The Marine Life Information Network

Temperature increase (local)

Benchmark. A 5°C increase in temperature for one month, or 2°C for one year (Temperature change pressure definition).

Evidence

There is limited information available about the tolerance of the characterizing sponges in this biotope. However, Kazanidis et al. (2019) reported how water temperature and salinity explained a significant amount of variation in sponge density. In the North Atlantic, the highest density values were found where temperature and salinity ranged from 6.52 to 8.98°C and 34.91 to 35.13 psu, respectively (Kazanidis et al., 2019). In Lough Hyne, a small (approx. 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). Long-term temperature increases may cause extension of the British Isles sponge populations, and decreases in temperature may result in population shrinkage. For example, 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). 

Broadribb, Bell & Rovellini (2021) studied the rapid acclimation in two sponges, Hymeniacidon perlevis and Halichondria panicea, to seasonal variation at two rocky shore sites on the west coast of Wales, north of Aberystwyth. Both sites are moderately exposed intertidal rocky areas with typical north-east Atlantic zonation patterns (Lewis, 1964 cited in Broadribb, Bell & Rovellini, 2021). They observed that both species had a higher proportion of inorganic tissue content in winter months, which correlated with higher levels of wave action and lower temperature, representing either an increase in spicule size/number or a loss of organic material (Broadribb, Bell & Rovellini, 2021). In addition, they also detected rapid decreases in organic content in some months, which corresponded with previously reported reproductive timings for the two species, and likely represent gamete release events rather than as a result of fluctuating abiotic factors (Broadribb, Bell & Rovellini, 2021).

To date, about 100 Halichondria species are accepted (Alvarez & Hooper, 2011, Van Soest et al., 2020, and Hooper & van Soest, 2020, cited in Goldstein & Funch, 2022). They occur in different types of marine habitats around the world, being widespread in Europe, North America, the Brazilian coastal waters of the Atlantic, but also in parts of the Baltic Sea, the Mediterranean Sea (Burton, 1930; Fell & Lewandrowski, 1981; Barthel, 1992; Vethaak et al., 1982, Carvalho & Hajdu, 2001; Erpenbeck & Van Soest, 2002, Gaino et al., 2007 and Picton & Goodwin, 2007 cited in Goldstein & Funch, 2022) and the White Sea (Khalaman & Komendantov, 2016). Halichondria species also occur in the North Pacific, including Alaska (Knowlton & Highsmith, 2000 and Wulff, 2012 cited in Goldstein & Funch, 2022), Japan (Hoshino, 2004 cited in Goldstein & Funch, 2022), Korea (Kang & Sim, 2008 and Jeon & Sim, 2009 cited in Goldstein & Funch, 2022), and the South China Sea (Li et al., 2007 cited in Goldstein & Funch, 2022).

The growth of Halichondria sponges is dependent on temperature (Kim et al., 1986 cited in Goldstein & Funch, 2022) and the concentration of available food, which mainly consists of bacteria and phytoplankton (Lüskow et al., 2019 cited in Goldstein & Funch, 2022). The pumping rates of Halichondria panicea increase linearly with temperature and require relatively low energy demands for filtering large volumes of seawater (Thomassen & Riisgård, 1995 and Riisgård et al., 1993 cited in Goldstein & Funch, 2022), as expressed by filtration/respiration-ratios ≥15.6 l H₂O (ml O₂), which are comparable to other filter-feeding marine invertebrates (Riisgård et al., 2019 cited in Goldstein & Funch, 2022). In contrast, the energetic cost of growth is high in sponges (Thomassen & Riisgård, 1995 and Koopmans et al., 2010 cited in Goldstein & Funch, 2022), with exponential growth at a maximum rate of 4% per day in Halichondria panicea under natural conditions (Riisgård & Larsen, 2022 cited in Goldstein & Funch, 2022). Tissue regression and high mortality during the colder months of the year have also been reported for temperate Halichondria sp. from the Mystic and Thames estuaries, USA (Fell & Lewandrowski, 1981; Fell et al., 1984 cited in Goldstein & Funch, 2022) and for Halichondria bowerbanki from New England, USA (Hartman, 1958 cited in Goldstein & Funch, 2022), respectively.

The sponge Pachymatisma johnstonia is one of the most common and well-known sponges throughout the North East Atlantic coasts, recorded from the Orkneys to Spain, and is ubiquitous across the western and southern coasts of Britain, being mainly found in the littoral and sublittoral zones (Ackers et al., 1992; Schiavo et al., 2024).

Amphilectus fucorum (syn. Esperiopsis fucorum) is found in a wide variety of situations and is distributed across the Atlantic coasts of Europe, from Norway to France (Ackers et al., 1992). 

Clathrina coriacea is a species of calcareous sponge, and Clathrina spp. are found in temperate and tropical seas, from Norway to the Mediterranean in Europe, and Central and South America, Australia, and French Polynesia (Klautau et al., 2020). Most Calcinean sponges are only known down to 200 m in depth, but in the deep-water regions of the eastern and central Great Australian Bight, Calcinean sponges were found to depths of 400 m (Klautau et al., 2025). Clathrina coriacea may be resilient to warming due to being observed in the same cave system in the Eastern Ligurian Riviera, Italy, approximately 55 years apart (1961-63 to 2016) from two surveys. During the half-century between surveys, the sponge communities inside the semi-submerged caves have changed since the historical dives, with positive thermal anomalies occurring in the Ligurian Sea in the last decade (2006 to 2016) likely causing a decrease in massive sponges in the caves, which were replaced by encrusting forms (Costa et al., 2018).

Lemoine et al. (2007) studied the effects of thermal stress on the holobiont of the sponge Halichondria bowerbanki collected from Virginia, USA. Whilst no apparent change in density or diversity of symbionts was detected over the range of temperatures (29°C, 30°C and 31°C), the presence of particular symbionts was temperature-dependent. Barthel (1986) reported that reproduction and growth in Halichondria panicea in the Kiel Bight were primarily driven by temperature, with higher temperatures corresponding with the highest growth.

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. Morphological changes most strongly correlated with a mixture of water visibility and temperature. Research by Webster et al. (2008, 2011), Webster & Taylor (2012) and Preston & Burton (2015) suggested that many sponges rely on a holobiont of many synergistic microbes. Webster et al. (2011) described a much higher thermal tolerance of sponge larval holobiont when compared with adult 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.

Sensitivity assessment

Typical surface water temperatures around the UK coast vary seasonally from 4 to 19°C (Huthnance, 2010). The biotope is considered to tolerate a 2°C increase in temperature for a year. The majority of characterizing species would be unaffected by an increase in temperature at the benchmark level. Morphological changes were observed in UK sponge communities, with temperature a factor, but the characterizing sponges assessed were not listed as the most highly contributing to these changes (Berman et al., 2013). Therefore, resistance is likely to be ‘Medium’, resilience is therefore ‘High’, and sensitivity is assessed as ‘Low’ at the benchmark level.

Temperature decrease (local)

Benchmark. A 5°C decrease in temperature for one month, or 2°C for one year (Temperature change pressure definition).

Evidence

All characterizing sponges, anemones, and the ascidian Dendrodoa grossularia are widely distributed across the coasts of the British Isles and are all found from the Channel Isles to Northern Scotland (NBN, 2015). 

There is limited information available about the tolerance of the characterizing sponges in this biotope. However, Kazanidis et al. (2019) reported how water temperature and salinity explained a significant amount of variation in sponge density. In the North Atlantic, the highest density values were found where temperature and salinity ranged from 6.52 to 8.98°C and 34.91 to 35.13 psu, respectively (Kazanidis et al., 2019). In Lough Hyne, a small (approx. 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). Long-term temperature increases may cause extension of the British Isles sponge populations, and decreases in temperature may result in population shrinkage. For example, 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). 

Broadribb, Bell & Rovellini (2021) studied the rapid acclimation in two sponges, Hymeniacidon perlevis and Halichondria panicea, to seasonal variation at two rocky shore sites on the west coast of Wales, north of Aberystwyth. Both sites are moderately exposed intertidal rocky areas with typical north-east Atlantic zonation patterns (Lewis, 1964 cited in Broadribb, Bell & Rovellini, 2021). They observed that both species had a higher proportion of inorganic tissue content in winter months, which correlated with higher levels of wave action and lower temperature, representing either an increase in spicule size/number or a loss of organic material (Broadribb, Bell & Rovellini, 2021). In addition, they also detected rapid decreases in organic content in some months, which corresponded with previously reported reproductive timings for the two species, and likely represent gamete release events rather than as a result of fluctuating abiotic factors (Broadribb, Bell & Rovellini, 2021).

Some sponges 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., 1982).

To date, about 100 Halichondria species are accepted (Alvarez & Hooper, 2011, Van Soest et al., 2020, and Hooper & van Soest, 2020, cited in Goldstein & Funch, 2022). They occur in different types of marine habitats around the world, being widespread in Europe, North America, the Brazilian coastal waters of the Atlantic, but also in parts of the Baltic Sea, the Mediterranean Sea (Burton, 1930; Fell & Lewandrowski, 1981; Barthel, 1992; Vethaak et al., 1982, Carvalho & Hajdu, 2001; Erpenbeck & Van Soest, 2002, Gaino et al., 2007 and Picton & Goodwin, 2007 cited in Goldstein & Funch, 2022) and the White Sea (Khalaman & Komendantov, 2016). Halichondria species also occur in the North Pacific, including Alaska (Knowlton & Highsmith, 2000 and Wulff, 2012 cited in Goldstein & Funch, 2022), Japan (Hoshino, 2004 cited in Goldstein & Funch, 2022), Korea (Kang & Sim, 2008 and Jeon & Sim, 2009 cited in Goldstein & Funch, 2022), and the South China Sea (Li et al., 2007 cited in Goldstein & Funch, 2022).

The growth of Halichondria sponges is dependent on temperature (Kim et al., 1986 cited in Goldstein & Funch, 2022) and the concentration of available food, which mainly consists of bacteria and phytoplankton (Lüskow et al., 2019 cited in Goldstein & Funch, 2022). The pumping rates of Halichondria panicea increase linearly with temperature and require relatively low energy demands for filtering large volumes of seawater (Thomassen & Riisgård, 1995 and Riisgård et al., 1993 cited in Goldstein & Funch, 2022), as expressed by filtration/respiration-ratios ≥15.6 l H₂O (ml O₂), which are comparable to other filter-feeding marine invertebrates (Riisgård et al., 2019 cited in Goldstein & Funch, 2022). In contrast, the energetic cost of growth is high in sponges (Thomassen & Riisgård, 1995 and Koopmans et al., 2010 cited in Goldstein & Funch, 2022), with exponential growth at a maximum rate of 4% per day in Halichondria panicea under natural conditions (Riisgård & Larsen, 2022 cited in Goldstein & Funch, 2022). Tissue regression and high mortality during the colder months of the year have also been reported for temperate Halichondria sp. from the Mystic and Thames estuaries, USA (Fell & Lewandrowski, 1981; Fell et al., 1984 cited in Goldstein & Funch, 2022) and for Halichondria bowerbanki from New England, USA (Hartman, 1958 cited in Goldstein & Funch, 2022), respectively.

Barthel (1986) reported that Halichondria panicea in the Kiel Bight degenerated and disintegrated after reproduction before winter, however, young colonies were observed from September. 

The sponge Pachymatisma johnstonia is one of the most common and well-known sponges throughout the North East Atlantic coasts, recorded from the Orkneys to Spain, and is ubiquitous across the western and southern coasts of Britain, being mainly found in the littoral and sublittoral zones (Ackers et al., 1992; Schiavo et al., 2024).

Amphilectus fucorum (syn. Esperiopsis fucorum) is found in a wide variety of situations and is distributed across the Atlantic coasts of Europe, from Norway to France (Ackers et al., 1992).

Clathrina coriacea is a species of calcareous sponge, and Clathrina spp. are found in temperate and tropical seas, from Norway to the Mediterranean in Europe, and Central and South America, Australia, and French Polynesia (Klautau et al., 2020). Most Calcinean sponges are only known down to 200 m in depth, but in the deep-water regions of the eastern and central Great Australian Bight, Calcinean sponges were found to depths of 400 m (Klautau et al., 2025).

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. Morphological changes most strongly correlated with a mixture of water visibility and temperature. Research by Webster et al. (2008, 2011), Webster & Taylor (2012) and Preston & Burton (2015) suggested that many sponges rely on a holobiont of many synergistic microbes. Webster et al. (2011) described a much higher thermal tolerance of sponge larval holobiont when compared with adult sponges. 

Crisp (1964) studied the effects of an unusually cold winter (1962 to 1963) on the marine life in Britain, including porifera in North Wales. Whilst difficulty in distinguishing between mortality and delayed development was noted, Crisp 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. It should be noted that Crisp’s general comments on all marine life state that damage decreased the deeper the habitat. 

Sensitivity assessment

Although the characterizing species have northern/boreal distributions and are unlikely to be affected at the benchmark level, there is evidence of sponge mortality at extremely low temperatures in the British Isles (Crisp, 1964a). Given this evidence, it is likely that a cooling of 5°C for a month could potentially affect the characterizing sponges in shallow examples of this biotope. Resistance has been assessed as ‘Medium’, with a resilience of ‘High’. Sensitivity has, therefore, been assessed as ‘Low’ at the benchmark level.

Salinity increase (local)

Benchmark. An increase in one MNCR salinity category above the usual range of the biotope or habitat (Salinity regime change pressure definition).

Evidence

The biotope is recorded at full salinity. An increase at the benchmark level would result in hypersaline conditions. 

Kazanidis et al. (2019) reported how water temperature and salinity explained a significant amount of variation in sponge density. In the North Atlantic, the highest density values were found where temperature and salinity ranged from 6.52 to 8.98°C and 34.91 to 35.13 psu, respectively (Kazanidis et al., 2019). 

The marine demosponge Halichondria panicea dominates across the Baltic Sea, where salinity ranges from high saline conditions (28.3 psu) in the Kattegat to the Eastern limit of its distribution at lower salinities (11.4 psu) (Hoy et al., 2026). Marin (1998) describes the presence of Dysidea fragilis in a hypersaline coastal lagoon (at 42 to 47 g/l; 42 to 47 ppt) in La Mar Menor, Spain. 

Sensitivity assessment

The high-energy (extremely wave-surged) environment characteristic of this biotope is likely to mix and dilute the hypersaline (brine) effluent rapidly. However, as limited evidence was found for other members of the community, ‘Insufficient evidence’ is recorded for the biotope as a whole. 

Salinity decrease (local)

Benchmark. A decrease in one MNCR salinity category above the usual range of the biotope or habitat (Salinity regime change pressure definition detail).

Evidence

Kazanidis et al. (2019) reported how water temperature and salinity explained a significant amount of variation in sponge density. In the North Atlantic, the highest density values were found where temperature and salinity ranged from 6.52 to 8.98°C and 34.91 to 35.13 psu, respectively (Kazanidis et al., 2019).

Some of the characterizing species occur in harbours and estuaries, including Halichondria spp. (Ackers et al., 1992). Of the characterizing sponges present in this biotope, all except Clathrina coriacea have been recorded in variable salinity or lower (Connor et al., 2004). For example, Halichondria panicea dominates across the Baltic Sea, where salinity ranges from high saline conditions (28.3 psu) in the Kattegat to the Eastern limit of its distribution at lower salinities (11.4 psu) (Hoy et al., 2026).

Although not characterizing, the encrusting sponge Hymeniacidon perlevis is similar to Halichondria panicea, and Hoeke, Wasson & Kahn (2025) studied Hymeniacidon perlevis abundance off the Californian coast, USA. Their analysis suggested that sponge cover decreased with increasing freshwater input. For example, after several years of consistent recruitment, sponges only recruited to a single site in 2023 in the wake of frequent storms with heavy rainfall in central California from November 2022 to April 2023, which generated the heaviest precipitation observed during the study period of two years. Hoeke, Wasson & Kahn (2025) concluded that intense precipitation may have in part created unsuitable conditions for Hymeniacidon perlevis reproduction or recruitment, but that the species was highly adaptable, responding to environmental conditions, such as an increase in cover after two to four months after an increase in salinity. 

Castric-Fey & Chassé (1991) conducted a factorial analysis of the subtidal rocky ecology near Brest, France, and rated the distribution of other sponge 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. Mean salinity difference between the two farthest zones 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 their study did not reach the lower benchmark level, and at least some of the characterizing sponges are likely to be affected at the benchmark level. 

Sensitivity assessment

This biotope is subtidal and characterized by high energy, and any low salinity events would probably be short-lived, but the lack of records within “reduced” salinity (18 to 30‰) suggests the community would not persist/be recognisable if salinity was reduced. Some mortality is likely, with some of the sponges (e.g. Clathrina coriacea) affected (Connor et al., 2004). Therefore, resistance is assessed as ‘Medium’, resilience as ‘High’ and sensitivity as ‘Low’.

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s and 0.2 m/s for more than one year (Water flow pressure definition). 

Evidence

These biotopes are characterized by their high-energy wave surge. There is a progression within this group of biotopes, with IR.FIR.SG.CrSp occurring in the areas most subject to wave surge (found in the narrowest gullies). In slightly less surge-affected areas, ascidians (primarily Dendrodoa grossularia) become more abundant and therefore important in defining the character of the biotopes CrSpAsAn, DenCcor and CrSpAsDenB.

Sponges and anthozoans are suspension feeders, relying on water currents to supply food (Hiscock, 1983; O’Reilly et al., 2022). These taxa therefore thrive in conditions of vigorous water flow, e.g. around Orkney and St Abbs, Scotland, where Alcyonium digitatum dominated biotopes may experience tidal currents of 3 and 4 knots (approximately 1.5 m/sec) during spring tides (De Kluijver, 1993; Coolen et al., 2015). 

Riisgard et al. (1993) discussed the low energy cost of filtration for sponges and concluded that passive current-induced filtration may be insignificant for sponges. Pumping and filtering occur in choanocyte cells, which generate water currents in sponges using flagella (de Vos et al., 1991). 

The sponges Pachymatisma johnstonia and Dysidea fragilis and the anemones Corynactis viridis and Metridium senile have been recorded in biotopes from very weak to very strong water flow (0 to >3 m/s). Some studies have also stated how massive and encrusting sponge morphotypes are more abundant at high-flow areas in the sublittoral zone, due to a high basal area to volume ratio, which decreases removal from the substrata (Kazanidis et al., 2019).

In Lough Hyne, a small (approx. 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 the recolonization rates in areas, and it is possible that very slow current speeds could reduce the larvae supply to sites (Micaroni et al., 2025).

Sensitivity assessment

All the biotopes within this complex are characterized by their high energy due to wave surge, and a change in water flow is unlikely to affect the hydrodynamics of the biotope. Change at the benchmark level (0.1 to 0.2 m/s) is unlikely to be significant enough to alter the biotope, and resistance is, therefore, assessed as ‘High’, resilience is ‘High’, and the biotope is ‘Not sensitive’ at the benchmark level.

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. (Emergence regime change pressure definition).

Evidence

This biotope occurs in the infralittoral, however, it has been recorded in the 0-5 m depth band and may therefore be subject to emergence.

Increased emergence may reduce habitat suitability for characterizing species through greater exposure to desiccation and reduced feeding opportunities for the faunal community. Given the subtidal nature of the biotope, emergence is likely to result in a significant decline in the characterizing species. However, the topography of vertical walls, caves and gullies may afford some resistance to areas within these biotopes and resistance is therefore assessed as ‘Low’. Recovery is assessed as ‘High’, and sensitivity as 'Low'.

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year (Wave action pressure definition). 

Evidence

All the biotopes within this complex are defined by their exposure to wave action. As sublittoral biotopes, a change in wave height would affect the wave surge experienced by characterizing species. A decrease in wave surge could result in the succession of the IR.FIR.SG.CrSp to biotopes more dominated by ascidians and/or anemones. Conversely, an increase in wave surge could result in loss of the IR.FIR.SG.CrSpAsAn, IR.FIR.SG.CrSpAsDenB and IR.FIR.SG.DenCcor to more impoverished or scoured biotopes. 

Sponges and anthozoans are suspension feeders, relying on water currents to supply food (Hiscock, 1983; O’Reilly et al., 2022). These taxa, therefore, thrive in conditions of vigorous water flow.

Sensitivity assessment

All the biotopes within this complex (IR.FIR.SG) are characterized by their high energy. This biotope is recorded from very wave-exposed and wave-exposed habitats, where the wave energy is concentrated by the walls of gullies and caves to create wave surge, and is dominated by low-growing or encrusting forms tolerant of the wave surge (JNCC, 2022). This biotope (IR.FIR.SG.CrSp) occurs at the middle or rear of caves and gullies where surge is intensified (JNCC, 2022). Hence, a reduction in wave exposure is likely to transition the biotope into other members of the complex typical of less exposed conditions, e.g. CrSpAsAn, DenCcor or CrSpAsDenB (JNCC, 2022). These ascidian and barnacle-dominated communities would probably develop quickly and possibly within a year (the benchmark duration), so the biotope could be replaced (lost), although its characteristic species would remain. Therefore, resistance is assessed as ‘Low’. However, the invading species would probably be removed quickly once wave exposure returned to its prior state and resilience is assessed as ‘High’ (within 2 years) and sensitivity is assessed as ‘Low’. The assessment is based on the biotope classification and, hence, expert judgement. An increase in wave exposure from very exposed to extremely exposed is unlikely to result in significant changes.

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant Transitional metal or organometal (e.g. TBT) contaminants via uncontrolled releases or incidental spills (Transitional metals and organometals pressure definition).

Evidence

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

Gentric et al. (2016) studied the bioaccumulation of metallic trace elements and organic pollutants in marine sponges from the South Brittany Coast, France. The sponges Raspailia ramosa and Hymeniacidon perlevis are commonly found on the seashore in the region. Their results indicated that Al, Co, Cr, Fe, Pb, and Ti particularly accumulated in both species at high levels (Gentric et al., 2016). Gentric et al. (2016) noted how Hymeniacidon perlevis showed impressive bioaccumulation properties, accumulating Pb at levels 50- fold higher compared to oysters, and concentrated almost all polycyclic aromatic hydrocarbons (PAHs) in higher amounts than the other sponge species or oysters collected, such as benzo(a)pyrene, which bioaccumulated up to 17 times higher in Hymeniacidon perlevis compared to oysters. Orani et al. (2018) also observed Hymeniacidon perlevis having high accumulation properties for most transition elements, which is higher than other sponge species. While some sponges, such as Cliona spp., have been used to monitor heavy metals by looking at the associated bacterial community (Marques et al., 2007; Bauvais et al., 2015), little literature exists on the effects of transition element or organo-metal pollutants on the characterizing sponges could be found.

Hydrocarbon & PAH contamination

Benchmark. Exposure of marine species or habitat to one or more relevant hydrocarbon or polyaromatic hydrocarbon (PAH) contaminants via uncontrolled releases or incidental spills (Hydrocarbon & PAH pressure definition).

Evidence

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

Oil pollution is mainly a surface phenomenon; its impact upon infralittoral 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 sublittoral 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-dwelling organisms in areas where oil components are deposited by sedimentation (Zahn et al., 1981). 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 were uncertain (Zahn et al., 1981).

Gentric et al. (2016) studied the bioaccumulation of metallic trace elements and organic pollutants in marine sponges from the South Brittany Coast, France. The sponge Hymeniacidon perlevis is commonly found on the seashore in the region. Gentric et al. (2016) noted how Hymeniacidon perlevis showed impressive bioaccumulation properties, accumulating Pb at levels 50- fold higher compared to oysters, and concentrated almost all polycyclic aromatic hydrocarbons (PAHs) in higher amounts than the other sponge species or oysters collected, such as benzo(a)pyrene, which bioaccumulated up to 17 times higher in Hymeniacidon perlevis compared to oysters.

Ignatiades & Becacos-Kontos (1970) found that the ascidian Ciona intestinalis can resist the toxicity of oil-polluted water, and ascidia are frequently found in polluted habitats such as marinas and harbours, etc. (Carver et al., 2006), as well as Ascidia mentula (Aneiros et al., 2015).

Ryland & de Putron (1998) found no detectable damage to under-boulder communities, which are similar to some overhang communities, in Watwick Bay, Pembrokeshire, following the Sea Empress oil spill. Part of the resistance to effects might be because oil does not settle onto overhanging surfaces. However, some species, especially gastropods, are likely to be narcotised and killed, and some damage is likely. The return to a previous species composition would occur from new settlement from larval sources, although some gastropods have no or only a short dispersal phase, so that recovery will be slow.

Synthetic compound contamination

Benchmark. Exposure of marine species or habitat to one or more synthetic compound contaminants via uncontrolled releases or incidental spills (Synthetic compound contamination pressure definition).

Evidence

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

Hoare & Hiscock (1974) reported that the anemone Urticina felina survived near to an acidified halogenated effluent discharge in a 'transition' zone where many other species were unable to survive, suggesting a tolerance to chemical contamination. However, Urticina felina was absent from stations closest to the effluent, which were dominated by pollution-tolerant species (such as polychaetes). Those specimens closest to the effluent discharge appeared generally unhealthy.

Radionuclide contamination

Benchmark. An increase in 10µGy/h above background levels (Radionuclides contamination pressure definition).

Evidence

‘No evidence’ was found.

Introduction of other substances

Benchmark. Exposure of marine species or habitat to one or more relevant "other" substances (solid, liquid or gas) contaminants via uncontrolled releases or incidental spills (Introduction of other substances pressure definition). 

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) (deoxygenation pressure definition).

Evidence

It is likely that as this biotope occurs in areas that experience significant wave surge, re-oxygenation is likely, which would limit the effects of any de-oxygenation events. However, this may mean that the species present have little exposure to low oxygen and may be sensitive to this pressure. 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; Vaquer-Sunyer & Duarte, 2008). Cole et al. (1999) suggested possible adverse effects on marine species below 4 mg/l and probable adverse effects below 2 mg/l.

Halichondria panicea has been reported to survive under oxygen levels as low as 0.5 to 4% saturation (ca 0.05 to 0.4 mg/l) for up to 10 days (Mills et al., 2014). Riisgård (2024) studied the oxygen extraction efficiency and tolerance to hypoxia in sponges. In the demosponge Halichondria panicea, respiration rate was constant down to about 1.5 ml O₂/l, which showed that its oxygen extraction efficiency increased with decreasing oxygen concentration. Riisgård (2024) argued that the relationship between the filtration rate and oxygen consumption in filter feeders was controlled by the resistance to the diffusion of oxygen across the boundary layer between the feeding current and the tissues of the body, and that a high tolerance to hypoxia was a consequence of the adaptation to filter feeding, and sponges do not have a special capacity to overcome hypoxic events.

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 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).

No information was found relating to the sensitivity of the other characterizing species to de-oxygenation.

Sensitivity assessment

The hydrodynamics of this biotope would likely mean that hypoxic events would be rare and short-lived. The available evidence suggests that Halichondria panicea is tolerant of de-oxygenation, and resistance is therefore recorded as ‘High’, resilience as ‘High’, and the biotope is ‘Not sensitive’ at the benchmark level. The lack of evidence for other characterizing species results in a ‘Low’ confidence score.

Nutrient enrichment

Benchmark. Increased levels of the elements nitrogen, phosphorus, silicon, and iron in the marine environment compared to background concentrations (Nutrient enrichment pressure definition).

Evidence

This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations. Sponges and anthozoans are suspension feeders, and nutrient enrichment of coastal waters that enhances the population of phytoplankton may be beneficial in terms of an increased food supply, but the effects are uncertain (Hartnoll, 1998). The survival of the characterizing species may be influenced indirectly. High primary productivity in the water column combined with high summer temperature and the development of thermal stratification (which prevents mixing of the water column) can lead to hypoxia of the bottom waters, which faunal species are likely to be highly intolerant of (see de-oxygenation pressure).

Although not characterizing, the encrusting sponge Hymeniacidon perlevis is similar to Halichondria panicea, and Hoeke, Wasson & Kahn (2025) studied Hymeniacidon perlevis abundance off the Californian coast, USA. Specifically, Elkhorn Slough in California has a distinct wet season where freshwater runoff reduces salinity in the slough and increases particulates and nutrients that cause eutrophication and hypoxia, particularly in the upper estuary (Hoeke, Wasson & Kahn, 2025). Whereas many sponges are sensitive to water quality changes such as suspended sediments, Hymeniacidon perlevis can survive being partially buried in mud in winter and thrives in Elkhorn Slough despite muddy substrate and high suspended particulate loads (Caffrey et al., 2002b and Cao et al., 2007 cited in Hoeke, Wasson & Kahn, 2025). Over the course of their two-year experiment, Hoeke, Wasson & Kahn (2025) estimated that Hymeniacidon perlevis biomass and potential for water filtration are greatest in fall, corresponding with peak cover, and weakest or non-existent in the spring. In addition, in their observations, Hoeke, Wasson & Kahn (2025) only observed recruitment in upper estuary sites; adults of this species are also only found in the upper estuary. The upper estuary is more eutrophic than the lower estuary, with higher chlorophyll, temperature, and longer residence time, and is even considered “hypereutrophic” (Hughes et al. 2011 cited in Hoeke, Wasson & Kahn, 2025), but has the greatest levels of Hymeniacidon perlevis cover (Hoeke, Wasson & Kahn, 2025). Hoeke, Wasson & Kahn (2025) concluded that Hymeniacidon perlevis may grow in greater densities in the upper estuary due to increased food availability from eutrophication, particularly in sites where eutrophic water is provided via daily tidal flow through narrow channels. 

Tolerance by Hymeniacidon perlevis to eutrophic waters have been observed elsewhere, with the sponge growing faster and more densely in the eutrophic outflow water near a fish farm (particularly in autumn and winter) compared to neighbouring areas (Mercurio et al., 2023). The highest increase in sponge volume was observed in conditions of reduced fish biomass, increased pellet amount, and high wastewater flow (Mercurio et al., 2023). Mercurio et al. (2023) suggests that Hymeniacidon perlevis could be used for wastewater treatment.

Sensitivity assessment

Limited evidence on the effects of nutrient enrichment on the characteristic species was found. The evidence suggests that the characteristic species vary in their response to nutrients. However, the evidence is ‘insufficient’ to form the basis of an assessment.

Organic enrichment

Benchmark. A deposit of 100 gC/m²/yr (Organic enrichment pressure definition).

Evidence

Organic enrichment leads to organisms no longer being limited by the availability of organic carbon. The consequent changes in ecosystem function can lead to the progression of eutrophic symptoms (Bricker et al., 2008), changes in species diversity and evenness (Johnston & Roberts, 2009) and decreases in dissolved oxygen and uncharacteristic microalgae blooms (Bricker et al., 1999, 2008). Indirect adverse effects associated with organic enrichment include increased turbidity, increased suspended sediment and the increased risk of deoxygenation. 

Although not characterizing, the encrusting sponge Hymeniacidon perlevis is similar to Halichondria panicea and is examined as a proxy species. Fu et al. (2007) described Hymeniacidon perleve in aquaculture ecosystems in sterilized natural seawater with different concentrations of total organic carbon (TOC), at several concentrations between 52.9 and 335.13 mg/l. Hymeniacidon perleve removed 44 to 61% TOC during 24 hours, with retention rates of ca. 0.19 to 1.06 mg/hr ·g-fresh sponge. Hymeniacidon perleve removed organic carbon excreted by Fugu rubripes with similar retention rates of ca. 0.15 mg/h g-fresh sponge, and the sponge biomass increased by 22.8%.

Some of the characterizing sponges occur in harbours and estuaries, including Halichondria spp. (Ackers et al., 1992) and may, therefore, tolerate high levels of organic carbon.

Tolerance by Hymeniacidon perlevis to eutrophic waters have been observed elsewhere, with the sponge growing faster and more densely in the eutrophic outflow water near a fish farm (particularly in autumn and winter) compared to neighbouring areas (Mercurio et al., 2023). The highest increase in sponge volume was observed in conditions of reduced fish biomass, increased pellet amount, and high wastewater flow (Mercurio et al., 2023). Mercurio et al. (2023) suggests that Hymeniacidon perlevis could be used for wastewater treatment.

Dissolved organic matter has been demonstrated to be an important component of nutrition in tropical encrusting and boring sponges, and a higher amount of organic nutrients may directly benefit boring sponges (Stubler et al., 2024).

Sensitivity assessment

These biotopes occur in high-energy conditions, and it is likely that the deposited organic content would be rapidly removed. There is also evidence that the filter-feeding characterizing species would tolerate an increase in organic content. Resistance is therefore assessed as ‘High’, resilience as ‘High’, and the biotope is ‘Not sensitive’ at the benchmark level but with ‘Low’ confidence due to the lack of direct evidence.

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site (Physical loss pressure definition). 

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 (Physical change in subtratum type pressure definition).

Evidence

If rock were replaced with sediment, this would represent a fundamental change to the physical character of the biotope, and the species would be unlikely to recover. The biotope would be lost. The characterizing species each require a hard substratum to attach to, such as rock, steel, and other coralligenous or hard formations.

Halichondria spp. occur on a variety of inorganic and organic hard substrata, including mussel banks, small stones and rocks, and macroalgae (Peattie & Hoare, 1981 and Barthel, 1986, 1988, 1991 cited in Goldstein & Funch, 2022). Hymeniacidon perlevis specifically is found in diverse habitats, including both the intertidal and the subtidal zones, and it can grow buried in sediment or on hard substrata (Erpenbeck and Van Soest, 2002, cited in Turner, 2020). In addition, Halichondria bowerbanki has been observed growing on cave walls (Ereskovsky et al., 2018). Kazanidis et al. (2019) found that the type of substratum explained a significant amount of variation in sponge density, with the highest densities found in cobble with boulders.

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) (Physical change in sediment type pressure definition). 

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) (Removal of substratum pressure definition). 

Evidence

The species characterizing this biotope are epifauna or epiflora occurring on rock and would be sensitive to the removal of the habitat. However, extraction of rock substratum is considered unlikely, and this pressure is considered to be ‘Not relevant’ to hard substratum habitats.

Abrasion / disturbance of the surface of the substratum or seabed

Benchmark. Damage to surface features (e.g. species and physical structures within the habitat) (Surface abrasion/disturbance pressure definition).

Evidence

Fishing disturbance is one of the largest pressures for epifaunal communities (Kaiser et al. 2018; Kazanidis et al., 2019; Graves et al., 2023). Although, they are also vulnerable to other anthropogenic physical disturbances, such as those from oil and gas exploration, deep-sea mining, and recreational SCUBA diving (Vad et al., 2018; Betti et al., 2019; Graves et al., 2023).

The sponge Halichondria panicea is compressible but crumbly in texture, easily broken (Ackers et al, 1992), and is typically found in cryptic or semi-cryptic areas (Hayward & Ryland, 1995b). Abrasion events are therefore likely to remove the sponge, but no r, specific evidence was found.

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, 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 other sponges, including Haliclona oculata and Ircina campana, were not significantly affected. The abundance of sponges had increased to pre-trawl densities or greater 12 months after trawling.

Tilmant (1979) found that, following a shrimp trawl in Florida, the US, 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 was provided.

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 (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 of the sponges assessed (14%), were damaged by being knocked over. 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 (1999) on warm shallow sponge communities. Impacts of trawling activity in Alaska were 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 was considered likely to take many years for deep sponge communities to recover if adversely affected by physical damage (Freese, 2001).

Deep sponges are sessile benthic organisms, and recovery would be aided by the exclusion of demersal towed fishing equipment (Pikesley et al., 2016; Chimienti, Nisio, & Lanzolla, 2020; Kaiser et al., 2018). Kaiser et al. (2018) specifically studied the recovery of sessile epifauna following the exclusion of towed mobile fishing gear in Lyme Bay, UK. 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, while Eunicella verrucosa and Pentapora foliacea increased in abundance, but had not fully recovered, with their projected recovery time being 17 to 20 years (Kaiser et al., 2018). Therefore, recovery rates of biota depend on life-history factors and habitat-specific requirements, with the longer-lived species that require specific habitats and have low dispersal potential taking longer to recover (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. Key sessile taxa (Pentapora foliacea and Phallusia mammillata) showed signs of early recovery between 2008 and 2013. 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 was thought to occur over decadal timescales, and measuring full recovery rates in-situ remained a priority for future research in Lyme Bay.

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 also indicated the presence of large sponges, mainly Axinella infundibuliformis (Picton & Goodwin, 2007).

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 sponge Pachymatisma johnstoni was highly damaged by the experimental trawl.  However, other members of the faunal turf community were not as vulnerable to damage from trawling as sedimentary fauna, and whilst damage to circalittoral rock fauna did occur, it was of an incremental nature, with the loss of faunal turf communities increasing with repeated trawls. The anthozoans and ascidians are epifaunal, and physical disturbance is likely to cause damage and mortality. Emergent epifauna are generally very intolerant of disturbance from fishing gear (Jennings & Kaiser, 1998). 

Hall-Spencer & Moore (2000a) reported that sessile epifauna, including sponges, where present, were significantly reduced in abundance in dredged areas for four years post-dredging.

Sensitivity assessment

These biotopes are defined by their high-energy wave surge, which prohibits succession. However, they occur above cave and gully floors, which experience significant scour, and biotopes containing scour-resistant barnacles and spirobids tend to prevail. Whilst a large proportion of the faunal crust community is likely to be affected by abrasion events and result in localized damage, there is some debate as to the level of effects depending on the size of the sponge and the type of abrasion effect (Coleman et al., 2013). An increase in scour or other abrasion events is likely to remove the sponge components. Therefore, resistance is assessed as ‘Low’, resilience as ‘Medium’, and sensitivity as ‘Medium’.

Penetration or disturbance of the substratum subsurface

Benchmark. Damage to sub-surface features (e.g. species and physical structures within the habitat) (Sub-surface penetration pressure definition).

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 (Suspended sediment pressure definition).

Evidence

An increase in suspended solids is unlikely to have direct effects on the characterizing species due to clogging. However, given the high energy associated with the biotope, suspended particulates could increase the scour experienced by the characterizing species.

Despite sediment being considered to have a negative impact on suspension feeders (Gerrodette & Flechsig, 1979; Jackson, 2004), many encrusting sponges appear to be able to survive in highly sedimented conditions, and in fact, many species prefer such habitats (Bell & Barnes, 2000; Bell & Smith, 2004; 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 than encrusting forms) and living, at least partially, embedded within the sediment (Bell et al., 2015; Schönberg, 2015). 

Halichondria spp. are known to occur in harbours and estuaries, waterbodies known for increased suspended sediments (Ackers et al., 1992). Although not characterizing, the encrusting sponge Hymeniacidon perlevis is similar to Halichondria panicea, and Hoeke, Wasson & Kahn (2025) studied Hymeniacidon perlevis abundance off the Californian coast, USA. Specifically, Elkhorn Slough in California has a distinct wet season where freshwater runoff reduces salinity in the slough and increases particulates and nutrients that cause eutrophication and hypoxia, particularly in the upper estuary (Hoeke, Wasson & Kahn, 2025). Whereas many sponges are sensitive to water quality changes such as suspended sediments, Hymeniacidon perlevis can survive being partially buried in mud in winter and thrives in Elkhorn Slough despite muddy substrata and high suspended particulate loads (Caffrey et al., 2002b and Cao et al., 2007 cited in Hoeke, Wasson & Kahn, 2025). 

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 aurantium, 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. Clathrina antofagastensis in the South-eastern Pacific has been observed in intertidal and subtidal rocky substrata, inhabiting areas protected from sunlight (underneath boulders) and with a moderate to high concentration of sediments (Polo et al., 2022). Dunlop et al. (2021) also noted how the sponges Polymastia spp. and Phakellia spp. showed significant declines in density with increasing sedimentation and were the principal taxa in communities at natural sedimentation levels. In addition, Kazanidis et al. (2018) noted that cold-water sponges exclusively feed on dissolved organic matter particles smaller than 10 μm, and that suspended Particulate Organic Matter (POM) concentration decreased, along with its quality, with depth. For example, in surface waters, 0 to 150 m deep, the average concentration of suspended POM is about four times higher than between 150 and 4,000 m water depth.

Sensitivity assessment

Whilst an increase in suspended sediment may result in extra energetic expenditure in cleaning, it is unlikely to increase mortality directly. An increase in suspended sediment could result in scour due to wave action and tidal flow. Overall, resistance is assessed as ‘Medium’, mainly due to the scour effect. Resilience is ‘High’, and sensitivity is ‘Low’.

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 (Smothering pressure definition).

Evidence

This biotope typically occurs on vertical or very steep walls and overhangs, with more scour-resistant biotopes occurring on cave and gully floors. The majority of the examples of this biotope are therefore unlikely to be affected by smothering following an increase in sediment deposition. This being said, where the biotope occurs on the upward-facing surfaces (15% of records, Connor et al., 2004), mortality could occur, depending on the level and rate of removal of the sediment Despite sediment being considered to have a negative impact on suspension feeders (Gerrodette & Flechsig, 1979; Jackson, 2004), many encrusting sponges appear to be able to survive in highly sedimented conditions, and in fact, many species prefer such habitats (Bell & Barnes, 2000; Bell & Smith, 2004; 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 than encrusting forms) and living, at least partially, embedded within the sediment (Bell et al., 2015; Schönberg, 2015). While studying coralligenous assemblages of the Apulian continental shelf in the Mediterranean Sea, Piazzi et al. (2019) found that sedimentation was higher on deep outcrops and suggested that it was the main driver of differences between shallow and deep assemblages.

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 and noted that upright forms intercepted a smaller amount of settling sediment than encrusting forms. Whilst many sponges are disadvantaged by sedimentation (as would be expected, being sessile filter feeders) (Gerrodette & Flechsig, 1979), many examples exist of sponges adapting to sediment presence (Bell et al., 2015; Schönberg, 2015) and 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). 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). 

Halichondria spp. are known to occur in harbours and estuaries, waterbodies known for increased suspended sediments (Ackers et al., 1992). Halichondria panicea is mainly found in shallow, protected coastal regions of the eastern parts of the North Atlantic, and shows adaptation to frequent air exposure, while Halichondria bowerbanki is most common in exposed habitats of the western parts, where it tolerates high levels of siltation (Vethaak et al., 1982 cited in Goldstein & Funch, 2022). Hymeniacidon perlevis is encrusting (up to 2.5 mm height) or cushion-forming up to 15 cm across and 2 to 5 cm thick (Hayward & Ryland, 1995b). It can tolerate smothering of the main body, with the papillae protruding through the silt layer (Hayward & Ryland, 1995b).

In a laboratory study on the impact of elevated sedimentation rates on deep-water sponges, a 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).

Sensitivity assessment

Whilst smothering by 5 cm of sediment may result in extra energetic expenditure in cleaning, it is unlikely to increase mortality, particularly if the thin layer of sediment is removed quickly, as the removal of the sediment is likely to be rapid (within one tidal cycle) in the wave-exposed conditions and resultant wave surge. Vertical, steep slopes and overhangs are unlikely to be affected. Therefore, resistance is assessed as ‘High’ resilience and ‘High’, and sensitivity as ‘Not sensitive’ at the 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 (Smothering pressure definition).

Evidence

This biotope typically occurs on vertical or very steep walls and overhangs, with more scour-resistant biotopes occurring on cave and gully floors. The majority of the examples of this biotope are therefore unlikely to be affected by smothering following an increase in sediment deposition. This being said, where the biotope occurs on the upward-facing surfaces (15% of records, Connor et al., 2004), mortality could occur, depending on the level and rate of removal of the sediment.

Despite sediment being considered to have a negative impact on suspension feeders (Gerrodette & Flechsig, 1979; Jackson, 2004), many encrusting sponges appear to be able to survive in highly sedimented conditions, and in fact, many species prefer such habitats (Bell & Barnes, 2000; Bell & Smith, 2004; 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 than encrusting forms) and living, at least partially, embedded within the sediment (Bell et al., 2015; Schönberg, 2015). While studying coralligenous assemblages of the Apulian continental shelf in the Mediterranean Sea, Piazzi et al. (2019) found that sedimentation was higher on deep outcrops and suggested that it was the main driver of differences between shallow and deep assemblages.

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 and noted that upright forms intercepted a smaller amount of settling sediment than encrusting forms. Whilst many sponges are disadvantaged by sedimentation (as would be expected, being sessile filter feeders) (Gerrodette & Flechsig, 1979), many examples exist of sponges adapting to sediment presence (Bell et al., 2015; Schönberg, 2015) and 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). 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). 

Halichondria spp. are known to occur in harbours and estuaries, waterbodies known for increased suspended sediments (Ackers et al., 1992). Halichondria panicea is mainly found in shallow, protected coastal regions of the eastern parts of the North Atlantic, and shows adaptation to frequent air exposure, while Halichondria bowerbanki is most common in exposed habitats of the western parts, where it tolerates high levels of siltation (Vethaak et al., 1982 cited in Goldstein & Funch, 2022). Hymeniacidon perlevis is encrusting (up to 2.5 mm height) or cushion-forming up to 15 cm across and 2 to 5 cm thick (Hayward & Ryland, 1995b). It can tolerate smothering of the main body, with the papillae protruding through the silt layer (Hayward & Ryland, 1995b).

In a laboratory study on the impact of elevated sedimentation rates on deep-water sponges, a 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).

Sensitivity assessment

Whilst smothering by 5 or 30 cm of fine sediment (the benchmark) may result in extra energetic expenditure in cleaning and could increase mortality in smaller species, it is unlikely, as the removal of the sediment is likely to be rapid (within one tidal cycle) in the wave-exposed conditions and resultant wave surge. Vertical, steep slopes and overhangs are unlikely to be affected. Overall, resistance to smothering by 30 cm of fine sediment on upward-facing surfaces in some examples of the biotope is assessed as ‘Medium’, mainly due to the possible scour effect as the sediment is removed, although there is the risk that some parts of the biotope that exist on flat or gently sloping areas may experience short-term burial. Resilience is ‘High’, and sensitivity is ‘Low’. Vertical, steep slopes and overhangs are unlikely to be affected.

Litter

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

Evidence

All characterizing species for this biotope are sessile epifauna, being either encrusting or branching. Physical disturbance by fishing gear has been shown to adversely affect sessile benthic and emergent epifaunal communities, with hydroid and bryozoan matrices reported to be greatly reduced in fished areas and increased when fishing activity is removed (Jennings & Kaiser, 1998; Sheehan et al., 2017; Kaiser et al., 2018). Both Sheehan et al. (2017) and Giusti et al. (2019) highlight how, in addition to the direct damage from fishing, ghost fishing may also be responsible for some coral mortality, Eunicella verrucosa in this case, either through direct damage or making them more vulnerable to removal from their anchorage to the sea floor, particularly during storms. Branching sponges would be similarly affected.

There are no records of ghost fishing affecting the characterizing species for this biotope. However, epifaunal communities are vulnerable to damage from fishing gear and are likely vulnerable to being dislodged or damaged through lost fishing gear, and possibly certain types of marine litter. At present, there is 'Insufficient evidence' to complete a sensitivity assessment on the effect of litter on this biotope.

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT (Electromagnetic pressure definition).

Evidence

Evidence on the effect of electromagnetic fields (EMFs) on benthic organisms is still severely lacking. 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. 

No studies have examined the effect of EMFs on the characterizing species, and there is ‘Insufficient evidence’ on which to base an assessment of the likely sensitivity of this biotope to EMFs.

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 could be found for the effect of noise or vibrations on the characterizing species of these biotopes, it is unlikely that these species have the facility for detecting noise or 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 (Introduced light or shade pressure definition).

Evidence

These biotopes have very limited algal abundance, although Connor et al. (2004) notes that kelps are unable to colonize the biotope primarily due to the level of wave surge and vertical topography. The impact of the introduction of light may therefore be of limited importance to increasing the algal component of the biotope.

Jones et al. (2012) compiled a report 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, presumably due to a lack of competition from algae. For example, the sponge Halichondria bowerbanki has been observed growing on cave walls, in the middle parts of caves, where there is a drastic reduction of incoming light (Ereskovsky et al., 2018).

Although not characterizing, the encrusting sponge Hymeniacidon perlevis is similar to Halichondria panicea, and Gastaldi et al. (2016) monitored Hymeniacidon perlevis in a stressful environment, San Antonio Bay, northern Patagonia. Sponges were exposed to a range of UV (mean daily dose from 200 to 1,500 kJ/m) and Photosynthetically Active Radiation (PAR) (mean daily dose from 200 to 1,400 kJ/m) for two years. They reported that seawater temperature, as well as UV and PAR solar radiation, was negatively correlated with subtidal Hymeniacidon abundance. Clathrina antofagastensis in the South-eastern Pacific has been observed in intertidal and subtidal rocky substrata, inhabiting areas protected from sunlight (underneath boulders) and with a moderate to high concentration of sediments (Polo et al., 2022).

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, 2015b; 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

Whilst little evidence was found for the effect of light on a few characteristic species of these biotopes, it is unlikely that these species would suffer mortality. As an infralittoral biotope focused on cave and gully walls which are naturally shaded from light, a decrease in light is unlikely to be important, and an increase at the benchmark level is unlikely to be significant, as growth ceases for a number of red algae (such as Chrondrus crispus) below ca 1.0 μmol m^-2l^-1 (ca 50 Lux), whereas this value is 2 μmol m^-2l^-1 (ca 100 Lux) for green algae (Leukart & Lüning, 1994). Whilst sponges seem to favour shaded areas in which to settle, it is unlikely that changes at the benchmark pressure would be significant.  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 (Barrier to species movement pressure definition).

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 (Death for collision pressure definition).

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 (Visual disturbance pressure definition). 

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 may result in changes in the genetic structure of local populations, hybridization, or a change in community structure (Translocation pressure definition).

Evidence

No evidence’ was 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) (pathogen or disease pressure definition).

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 to 95% disappearance of sponge specimens) (Galstoff,1942) and 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 1989 (Webster, 2007). However, Longo et al. (2016) studied the ability of the similar demosponge, Hymeniacidon perlevis, to act as a bacterial mitigation tool in aquaculture. Specimens of Hymeniacidon perlevis were collected from the Mar Grande (Northern Ionian Sea, Italy), and laboratory experiments showed Hymeniacidon perlevis had a high efficiency to accumulate bacteria within 24 hours (Longo et al., 2016). In particular, the abundances of culturable bacteria at 37°C and heterotrophic bacteria at 22°C were lower already after 4 hours from the beginning of the experiment, whilst the same trend was recorded after 24 hours for vibrios, total coliforms, Escherichia coli and intestinal streptococci (Longo et al., 2016). In addition, Hymeniacidon perlevis accumulated not only bacteria of marine origin but also non-indigenous pathogenic microorganisms (Longo et al., 2016). Longo et al. (2016) concluded that sponges demonstrate a high efficiency in removing all the considered bacterial groups compared to other filter feeders, such as mussels, and due to the conspicuous bacterial accumulation by the sponge, Hymeniacidon perlevis is a powerful tool in reducing the bacterial load in shellfish culture areas.

Sensitivity assessment

Whilst mass mortality and even extinction have been reported further afield, sponge diseases have caused limited mortality in some species in the British Isles. Due to 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’.

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale (targeted removal pressure definition).

Evidence

No evidence of targeting of the characterizing species could found and the pressure is ‘Not relevant’ to this biotope group.

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 (non-targeted removed pressure definition).

Evidence

The characteristic species probably compete for space within the biotope, so that loss of one species would probably have little if any effect on the other members of the community. However, the incidental removal of the characteristic epifauna is likely to remove a proportion of the biotope and change the biological character of the biotope. 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 on this biotope. 

Sensitivity assessment

Based on broad agreement  of trawl impacts on sponge communities and the likely disturbance to the sessile epifaunal species, resistance is recorded as ‘Low’, resilience is recorded as ‘High’, and Sensitivity is ‘Low’

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 & Callaway, 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 to 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 infralittoral rock characterizing this biotope could be suitable for colonization by Crepidula fornicata where less exposed wave conditions are found, where water is mediated by tidal flow rather than wave action. Crepidula has been recorded from areas of strong tidal streams (Hinz et al., 2011), and has been recorded from the lower intertidal to ca 160 m in depth, but it 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). However, no evidence was found of the effect of Crepidula populations on faunal turf-dominated habitats or infralittoral rock habitats. At present, there is 'Insufficient evidence' to suggest that the infralittoral biotopes are sensitive to colonization by Crepidula fornicata; 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-mediated 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 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 and 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 & Grosholz, 2015; Vercaemer et al., 2015).

Furthermore, the characterizing ascidian Clavelina lepadiformis was observed growing in abundance in Fangar Bay, at the northern side of the Ebro Delta in the NE Iberian Coast on PVC plates (between May and September 2015 at 0.2, 1, and 2 m deep) (Casso et al., 2018). Didemnum vexillum was also reported growing alongside it on the PVC plates and, therefore, likely competes with Clavelina lepadiformis for space and food (Casso et al., 2018).

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 are enough to cause dislodgement (Reinhardt et al., 2012).  

Sensitivity assessment

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 biotope occurs on infralittoral rock, 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 weak water flow (<1 kn; <0.5 m/s) but exposed to very exposed wave exposure that results in extreme wave surge. The community is dominated by thin and encrusting surge-resistant forms, and it is unclear if Didemnum could colonize the biotope. If Didemnum sp. could gain a 'foothold', it might overgrow, smother or cause mortality of epifauna. 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 biotope is probably 'Not sensitive to this INIS.

Wireweed, Sargassum muticum

Evidence

The wave surge, combined with the low levels of light within this biotope, particularly the rear walls of caves, are considered to inhibit colonization of invasive algal species. Hence, it is unlikely to be colonized by Sargassum. Therefore, this biotope is probably 'Not sensitive to this INIS.

Wakame, Undaria pinnatifida

Evidence

The wave surge, combined with the low levels of light within this biotope, particularly the rear walls of caves, are considered to inhibit colonization of invasive algal species. Hence, it is unlikely to be colonized by Undaria. Therefore, this biotope is probably 'Not sensitive’ to this INIS.

Other INIS

Evidence

Wave surge in this biotope will probably limit the establishment of all but the most surge-resistant invasive non-indigenous species (INIS), and no direct evidence was found for the effects of INIS on this biotope. The low levels of light within this biotope, particularly the rear walls of caves, are also considered to inhibit invasive algal species.

The Australasian barnacle Elminius modestus was introduced to British waters on ships during the Second World War. The species does well in estuaries and bays, where it can displace Semibalanus balanoides. However, on exposed shores, the native species outcompete this invasive species (Raffaelli & Hawkins, 1999).

Two non-native spirobids, Dexiospira oshoroensis and Pileolaria rosepigmentata, were found on the non-native algae Sargassum muticum in Portsmouth (Knight-Jones et al., 1975). Invasive tubeworms are reported from UK harbours (Thorp et al., 1986) and are likely to be well established in areas with large volumes of ship traffic.

There was ‘No evidence’ regarding other known invasive species that may pose a threat to IR.FIR.SG.CrSpAsAn, IR.FIR.SG.CrSpAsDenB, IR.FIR.SG.DenCcor or IR.FIR.SG.CrSp. Due to the constant risk of new invasive species, the literature for this pressure should be revisited.