may be used as a food resource. Brissopsis lyrifera has been reported to increase surface deposit feeding activity after the addition of organic matter, which resulted in an increase in growth (Hollertz et al., 1998). The species is also capable of filter feeding, although ventilation rates are not high enough to sustain the animal on filter feeding alone (Hollertz, 2002). However, a study by Kutti et al. (2008) of the effects of organic effluents from a salmon farm on a Norwegian fjord system, indicated that the threshold for increased infauna production for a benthic ecosystem at the depth of 230 m was reached at an annual flux of 300 to 500 gC m² and that continuous loadings at this magnitude over time might cause overloading of fish farm localities. Brissopsis lyrifera was recorded as one of the three dominating infauna species within close proximity of the farm (within 250 m). Pearson et al. (1985) found that Brissopsis lyrifera showed little to no change in comparative studies across stations in the Kattegatt despite increasing eutrophication.

Nevertheless, eutrophication resulting from high pelagic production, in combination with thermal stratification of the water column in summer is likely to cause mortality of Brissopsis lyrifera indirectly, owing to the effects of hypoxia. Although hypoxia or even anoxia is likely to occur as a result of increased organic matter deposition, macrofauna, such as the characterizing species of this biotope are likely to enhance utilization of organic matter by aerobic bacteria in the sediment by bioturbation (active mixing of sediment) and bioirrigation (active flushing of solutes) (Sanz-Lázaro & Marín, 2011). Rosenberg (1972) found that Brissopsis lyrifera recolonized a benthic community in a Swedish fjord, after the discontinued use of a sulphite mill, which reduced the amount of wastewater discharge. It was suggested that the conditions of the fjord had improved and recovered in health, as seen in the increase in growth rates of Brissopsis lyrifera (Rosenberg, 1972). However, in areas of Brissopsis lyrifera may be outcompeted by a more mature community, able to tolerate slightly polluted conditions, as conditions improve further (Rosenberg, 1972).

Borja et al. (2000) and Gittenberger & van Loon (2011) in the development of an AMBI index to assess disturbance (including organic enrichment) both assigned Brissopsis lyrifera to their Ecological Group I ‘species very sensitive to organic enrichment and present under unpolluted conditions’. However, for Amphiura chiajei, while Borja et al. (2000) assigned the species to Ecological Group I ‘species very sensitive to organic enrichment and present under unpolluted conditions’, Gittenberger & van Loon (2011) assigned the species to Ecological Group II ‘species indifferent to enrichment, always present in low densities with non-significant variations with time (from initial state, to slight unbalance)’.

Sensitivity assessment. The evidence presented based on the AMBI scores conflicts with the other evidence reviewed and is not directly comparable with the benchmark, so it is considered with caution. Forrest et al. (2009) identified that the recovery of muddy sediments beneath fish farms from enrichment can be highly variable and may be many years at poorly flushed sites, such as those where this biotope tends to occur. The evidence above indicates that the Amphiura sp. and Brissopsis are resistant to ‘moderate’ enrichment and may benefit from increased input of organic material, but are lost from areas subject to high levels of enrichment, depending on how ‘moderate’ and ‘high’ are defined in individual studies. Nilsson (1999) reported that the growth of Amphiura chaejei increased in areas subject to the addition 27 to 55 gC/m² (over an eight-week period), a value higher than the benchmark of 100 gC/m²/year. Similarly, Kutti et al. (2008) reported that Brissopsis was a dominate member of the infauna within 250 m of a fish farms that deposited 300 to 500 gC/m2 per year, much higher than the benchmark. However, the remining studies indicate that Brissopsis and both species of Amphiura are lost as enrichment increases.

Overall, the characterizing species of this biotope may resist or benefit from a deposit of 100 gC/m² over the period of one year. However, the remaining evidence, including AMBI scores suggest the characteristic species are sensitive or indifferent to enrichment. Therefore, resistance is assessed as ‘Medium’ as a worst-case precaution to represent the potential variation in response between sites and the type or source of organic enrichment. Resilience is assessed as ‘Medium’ to represent the time taken for the pollution to flush from the system, itself dependent on the site. Thus, the biotope is assessed as ‘Medium’ sensitivity to organic enrichment at the pressure benchmark level, with ‘Low’ confidence due to the variation in evidence.

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site. Further detail

Evidence

All marine habitats and benthic species are considered to have a resistance of 'None' to this pressure and to be unable to recover from a permanent loss of habitat (resilience is 'Very Low'). Sensitivity within the direct spatial footprint of this pressure is, therefore, ​​​​​​'High'. Although no specific evidence is described, confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.

Physical change (to another seabed type)

Benchmark. Permanent change from sedimentary or soft rock substrata to hard rock or artificial substrata or vice-versa. Further detail

Evidence

If the silty mud that characterizes this biotope was replaced with soft or hard rock substrata, this would represent a fundamental change to the physical character of the biotope. Additionally, the biological community that occurs and characterizes the biotope would no longer be supported. The biotope would, therefore, be lost.

Sensitivity assessment: Resistance to the pressure is considered 'None', and resilience is 'Very Low'. Sensitivity has been assessed as 'High'.

Physical change (to another sediment type)

Benchmark. Permanent change in one Folk class (based on UK SeaMap simplified classification). Further detail

Evidence

Records indicate that SS.SMu.CFiMu.BlyrAchi is limited to silty muds (Connor et al., 2004), and particularly the characterizing species that it supports exhibit specific preferences for fine sediment substrata. Brissopsis lyrifera has a preference for cohesive sandy mud and silty mud (Tillin & Tyler-Walters, 2014); and Amphiura chiajei is found in mud and muddy sand (Budd, 2006). The remaining species that compose the biological community demonstrate higher plasticity in terms of sediment preferences. Although Nephrops norvegicus has been recorded across a range of substrata, creating permanent burrows will typically have specific sediment requirements which have been seen to relate to the maintenance of burrow structures, and to affect population densities, with preferences seeming to focus on sandy-muds than muds and In medium-grained mud sediments (Daly & Mathieson, 1977; Afonso-Dias, 1998).

Sensitivity assessment: The characterizing species of this biotope are associated with a narrow range of sediments types, including silty muds and muddy sands. For this biotope, a change in Folk class would mean a change from mud and sandy mud to sand or muddy sand, or to gravelly mixed sediment. The characterizing species are unlikely to be resistant of such a change in sediment type, no longer being supported, and hence lost from the biotope. The biotope is likely to be lost, so resistance is therefore assessed as 'None' and resilience as 'Very Low', given the permanent character of this pressure. The biotope is considered to have a 'High' sensitivity to a change in seabed type by one Folk class.

Habitat structure changes - removal of substratum (extraction)

Benchmark. The extraction of substratum to 30 cm (where substratum includes sediments and soft rock but excludes hard bedrock). Further detail

Evidence

SS.SMu.CFiMu.BlyrAchi is highly unlikely to be resistant of substratum loss because most species are infaunal and extraction of substratum to 30 cm is likely to result in the removal of the biological community along with substrata, including the characterizing species. Newell et al. (1998) stated that removal of 0.5 m (50 cm) depth of sediment is likely to eliminate benthos from the affected area. Brissopsis lyrifera burrows to a depth of up to 10 cm; Amphiura chiajei lives partially buried with its disc at a depth of 6 cm. These species are not sufficiently mobile to avoid substratum removal. Although some species are mobile, e.g. Nephrops norvegicus, if disturbed they are likely to seek refuge within a burrow within the substratum and so are also likely to be removed. Shallow and deep disturbance can injure, kill and displace benthic organisms and, in the case of fisheries, target and non-target species can be removed from the habitat. Through these effects, fisheries can alter the biomass, production and species richness of benthic invertebrate communities (Hiddink et al., 2006). Otter boards plough a groove in the seabed, which can vary from a few cm to 0.3 m deep (Jones, 1992, references therein). The trawl may remove or damage sedentary organisms and displace stones. Bobbins and chains can also leave tracks (Krost et al., 1989) and remove surface sediment. The disturbance depth depends on board weight, the angle of tow and the nature of the substratum (Jones, 1992). Sediment recovery time and infilling will depend on local hydrodynamics and the substratum. Neither Brissopsis lyrifera or Amphiura chiajei are targeted for collection or harvesting, but Nephrops norvegicus is the target of a large commercial fishery. For example, dredging operations were shown to affect large infaunal and epifaunal species, decrease sessile polychaetes and reduce numbers of burrowing heart urchins (Eleftheriou & Robertson, 1992). In a study on the effects of otter trawling for Nephrops norvegicus on the benthos of locations in the Irish Sea and Scottish sea lochs, Ball et al. (2000a) reported a reduction in the abundance of large-bodied and fragile organisms such as Brissopsis lyrifera and Amphiura chiajei and suggested that these species are particularly unlikely to be resistant of trawling disturbance. An altered but stable community resulted, composed of fewer species and reduced faunal diversity, and primarily of small polychaetes. In a study comparing the responses of marine benthic communities within a variety of sediment types to physical disturbance, Dernie et al. (2003) found that mud habitats had an ‘intermediate’ recovery time (compared to clean sand communities, which had the most rapid recovery rate, and muddy sand habitats had the longest recovery times).

Sensitivity assessment. Due to the nature of this pressure, it is highly likely that a large amount of the sediment would be removed along with the biological community, resulting in the removal of the biotope. Disturbance effects may be particularly apparent in more sheltered, stable habitats, than in more disturbed mobile sediments (Kaiser & Spencer, 1996). Resistance is, therefore, assessed as 'None' and resilience as 'Medium' with a sensitivity of 'Medium' to the extraction of substratum to 30 cm.

Abrasion / disturbance of the surface of the substratum or seabed

Benchmark. Damage to surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

SS.SMu.CFiMu.BlyrAchi may be affected by fishing activity in areas such as the northern Irish Sea, where the community may also contain Nephrops norvegicus (Mackie et al., 1995).

Populations of Brissopsis lyrifera are likely to be reduced owing to damage inflicted on the 'test' by the fishing gear, and broken tests may be seen on the seabed (E.I.S. Rees, M. Costello, pers comm. to Connor et al., 2004). Similar evidence has been reported for other heart urchins. For example, Houghton et al. (1971), Graham (1955), de Groot & Apeldoorn (1971) and Rauck (1988) refer to significant trawl-induced mortality of heart urchin Echinocardium cordatum. A substantial reduction in the numbers of Brissopsis lyrifera due to physical damage from scallop dredging was reported by Eleftheriou & Robertson (1992). Overall, species with brittle, hard tests are regarded as sensitive to impact with scallop dredges (Kaiser & Spencer, 1995; Bradshaw et al., 2000; Bergman & van Santbrink, 2000).

Brittlestars have fragile arms that are likely to be damaged by abrasion or physical disturbance. However, brittlestars can tolerate considerable damage to arms and even the disk without suffering mortality and are capable of arm and even some disk regeneration (Sköld, 1998). Amphiura chiajei burrows in the sediment and extends its arms across the sediment surface to feed. Ramsay et al. (1998) suggest that Amphiura species may be less susceptible to beam trawl damage than other species of echinoid or tube dwelling amphipods and polychaetes. Bergman & Hup (1992) for example, found that beam trawling in the North Sea had no significant direct effect on small brittlestars. Bradshaw et al. (2002) noted that the brittlestars Ophiocomina nigra, Ophiura albida and Amphiura filiformis had increased in abundance in a long-term study of the effects of scallop dredging in the Irish Sea.

Sköld et al. (2018) showed that Amphiura chiajei abundance increased to an extent from trawling intensities of up to five trawls per year, beyond which its abundance declined. In the five years following the establishment of a nearby MPA where trawling was prohibited, the abundance of Amphiura chiajei declined from more than 20 individuals/m² to around 10 individuals/m² (Sköld et al., 2018). While this overall decline was statistically insignificant, the authors suggest it is still a meaningful decline due to the stability of the abundances of these two species in the trawled sites. It was presumed that this was due to reduced pressure on predatory fish and crustaceans, which were target species of the fishery. Sköld et al. (2025) later reported significant declines in Amphiura abundance and biomass in the 12 years following the cessation of trawling in the Kattegat, with reductions in abundances estimated at 45% for Amphiura chiajei.

The infaunal position occupied by species in this biotope may provide some protection from abrasion at the surface only. However, burrow structures may collapse and flatten other small-scale habitat features, recovery from which may result in some subsequent energy costs associated (Tillin & Hull, 2013a).

Kaiser et al. (2006) undertook a meta-analysis of different fishing gears on a range of habitats. The authors concluded that the footprint of the impact and the recovery of communities varied with gear and habitat types. For example, beam trawling and scallop dredging had significant negative short-term impacts in sand and muddy-sand habitats; and mud habitats were shown to have substantial initial impacts by otter trawling, but the effects tended to be short-lived with an apparent long-term positive post-trawl disturbance response from the increase of small-bodied fauna.

Furthermore, SS.SMu.CFiMu.BlyrAchi occurs in silty muds (Connor et al., 2004: JNCC, 2015, 2022). Abrasion events caused by a passing fishing gear or scour by objects on the seabed surface are likely to have marked impacts on the substratum and cause turbulent re-suspension of surface sediments. When used over fine muddy sediments, trawls are often fitted with shoes designed to prevent the boards digging too far into the sediment (M.J. Kaiser, pers. obs., cited in Jennings & Kaiser, 1998). The effects may persist for variable lengths of time depending on tidal strength and currents and may result in a loss of biological organization and reduce species richness (Hall, 1994; Bergman & van Santbrink, 2000; Reiss et al., 2009) (see 'change in suspended solids' and 'smothering' pressures).

Sensitivity assessment. Although burrowing life habits may provide some protection from damage by abrasion at the surface, the fragile tests and arms of Brissopsis lyrifera and Amphiura chiajei, respectively, the characterizing species in this biotope, are likely to be adversely affected during abrasion events. Furthermore, the nature of the soft sediment where they occur means that objects causing abrasion, such as fishing gears (including pots and creels) are likely to penetrate the surface and cause further damage to the characterizing species. Resistance is therefore assessed as ‘Low’ and resilience as ‘Medium’, so sensitivity is assessed as ‘Medium’.

Penetration or disturbance of the substratum subsurface

Benchmark. Damage to sub-surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

The two key species in the biotope, Brissopsis lyrifera and Amphiura chiajei, are infauna found close to the sediment surface. The biotope occurs in silty muds (Connor et al., 2004) so penetrative activities (e.g. anchoring, scallop or suction dredging) and damage to the seabed’s sub-surface is likely to remove and/or damage the infaunal community, including the characterizing species, given the fragility of the tests and that bottom fishing gears penetrate deeper into softer sediments (Bergman & van Santbrink, 2000). Overall, species with brittle, hard tests are regarded to be sensitive to impact with scallop dredges (Kaiser & Spencer, 1995; Bradshaw et al., 2000; Bergman & van Santbrink, 2000).

Furthermore, penetrative events caused by a passing fishing gear are also likely to have marked impacts on the substratum and cause turbulent re-suspension of surface sediments (see abrasion pressure). When used over fine muddy sediments, trawls are often fitted with shoes designed to prevent the boards digging too far into the sediment (M.J. Kaiser, pers. obs., cited in Jennings & Kaiser, 1998). The effects may persist for variable lengths of time depending on tidal strength and currents and may result in a loss of biological organization and reduce species richness (Hall, 1994; Bergman & van Santbrink, 2000; Reiss et al., 2009) (see 'change in suspended solids' and 'smothering' pressures).

Sensitivity assessment: The biotope could be lost or severely damaged, depending on the scale of the activity (see abrasion pressure). Therefore, a resistance of 'None' is suggested. Resilience is probably 'Medium', and therefore the biotope’s sensitivity to this pressure is likely to be 'Medium'.

Changes in suspended solids (water clarity)

Benchmark. A change in one rank on the WFD (Water Framework Directive) scale e.g. from clear to intermediate for one year. Further detail

Evidence

Clogging of feeding apparatus by suspended sediment is not a consideration for the characterizing species of this biotope. Brissopsis lyrifera and Amphiura chiajei are burrowing infauna and non-selective surface and sub-surface deposit feeders. For most benthic deposit feeders, food is suggested to be a limiting factor for body and gonad growth, at least between events of sedimentation of fresh organic matter (Hargrave, 1980; Tenore, 1988). Consequently, increased organic matter in suspension that is deposited may become incorporated into sediments via bioturbation and may enhance food supply. A decrease in the suspended sediment and hence siltation may reduce the flux of particulate material to the seabed. Since this includes organic matter the supply of food to the biotope would probably also be reduced. Although characteristically a sub-surface deposit feeder, Brissopsis lyrifera has been observed to increase its surface feeding (apical tuft becomes visible) activity after addition of organic matter to the sediment surface and utilized the material for growth (Hollertz et al., 1998; Hollertz, 1998). This suggests that an increase in siltation may be beneficial to the population.

Where a change in suspended solids results in increased turbidity and change of light, the community is unlikely to be directly affected. However, increased turbidity may hinder predation by visual predators such as Nephrops norvegicus, dab Limanda limanda, and haddock Melanogrammus aeglefinus upon Amphiura chiajei. There may be some increased energetic costs experienced by certain species, associated with increased turbidity, but effects are not likely to be significant. The community is also unlikely to be directly affected by increased light penetration of the water column caused by a decrease in turbidity. Greater light penetration of the water column may improve primary production by phytoplankton in the water column and contribute to secondary productivity via the production of detritus from which the community may benefit.

Sensitivity assessment: An increase in the suspended matter settling out from the water column to the substratum may increase food availability. On the other hand, decreased siltation is unlikely to affect the mainly deposit feeding community that occurs in SS.SMu.CFiMu.BlyrAchi. Resistance and resilience are assessed as 'High', and the biotope, therefore, is considered 'Not Sensitive' at the pressure benchmark.

Smothering and siltation rate changes (light)

Benchmark. ‘Light’ deposition of up to 5 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

Little or no direct evidence for the effects of siltation on the characterizing species of this biotope, Amphiura chiajei and Brissopsis lyrifera, was found.

Last et al. (2011) buried Ophiura ophiura individuals under three different depths of sediment; shallow (2 cm), medium (5 cm) and deep (7 cm). The results indicated that Ophiura ophiura is highly tolerant of short-term (32 days) burial events, with less than 10% mortality of all buried specimens. This is largely a reflection of the ability of the species to re-emerge from all depths across all sediment fractions tested. Survival of specimens that remained buried was low, with 100% mortality of individuals that remained buried after 32 days. The experiments utilized three different fractions of kiln dried, commercially obtained marine sediment: coarse (1.2 to 2.0 mm diameter), medium fine (0.25 to 0.95 mm diameter) and fine (0.1 to 0.25 mm diameter).

Bijkerk (1988, results cited from Essink, 1999) indicated that the maximal overburden through which Echinocardium cordatum could migrate was approximately 30 cm in sand. No further information was available on the rates of survivorship or the time taken to reach the surface. Brissopsis is a burrower, adapted to life within sediments and therefore likely to be able to move within sediments, although the character of the overburden will determine some degree of the impact. However, there may be some energetic cost expended to either re-establish burrow openings in the case of Nephrops norvegicus or to self-clean feeding apparatus because of this pressure, though this is not likely to be significant. Both Brissopsis lyrifera and Amphiura chiajei live buried in muddy sediments up to 2 to 5 and 6 cm deep, respectively. Being adapted for burrowing means these species are likely to resist additional fine sediment. However, it should be remembered that smothering by impermeable or viscous materials is likely to have some effect on the animals, e.g. by causing de-oxygenation.

Puig et al. (2015) demonstrated that bottom trawling in the La Fonera Canyon, Mediterranean, has increased sedimentation rates to around 2.4 cm per year. The study observed large densities of Brissopsis lyrifera, which had colonized the newly deposited fine-grained material in the lower canyon, indicating a high tolerance to sustained sedimentation. Additional paleontological evidence found fossil material of Brissopsis pohangensis, recovered as a dense assemblage (428 Brissopsis pohangensis specimens) preserved in a two-dimensionally flattened state, indicating a rapid, catastrophic sedimentation event, in which death and burial of Brissopsis pohangensis occurred simultaneously (Lee & Kong, 2025). Lee & Kong (2025) suggest that while Brissopsis spp. can usually tolerate gradual burial, it is sensitive to sudden mass sedimentation, where the species may be buried too quickly to respond or to re-establish blocked respiratory funnels, leading to mortality. However, the nature of the mass sedimentation event is unclear.

Sensitivity assessment. The characterizing species in this biotope are burrowers and therefore likely to be able to move within the sediment deposited as a result of a deposition of 5 cm of sediment. Resistance is assessed as 'High' and resilience as 'High' (by default), and the biotope is considered 'Not Sensitive' at the pressure benchmark.

Smothering and siltation rate changes (heavy)

Benchmark. ‘Heavy’ deposition of up to 30 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

Little or no direct evidence for the effects of siltation on the characterizing species of this biotope, Amphiura chiajei and Brissopsis lyrifera, was found.

 Last et al. (2011) buried Ophiura ophiura individuals under three different depths of sediment; shallow (2 cm), medium (5 cm) and deep (7 cm). The results indicated that Ophiura ophiura is highly tolerant of short-term (32 days) burial events, with less than 10% mortality of all buried specimens. This is largely a reflection of the ability of the species to re-emerge from all depths across all sediment fractions tested. Survival of specimens that remained buried was low, with 100% mortality of individuals that remained buried after 32 days. The experiments utilized three different fractions of kiln dried, commercially obtained marine sediment: coarse (1.2 to 2.0 mm diameter), medium fine (0.25 to 0.95 mm diameter) and fine (0.1 to 0.25 mm diameter).

Bijkerk (1988, results cited from Essink, 1999) indicated that the maximal overburden through which Echinocardium cordatum could migrate was approximately 30 cm in sand. No further information was available on the rates of survivorship or the time taken to reach the surface. Brissopsis is a burrower, adapted to life within sediments and therefore likely to be able to move within sediments although the character of the overburden will determine some degree of the impact. However, there may be some energetic cost expended to either re-establish burrow openings in the case of Nephrops norvegicus or to self-clean feeding apparatus as a result of this pressure, though this is not likely to be significant. Both Brissopsis lyrifera and Amphiura chiajei live buried in muddy sediments up to 2 to 5 and 6 cm deep, respectively. Being adapted for burrowing means these species are likely to resist additional fine sediment. However, it should be remembered that smothering by impermeable or viscous materials is likely to have some effect on the animals, e.g. by causing de-oxygenation.

Studies by Maurer et al. (1986) analysed the ability to vertically migrate and survival responses of three major taxa (polychaetes, crustacean and molluscs) when exposed to a simulated disposition of dredged materials (0 to 40 cm). Their results suggested that there was evidence of synergistic effects on burrowing activity and mortality with changes in time of burial, sediment depth, sediment type and temperature. Significant mortality was observed among all taxa under the maximum overburden by sand or fine sediment with varying contents of silt-clay.

Puig et al. (2015) demonstrated that bottom trawling in the Le Fonera canyon, Mediterranean, has increased sedimentation rates to around 2.4 cm per year. The study observed large densities of Brissopsis lyrifera, which had colonized the newly deposited fine-grained material in the lower canyon, indicating a high tolerance to sustained sedimentation. Additional paleontological evidence found fossil material of Brissopsis pohangensis, recovered as a dense assemblage (428 Brissopsis pohangensis specimens) preserved in a two-dimensionally flattened state, indicating a rapid, catastrophic sedimentation event, in which death and burial of Brissopsis pohangensis occurred simultaneously (Lee & Kong, 2025). Lee & Kong (2025) suggest that while Brissopsis spp. can usually tolerate gradual burial, it is sensitive to sudden mass sedimentation, where the species may be buried too quickly to respond or to re-establish blocked respiratory funnels, leading to mortality. However, the nature of the mass sedimentation event is unclear.

Sensitivity assessment: The characterizing species in this biotope are burrowers and therefore likely to be able to move within the deposited sediment. However, a deposition of 30 cm of fine sediment is likely to result in a significant overburden of the infaunal species and, as a result, there may be some mortality of the characterizing species. Resistance is therefore assessed as 'Medium' and resilience as 'High' and the biotope is considered to have a 'Low' sensitivity at this pressure benchmark.

Litter

Benchmark. The introduction of man-made objects able to cause physical harm (surface, water column, seafloor or strandline). Further detail

Evidence

Not assessed.

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT. Further detail

Evidence

Evidence on the effect of electromagnetic fields (EMFs) on benthic organisms is still severely lacking. There have been no studies examining the effect of EMFs on macroalgae. Some studies have investigated the effect of anthropogenically induced EMFs on benthic invertebrates at intensities ranging between 2 nT and 40 mT, which is often much higher than in-situ measurements from subsea cables. While some report changes to behaviour, physiology, reproduction, development, immunology, cytotoxicity and orientation, others demonstrate no effect from exposure to the EMF (Albert et al., 2020; Hutchison et al., 2020), depending on the study species and duration and intensity of exposure. There have been no studies investigating the effect of EMFs at the population or community level for benthic organisms.  

Sensitivity assessment. Given the lack of data at the level of individual biotopes, resistance and resilience to EMFs cannot be robustly assessed. Sensitivity is therefore recorded as 'Insufficient evidence'. 

Underwater noise changes

Benchmark. MSFD indicator levels (SEL or peak SPL) exceeded for 20% of days in a calendar year. Further detail

Evidence

No Evidence was available on which to assess this pressure.

Introduction of light or shading

Benchmark. A change in incident light via anthropogenic means. Further detail

Evidence

Since 2016, research on artificial light at night (ALAN) has expanded considerably in the marine and coastal environment. Light was previously assumed to be of low ecological significance in subtidal and intertidal habitats, but there is now evidence that ALAN is widespread in the marine environment, with biologically relevant levels of light penetrating to depths of up to 50m (Davies et al., 2020; Smyth et al., 2021). ALAN can alter biological processes across taxa and at multiple levels of organisation. Documented responses include disruption of diel and circalunar rhythms, changes in activity and foraging, altered predator–prey interactions, shifts in community composition, and impacts on algal growth and phenology (Davies et al., 2014, 2015; Gaston et al., 2017; Tidau et al., 2021; Lynn et al., 2022; Marangoni et al., 2022; Miller & Rice, 2023; Ferretti et al., 2025). Evidence for benthic habitats and assemblages specifically is beginning to emerge (e.g. Trethewy et al., 2023; Schaefer et al., 2025), but remains limited and fragmented, often focusing on single taxa or short-term experiments. Mortality thresholds, long-term consequences, and responses at the biotope scale are rarely addressed, and there are major gaps around indirect effects such as trophic cascades or habitat modification. 

Sensitivity assessment. SS.SMu.CFiMu.BlyrAchi is a sublittoral biotope (Connor et al., 2004) not characterized by the presence of primary producers and is, therefore, not directly dependent on sunlight. Given the rapid expansion of the evidence base but the continuing lack of data at the level of individual biotopes, resistance and resilience cannot be robustly assessed. Sensitivity is therefore recorded as ‘Insufficient evidence’. 

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion. Further detail

Evidence

Not relevant. Barriers and changes in tidal excursion are not relevant to biotopes restricted to open waters.

Death or injury by collision

Benchmark. Injury or mortality from collisions of biota with both static or moving structures due to 0.1% of tidal volume on an average tide, passing through an artificial structure. Further detail

Evidence

Not relevant to seabed habitats.

Visual disturbance

Benchmark. The daily duration of transient visual cues exceeds 10% of the period of site occupancy by the feature. Further detail

Evidence

Although some species within the community have visual perception e.g. Nephrops norvegicus, detecting the presence of boats or machinery, is likely to be beyond their visual acuity. Additionally, some response to visual disturbance has been detected in echinoderms. However, Brissopsis lyrifera lives buried in muddy substrata up to 10 cm deep thus visual disturbance was not considered relevant to this species.

Brittlestars exhibit a wide range of responses to light intensity, from a largely indifferent behaviour to pronounced colour changes and rapid escape behaviour. Aizenberg et al. (2001) reported that certain calcite crystals used by brittlestars for skeletal construction are also a component of a specialised photosensory organ. However, these structures are absent in light indifferent species. Thus, Amphiura chiajei may have visual perception but is likely to have poor visual acuity.

Sensitivity assessment: No adverse effects are expected for the characterizing species of this biotope as a result of this pressure. Resistance and resilience are, therefore, assessed as High and the biotope considered Not Sensitive to visual disturbance at the pressure benchmark level.

Genetic modification & translocation of indigenous species

Benchmark. Translocation of indigenous species or the introduction of genetically modified or genetically different populations of indigenous species that may result in changes in the genetic structure of local populations, hybridization, or change in community structure. Further detail

Evidence

The key characterizing species in the biotope are not cultivated in the British Isles or likely to be translocated. This pressure is therefore considered Not Relevant.

Introduction of microbial pathogens

Benchmark. The introduction of relevant microbial pathogens or metazoan disease vectors to an area where they are currently not present (e.g. Martelia refringens and Bonamia, Avian influenza virus, viral Haemorrhagic Septicaemia virus). Further detail

Evidence

The occurrence of the ascothoracidan parasite Ulophysema öresundense (Brattström) has been observed in the body cavity of Brissopsis lyrifera (Brattström, 1946). This parasite may cause sexual castration but no further information concerning the effect of this parasite on the population was found. Takano et al., (2020) described a new parasitic gastropod species Heliella seisuimaruae attached to Brissopsis luzonica in Japan, representing the first host of the genus Heliella. No evidence of impacts to Brissopsis has been reported.

The only major biological agent known to affect a species in this biotope is the dinoflagellate parasite, Hematodinium spp., now prevalent in Nephrops norvegicus populations from the west of Scotland, Irish Sea and North Sea (Field et al., 1992). The Hematodinium parasite occurs in the blood and connective tissue spaces and appears to cause death in the host by blocking the delivery of oxygen to the host's tissues (Taylor et al., 1996). Heavily infested animals become moribund, spend more time out of their burrows and are probably less able to evade capture by predators or fishing gear. However, the ecological consequences of this infestation are unknown, and evidence suggests that the Nephrops stocks have not been seriously affected (Hughes, 1999b).

No pathogens are known to affect Amphiura chiajei.

Sensitivity assessment. No evidence of losses of this biotope due to disease was found, and it is likely that microbial pathogens will have only a minor possible impact on this biotope. Resistance and resilience are therefore assessed as 'High' and the biotope assessed as 'Not Sensitive'.

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

Direct, physical impacts from harvesting are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of target species on this biotope. Although Brissopsis lyrifera and Amphiura chiajei are not targeted by fisheries, dredging operations may adversely affect these species by removal or damage.

Nephrops norvegicus is now one of the most valuable shellfish resources in the north-eastern Atlantic (Hughes, 1998b), and is harvested by static and mobile gears. No evidence was found for the proportion of the population that is removed by targeted harvesting. Video studies have found that only a low proportion (circa 5%) of Nephrops that approached creels entered them (Bjordal, 1986; Adey, 2007, cited in Ungfors et al., 2013). Factors that govern emergence will influence catch rates as only individuals that have emerged from burrows will be caught by trawl hauls. The degree of emergence from burrows for feeding or mating appears to be mainly governed by light intensities and therefore depends on factors such as time of day and season and varies between populations at different depths (Katoh et al., 2013). Experimental trawling (Ameyaw-Akumfi & Naylor, 1987) to evaluate catch rates showed that catchability varied between vessels in the same area and that catch rates were strongly linked to tidal cycles with increased catch rates at spring rather than neap tides. Catch rates differed between genders (Ungfors et al., 2013 and references therein), berried females tend to stay within burrows and are rarely caught in trawls (Aguzzi & Sarda, 2008, cited in Katoh et al., 2013).

Furthermore, commercial fisheries may result in discarding of damaged or undersized target species. This will increase the available food supply but may also attract mobile predators and scavengers including fish and crustaceans, which may alter predation rates in the biotope.

Sensitivity assessment: The biotope supports Nephrops norvegicus, a species targeted by fisheries, lending the characterizing species of this biotope vulnerable to damage or removal by fishing gears. However, only the biological and ecological effects on the biotope are assessed under this pressure. Records indicate that Nephrops norvegicus does not occur in all examples of the biotope (Connor et al., 2004) suggesting that that biotope’s main character and component species are unlikely to be adversely affected by the targeted removal of Nephrops norvegicus. Resistance is therefore assessed as 'High' and resilience as 'High' (by default), so the biotope is considered 'Not Sensitive' to removal of targeted species.

Removal of non-target species

Benchmark. Removal of features or incidental non-targeted catch (by-catch) through targeted fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

Direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures, while this pressure considers the ecological or biological effects of by-catch. Species in this biotope, including the characterizing species Brissopsis lyrifera and Amphiura chiajei, may be damaged or directly removed by static or mobile gears that are targeting other species, with reports of high levels of mortality of both characterizing species (see abrasion and penetration of the seabed pressures). Where heavy demersal fishing occurs populations of Brissopsis lyrifera may be reduced owing to damage inflicted to the 'test' by the fishing gear. Broken tests may be seen on the sea bed (E.I.S. Rees, M. Costello pers. comm. in Connor et al., 2004). Munday (1993) observed that 99% of Amphiura chiajei showed evidence of arm tip regeneration in the population off Killary Harbour. Whilst benthic trawling may contribute to arm damage, sub-lethal levels of predation appeared to be the main causative factor for regeneration and was a persistent experience.

Furthermore, commercial fisheries may discard damaged or dead non-target species, which could result in increased available food supply to detritus feeding such as the characterizing species of this biotope that may have survived in the area targeted by fisheries, but may also attract mobile predators and scavengers including fish and crustaceans which may alter predation rates in the biotope.

Sensitivity assessment. The evidence suggests that some loss of the characterizing species is likely to occur as a result of unintentional removal. Removal of Brissopsis lyrifera and Amphiura chiajei, the characterizing species of this biotope, would result in the biotope being lost. Thus, the biotope is considered to have 'Low' resistance to this pressure and to have 'Medium' resilience, resulting in the sensitivity being judged as 'Medium'.

The American slipper limpet, Crepidula fornicata

Evidence

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and the Essex coast in 1887 to 1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 1999, 2018; Hinz et al., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). It ranges from the Baltic Sea, the Kattegat and Skagerrak, the North Sea coasts of the UK, Germany, and Belgium, through the English Channels and into the Irish sea coasts of Ireland and south Wales with records in east and west Scotland, Northern Ireland, northwest France, Spain and south into the Mediterranean (NBN, 2023; OBIS, 2023).

Abundances at its northern and southern extremes may be low but densities in UK and France are often over 1000/m2 and it may carpet the seafloor in the Solent and Essex. In the UK, it was reported to reach abundances of >1000/m2 (max. 2,748/ m²) in the Milford Harbour Waterway (MHW) (Bohn et al., 2012), 84 /m² in Portsmouth, 174/m² in Langstone and 306/m² in Chichester harbours in 2017 (Helmer et al., 2019). In France, it has been reported to reach >4,700/m2 in the Bay of Marennes-Oleron, 11.6 tonnes/ha in Bay of Mont-Saint-Michel, 8.2 tonnes/ha in the Bay of Brest and 2.8 tonnes/ha in the Bay of Saint-Brieuc (Blanchard, 2009; Bohn et al., 2012, 2015; Powell-Jennings & Calloway, 2018).

Crepidula fornicata is recorded from shallow, sheltered bays, lagoons and estuaries or the sheltered sides of islands, in variable salinity (from 18 to 40) although it prefers ca 30 (Tillin et al., 2020). Larvae require hard substrata for settlement. It prefers muddy gravelly, shell-rich, substrata that include gravel, or shells of other Crepidula, or other species e.g., oysters, and mussels. It is highly gregarious and seeks out adult shells for settlement, forming characteristic ‘stacks’ of adults. But it also recorded from rock, artificial substrata, and Sabellaria alveolata reefs (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015; Tillin et al., 2020).   

It is recorded from the lower intertidal to ca 160 m in depth but it most common in the shallow subtidal and low water springs (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 2023; Tillin et al., 2020). Bohn et al. (2012, 2013a, 2013b, 2015) suggested that extreme conditions in intertidal limited its upward distribution due to early post-settlement mortality. It reached its highest densities in the lower shore (below ca 0.7 m) and was absent from high tidal level (ca 1.8 m) in the MHW (Bohn et al., 2015). Bohn et al. (2013b) noted that Crepidula spat in their experimental intertidal panels suffered high mortality 78 to 100% during emersion by low water spring tides. Thieltges et al. (2003) noted that Crepidula abundance at the intertidal to subtidal transition zone (ca 21 / m²) was significantly higher than in the upper, mid, and lower intertidal ca <3 / m²). Similarly, Diederich & Pechenik (2013) noted that Crepidula densities were not significantly different in the low intertidal (+0.2 m) and shallow subtidal (-1 m) but became lower at +0.4 and were absent above +0.6 m in Bissel Cove, Rhode Island where the mean high water was +1.38 m. They reported that intertidal adults experienced temperatures of ca 42°C, which were 15°C higher than subtidal adults. However, there was no significant difference in the tolerance of subtidal and intertidal adults with a lethal range of 33 to 37°C after 3 hours in the laboratory. Diederich & Pechenik, (2013) suggested that adult Crepidula were living close to their upper thermal limit in Rhode Island and would be driven into the subtidal due to climate change. Diederich et al. (2015) reported that most juvenile Crepidula died after aerial exposure under laboratory conditions (20°C, 75% relative humidity), while adults from the intertidal and subtidal survived (26°C, 75% relative humidity). Franklin et al. (2023) noted that the body mass index of adult Crepidula did not decrease significantly in winter months in New Hampshire, USA, but did decrease in spring and summer, probably due to its investment in reproduction.

The availability of hard substrata (e.g., gravel) may only restrict initial colonization as higher densities of Crepidula function as substrata for subsequent colonization (Thieltges et al., 2004; Blanchard, 2009). However, Bohn et al. (2015) noted that Crepidula occurred at low density or was absent in areas of homogenous fine sediment and areas dominated by boulders. Bohn et al. (2015) suggested that wave action (exposure) probably prevented the establishment of large numbers of Crepidula in high-energy areas. Blanchard (2009) noted that sandy areas in the Bay of Saint-Mont Michel were not colonized by Crepidula because of surface sand mobility. Thieltges et al. (2003) also noted that storm events removed some clumps of mussels and presumably Crepidula onto tidal flats where they disappeared, which caused their abundance to fluctuate. Similarly, Crepidula was absent from sandy substrata in Swansea Bay but was most abundant in the shelter of the breakwater at the Swansea east site (Powell-Jennings & Calloway, 2018). Powell-Jennings & Calloway (2018) noted that Crepidula is killed by sudden burial and, possibly, burial due to deposition, which could mitigate Crepidula density.

Crepidula fornicata larvae require hard substrata for settlement. It prefers muddy gravelly, shell-rich, substrata that include gravel, or shells of other Crepidula, or other species e.g., oysters, and mussels. It is highly gregarious and seeks out adult shells for settlement, forming characteristic ‘stacks’ of adults. But it also recorded from rock, artificial substrata, and Sabellaria alveolata reefs (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Tillin et al., 2020). Close examination of the literature shows that evidence of its colonization and density on bedrock in the infralittoral or circalittoral was lacking. Tillin et al. (2020) suggested that Crepidula could colonize circalittoral rock due to its presence on tide-swept rough grounds in the English Channel (Hinz et al., 2011). However, Hinz et al. (2011) reported that Crepidula fornicata only dominated one assemblage (with an average of 181 individuals per trawl) on gravel substratum with boulders. Bohn et al. (2015) noted that Crepidula occurred at low density or was absent in areas dominated by boulders, and Bohn et al. (2013a, 2013b, 2015) and Preston et al. (2020) showed that while Crepidula could settle on slate panels or ‘stone’ it preferred shell, especially that of conspecifics.   

Sensitivity assessment. The circalittoral mud characterizing this biotope is likely to be unsuitable for the colonization by Crepidula fornicata due to substratum (fine mud) and depth (Tillin et al., 2020). Crepidula has been recorded from the lower intertidal to ca 160 m in depth but is most common in the shallow subtidal above 50 m (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 2023; Tillin et al., 2020). In addition, Crepidula requires some hard substratum (stones, gravel or shells) to successfully settle, which is lacking or rare in this fine mud biotope. Therefore, it is unlikely to colonize the deeper and more wave exposed examples of the biotope. However, boulders may occur in a few examples of the biotope and may allow Crepidula to obtain a foothold. The silted conditions would probably limit its abundance, so resistance is assessed as ‘Medium’ as a precaution. Hence, resilience is assessed as ‘Very low’ and sensitivity as ‘Medium’, but with ‘Low’ confidence due to the lack of direct evidence of colonization.

The carpet sea squirt, Didemnum vexillum

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009).

Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024).

In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Michin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

Although a widespread invader, Didemnum vexillum has a limited ability for natural dispersal since the pelagic larvae remain in the water column for a short time (up to 36 hours). Therefore, it has a short dispersal phase that can allow the species to build localized populations (Herborg et al., 2009; Vercaemer et al., 2015; Holt, 2024). However, Bullard et al. (2007) suggested that Didemnum vexillum can form new colonies asexually by fragmentation. Colonies can produce long tendrils from an encrusting colony, which can fragment, disperse and settle, attaching to suitable hard substrata elsewhere (Bullard et al., 2007; Lambert, 2009; Stefaniak & Whitlatch, 2014). A fragmented colony can spread naturally for up to three weeks transported by ocean currents, attached to floating seaweed, seagrass or other floating biota, or as free-floating spherical colonies (Bullard et al., 2007; Lengyel et al., 2009; Stefaniak & Whitlatch, 2014; Holt, 2024). Fragments can reattach to suitable substrata within six hours of contact. Fragments have the potential to disperse around 20 km before reattachment (Lengyel et al., 2009). Valentine et al. (2007a) reported that colonies of Didemnum vexillum enlarged by 6 to 11 times by asexual budding after 15 days and enlarged from 11 to 19 times after 30 days. Valentine et al. (2007a) concluded fragments could successfully grow, survive, and help to spread Didemnum vexillum.

While natural fragmentation of tendrils is thought to allow Didemnum vexillum to invade longer distances and increase its dispersal potential, Stefaniak & Whitlatch (2014) found that only a one tendril out of 80 reattached to the flat, bare substrata used in their study, because tendrils required an extensive (at least eight hour) period of contact to reattach. Stefaniak & Whitlatch (2014) suggested that once fragmented from a colony, the success of tendril reattachment was limited and reattachment was not a major contributor to the invasive success of Didemnum vexillum. However, Stefaniak & Whitlatch (2014) also found that larvae-packed tendril fragments may increase natural dispersal distance, reproduction and invasive success of Didemnum vexillum, and increase the distance larvae can travel. Not all colonies produce tendrils at all locations.

Human-meditated transport via aquaculture facilities, boat hulls, commercial fishing vessels, ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijstra et al., 2007; Griffiths et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009).

Didemnum vexillum was likely introduced into the UK from northern Europe or Ireland via poorly maintained or not antifouled vessels, movement of contaminated shellfish stock and aquaculture equipment, or via marine industries such as oil, gas, renewables and dredging (Holt, 2024). Recent evidence from genetic material suggests human-mediated dispersal, between marinas and shellfish culture sites, is the most likely pathway for connectivity of Didemnum vexillum populations throughout Ireland and Britain (Prentice et al., 2021; Holt, 2024). Didemnum vexillum can disperse away from artificial substrata, invading and colonizing natural substrata in surrounding areas (Tillin et al., 2020). Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. The current evidence of Didemnum vexillum’s ability to spread on natural habitats in this area is sparse and often conflicting, complicated by genetics and its apparent variable habitat preferences and tolerances and its variable ability to adapt to ‘new’ conditions (Holt 2024).

Didemnum vexillum has a seasonal growth cycle that is influenced by temperature (Valentine et al., 2007a). In warmer months (June and July) colonies may be large and well-developed encrusting mats. Populations experience more rapid growth from July to September sometimes continuing into December. Colonies begin to decline in health and ‘die-off’ when temperatures drop below 5°C during winter months from around October to April (Gittenberger, 2007; Valentine et al., 2007a; Herborg et al., 2009). Cold winter months cause colonies to regress and reduce in size, yet they often regenerate as temperatures warm (Griffith et al., 2009; Kleeman, 2009, Mercer et al., 2009), although some populations may not survive winter at all (Dijkstra et al., 2007). The early growth phase, from May to July, is initiated by smaller colonies developing from remnants of colonies that survived the cold winter (Valentine et al., 2007a). The seasonal growth cycle is also likely influenced by location. For example, the Didemnum sp. growth cycle for colonies in Sandwich tide pool (temperature range from -1°C to 24°C, with daily fluctuations), probably does not occur in deep offshore subtidal habitats in Georges Bank (annual temperature range from 4°C to 15°C, and daily fluctuations are minimal) (Valentine et al., 2007a).  Larval release and recruitment typically occur between 14 to 20°C and slow or cease below 9 to 11°C as summer ends (Griffith et al., 2009; Mckenzie et al., 2017). In New Zealand, recruitment occurs from November to July, where highest average temperatures were recorded in February (18 to 22°C) and the lowest average temperatures were recorded in July (9 to 10°C) (Fletcher et al., 2013a). In this New Zealand study, higher water temperatures were associated with a higher level of recruitment (Fletcher et al., 2013a).

Didemnum vexillum requires suitable hard substrata for successful settlement and the establishment of colonies. It can grow quickly and can establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems by its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock gravel, pebble, cobble, or boulders (Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020). 

There are few observations of Didemnum vexillum on soft bottom habitats as evidence suggests it is unable to establish or grow easily on mud, mobile sand or other unstable substrata, and it is vulnerable to smothering by fine sediment (Bullard et al., 2007; Valentine et al., 2007a; Griffith et al., 2009). The species is usually found established in areas where the colony is protected from sedimentation and wave action (Valentine et al., 2007b; Mckenzie et al., 2017; Tillin et al., 2020). For example, at Georges Bank, USA the Didemnum vexillum mats were limited to gravelly areas and unable to colonize the sand ridges that bounded the site, which have a mobile surface that is moved daily by the strong tidal currents (Valentine et al., 2007b). In addition, evidence found the species can also not survive being buried or smothered by coarse or fine grained sediment. Furthermore, in Holyhead marina, Didemnum vexillum colonies were contained in the harbour and established on artificial pontoons, and they were not present on the natural seabed under the pontoon, which is composed of silty mud or on deeper sections of mooring chains that are immersed in mud at low spring tides (Griffiths et al., 2009). 

However, some studies on Georges Bank, USA and Sandwich, Massachusetts observed colonies were able to survive partial covering by sand (Bullard et al., 2007; Valentine et al., 2007a). Gittenberger et al. (2015) reported that Didemnum vexillum was able to overgrow sandy bottom (cited Gittenberger, 2007). In the Netherlands the coastal zone is composed of mud and sand, with only shells as hard substrata. Didemnum sp. remained rare until 1996 when populations quickly expanded and it became a dominant invasive species because of an increase in available hard substrata for colonization after a cold winter between 1995 and 1996 caused a decrease in the abundance of many marine animals (Gittenberger, 2007). Thus, Didemnum vexillum was able to colonize and establish in mud and sand habitats where hard substrata was present.  

Didemnum vexillum has been recorded from less than 1 m to at least 81 m deep (Bullard et al., 2007; Tagliapietra et al., 2012; Tillin et al., 2020). It is abundant across various shore heights, thriving in both nearshore and offshore sites, particularly in subtidal areas. For example, colonies of Didemnum vexillum were dominant at depths between 45 to 60 m, occupying 50 to 90% of available space in two gravelly areas (more than 230 km²) composed of immobile pebble and cobble pavement on Georges Bank fishing ground, USA (Bullard et al., 2007; Valentine et al., 2007b; Lengyel et al., 2009). In addition, patchy mats have been observed covering approximately 1 to 1.5 km² of the pebble cobble seabed, which is interspersed with large boulders and 30 m deep in Long Island Sound, USA (Mercer et al., 2009). In an offshore scallop dredge survey, Didemnum sp. was found attached to cobbles and boulders at 10 to 34 m (Vercaemer et al., 2015). 

Didemnum vexillum’s preference to sheltered conditions, established colonies observed in Georges Bank and Long Island Sound were exposed to moderately strong tidal currents (1 to 2 knots; ca 0.5 to 1 m/s recorded at both sites) that may mobilise sediment (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020). However, Valentine et al. (2007b) describe the substratum as immobile, presumably consolidated gravel, cobbles and pebbles.  

Sensitivity assessment. The presence of silty mud and the lack of hard substrata in this biotope may be unsuitable for the colonization of Didemnum. The moderately wave-exposed examples of the biotope may mitigate or prevent colonization by Didemnum, but the sheltered or very sheltered wave-exposed examples of the biotope may be suitable. However, boulders may occur in a few examples of the biotope and may allow Didemnum to obtain a foothold. The silted conditions would probably limit its abundance, so resistance is assessed as ‘Medium’ as a precaution. Hence, resilience is assessed as ‘Very low’ and sensitivity as ‘Medium’, but with ‘Low’ confidence due to the lack of direct evidence of colonization.

The Pacific oyster, Magallana gigas

Evidence

The Pacific oyster, Magallana (syn. Crassostrea) gigas, is native to warm temperate regions from the northwest Pacific to Japan and northeast Asia, including Cape Mariya (Russia) to Hong Kong (China) (Carrasco & Baron, 2010; GBNNSIP, 2011, 2012). It is a fast-growing and tolerant species that has become a successful invader in the coastal waters of all continents, aside from Antarctica (Wrange et al., 2010; Carrasco & Baron, 2010; Padilla, 2010). Magallana gigas is recognised as a beneficial and important species in aquaculture worldwide (Padilla, 2010). It was initially introduced for aquaculture in Europe and the UK in the 1960s due to a decline in the Portuguese oyster (Crassostrea angulata) and the European flat oyster (Ostrea edulis) (Spencer et al., 1994; GBNNSIP, 2011, 2012; Humphreys et al., 2014 cited in Alves et al., 2021; Hansen et al., 2023).  

Since introduction, the species has invaded and established self-sustaining natural populations throughout Europe from the North Sea, Wadden Sea and Scandinavian coastlines to the Atlantic coastlines of Spain and Portugal, as well as the Mediterranean and Adriatic Sea (Wrange et al., 2010; GBNNSIP, 2011, 2012; Ezgeta-Balic et al., 2019; Spagnolo et al., 2019; Bergstrom et al., 2021; Hansen et al., 2023). In the UK, the species predominantly occurs around the southern and western coastlines (OBIS, 2024; NBN, 2024).  

Shipping activity has also been associated with the introduction of Magallana gigas in the northeastern Adriatic Sea, where it was not introduced for aquaculture (Ezgeta-Balic et al., 2019). It was also suggested that some Magallana gigas populations were established in southwest England from France possibly via fouling on ships (GBNNSIP, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).  

Magallana gigas requires hard substrata for successful settlement and establishment, including littoral rock, bedrock, chalk, bare boulders, cobbles and pebbles and shells (Kochmann et al., 2012, 2013; Mckinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020) because its larvae require hard substrata for successful settlement and development (Mckinstry & Jensen, 2013; Tillin et al., 2020). It also prefers mudflats with mixed sediment composed of shingle and sand, attaching to whatever hard substrata are available within otherwise unsuitable fine muddy sediment (Spencer et al., 1994; Mckinstry & Jensen, 2013; Tillin et al., 2020). Therefore, fine mud sediments without hard substrata (such as small stones, gravel, and shell) are unlikely to be suitable (Tillin et al., 2020).

Invasive populations of Magallana gigas have been found wave-exposed rocky shores to wave-sheltered soft sediment environments and it has been described as a habitat generalist (Troost, 2010; Kochmann et al., 2012, 2013). It has been suggested that recruitment is enhanced, and abundances are higher in wave-sheltered conditions (Robinson et al., 2005; Ruesink, 2007 cited in Teschke et al., 2020; Tillin et al., 2020). Teschke et al. (2020) found the abundance of Magallana gigas was significantly higher at wave-protected sites within the artificial harbours of Helgoland, North Sea, compared to wave exposed sites outside the harbours. The authors suggested that the successful colonization in wave-protected sites could be due to the relative retention of water masses in the harbours that reduces larval drift and whiplash effect on newly settled larvae. In addition, better growth and higher survival rates were observed at wave-protected sites, whereas mortality rates increased at wave exposed sites, due to the wave exposure causing dislodgement or detachment from the settlement substratum (Teschke et al., 2020; Tillin et al., 2020). Similarly, Bergstrom et al. (2021) noted that the occurrence of high densities of both Ostrea edulis and Magallana gigas decreased with increasing wave exposure.  

The majority of the evidence indicates that infralittoral rock and other habitats that occur at depths more than 10 m are unlikely to be suitable for Magallana gigas because it is considered an intertidal and shallow subtidal species rarely recorded below extreme low water (Herbert et al., 2012, 2016; Tillin et al., 2020). However, in suitable situations (e.g. Oosterschelde) it may form beds down to 42 m.  

Sensitivity assessment. The presence of silty mud and lack of hard substrata in this biotope may be unsuitable for the colonisation of Magallana gigas. This biotopes depth range (20 to 100 m) is likely  too deep for the colonization of Magallana gigas, as most evidence indicates that habitats that occur at depths more than 10 m are unlikely to be suitable for colonization. Therefore, resistance is assessed as ‘High’, albeit with low confidence due to no direct evidence of colonization in this biotope. Hence, resilience is assessed as ‘High’ and sensitivity is assessed as ‘Not sensitive’.

Wireweed, Sargassum muticum

Evidence

The depth and sedimentation probably exclude macroalgae from this biotope. Hence, it is unlikely to be colonized by Sargassum. 

Wakame, Undaria pinnatifida

Evidence

The depth and sedimentation probably exclude macroalgae from this biotope. Hence, it is unlikely to be colonized by Undaria.