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

There is no information available about the tolerance of the characterizing species of deep sponges Axinella dissimilis, Axinella infundibuliformis, Phakellia ventilabrum, or Stelligera stuposa to changes in temperature. In the British Isles, each of these species has a mainly southern and western distribution (OBIS, 2025), and in Lough Hyne, a small (∼0.5 km²) lough on the southwest coast of Ireland, sponge communities are observed in waters where temperatures range from 7 to 9°C in winter and 14–18°C in summer (Micaroni et al., 2025). Outside of British waters, all species are found as far south as Spain or in limited areas of the Mediterranean, and as north as southern Norway. Only Phakellia ventilabrum is observed further, as far north as northern Norway and as south as southern Argentina (OBIS, 2025; Van Soest, 2004). In addition, each species was recorded from a sea surface temperature range of 10 to 15°C (OBIS, 2025). All of these species seem to be replaced in the Mediterranean by the very similar species, Axinella polypoides (Howson & Picton, 1997). 

Long-term increases in temperature may cause extension of the British Isles populations, and decreases in temperature may result in population shrinkage. Goodwin et al (2013) noted increases in the abundance of Axinella damicornis and Axinella dissimilis in Northern Ireland over a 20-year period and suggested the increase was due to sea temperature warming (relating to a 0.3 to 0.5°C increase in Northern Irish sea surface temperature between 1850 and 2007).  Berman et al. (2013) monitored sponge communities off Skomer Island, UK, over four years, with all characterizing sponges for this biotope assessed.  Seawater temperature, turbidity, photosynthetically active radiation and wind speed were all recorded during the study. They concluded that, despite changes in species composition, primarily driven by the non-characterizing Hymeraphia stellifera and Halicnemia patera, no significant difference in sponge density was recorded in all sites studied. 

Davison et al. (2019) highlighted the relationship between Porifera aggregations and temperature in the Faroe-Shetland Channel and noted a statistically significant positive relationship between sponge density and temperature range between 400 and 600 m, with the highest sponge densities occurring at depths of greatest temperature range. Aggregations of sponges in the Faroe-Shetland Channel are frequently subjected to water <0°C due to the movement of the pycnocline, but temperature varies between approximately −1 and 9°C and is on average around 3.5°C, and consistently experience 6°C shifts in temperature, with the highest variation being over 9°C (Davison et al. 2019). This temperature range approximately lines up with the evidence found by Graves (2022), who recorded a temperature range of 3 to 8.5°C between 800 and 2,200 m deep at the Rosemary Bank (UK) when studying the sessile benthic Xenophyophore, Syringammina fragilissima. Graves (pers comm, 2022) characterized the thermal niche of Reteporella and Axinellid sponge assemblages across the UK and Ireland as 9.3 to 9.5°C based on their biotope distribution.

Cebrian et al. (2011) conducted four-year surveys of two shallow-water sponge species, Ircinia fasciculata and Sarcotragus spinosulum, in the western Mediterranean Sea. Two severe sponge die-offs (total mortality ranging from 80 to 95% of specimens) occurred in the summers of 2008 and 2009. These events primarily affected Ircinia fasciculata, and a significant positive correlation was observed between elevated temperature and injured specimens.  It was suggested, following in vitro studies of the associated cyanobacteria in increasing temperatures up to those experienced in ‘extreme summers’ of 27°C, that heat-related disappearance of the cyanobacteria in Ircinia fasciculata (a bacteriosponge) was important when considering sponge mortality. 

Puerta et al. (2020) observed the impact of water masses on deep-sea benthic communities in the North Atlantic and noted how marine heatwaves, in their case, a natural increase of 4°C in 24 hrs (up to 12°C) at the Tisler reef, Norway, was followed by a mass mortality event of the sponge Geodia baretti. However, when a subsequent ex situ experiment exposed the same sponge to acute thermal conditions (up to 5°C above ambient temperature for 14 days), it did not induce any mortality. These results suggested that other processes (e.g., low oxygen concentrations, elevated nutrient levels, reduced salinity, and disease) in combination with the heatwave could be responsible for the mortality event in those sponges. Research by Webster et al. (2008, 2011), Webster & Taylor (2012) and Preston & Burton (2015) suggested that many sponges relied on a holobiont of synergistic microbes.  Webster et al. (2011) described a much higher thermal tolerance to sponge larval holobiont when compared with adult sponges.  For example, adult Rhopaloeides odorabile from the Great Barrier Reef has been shown to have a strict thermal threshold of between 31-33°C (Webster et al., 2008) whereas the larvae could tolerate temperatures of up to 36°C with no adverse effects (Webster et al., 2011).

Stevenson et al. (2020) studied the response of the glass sponge Aphrocallistes vastus in response to warming in the Northeast Pacific Ocean. Stevenson et al. (2020) observed that, within one month of warming, sponges ceased pumping (50-60%) and exhibited tissue withdrawal (10-25%). In addition, thermal stress, alongside acidification stress, significantly reduced skeletal stiffness, and warming weakened the skeleton, potentially curtailing reef formation. The data suggested that conditions causing irreversible damage were possible in the field at +0.5°C above current conditions, which indicated that ongoing climate change was a serious and immediate threat to Aphrocallistes vastus, reef-dependent communities, and potentially other sponges.  

Bell et al. (2018) investigated the potential response of sponges to climate change. Although numerous mass sponge mortalities have been reported in association with abnormally high seawater temperatures, it was unclear if these resulted from exceeding the host's thermal threshold, or because of the disruption of functionally important symbiotic partnerships, or infection by opportunistic pathogens. In contrast, Bell et al. (2018) highlighted that other studies had shown sponges to be more tolerant to increased temperature than other benthic organisms. For example, sponge assemblages in Brazil were highly stable before and after the El Niño Southern Oscillation (a 2°C increase in temperature during the El Niño event), despite massive declines in corals and other benthic organisms. However, ocean warming could induce bleaching, such as in the bioeroding sponge Cliona orientalis (an increase of 2.7°C above the local maximum monthly mean), or reduce flow rates and feeding efficiency, such as in the sponge Rhopaloeides odorabile (exposed to a 2°C higher than the average local ambient seawater temperature). Therefore, temperature tolerance appears to vary among sponge species.

Whalan et al. (2008) noted that while the larvae of Rhopaloeides odorabile survived elevated temperatures, the planktonic duration of the larvae was reduced markedly.  Whalan et al. (2008) suggested that the increases in temperature predicted under climate change on the Great Barrier Reef may reduce planktonic duration but result in reduced dispersal and increased population subdivision. However, little research had been undertaken on the larvae of temperate species (Goodwin, pers. comm., 2017).

Sensitivity assessment: No direct evidence was found on mortality due to increases in temperature in the characterizing sponge species. Evidence from similar sponge communities (Goodwin et al., 2013; Berman et al., 2013) detected no effects of seasonal changes in temperature, while increased temperature over 20 years may have been beneficial. The evidence from other sponge communities, including the deep-sea, varied depending on species, and may not be due to the effects of temperature alone, but rather the effects on sponge symbionts or pathogens.  However, it is possible that localised short-term acute changes in temperature could result in some mortality. Therefore, a cautious resistance assessment of ‘Medium’ is applied, albeit with a ‘Low’ confidence due to the lack of evidence. In the event of any mortality, resilience is assessed as ‘Very Low’. Therefore, sensitivity is assessed as ‘Medium’.

Temperature decrease (local)

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

Evidence

There is no information available about the tolerance of the characterizing species of deep sponges Axinella dissimilis, Axinella infundibuliformis, Phakellia ventilabrum, or Stelligera stuposa to changes in temperature. In the British Isles, each of these species mainly has a southern and western distribution (OBIS, 2025), and in Lough Hyne, a small (∼0.5 km²) lough on the southwest coast of Ireland, sponge communities are observed in waters where temperatures range from 7 to 9°C in winter and 14 to 18°C in summer (Micaroni et al., 2025). Outside of British waters, all species are found as far south as Spain or in limited areas of the Mediterranean, and as north as southern Norway. Only Phakellia ventilabrum is observed further, as far north as northern Norway and as south as southern Argentina (OBIS, 2025; Van Soest, 2004). In addition, each species has a sea surface temperature range of 10 to 15°C (OBIS, 2025). All of these species seem to be replaced in the Mediterranean by the very similar species, Axinella polypoides (Howson & Picton, 1997). 

The British Isles are at the northern distribution limit of Axinella dissimilis (Ackers et al., 1992).  Apparent shrinkage of individual sponges (negative growth rate) observed in Lundy in some years was attributed to particularly cold winters, notably between 1985 and 1986 (Hiscock, 1993). Phakellia ventilabrum is distributed from the Arctic to the coast of North Africa (Van Soest, 2004).  However, it occurs only in deeper, colder waters at the Southern part of its range and has a northern / deeper distribution, only occurring in diving depths commonly off Scotland, with scattered records on the west coast of Ireland (Goodwin pers. comm., 2017).

Berman et al. (2013) monitored sponge communities off Skomer Island, UK, over four years, with all characterizing sponges for this biotope assessed. Seawater temperature, turbidity, photosynthetically active radiation and wind speed were all recorded during the study. It was concluded that, despite changes in species composition, primarily driven by the non-characterizing Hymeraphia stellifera and Halicnemia patera, no significant difference in sponge density was recorded in all sites studied. 

Some sponges do exhibit morphological strategies to cope with winter temperatures, e.g. Halichondria bowerbanki goes into a dormant state below 4°C, characterized by major disintegration and loss of choanocyte chambers with many sponges surviving mild winters in more protected areas from where it can recolonize (Vethaak et al., 1992).  Crisp (1964a) studied the effects of an unusually cold winter (1962-3) on the marine life in Britain, including Porifera in North Wales. Whilst difficulty distinguishing between mortality and delayed development was noted, Crisp (1964a) found that Pachymastia johnstonia and Halichondria panicea were wholly or partly killed by frost, and several species appeared to be missing, including Amphilectus fucorum. Others, including Hymeniacidon perleve, were unusually rare, and a few species, including Polymastia boletiformis, were not seriously affected.  No mention was made of the characterizing sponges assessed in this review.  It should be noted that Crisp’s general comments on all marine life state that damage decreased the deeper the habitat. In addition, the extremely cold temperatures recorded in 1962/63 (sea temperatures between 4-6°C colder than the five-year mean over a period of 2 months) are more extreme than the benchmark level for assessment. 

Davison et al. (2019) highlighted the relationship between Porifera aggregations and temperature in the Faroe-Shetland Channel and noted a statistically significant positive relationship between sponge density and temperature range between 400 and 600 m, with the highest sponge densities occurring at depths of greatest temperature range. Aggregations of sponges in the Faroe-Shetland Channel are frequently subjected to water <0°C due to the movement of the pycnocline, but temperature varies between approximately −1 and 9 °C and is on average around 3.5 °C, and consistently experience 6°C shifts in temperature, with the highest variation being over 9 °C (Davison et al. 2019). This temperature range approximately lines up with the evidence found by Graves (2022), who recorded a temperature range of 3 to 8.5°C between 800 and 2,200 m deep at the Rosemary Bank (UK) when studying the sessile benthic Xenophyophore, Syringammina fragilissima. Graves (pers comm, 2022) characterized the thermal niche of Reteporella and Axinellid sponge assemblages across the UK and Ireland as 9.3 to 9.5°C based on their biotope distribution.

Sensitivity assessment: There is evidence of sponge mortality at extremely low temperatures in the British Isles, and shrinkage (negative growth rate in individuals) of Axinella dissimilis has been attributed to particularly cold winters.  It is possible that rapid cooling of 5°C would affect the characterizing sponges. However, this biotope is protected from the effects of acute temperature change due to its depth. Therefore, a cautious resistance assessment of ‘Medium’ is applied, albeit with a ‘Low’ confidence due to the lack of evidence. In the event of any mortality, resilience is assessed as ‘Very Low’.  Sensitivity is therefore assessed as ‘Medium’.

Salinity increase (local)

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

Evidence

CR.HCR.DpSp is a deep circalittoral biotope, and given that Axinella dissimilis is recorded as having a preference for full salinity of 30-40 psu (Jackson, 2008c; Puerta et al., 2020; OBIS, 2025; Micaroni et al., 2025), it is possible that the characterizing species are intolerant of an increase in salinity at the benchmark level (>40). However, no evidence was found.

Salinity decrease (local)

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

Evidence

CR.HCR.DpSp is a deep circalittoral biotope, and given that Axinella dissimilis is recorded as having a preference for full salinity of 30-40 psu (Jackson, 2008c; Puerta et al., 2020; OBIS, 2025; Micaroni et al., 2025), it is likely that the characterizing species are intolerant of a decrease in salinity.

Castric-Fey & Chassé (1991) conducted a factorial analysis of the subtidal rocky ecology near Brest, France and rated the distribution of species from estuarine to offshore conditions.  Dysidea fragilis and Raspailia ramosa were rated as indifferent to this range.  Cliona celata and Pachymatisma johnstonia had a slight preference for more estuarine conditions, while Polymastia mamillaris and Tethya citrina had a slight preference for offshore conditions. Stelligera rigida and Polymastia boletiformis (as Polymastia robusta) were intolerant of the more estuarine conditions.  Mean salinity difference was low (35.1 and 33.8‰, respectively) but with a greater range being experienced in the Inner Rade (±2.4‰ compared with ±0.1‰).  It should be noted that the range of salinities identified in this study does not reach the lower benchmark level.

Sensitivity Assessment: CR.HCR.DpSp is a deep circalittoral group biotope and, combined with evidence of low salinity intolerance in some sponge species, it is likely that the characterizing sponges would be intolerant of a salinity decrease at the benchmark level, however unlikely, e.g. due to localised freshwater effluent. Therefore, resistance is assessed as ‘Low’. In the event of any mortality, resilience is assessed as ‘Very Low’, and sensitivity is assessed as ‘High’. 

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

Riisgard et al. (1993) discussed the low energy cost of filtration for sponges and concluded that passive current-induced filtration may be of insignificant importance for sponges. However, water movement is probably required to ensure the supply of food (particulates and dissolved organic matter) as well as oxygen. The sponges Axinella spp. and Phakellia ventilabrum were recorded in biotopes that experienced moderate to very weak flow (0 to 1.5 m/s), whereas Stelligera stuposa was recorded in biotopes from strong to very weak (0 to 3 m/s) (Connor et al., 2004). In Norway, both Phakellia ventilabrum and Axinella infundibuliformis were primarily observed at sites with a relatively slower horizontal current velocity (0.02 to 0.03 m/s) (Dunlop et al., 2020). In Lough Hyne, a small (∼0.5 km²) lough on the southwest coast of Ireland, sponge communities experience currents reaching >300 cm/s (>3 m/s), and greater sponge growth/recovery was observed at sites with greater water movement, possibly due to sponges using ambient currents to reduce the high energy costs associated with their filtration activity (Micaroni et al., 2025). Furthermore, the dispersal ability of sponge larvae is generally low, which can slow recolonization rates in areas, and very slow current speeds could reduce larval supply to sites (Micaroni et al., 2025). However, this biotope occurs in wave exposed conditions, and although ameliorated by depth, wave action might be a more important source of water movement than tidal streams.

Sensitivity assessment: The biotope is recorded from sites that experience very weak to moderately strong water flow (0 to 1.5 m/s).  It is unlikely that a change at the benchmark level (increase or decrease) would cause mortality in the characterizing sponges. Resistance is therefore assessed as ‘High’, resilience as ‘High’ and sensitivity as ‘Not Sensitive’.

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 this biotope as it is 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

In northern Norway, Phakellia ventilabrum and Axinella infundibuliformis have been observed to dominate sponge communities on wave-exposed circalittoral rock habitat (Dunlop et al., 2020). Similarly, in the Faroe Shetland Channel, sponge abundance was noted to be highest in the region of internal wave activity at the seabed (Eerkes-Medrano et al., 2020). High sponge diversity and aggregations are likely found in areas of high-wave activity, such as near shelf breaks, due to the wave activity providing an abundant and stable food supply to them, with diversity and densities of sponges decreasing away from such areas (Santín et al., 2018).

Roberts et al. (2006) studied deep sponge reef communities (18-20 m) in sheltered and exposed locations in Australia. They reported greater diversity and cover (>40% cover) of sponges in wave-sheltered areas compared with a sparser and more temporal cover in exposed sites (25% cover). Erect sponges dominated the sheltered sites, while encrusting sponges dominated in exposed locations. Erect or massive forms possessing a relatively small basal area relative to volume do poorly in high-energy environments (Wulff, 1995; Bell & Barnes, 2000).

Whilst little evidence for the characterizing sponges could be found, Connor et al. (2004) noted that in shallower conditions with increased wave action, water mixing is more prevalent and the CarSp.PenPor biotope occurs, while DpSp.PhaAxi develops in deeper waters where wave-induced mixing is reduced. CR.HCR.DpSp is exposed to the highest levels of wave exposure (exposed to extremely exposed) (Connor et al. 2004), but the effects of wave action decrease with depth (Hiscock, 1983).  Hiscock (2002) suggested that ‘prolonged Easterly gales in 1985’ might account for the loss of Axinella dissimilis specimens at Lundy.

Sensitivity Assessment: CR.HCR.DpSp is a deep circalittoral biotope complex recorded in extremely wave exposed to wave exposed conditions.  Wave action is probably an important source of water movement energy in the biotope, and may drive internal waves and vertical mixing. However, the direct effects of wave action decrease with depth.

A decrease in wave action may reduce water movement further.  It is uncertain what effect, if any, would result.  Connor et al. (2004) note that an increase in mixing would probably replace the biotope with a CaSp.PenPor biotope.  Hiscock (2002) also noted the mortality of axinellids after storms at Lundy. However, an increase in wave action above extremely exposed is unlikely. In addition, a change in wave action at the benchmark level is not significant compared with the biotope’s natural range. Hence, mortality at the benchmark level (3-5% change in significant wave height) is unlikely, and resistance is therefore assessed as ‘High’, resilience as ‘High’, and the biotope is ‘Not sensitive’ at the benchmark level. 

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

Whilst some sponges, such as Cliona spp. have been used to monitor heavy metals by looking at the associated bacterial community (Marques et al., 2006; Bauvis et al., 2015), no literature on the effects of transition element or organo-metal pollutants on the characterizing sponges could be found. Nevertheless, this pressure is Not assessed but evidence is presented where available.

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. CR.HCR.DpSp is a sub-tidal biotope complex (Connor et al., 2004). Oil pollution is mainly a surface phenomenon; its impact upon circalittoral turf communities is likely to be limited. However, as in the case of the Prestige oil spill off the coast of France, high swell and winds can cause oil pollutants to mix with the seawater and potentially negatively affect sub-littoral habitats (Castège et al., 2014).

Filter feeders are highly sensitive to oil pollution, particularly those inhabiting the tidal zones, which experience high exposure and show correspondingly high mortality, as are bottom bottom-dwelling organisms in areas where oil components are deposited by sedimentation (Zahn et al., 1981). Zahn et al. (1981) found that Tethya lyncurium concentrated BaP (benzo[a]pyrene) to 40 times the external concentration and no significant repair of DNA was observed in the sponges, which, in higher animals, would likely lead to cancers. As sponge cells are not organized into organs, the long-term effects are uncertain (Zahn et al., 1981).

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.

Radionuclide contamination

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

Evidence

'No evidence'.

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

In general, respiration in most marine invertebrates does not appear to be significantly affected until extremely low concentrations are reached. For many benthic invertebrates, this concentration is about 2 ml/l (Herreid, 1980; Rosenberg et al., 1991; Diaz & Rosenberg, 1995). Cole et al. (1999) suggested possible adverse effects on marine species below 4 mg/l and probable adverse effects below 2 mg/l.

Hiscock & Hoare (1975) reported an oxycline forming in the summer months (Jun-Sep) in a quarry lake (Abereiddy, Pembrokeshire) from close to full oxygen saturation at the surface to <5% saturation below ca 10 m.  No Tethya citrina, Kirchenpaueria pinnata, Hymeniacidon pereleve, Polymastia boletiformis or Ascidia mentula were recorded at depths below 10 to 11 m.  Demosponges maintained under laboratory conditions can tolerate hypoxic conditions for brief periods. Gunda & Janapala (2009) investigated the effects of variable oxygen levels on the survival of the marine sponge, Haliclona pigmentifera. Under hypoxic conditions (1.5 to 2.0 ppm O₂), Haliclona pigmentifera with intact ectodermal layers and subtle oscula survived for 42 ± 3 days.  Sponges with prominent oscula, foreign material, and damaged pinacoderm exhibited poor survival (of 1-9 days) under similar conditions. Complete mortality of the sponges occurred within two days under anoxic conditions (<0.3 ppm O₂ = <0.3 mg/l). Bell et al. (2024) studied the stability of shallow water sponges at Lough Hyne, Ireland, and concluded that changes to the deeper subtidal sponge assemblages were possibly driven by local processes associated with deeper water, potentially related to the seasonal oxythermocline (development of a colder, oxygen-poor layer in the deeper areas from northern hemisphere spring) that forms within Lough Hyne. This low-oxygen layer is thought to have a strong influence on the ecology and biology of organisms in the deeper areas of the lough, with a marked decline in the biodiversity of sponges and other organisms below approximately 25 m (Bell et al., 2024). However, explicit testing of Lough Hyne sponges’ oxygen tolerance found sponges to be resilient to short-term oxygen stress, with the focus now being on the presence of hydrogen sulphide as the main driver of change (Bell et al., 2024).

Sensitivity assessment: Whilst some sponges have demonstrated tolerance to short-term hypoxic events, others were excluded below the oxycline at Abereiddy Quarry (Hiscock & Hoare, 1975). Therefore, some members of the community may be lost, and a precautionary resistance assessment of ‘Medium’ is justified, albeit with ‘Low’ confidence. Resilience is ‘Very Low’, and sensitivity is therefore assessed as ‘Medium’.

Nutrient enrichment

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

Evidence

Gochfeld et al. (2012) studied the effect of nutrient enrichment (≤0.05 to 0.07 μM for nitrate and ≤0.5 μM for phosphate) as a potential stressor in the sponge Aplysina caulifornis and its bacterial symbionts and found that nutrient enrichment had no effects on sponge or symbiont physiology when compared to control conditions. This study contradicts findings in Gochfeld et al. (2007), in which Aplysina spp. sponges were virtually absent from a site of anthropogenic stress in Bocas del Toro, Panama, which experienced high rainfall and terrestrial runoff.  The author suggested that whilst this site did include elevated nutrient concentrations, other pressures and stresses could be contributing. Rose & Risk (1985) described an increase in abundance of Cliona delitrix in an organically polluted section of Grand Cayman fringing reef affected by the discharge of untreated faecal sewage and reported a positive correlation between the two. Ward-Paige et al. (2005) noted that the greatest size and biomass of Clionids corresponded with areas with the highest nitrogen, ammonia and δ¹⁵N levels. However, no evidence of the effects of nutrient enrichment in this or similar biotopes was found.

Organic enrichment

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

Evidence

Rose & Risk (1985) described an increase in abundance of the sponge Cliona delitrix in an organically polluted section of Grand Cayman fringing reef affected by the discharge of untreated faecal sewage. De Goeij et al. (2008) used ¹³C to trace the fate of dissolved organic matter in the coral reef sponge Halisarca caerulea.  Biomarkers revealed that the sponge incorporated dissolved organic matter through both bacteria-mediated and direct pathways, suggesting that it feeds, directly and indirectly, on the dissolved organic matter. However, no evidence of the effect of organic enrichment in this or similar biotopes was found.

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

In northern Norway, Phakellia ventilabrum and Axinella infundibuliformis were primarily observed at sites with hard substratum, including sandy-gravels and cobbles, and dominate sponge communities on wave-exposed circalittoral rock habitat (Dunlop et al. 2020). This biotope is also characteristic of circalittoral rock (JNCC, 2022).

If rock were replaced with sediment, this would represent a fundamental change to the physical characteristics of the biotope, and the species would be unlikely to recover. The biotope would be lost. Therefore, resistance to the pressure is considered ‘None’, resilience ‘Very low’, and sensitivity is 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

‘Not relevant’ to biotopes occurring on bedrock.

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

The species characterizing this biotope, such as Phakellia ventilabrum and Axinella infundibuliformis, are epifauna or epiflora occurring on rock and other hard substrata (Dunlop et al., 2020) and would be sensitive to the removal of the habitat. Extraction of bedrock substratum is considered unlikely, and this pressure is usually considered to be ‘Not relevant’ to hard substratum habitats. However, Picton & Goodwin (2007) noted that an area of boulders with a rich fauna of sponges and hydroids on the east coast of Rathlin Island, Northern Ireland, was significantly altered since the 1980s. Scallop dredging began in 1989, and boulders were observed to have been turned and the gravel harrowed. In addition, many of the boulders had disappeared, and rare hydroid communities were greatly reduced (Picton & Goodwin, 2007). Prior records indicated the presence of large sponges, mainly Axinella infundibuliformis (Picton & Goodwin, 2007).  Freese et al. (1999) also noted that trawling could remove important substrata such as boulders. Therefore, where this biotope occurs on boulders that could be subject to removal or extraction, resistance is likely to be 'Low'. Hence, as resilience is probably 'Very low', sensitivity is assessed as 'High'. 

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

All characterizing sponges for this biotope are sessile epifauna, being either branching or cup-like. Phakellia ventilabrum is firm, quite elastic, fairly tough, but it becomes softer in older specimens, when it can become easily torn (Ackers et al., 1992). Stelligera stuposa is branching, moderately firm, elastic, with a soft outer layer (Ackers et al., 1992). Axinella infundibuliformis is moderately firm and resilient, but pieces break off if bent through 90° (Ackers et al., 1992). Axinella dissimilis is quite elastic and flexible (Moss & Ackers, 1982). However, if the sponge is bent through more than 90°, the surface will crack (Ackers et al., 1992). The species that create biogenic habitats, such as sponges and corals, often form complex ecological associations and tend to be long-lived, slow-growing and fragile, sensitive to disturbance and vulnerable to damage (Kaiser et al. 2018). Picton & Goodwin (2007) noted that an area of boulders with a rich fauna of sponges and hydroids on the east coast of Rathlin Island, Northern Ireland, was significantly altered since the 1980s. Scallop dredging had begun in 1989, and boulders were observed to have been turned and the gravel harrowed. In addition, many of the boulders had disappeared, and rare hydroid communities were greatly reduced (Picton & Goodwin, 2007). Prior records indicated the presence of large sponges, mainly Axinella infundibuliformis (Picton & Goodwin, 2007). Hiscock (2014) identified Axinella dissimilis as being very susceptible to towed fishing gear. Hinz et al. (2011) studied the effects of scallop dredging in Lyme Bay, UK, and found that the presence of the erect sponge Axinella dissimilis was significantly higher at non-fished sites (33% occurrence) compared to fished sites (15% occurrence). 

Freese et al. (1999) studied the effects of trawling on seafloor habitats and associated invertebrates in the Gulf of Alaska.  They found that a transect following a single trawling event showed a significant reduction in ‘vase’ sponges (67% expressed damage) and ‘morel’ sponges, although total damage could not be quantified as their brittle nature meant that these sponges were completely torn apart and scattered. The ‘finger’ sponges, the smallest and least damaged (only 14%) of the sponges assessed, were damaged by being knocked over. Van Dolah et al. (1987) studied the effects on sponges and corals of one trawl event over a low-relief hard bottom habitat off Georgia, the USA. The densities of individuals taller than 10 cm of three species of sponges in the trawl path and in the adjacent control area were assessed by divers and were compared before, immediately after and 12 months after trawling. Of the total number of sponges remaining in the trawled area, 32% were damaged. Most of the affected sponges were the barrel sponges Cliona spp., whereas Haliclona oculta and Ircina campana were not significantly affected. Freese (2001) studied deep cold-water sponges in Alaska a year after a trawl event; 46.8% of sponges exhibited damage, with 32.1% having been torn loose. None of the damaged sponges displayed signs of regrowth or recovery. This was in stark contrast to early work by Freese et al. (1999) on shallow sponge communities, with impacts of trawling activity being much more persistent due to the slower growth/regeneration rates of deep, cold-water sponges. Given the slow growth rates and long lifespans of the rich, diverse fauna, it is likely to take many years for deep sponge communities to recover if adversely affected by physical damage.

Tilmant (1979) found that, following a shrimp trawl in Florida, the USA, over 50% of sponges, including Neopetrosia, Spheciospongia, Spongia and Hippiospongia, were torn loose from the bottom. The highest damage incidence occurred to the finger sponge Neopetrosia longleyi. Size did not appear to be important in determining whether a sponge was affected by the trawl. Recovery was ongoing, but not complete, 11 months after the trawl, although no specific data relating to the sponges was provided. Boulcott & Howell (2011) conducted experimental Newhaven scallop dredging over a circalittoral rock habitat in the Sound of Jura, Scotland and recorded the damage to the resident community. The results indicated that epifaunal species, including the sponge Pachymatisma johnstoni, were highly damaged by the experimental trawl. Coleman et al. (2013) described a four-year study on the differences between a commercially potted area in Lundy with a no-take zone. No significant difference in Axinellid populations was observed. The authors concluded that the study indicated that lighter abrasion pressures, such as potting, were far less damaging than heavier gears, such as trawls.

In the Faroe-Shetland Channel, sessile species, such as sponges, Cirripedia, and Hydrocorals, were associated with areas characterized by low fishing activities, and fishing effort is one of the strongest factors driving the distribution of sponges (Vad et al., 2019). For example, Vieira et al. (2020) observed the effect of commercial bottom trawl fishing on deep-sea sponge aggregations (at a depth range of 1210-1250 m) and observed a sponge density decline from 1.09 to 0.03 ind/m², and biomass density from 246 to 4 gwwt/m², between the pre- and post-fishing surveys.

Deep-sea sponge recovery from physical disturbances like fishing can be slow and take multiple years to return to previous community levels (Samuelsen et al. 2022). Althaus et al. (2009), in a study of seamounts off Tasmania, reported no significant recovery when trawling was reduced to less than 5% for a decade or ceased for five years. Similarly, Morrison et al. (2020) observed the effect of trawling on deep-sea sponge grounds and recorded that four years post-disturbance, megafaunal densities of the shallow (∼600 m depth) and deep (∼1,400 m depth) sites were still significantly lower on the disturbed patches compared to the control areas. Although few studies exist on the growth and reproduction of most deep-sea sponges (e.g. Geodia grounds) to adequately predict post-disturbance trajectories, available studies indicate that deep-sea sponge grounds have comparatively low potential for recovery from physical disturbance events, and that recovery following impacts is considered more than temporary if recovery takes more than five to 20 years (Pham et al.,2019).

Kaiser et al. (2018) specifically studied the recovery of sessile epifauna following the exclusion of towed mobile fishing gear in Lyme Bay, UK, in 2008. Their estimates suggest that no recovery occurred within the timescale of the study (10 years), and that some biogenic habitats (particularly sponges and soft corals) could require up to, or more than, 20 to 30 years before signs of recolonization and recovery may occur. The maximum recovery time modelled was 51 years for yellow branched sponges (Kaiser et al., 2018). A 15-year review of the Lyme Bay trawling ban by Renn et al. (2024) highlighted definitive evidence of recovery, in terms of increased species richness, with key sessile taxa (Pentapora foliacea and Phallusia mammillata) showing signs of early recovery between 2008 and 2013. In terms of exploited species, between 2008 and 2019, fish experienced a 430% increase in taxon richness and a 370% increase in total abundance inside the Marine Protected Area (MPA), but invertebrates (crab, lobster, cuttlefish, and whelk) exhibited no signs of recovery (Renn et al., 2024). Renn et al. (2024) concluded that the evidence of recovery recorded in Lyme Bay broadly aligned with the wider literature by detecting early stages of recovery within the first few years of MPA establishment. However, full recovery is thought to occur over decadal timescales, and measuring full recovery rates in-situ remains a priority for future research in Lyme Bay.

Sensitivity assessment: Whilst some of the characterizing sponges can be quite elastic, abrasion pressures, especially by heavy gears, have been shown to cause significant damage to the sessile epifaunal sponges.  Therefore, resistance is assessed as 'Low'.  Hence, resilience is assessed as 'Very Low' and sensitivity as 'High'.

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 species characterizing this biotope group are epifauna or epiflora occurring on rock which is resistant to subsurface penetration.  The assessment for abrasion at the surface only is, therefore considered to equally represent sensitivity to this pressure. This pressure is thought ‘Not Relevant’ to hard rock biotopes

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

Schönberg (2015) reviewed and observed the interactions between sediments and marine sponges in Australia and described the lack of research on Porifera.  Bell et al. (2015) reviewed the effects and interactions of sponges with sediment in suspension and after deposition. Whilst many sponges are disadvantaged by sedimentation (as would be expected, being sessile filter feeders), many examples exist of sponges adapting to sediment presence (Bell et al., 2015; Schönberg, 2015). Adaptations included sediment incorporation, sediment encrusting, structural modification (such as reduction in numbers of oscula, or repositioning of inhalant and exhalant openings), soft sediment anchoring using spicules, modification of spicules to shield the body from sediment, backwashing, mucus production, morphology (e.g. upright forms intercept less settling sediment) and living, at least partially, embedded within the sediment (Bell et al., 2015; Schönberg, 2015). Despite sediment being considered to have a negative impact on suspension feeders (Gerrodette & Flechsig 1979), many encrusting sponges appear to be able to survive in highly sedimented conditions, and, in fact, many species prefer such habitats (Bell & Barnes 2001; Bell & Smith 2004). 

Among the characterizing species, Schönberg (2015) found that Axinellids frequently formed external crusts and sediment interaction was observed in 5.8 ± 4.8% of observations, but required rock substrata under the sediment for attachment.  Ackers et al. (1992) describe Axinella dissimilis as preferring clean oceanic water but tolerating silt. Sanchez et al. (2009) described finding communities composed primarily of Phakellia ventilabrum and Dendrophyllia cornigera in circalittoral rocky habitats in the Cantabrian Sea, northern Spain.  Phakellia ventilabrum showed greater tolerance to sedimentation pressures than the coral. The authors concluded that Phakellia ventilabrum preferred a mixed rock–sand habitat where deposition processes predominate, and hence sedimentation, together with hard substrata where it settles (Sanchez et al., 2009). Axinella dissimilis is mainly found on upward-facing clean or silty rock, and whilst it tends to prefer clean oceanic water, it is tolerant of silt (Ackers et al., 1992).

Castric-Fey & Chassé (1991) conducted a factorial analysis of the subtidal rocky ecology near Brest, France and rated the distribution of species in varying turbidity (corroborated by the depth at which laminarians disappeared).  Cliona celata and Stelligera rigida were classed as indifferent to turbidity, Tethya citrina, Pachymatisma johnstonia and Polymastia boletiformis (as Polymastia robusta) had a slight preference for clearer water, while Dysidea fragilis, Polymastia mamillaris, and Raspailia ramosa had a strong preference for turbid water.  None of the important characterizing sponges in this biotope were assessed. Bell et al. (2015) noted that upright forms intercepted a smaller amount of settling sediment than encrusting forms. For example, Bell & Barnes (2002; cited in Bell et al., 2015) reported considerable variation in the branching characteristics of Raspaillia ramosa and Stelligera stuposa across a sediment gradient in Lough Hyne, Northern Ireland, although the patterns were due to the interaction between sedimentation and water flow. Raspailia ramosa and Stelligera stuposa have a reduced maximum size in areas of high sedimentation (Bell et al., 2002). 

Storr (1976) observed the sponge Sphecispongia vesparium backwashing to eject sediment and noted that other sponges (such as Condrilla nucula) use secretions to remove settled material. Tjensvoll et al. (2013) found that Geodia barretti physiologically shuts down when exposed to sediment concentrations of 100 mg /l (86% reduction in respiration).  Rapid recovery to initial respiration levels directly after the exposure indicated that Geodia barretti can cope with a single short exposure to elevated sediment concentrations.  However, it should be noted that a laboratory study on the impact of elevated sedimentation rates on deep-water sponges found that sediment load of 30 mg sed./l resulted in significantly higher sponge mortality compared with sponges exposed to 5 and 10 mg sed./l, although no additional information was provided (Hoffman & Tore Rapp, pers comm. cited in Lancaster et al., 2014). Pineda et al. (2017a) examined the effect of suspended sediments in three species of sponge from New Zealand; two phototrophic (due to symbiotic algae) (Cliona orientalis and Carteriospongia foliascens) and one heterotrophic (Ianthella basta) under laboratory conditions. All sponges exhibited a short-term response to suspended sediment, e.g. closed ocsula, mucus production, and tissue regression. Most survived low to medium turbidity (≤33 mg/l) for up to 28 days, but at high turbidity (≤76 mg/l) Cliona orientalis and Carteriospongia foliascens experienced 20-90% mortality, and Ianthella basta showed tissue regression. Pineda et al. (2017a) suggested that suspended sediment combined with low light due to turbidity increased mortality in the phototrophic species but noted that there was considerable interspecies variation in their response.  

Sensitivity assessment: Despite reports that demonstrated increased sponge mortality at the benchmark level (see Lancaster et al., 2014, Pineda et al., 2017), the majority of the literature reviewed suggests that a change at the benchmark level is unlikely to cause significant mortality of the species considered characteristic of this biotope.  Therefore, resistance at the benchmark has been assessed as ‘High’, resilience as ‘High’, and the biotope is ‘Not sensitive’ at the benchmark level. 

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

Schönberg (2015) reviewed and observed the interactions between sediments and marine sponges in Australia and described the lack of research on Porifera. Bell et al. (2015) reviewed the effects and interactions of sponges with sediment in suspension and after deposition. Whilst many sponges are disadvantaged by sedimentation (as would be expected, being sessile filter feeders), many examples exist of sponges adapting to sediment presence (Bell et al., 2015; Schönberg, 2015). Adaptations included sediment incorporation, sediment encrusting, structural modification (such as reduction in numbers of oscula, or repositioning of inhalant and exhalant openings), soft sediment anchoring using spicules, modification of spicules to shield the body from sediment, backwashing, mucus production, morphology (e.g. upright forms intercept less settling sediment) and living, at least partially, embedded within the sediment (Bell et al., 2015; Schönberg, 2015). Despite sediment being considered to have a negative impact on suspension feeders (Gerrodette & Flechsig 1979), many encrusting sponges appear to be able to survive in highly sedimented conditions, and, in fact, many species prefer such habitats (Bell & Barnes 2001; Bell & Smith 2004). Pineda et al. (2017b) exposed three phototrophic (due to symbiotic algae) and two heterotrophic sponges from New Zealand to repeated deposition events and sediment cover over 80-100% of sponge surface to a depth of ca 0.5 mm for up to 30 days in laboratory conditions. All five species survived with minimal physiological effects. However, Wulff (2006) described mortality in three sponge groups following four weeks of complete burial under sediment; 16% of Amphimedon biomass died compared with 40% and 47% in Iotrochota and Aplysina, respectively. In Lough Hyne, a small (∼0.5 km²) lough on the southwest coast of Ireland, similar sponge communities experience varying levels of sedimentation, from negligible levels (3 ± 0.2 mm) to higher rates (18 to 34 g/m/day), and sponges have been continually recorded in these areas under these levels of sedimentation (Micaroni et al., 2025).

Tidal fluctuations, mixing by internal waves, and storms (particularly in shallower waters) are natural ways in which sediments are periodically resuspended within oceans, and help to keep deep-sea sponges fed with organic material (Samuelsen et al. 2022). However, one understood source of sedimentation within the marine environment is from offshore oil and gas activities, mainly via drilling (Vad et al., 2018). Vad et al. (2018) studied the impacts of oil and gas drilling on deep-sea sponges and observed that physical disruption and increased sedimentation during well drilling and infrastructure installations can locally diminish benthic communities by more than 90% in terms of megafaunal density within sponge grounds. Major reductions in sponge densities and reduced diversity were seen close to drilling activity, within 100 to 200 m, and persisted for several years (Vad et al. 2018).

Vad et al. (2018) concluded that effects on deep-sea sponge grounds from such physical disturbance were still detectable up to 10 years post-drilling, and this slow, partial recovery was inversely related to the distance to the well and the time after drilling, resulting from the long-lived nature, slow growth rates and low reproduction rates of most deep-sea organisms. Furthermore, if oil and gas drilling used synthetic and water-based muds, the decrease in community diversity and abundance was detected up to 1,000 m away from the release (Vad et al. 2018). Functional changes in benthic communities, associated with a loss of suspension-feeding species and an increase in deposit feeders, have also been detected at drill release sites (Vad et al. 2018). Conversely, Durden et al. (2023) also observed the effects of industrial sedimentation on sponge communities; however, once sedimentation accumulated on sponges, it cleared mostly from them gradually over time, but sometimes sharply. Yet, sponges never returned to their original state, and this partial recovery likely involved a combination of active and passive removal of the sediment.

Among the characterizing species, Schönberg (2015) found that Axinellids frequently formed external crusts and sediment interaction was observed in 5.8 ± 4.8% of observations, but required rock substrata under the sediment for attachment.  Ackers et al. (1992) describe Axinella dissimilis as preferring clean oceanic water but tolerating silt. Sanchez et al. (2009) described finding communities composed primarily of Phakellia ventilabrum and Dendrophyllia cornigera in circalittoral rocky habitats in the Cantabrian Sea, northern Spain.  Phakellia ventilabrum showed greater tolerance to sedimentation pressures than the coral. The authors concluded that Phakellia ventilabrum preferred a mixed rock–sand habitat where deposition processes predominate, and hence sedimentation, together with hard substrata where it settles (Sanchez et al., 2009). Axinella dissimilis is mainly found on upward-facing clean or silty rock, and whilst it tends to prefer clean oceanic water, it is tolerant of silt (Ackers et al., 1992).

Castric-Fey & Chassé (1991) conducted a factorial analysis of the subtidal rocky ecology near Brest, France and rated the distribution of species in varying turbidity (corroborated by the depth at which laminarians disappeared).  Cliona celata and Stelligera rigida were classed as indifferent to turbidity, Tethya citrina, Pachymatisma johnstonia and Polymastia boletiformis (as Polymastia robusta) had a slight preference for clearer water, while Dysidea fragilis, Polymastia mamillaris, and Raspailia ramosa had a strong preference for turbid water.  None of the important characterizing sponges in this biotope were assessed. Bell et al. (2015) noted that upright forms intercepted a smaller amount of settling sediment than encrusting forms. For example, Bell & Barnes (2002; cited in Bell et al., 2015) reported considerable variation in the branching characteristics of Raspaillia ramosa and Stelligera stuposa across a sediment gradient in Lough Hyne, Northern Ireland, although the patterns were due to the interaction between sedimentation and water flow. However, Raspailia ramosa and Stelligera stuposa have a reduced maximum size in areas of high sedimentation (Bell et al., 2002).  Storr (1976) observed the sponge Sphecispongia vesparium backwashing to eject sediment and noted that other sponges (such as Condrilla nucula) use secretions to remove settled material.   The complete disappearance of the sea squirt Ascidiella aspera biocoenosis and ‘associated sponges’ in the Black Sea near the Kerch Strait was attributed to siltation (Terent'ev, 2008, cited in Tillin & Tyler-Walters, 2014).

Sensitivity assessment: The characterizing sponges are all large, erect sponges, while some of the characterizing sponges have been reported to cope with sediment occurring on rock (Sanchez et al., 2009).  It should also be noted that some of the characterizing sponges are likely to be buried in 5 cm of sediment deposition.  This is a high-energy biotope due to wave action, which is probably attenuated at the depths where the biotope occurs, 30 to 50 m. Any deposited sediment is unlikely to be removed rapidly (a few tidal cycles) but is likely to be removed in stormy weather. Therefore, resistance (at the benchmark level) has been assessed as ‘Medium’, resilience as ‘Very Low’ and sensitivity as ‘Medium’.

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

Schönberg (2015) reviewed and observed the interactions between sediments and marine sponges in Australia and described the lack of research on Porifera. Bell et al. (2015) reviewed the effects and interactions of sponges with sediment in suspension and after deposition. Whilst many sponges are disadvantaged by sedimentation (as would be expected, being sessile filter feeders), many examples exist of sponges adapting to sediment presence (Bell et al., 2015; Schönberg, 2015). Adaptations included sediment incorporation, sediment encrusting, structural modification (such as reduction in numbers of oscula, or repositioning of inhalant and exhalant openings), soft sediment anchoring using spicules, modification of spicules to shield the body from sediment, backwashing, mucus production, morphology (e.g. upright forms intercept less settling sediment) and living, at least partially, embedded within the sediment (Bell et al., 2015; Schönberg, 2015). Despite sediment being considered to have a negative impact on suspension feeders (Gerrodette & Flechsig 1979), many encrusting sponges appear to be able to survive in highly sedimented conditions, and, in fact, many species prefer such habitats (Bell & Barnes 2001; Bell & Smith 2004). Pineda et al. (2017b) exposed three phototrophic (due to symbiotic algae) and two heterotrophic sponges from New Zealand to repeated deposition events and sediment cover over 80-100% of sponge surface to a depth of ca 0.5 mm for up to 30 days in laboratory conditions. All five species survived with minimal physiological effects. However, Wulff (2006) described mortality in three sponge groups following four weeks of complete burial under sediment; 16% of Amphimedon biomass died compared with 40% and 47% in Iotrochota and Aplysina, respectively. In Lough Hyne, a small (∼0.5 km²) lough on the southwest coast of Ireland, sponge communities experience varying levels of sedimentation, from negligible levels (3 ± 0.2 mm) to higher rates (18 to 34 g/m/day), and sponges have been continually recorded in these areas under these levels of sedimentation (Micaroni et al., 2025).

Tidal fluctuations, mixing by internal waves, and storms (particularly in shallower waters) are natural ways in which sediments are periodically resuspended within oceans, and help to keep deep-sea sponges fed with organic material (Samuelsen et al. 2022). However, one understood source of sedimentation within the marine environment is from offshore oil and gas activities, mainly via drilling (Vad et al., 2018). Vad et al. (2018) studied the impacts of oil and gas drilling on deep-sea sponges and observed that physical disruption and increased sedimentation during well drilling and infrastructure installations can locally diminish benthic communities by more than 90% in terms of megafaunal density within sponge grounds. Major reductions in sponge densities and reduced diversity were seen close to drilling activity, within 100 to 200 m, and persisted for several years (Vad et al. 2018).

Vad et al. (2018) concluded that effects on deep-sea sponge grounds from such physical disturbance were still detectable up to 10 years post-drilling, and this slow, partial recovery was inversely related to the distance to the well and the time after drilling, resulting from the long-lived nature, slow growth rates and low reproduction rates of most deep-sea organisms. Furthermore, if oil and gas drilling used synthetic and water-based muds, the decrease in community diversity and abundance was detected up to 1,000 m away from the release (Vad et al. 2018). Functional changes in benthic communities, associated with a loss of suspension-feeding species and an increase in deposit feeders, have also been detected at drill release sites (Vad et al. 2018). Conversely, Durden et al. (2023) also observed the effects of industrial sedimentation on sponge communities; however, once sedimentation accumulated on sponges, it cleared mostly from them gradually over time, but sometimes sharply. Yet, sponges never returned to their original state, and this partial recovery likely involved a combination of active and passive removal of the sediment.

Among the characterizing species, Schönberg (2015) found that Axinellids frequently formed external crusts and sediment interaction was observed in 5.8 ± 4.8% of observations, but required rock substrata under the sediment for attachment.  Ackers et al. (1992) describe Axinella dissimilis as preferring clean oceanic water but tolerating silt. Sanchez et al. (2009) described finding communities composed primarily of Phakellia ventilabrum and Dendrophyllia cornigera in circalittoral rocky habitats in the Cantabrian Sea, northern Spain.  Phakellia ventilabrum showed greater tolerance to sedimentation pressures than the coral. The authors concluded that Phakellia ventilabrum preferred a mixed rock–sand habitat where deposition processes predominate, and hence sedimentation, together with hard substrata where it settles (Sanchez et al., 2009). Axinella dissimilis is mainly found on upward-facing clean or silty rock, and whilst it tends to prefer clean oceanic water, it is tolerant of silt (Ackers et al., 1992).

Castric-Fey & Chassé (1991) conducted a factorial analysis of the subtidal rocky ecology near Brest, France and rated the distribution of species in varying turbidity (corroborated by the depth at which laminarians disappeared).  Cliona celata and Stelligera rigida were classed as indifferent to turbidity, Tethya citrina, Pachymatisma johnstonia and Polymastia boletiformis (as Polymastia robusta) had a slight preference for clearer water, while Dysidea fragilis, Polymastia mamillaris, and Raspailia ramosa had a strong preference for turbid water.  None of the important characterizing sponges in this biotope were assessed. Bell et al. (2015) noted that upright forms intercepted a smaller amount of settling sediment than encrusting forms. For example, Bell & Barnes (2002; cited in Bell et al., 2015) reported considerable variation in the branching characteristics of Raspaillia ramosa and Stelligera stuposa across a sediment gradient in Lough Hyne, Northern Ireland, although the patterns were due to the interaction between sedimentation and water flow. However, Raspailia ramosa and Stelligera stuposa have a reduced maximum size in areas of high sedimentation (Bell et al., 2002).  Storr (1976) observed the sponge Sphecispongia vesparium backwashing to eject sediment and noted that other sponges (such as Condrilla nucula) use secretions to remove settled material.   The complete disappearance of the sea squirt Ascidiella aspera biocoenosis and ‘associated sponges’ in the Black Sea near the Kerch Strait was attributed to siltation (Terent'ev, 2008, cited in Tillin & Tyler-Walters, 2014).

Sensitivity assessment: In 30 cm of deposition, the majority of sponges (whose growth is up to ca 30 cm) are likely to be buried, unless the topography of the biotope includes many vertical surfaces. Hiscock & Jones (2004) reported that Axinella dissimilis (as Axinella polypoides) and Homaxinella subdola grew up to a height of ca 30 cm.  The benchmark level is, therefore, at the upper limit of the growth of the characterizing sponges. As this biotope experiences negligible water flow, it is unlikely that this sediment would be removed rapidly.  Therefore, resistance at the benchmark has been assessed as ‘Low’, resilience as ‘Very Low’ and sensitivity as ‘High’. 

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

'No evidence' was found.

Underwater noise changes

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

Evidence

Whilst no evidence was found on the effect of noise or vibrations on the characterizing species of these biotopes, it is unlikely that these species have the facility for detecting or noise vibrations.

Sensitivity assessment: The characterizing sponges are unlikely to respond to noise or vibrations and resistance is, therefore assessed as ‘High’, Resilience as ‘High’ and Sensitivity as ‘Not Sensitive’.

Introduction of light or shading

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

Evidence

Jones et al.(2012) reported on the monitoring of sponges around Skomer Island and found that many sponges, particularly encrusting species, preferred vertical or shaded bedrock to open, light surfaces. Bell & Barnes (2000; cited in Bell et al., 2015) noted higher sponge diversity and abundance at areas subject to sedimentation in Lough Hyne, Northern Ireland and suggested reduced competition with macroalgae was a factor. However, Cárdenas et al. (2016) reported high sponge diversity associated with canopy-forming macroalgae in the Antarctic. Nevertheless, whilst no evidence could be found for the effect of light on the characterizing species of these biotopes, we know that within the first 200 m of ocean depth, light, both natural and artificial, would reach the seabed. In addition, shading of light or the introduction of light within the first 50 m could have an effect on marine organisms, such as triggering early coral spawning or affecting the opening and reproduction rhythm of bivalves (Charifi et al., 2023; Davies et al., 2023; Smyth et al.,2021). Below 200 m, it is unlikely that these species would be impacted, as the light level that reaches beyond this point is very low and unsuitable for photosynthesis.

Sensitivity assessment: Changes in light (introduction or shading) are unlikely to affect the adults of the characteristic species. However, reproductive cues (e.g. spawning of propagules) may become out of phase with seasonal conditions or food supply, resulting in the disruption of fertilization, larval development or settlement behaviour, and hence recruitment (Charifi et al., 2023; Davies et al., 2023; Smyth et al.,2021). However, no direct evidence of these effects on the characteristic species was found and is 'insufficient' to form the basis of an assessment. 

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

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.  NB. Collision by grounding vessels is addressed under ‘surface abrasion’.

Visual disturbance

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

Evidence

'Not relevant'.

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

'No evidence' for the characterizing sponges could be found.

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

Gochfeld et al. (2012) found that diseased sponges hosted significantly different bacterial assemblages compared to healthy sponges, with diseased sponges also exhibiting a significant decline in sponge mass and protein content.  Sponge disease epidemics can have serious long-term effects on sponge populations, especially in long-lived, slow-growing species (Webster, 2007).  Numerous sponge populations have been brought to the brink of extinction, including cases in the Caribbean with 70-95% disappearance of sponge specimens (Galstoff, 1942), the Mediterranean (Vacelet, 1994; Gaino et al.,1992).  Decaying patches and white bacterial film were reported in Haliclona oculata and Halichondria panicea in North Wales, 1988 to 89 (Webster, 2007). Specimens of Cliona spp. have exhibited blackened damage since 2013 in Skomer. Preliminary results have shown that clean, fouled and blackened Cliona all have very different bacterial communities. The blackened Cliona are effectively dead and have a bacterial community similar to marine sediments. The fouled Cliona have a very distinct bacterial community, which may suggest a specific pathogen caused the effect (Burton, pers comm; Preston & Burton, 2015). 

Sensitivity assessment: 'No evidence' of diseases affecting the important characterizing sponges has been recorded.  Sponge diseases have caused limited mortality in some species in the British Isles, although mass mortality and even extinction have been reported further afield. 

Removal of target species

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

Evidence

Hiscock (2002) stated that the greatest loss of Axinella dissimilis at Lundy might have been due to collecting during scientific studies in the 1970s. No indication of recovery was evident. Axinella damicornis was harvested in Lough Hyne during the 1980s (for molecular investigations) and the populations were reduced to very low densities, which subsequently recovered very slowly, although they are now considered to be back to their original densities (Bell, 2007).

Sensitivity assessment: Based on the above observations, resistance is assessed as ‘None’ and resilience as ‘Very Low’ with a resultant sensitivity of ‘High’.

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

This biotope may be removed or damaged by static or mobile gears that are targeting other species. These direct, physical impacts 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 non-target species in this biotope.  The unintentional removal of the important characterizing species will result in loss of the biotope.  Therefore, resistance is recorded as ‘Low’, resilience as ‘Very Low’ and sensitivity as ‘High’.

The American slipper limpet, Crepidula fornicata

Evidence

Crepidula fornicata larvae require hard substrata for settlement. It prefers muddy, gravelly, shell-rich substrata that include gravel, the 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 (2023) 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 at 60 metres 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 a gravel substratum with boulders. Bohn et al. (2015) noted that Crepidula occurred at low density or was absent in areas dominated by boulders. 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. In addition, no evidence was found of the effect of Crepidula populations on faunal turf-dominated habitats. It was only recorded at low density (0.1-0.9/m²) in one faunal turf biotope (CR.MCR.CFaVS.CuSpH.As) (JNCC, 2015). Faunal turfs are dominated by suspension feeders, so larval predation is probably high, which may prevent colonization by Crepidula. Also, faunal turf species actively compete for space, and many are fast-growing and opportunistic, so they may out-compete Crepidula for space even if it gained a foothold in the community. 

Sensitivity assessment. The circalittoral rock characterizing this biotope is likely to be unsuitable for the colonization by Crepidula fornicata due to the extremely wave exposed to wave-exposed conditions, in which wave action and storms may mitigate or prevent the colonization by Crepidula at high densities, although Crepidula has been recorded from areas of strong tidal streams (Hinz et al., 2011). 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). Therefore, colonization of Crepidula would be limited to low densities in deeper examples of the biotope. In addition, no evidence was found of the effect of Crepidula populations on faunal turf-dominated habitats or infralittoral or circalittoral rock habitats. At present, there is 'Insufficient evidence' to suggest that the circalittoral rock biotopes are sensitive to colonization by Crepidula fornicata or other invasive species; further evidence is required. 

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; Minchin & 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 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 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, and 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; Dijkstra et al., 2007; Griffith 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). For example, in oyster farms in British Columbia, large fragments of Didemnum sp. come off oyster strings when they are pulled out of water, and other fragments can be pulled off oysters and mussels and thrown back into the water, which is likely to aid dispersal of the invasive species (Bullard et al., 2007). Dijkstra et al. (2007) hypothesised that Didemnum sp. was introduced to the Gulf of Maine with oyster aquaculture in the Damariscotta River and transported via Pacific oysters.

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 that 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 water 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 water (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 pools (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 the 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 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). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems because of its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock, gravel, pebble, cobble, or boulders or artificial substrata such as a variety of maritime structures, such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ship hulls and at aquaculture facilities (Valentine et al., 2007a&b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012; 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).

Didemnum vexillum has the ability to rapidly overgrow and displace on other sessile organisms such as other colonial ascidians (Ciona intestinalis, Styela clava, Ascidiella aspera, Botrylloides violaceus, Botryllus schlosseri, Diplosoma listerianium and Aplidium spp.), bryozoan, hydroids, sponges (Clione celata and Halichrondria sp.), anemone (Diadumene cincta), calcareous tube worms, eelgrass (Zostera marina), kelp (Laminaria spp. and Agarum sp.), green algae (Codium fragile subsp. fragile), red algae (Plocamium, Chondrus crispus and bush weed Agardhiella subulata), brown algae (Ascophyllum nodosum, Sargassum, Halidrys, Fucus evanescens and Fucus serratus), calcareous algae (Corallina officinalis), mussels (Mytilus galloprovincialis, Perna canaliculus  and Mytilus edulis), barnacles, oysters (Magallana gigas, Ostrea edulis and Crassostrea virginica), sea scallops (Placopecten magellanicus), or dead shells (Dijkstra et al., 2007; Gittenberger, 2007; Valentine et al., 2007a; Valentine et al., 2007b; Griffith et al., 2009; Carman & Grunden, 2010; Dijkstra & Nolan, 2011; Groner et al., 2011; Hitchin, 2012; Tagliapietra et al., 2012; Minchin & Nunn, 2013; Gittenberger et al., 2015; Long & Groholz, 2015; Vercaemer et al., 2015).

In contrast, Didemnum vexillum’s preference for 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. Kleeman (2009), stated that the presence of a consistent mild wave action or ‘swash zone’ appears to favour Didemnum sp. establishment in the intertidal. Although some evidence suggests that waves and currents can facilitate the fragmentation and spread of Didemnum vexillum (Mckenzie et al., 2017), the tidal current velocities at some sites where Didemnum vexillum has been reported (for example, New England, where current velocities reach up to around 3 m/s) is lower than the current velocity required for the dislodgement of Didemnum vexillum fragments (around 7.6 m/s) (Reinhardt et al., 2012). This suggests that not all tidal currents are likely to dislodge Didemnum vexillum fragments. When on boat hulls the species can experience higher current velocities, which is enough to cause dislodgement (Reinhardt et al., 2012).  

Didemnum vexillum has not been reported to colonize deep sponge communities. Nevertheless, Didemnum vexillum has been recorded in the sublittoral to depths of 81 m in Georges Bank and 30 m in Long Island, USA (Bullard et al., 2007; Valentine et al., 2007b; Mercer et al., 2009). This deep sponge biotope occurs on bedrock, which could provide a suitable hard substratum for colonization by Didemnum sp. Didemnum vexillum is reported to prefer sheltered conditions but has also been recorded in moderately strong currents (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020) and is predicted to survive stronger currents, as the current velocity which will dislodge Didemnum vexillum is around 7.6 m/s (Reinhardt et al., 2012). This biotope experiences very weak to moderately strong water flow (0 to 1.5 m/s) but moderate to extreme wave exposure. However, the effect of wave action reduces with depth, so it is possible that only the most wave exposed examples of the biotope could be unsuitable for Didemnum. Didemnum vexillum regresses as temperatures decline in winter, so shallow examples may be able to recover their condition in winter (Gittenberger, 2007; Valentine et al., 2007a; Herborg et al., 2009). However, deeper examples may not experience enough temperature change to trigger the decline in Didemnum vexillum (Valentine et al., 2007a). If Didemnum sp. could gain a 'foothold', it might overgrow, smother or cause mortality of deep sponges. Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. Therefore, a resistance of 'Medium' (some, <25% mortality) is suggested as a precaution in case Didemnum vexillum could colonize the biotope, but with 'Low' confidence due to the lack of direct evidence. Resilience is assessed as 'Very low' as recovery would require the physical removal of Didemnum sp., so sensitivity is assessed as 'Medium'. 

The Pacific oyster, Magallana gigas

Evidence

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). Therefore, this INIS is probably 'Not relevant' in this biotope. 

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. 

Other INIS

Evidence

The non-native sponge, the cauliflower sponge Celtodoryx ciocalyptoides is thought to have been introduced to the North East Atlantic from Japan via Magallana gigas aquaculture. It has not been recorded in UK waters to date but has become a major space occupier in the Oosterschelde, Netherlands and Gulf of Morbihan, France (Van Soest et al., 2007; Henkel & Janussen, 2011; GBNNSIP, 2017).