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

All characterizing sponges (Cliona celata, Halichondria bowerbanki, Halichondria panicea, Hymeniacidon perleve, Leucosolenia botryoides, Suberites ficus, Raspailia ramosa, Amphilectus fucorum) are widely distributed across the coasts of the British Isles and are all found from the Channel Isles to northern Scotland (Ackers, 1993; NBN, 2015).

As well as its wide British Isles distribution, Hymeniacidon perlevis (syn. Hymeniacidon perleve) have been reported more widely and as southerly as Cape Verde (Van Soest, 1993) and the Azores (Boury-Esnault & Lopes, 1985) respectively. Hymeniacidon perlevis specifically has a substantial range from Norway in the north to the Macaronesian Islands off Africa in the south (Erpenbeck and Van Soest 2002 cited in Turner, 2020).

In southern Italy, Hymeniacidon perlevis near fish farms (in the farms drainage conduits) live, grow and reproduce in constant temperatures (18°C). As a result, individuals display an active state during the entire year, while avoiding stages of decline and long dormancy usually observed in wild populations, although Hymeniacidon perlevis at the fish farm did seem to grow more in autumn and winter (Mercurio et al., 2023). In addition, it seems plausible that stable environmental conditions induce an almost continuous sexual phase, probably under the control of endogenous factors (Mercurio et al., 2023). During their observations, Mercurio et al. (2023) saw no asexual reproduction take place, however, it was impossible to exclude the contribution of asexual reproduction in the origin of the newly settled sponges, which were repeatedly detected throughout the study. Mercurio et al. (2023) concluded that the absence of a typical four-stage life cycle is probably due to the constant water temperature experienced here, which aligned with the optimal temperature range (10 to 20°C) for the vital maintenance stage in Hymeniacidon perlevis (Stone, 1970, Cao et al., 2007, and Gaino, Cardone & Corriero, 2010 cited in Mercurio et al., 2023).

In comparison, in wild populations of Hymeniacidon perlevis, the onset of gametogenesis seems to be triggered by rapid spring changes in water temperature. At Mar Piccolo of Taranto, Mercurio et al. (2023) cited that Gaino, Cardone & Corriero (2010) reported the water temperature increase in April promoted female gamete differentiation, which precedes the presence of spermatic cysts by one month. A similar trend was also observed at the Étang de Thau, where gamete differentiation in spring was triggered by increased water temperature, and the species showed successive hermaphroditism in which oogenesis preceded spermatogenesis (Diaz, 1973 cited in Mercurio et al., 2023).

Gastaldi et al. (2016) monitored Hymeniacidon perlevis in a stressful environment, San Antonio Bay, northern Patagonia, where the sponge is affected by a semidiurnal macrotidal regime (up to 9 m) and extreme atmospheric temperatures range from -7.7°C in July (austral winter) and 41.4°C in February (austral summer); seawater temperature ranged from 5°C in August (austral winter) and 28°C in February (austral summer). They noted how seawater temperature, as well as UV and PAR solar radiation, was negatively correlated with subtidal Hymeniacidon abundance.

Hoeke, Wasson & Kahn (2025) studied Hymeniacidon perlevis abundance off the Californian coast, USA. Their analysis suggested that sponge cover increased and correlated with warmer temperatures and lower dissolved oxygen at all sites, with a time lag of two to four months. In addition, Hymeniacidon perlevis may fare particularly well in relatively warm estuarine intertidal zones, as newly settled juveniles were noted growing faster when incubated in warmer water, up to 23 °C (Xue & Zhang 2009 cited in Hoeke, Wasson & Kahn, 2025). However, sudden population declines have been observed at high temperatures, such as those in the Ionian Sea just after temperatures peaked at 26.8°C (Gaino et al., 2010 cited in Hoeke, Wasson & Kahn, 2025). Feeding rates of Hymeniacidon perlevis also vary with water temperature, with peak removal rates previously documented at 15°C and less removal at higher and lower temperatures (Zhang et al., 2010 cited in Hoeke, Wasson & Kahn, 2025).

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).

The sponge Cliona celata is found throughout the Atlantic between 1 and 200 m deep and has been recorded from Sweden to the Mediterranean in Europe, across the Southeastern USA in North America, and off Argentina in South America (Ackers et al., 1992; Stubler et al., 2024; Novarin et al., 2025). Cliona celata have been recorded in waters with a temperature range of 23.19 to 35.11°C off Southeastern USA (Stubler et al., 2024), and in a similar boring sponge, Pione truitti, growth was documented to decrease as temperatures dropped below 20°C (Pomponi & Meritt, 1985 cited in Stubler et al., 2024). One way in which sponges respond to increases in temperature is with higher respiration rates, and Cliona celata has been recorded as having a significant change in respiration rate as a stress response (Wood et al., 2025). However, Cliona celata have been reported to be resistant to temperatures of up to 4 to 5°C higher than ambient temperatures (Bosch-Belmar et al., 2024). Duckworth & Peterson (2013) demonstrated that an increase in water temperature to 26 and 31°C had no detectable effect on the boring activity of Cliona celata. Whereas cold winter temperatures (<5°C) can cause boring activity to cease, the incurrent papillae to withdraw and the excurrent papillae constrict (Fell et al., 1984; Carver et al., 2010). Indicating cold temperatures are more limiting to Cliona celata than hot temperatures. Piscitelli et al. (2011) observed an annual peak in reproductive activity in April to May from individuals in the Mediterranean and suggested this was a result of a sharp seasonal increase in water temperature. This suggests variable temperatures could affect larval recruitment processes, but not otherwise negatively affect Cliona celata.

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). In particular, 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 (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.

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. 

Leucosolenia spp. are Boreal and Arctic calcareous sponges, and populations in the White Sea, Russia, experience seawater temperatures between 6 and 10°C from June to September (Lavrov et al., 2018; 2024). However, Leucosolenia sp. have been recorded in bottom water temperatures as low as 1.91 to -1.38°C in Prydz Bay, East Antarctica, where it was, along with Halichondria sp., the most prominent sponge present (Liu et al., 2020).

There is limited information available about the tolerance of the other 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, and 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. 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).

Berman et al. (2013) monitored sponge communities off Skomer Island, UK, over three years, with all characterizing sponges for this biotope assessed. Seawater temperature, turbidity, photosynthetically active radiation and wind speed were all recorded during the study. It was concluded that, despite changes in species composition, primarily driven by the non-characterizing Hymeraphia, Stellifera and Halicnemia patera, no significant difference in sponge density was recorded in all sites studied. Morphological changes most strongly correlated with a mixture of water visibility and temperature. In addition, Goodwin et al. (2013) found little evidence to suggest that rising seawater temperatures (ca 1-2°C) had an effect on subtidal benthic assemblages in Northern Ireland between pre-1986 and post-2006 surveys. However, significant effects were noted in rarer species at the edge of the biogeographic ranges (Goodwin et al., 2013). 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. 

Nemertesia ramosa is found inshore to deeper water, being common throughout the British Isles and is distributed from Iceland to north-west Africa (Hayward & Ryland, 1994). In the Azores, for example, it is well known in the sublittoral at 15 to 158 m deep in the central group of islands, but historical records exist from seamounts down to bathyal grounds of nearly 1000 m (Gomes-Pereira & Tempera, 2016). The hydroid Plumularia setacea is often described as cosmopolitan and has been frequently reported along the Atlantic coasts of West Africa, including Guinea-Bissau (cf. Ansín Agís et al., 2001 cited in Moura et al., 2025). Cantero et al. (2002) described the presence and year-round fertility of Obelia dichotoma, Kirchenpaureria pinnata, Nemertesia ramosa and Halecium spp. in the Mediterranean, indicating probable tolerance to temperature increases at the benchmark figure. The hydroid genus Eudendrium is a monophyletic taxon with a worldwide distribution that includes more than 70 nominal species, and in the Mediterranean, a total of 15 species of Eudendrium are listed (González-Duarte, Megina & López-González, 2023). Eudendriums spp. are also found in cooler waters such as the Bay of Fundy, Canada (Calder, 2017).

Moglia et al. (2025) monitored the long-term shifts in hydrozoan communities by replicating(in 2024) an annual survey originally conducted in 1980 and repeated in 2004 along a depth gradient along the rocky cliff of Portofino Promontory (Ligurian Sea). They noted that species richness declined significantly from 83 (1980) to 72 (2004) and 43 (2024), with the greatest losses in cold-affinity species. Several winter-dominant species, such as Eudendrium glomeratum, Sertularella crassicaulis, and Ectopleura larynx, have strongly reduced or disappeared (Moglia et al., 2025). In contrast, summer species like Eudendrium racemosum are now present year-round (Moglia et al., 2025). The appearance of southern species like Corydendrium parasiticum and Pennaria disticha, along with a reduction in sexual reproduction, reflects a community-level response to ocean warming, leading to seasonal and bathymetric homogenization (Moglia et al., 2025). Moglia et al. (2025) concluded that this 44-year dataset provided a rare long-term benchmark and supported the importance of accurate taxonomic analysis for understanding climate-driven transformations in Mediterranean benthic ecosystems, particularly for hydrozoans, where data is scarce.

Clavelina lepadiformis is native to the Mediterranean Sea, Atlantic subarctic regions (i.e., Norwegian Sea and Greenland), and the Bay of Biscay, and has successfully spread to the coastline of South Africa, the coasts of North and Central America, and Brazil (Van Name, 1945, Turon et al., 2003, Pyo & Shin, 2011, and Reinhardt et al., 2010 cited in Hiebert et al., 2022). Reproduction of Clavelina lepadiformis is temperature-dependent (Millar, 1970; Berrill, 1975). A change in temperature could affect the time and duration of spawning, however, the distribution of Clavelina lepadiformis extends to the north and south of the British Isles from Norway to the Adriatic (Hayward & Ryland, 1996).

In lab experiments involving Clavelina lepadiformis collected from the harbour of La Spezia, Italy, Hiebert et al. (2022) noted that the production of their dormant form, the ‘winter buds’, occurred at 18°C. Adult zooids of Clavelina lepadiformis were also exposed to abrupt shifts in temperature (−20°C to 37°C) and salinity (10 to 44 ppt) for 24 hours, then transferred to 24°C, and the viability of zooids and the germination of dormant forms were observed after one week. Clavelina lepadiformis zooids were more resistant to lower temperatures (between 0 and 28°C) and higher salinity (30 to 45 ppt), and winter buds resisted a wider range of conditions compared to zooids (temperatures between 0 and 32°C and salinity between 15 and 45 ppt) (Hiebert et al., 2022). Clavelina lepadiformis have also been found to exist in the absence of zooids, and winter buds without zooids have been documented in the North Sea in the winter months (Orton, 1914 and Orton, 1921 and cited in Hiebert et al., 2022), in the summer months in the eastern Mediterranean, and Clavelina regress during summer months (called aestivation) in the western Mediterranean (Caralt et al., 2002; Mukai, 1977 cited in Hiebert et al., 2022). Hiebert et al. (2022) concluded that their results showed that the dormant stages of Clavelina lepadiformis were resistant to extremes in temperature and salinity. Thus, this linked capacity for dormancy and regeneration is likely an important adaptation that allows these species to survive drastic seasonal changes.

Casso et al. (2018) observed Clavelina lepadiformis growing in Fangar Bay, at the northern side of the Ebro Delta in the North East Iberian Coast and noted how Clavelina lepadiformis showed a slightly higher occurrence during colder seasons but was present throughout the year. Reproduction of Clavelina lepadiformis is highly seasonal, with larval occurrence restricted to two to three months during winter and spring (Caralt et al., 2002). Colonies disappear during summer (aestivation), and only survive as dormant stolons, with re-growth occurring in spring (Caralt et al., 2002; Picton & Morrow, 2004c).

In the ascidian Dendrodoa grossularia, gamete release occurs from spring to autumn, with peaks in early spring and another in late summer. Gamete release is reduced at temperatures above 15°C and totally suppressed above ca 20°C (Millar, 1954). No information was found on the upper temperature threshold of mature Dendrodoa grossularia. Whilst widespread throughout the British Isles (NBN, 2015), a dramatic increase in temperature may cause mortality. 

The ascidian Diplosoma listerianum has been observed spreading along the south-western coast of Iceland from the temperate zone where it is considered non-indigenous (Micael et al., 2022). It is suspected that this northward expansion is due to warming oceans, and the proliferation of Diplosoma listerianum in newly colonized areas will benefit from the further projected warming of sea-surface temperatures (Micael et al., 2022). As well as a warming-induced northward expansion, Diplosoma listerianum has been observed colonizing surfaces more quickly at warmer temperatures, in this case 12 to 16, 16 to 20, and 20 to 24 °C (Lord, 2017b), with growth rate peaking for the species around 20°C (Lord, 2017).

Species belonging to the bryozoan genus Bowerbankia are found fouling in the warm waters around the Galapagos Islands in the Pacific Ocean (McCann et al., 2019).

Bugula spp. have been recorded as far north as Trondheim, Norway (Christie et al., 2003). In the south, Bugula neritina is recorded in the Arabian Sea (Molnar et al., 2008), and Bugulina turbinata (syn. Bugula turbinata) and Crisularia plumosa (syn. Bugula plumosa) are recorded in the Iberian Peninsula (Ramos, 2010). Bugulina turbinata is a predominantly southern species in British waters (Lewis, 1964; Hayward & Ryland, 1998) but has been recorded as far north as Shetland (NBN, 2016). Furthermore, Bugula spp. and Crisularia spp. are recorded in the Eastern Mediterranean and are found in the Madeira Archipelago (Ramalhosa, Souto, & Canning-Clode, 2017; Koçak & Bakal, 2019). A long-term increase in temperature may increase the abundance of Bugula spp. in northern British waters and allow species to extend their range. Due to Bugula spp. occurring in warmer waters, they are likely to tolerate increases in temperature, at the benchmark level. However, Cocito & Sgorbini (2014) studied spatial and temporal patterns of a colonial bryozoan (Pentapora fascialis) in the Ligurian Sea over nine years. High temperature events were recorded, with the first causing mass mortality among a number of species, including bryozoans, sponges, corals, gorgonians, and encrusting coralline algae. 

In particular, Bugulina flabellata is considered native and widely distributed in western Europe and was recently documented as introduced in Australia and New Zealand, presumably having travelled on ships (Gordon & Mawatari, 1992 and Ryland et al., 2011 cited in Ramalhosa, Souto, & Canning-Clode, 2017). Along the Pacific coast of Chile (18.4°S to 41.9°S), Bugulina flabellata was associated with tidepools in the upper and lower tidal zone at both high (41.8°S) and low (22.1°S) latitudes, but was more common towards low latitudes, where high temperatures (fluctuating between 13 and 33°C) were registered in tidepools (López-Gappa et al., 2022). Bugula neritina was associated with low latitudes (~20.2°S and 23.7°S) and the upper intertidal zone, where tidepools are distant from the sea and associated with extreme abiotic conditions of temperature and salinity (López-Gappa et al., 2022). López-Gappa et al. (2022) concluded that both temperature and salinity influenced the ecological niches of the studied Bugula spp. In Otago Harbour, Aotearoa, New Zealand, Bryozoa exhibited both a susceptibility and tolerances to increases in heat; significance was found between increases of +1 or +2°C and the colony and zooid characteristics of Beania sp., and in the composition of Caberea zelandica, however, no significance was found in any parameter of Bugulina flabellate (Feary, 2024). Feary (2024) concluded that Bugulina flabellata is likely relatively robust to temperature changes, and that the increase in temperature (+1 to +2°C) was not high enough to elicit a response from those taxa. Therefore, Bugulina flabellata is tolerant of small increases in temperature, though there is still likely to be a thermal upper limit for them, such as experiencing necrosis in the range of 26 to 29°C as has been documented in other bryozoans (Feary, 2024).

Bugula spp. grow and reproduce in the summer months. However, day length and/or the phytoplankton bloom characteristic of temperate waters are probably more important than temperature (Ryland, 1967; 1970). Recruitment for Bugula spp. is known to occur between 19 and 24°C, and it can survive temperatures between 6 and 30°C (Lord, 2017; Jurgens et al., 2018; Kocak, Kucuksezgin & Bakal, 2019; Kitamura & Hirayama, 1984 cited in Fortic et al., 2025). For example, at Zikim beach, in the eastern Mediterranean, Bugula neritina and other Bugula spp. were observed in mean sea water surface temperatures of 20.27 ± 1.81°C (Ros, Guerra-García & Hoffman, 2016).

Lange & Marshall (2017) conducted Bugula larvae settlement experiments against multiple stressors: larval duration prolonged by two hours, copper exposure at 65 μg/L, salinity at 30 psu, and temperature at 22°C. Individuals in the control group (larval duration not prolonged, no exposure to any additional copper other than trace amounts naturally occurring in seawater, salinity at 36 psu, and temperature at 17.5°C) survived better than those exposed to the multiple stressors (Lange & Marshall, 2017). However, those individuals/colonies that survived the stressor treatments showed slight carry-over effects to later developmental stages, such that colonies that had experienced warmer temperatures as a single stressor produced slightly more offspring (Lange & Marshall, 2017). Lange & Marshall (2017) concluded that stressor exposure had strong effects on early survival in Bugula, but only a few on subsequent fecundity and growth.

In a follow-up study, Marshall (2021) studied how temperature selected offspring size in Bugula neritina over four years. Temperature was kept at 18°C (±0.5) during settlement but then varied between 10.5 and 22.5°C during the experimental period. At temperatures cooler than 17°C, selection on both mothers and offspring favoured larger offspring, and at the hottest temperature (24°C), selection on both mothers and offspring favoured smaller offspring (Marshall, 2021). However, between these two temperatures (18 to 22°C), selection on offspring favoured larger offspring sizes, but selection on mothers favoured smaller offspring sizes (Marshall, 2021). Marshall (2021) concluded that offspring size closely tracked the local environmental temperature across cohorts, with this offspring size-temperature covariance appearing to be adaptive, at least from the perspective of mothers. When temperatures were warmer, the relationship between offspring size and performance was weak; when temperatures were cooler, the relationship was strongly positive.

Gauff et al. (2025) observed the effect of an eight-day stimulated heatwave (5°C above the threshold of 18°C) on Bugula neritina. The heatwave simulation followed the daily sequence +0.5°C, +1.5°C, +3.5°C, +5°C, +5°C, +3.5°C, +1.5°C, +0.5°C until the max temperature of 23°C was reached on day eight. Gauff et al. (2025) noted how the community structure remained vastly unaffected, as did the total metabolome (metabolism measured via the complete collection of small-molecule chemicals, or metabolites, found within a biological sample) of Bugula neritina, showing that the community was able to resist the disturbance of the heatwave. Bugula neritina additionally showed a certain metabolic resilience as the already minor differences in the metabolome between control and the heatwave treatment diminished even further between the tested time steps (Gauff et al., 2025). However, despite showing resilience at 23°C, Bugula neritina has been observed to be sensitive to high-intensity Mediterranean heatwaves at 28°C (Gauff et al., 2023 cited in Gauff et al., 2025). Furthermore, Bugula neritina has been shown to have a reduced cover and a modified metabolism, expressing molecules that can be linked to heat stress, in a long-term heating experiment (+3 °C above ambient temperature) in the same studied marina (Chateau Marina, Brest, France) (Gauff et al., 2022b cited in Gauff et al., 2025). 

Lord (2017) studied the impact of seawater temperature on the growth and recruitment of Bugula neritina by measuring one-week per cent cover in four 3°C temperature ranges (>12 to 15, >15 to 18, >18 to 21 and >21 to 24°C) in which the species is found. Bugula neritina experienced a 9% average recruitment after one week, with maximum recruitment (65%) observed at 19.5°C, and exhibited two peaks in growth; one around 15°C and another at approximately 20°C (Lord, 2017). Lord (2017) concluded that recruitment was linked to temperature, with generally higher recruitment at warmer seawater temperatures, and there is an increasing prevalence of warm-water species like Bugula neritina with warming seawater temperatures. For example, Lord (2017) highlights how in this study, experimental panels filled up twice as quickly at 24 as 18°C due to a combination of high recruitment and higher growth rates, signifying that higher seawater temperatures may also lead to a reduction in the amount of open space, and that space is the limiting factor in fouling communities (Sebens, 1986), so increased growth or recruitment by one species is likely to result in corresponding decreases for other species.

Powell & Burgess (2024) studied the effect of temperature on the filter-feeding behaviour of Bugula neritina. Clearance rate increased linearly from 18 to 32°C, a temperature range that the population experiences most of the year, however, temperature increased algal cell size and decreased the proportion of feeding zooids, suggesting indirect effects of temperature on clearance rates. Temperature increased polypide regression, possibly as a stress response because satiation occurred quicker, or because phytoplankton quality declined, and temperature had a greater effect on clearance rate per feeding zooid than it did per total zooids (Powell & Burgess, 2024). Powell & Burgess (2024) concluded that the effect of temperature on clearance rate at the colony level is not just the outcome of individual zooids feeding more in direct response to temperature but also emerges from temperature increasing polypide regression and the remaining zooids increasing their feeding rates in response. 

In a follow-up study, Powell & Burgess (2025) studied the effect of temperature on the reproduction of Bugula neritina under the same conditions (a thermal gradient from 23 to 32°C), which reflected the upper thermal range of seasonal variation in the field. The effect of temperature was more acute on zooid states rather than whole-colony survival, and increased temperature increased the frequency of polypide regression. Most colonies reached reproductive maturity up to 30°C, but growth rate and reproduction decreased at temperatures beyond 25°C (Powell & Burgess, 2025). The decline in reproductive capacity over temperatures above 25°C was then due to the decline in the production of zooids capable of brooding embryos and zooids transitioning to regressed states up until about 30°C and transitioning to a dead state beyond that (Powell & Burgess, 2025). Powell & Burgess (2025) concluded that temperature has a role in driving the sampled population's dynamics by setting the number of generations that occur during the time window when temperatures are conducive to reproduction.

Sensitivity assessment

The important characterizing species are distributed to the north and south of the British Isles and are unlikely to be sensitive to change at the benchmark level. Morphological changes were observed in UK sponge communities, with the temperature a factor, but the characterizing sponges assessed were not listed as the most highly contributing to these changes (Berman et al., 2013; Bosch-Belmar et al., 2024). Bryozoa have been shown to acclimate to 5°C temperature increases before significant mortality occurs. Therefore, resistance is assessed as ‘High’, resilience as ‘High’, and the biotope is assessed as ‘Not Sensitive’ 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 (Cliona celata, Halichondria bowerbanki, Halichondria panicea, Hymeniacidon perleve, Leucosolenia botryoides, Suberites ficus, Raspailia ramosa, Amphilectus fucorum) are widely distributed across the coasts of the British Isles and are all found from the Channel Isles to northern Scotland (Ackers, 1993; NBN, 2015).

As well as its wide British Isles distribution, Hymeniacidon perlevis (syn. Hymeniacidon perleve) have been reported more widely and as southerly as Cape Verde (Van Soest, 1993) and the Azores (Boury-Esnault & Lopes, 1985) respectively. Hymeniacidon perlevis specifically has a substantial range from Norway in the north to the Macronesian Islands off Africa in the south (Erpenbeck and Van Soest 2002 cited in Turner, 2020).

In southern Italy, Hymeniacidon perlevis near fish farms (in the farms drainage conduits) live, grow and reproduce in constant temperatures (18°C). As a result, individuals display an active state during the entire year, while avoiding stages of decline and long dormancy usually observed in wild populations, although Hymeniacidon perlevis at the fish farm did seem to grow more in autumn and winter (Mercurio et al., 2023). In addition, it seems plausible that stable environmental conditions induce an almost continuous sexual phase, probably under the control of endogenous factors (Mercurio et al., 2023). During their observations, Mercurio et al. (2023) saw no asexual reproduction take place, however, it was impossible to exclude the contribution of asexual reproduction in the origin of the newly settled sponges, which were repeatedly detected throughout the study. Mercurio et al. (2023) concluded that the absence of a typical four-stage life cycle is probably due to the constant water temperature experienced here, which aligned with the optimal temperature range (10 to 20°C) for the vital maintenance stage in Hymeniacidon perlevis (Stone, 1970, Cao et al., 2007, and Gaino, Cardone & Corriero, 2010 cited in Mercurio et al., 2023). 

Gastaldi et al. (2016) monitored Hymeniacidon perlevis in a stressful environment, San Antonio Bay, northern Patagonia, where the sponge is affected by a semidiurnal macrotidal regime (up to 9 m) and extreme atmospheric temperatures range from -7.7°C in July (austral winter) and 41.4°C in February (austral summer); seawater temperature ranged from 5°C in August (austral winter) and 28°C in February (austral summer). They noted how seawater temperature, as well as UV and PAR solar radiation, was negatively correlated with subtidal Hymeniacidon abundance.

Hoeke, Wasson & Kahn (2025) noted how temperature may, in part, explain differences in Hymeniacidon perlevis population dynamics across various regions. Water temperatures in the English Channel (within Hymeniacidon perlevis’ presumed native range) only dropped to 5°C during the winters of 1967 and 1968 when sponge cover was measured (Stone, 1970 cited in Hoeke, Wasson & Kahn, 2025), while in the Yellow Sea where sponges regress completely, the water temperature was as low as -1.1°C in the winter (Cao et al., 2012 cited in Hoeke, Wasson & Kahn, 2025). Hoeke, Wasson & Kahn (2025) noted that the minimum water surface temperature they observed in Elkhorn Slough (California, USA) across the study period (two years) was 4.5°C, much closer to the recorded temperatures in the English Channel.

The sponge Cliona celata is found throughout the Atlantic between 1 and 200 m deep and has been recorded from Sweden to the Mediterranean in Europe, across the Southeastern USA in North America, and off Argentina in South America (Ackers et al., 1992; Stubler et al., 2024; Novarin et al., 2025). Cliona celata have been recorded in waters with a temperature range of 23.19 to 35.11°C off Southeastern USA (Stubler et al., 2024), and in a similar boring sponge, Pione truitti, growth was documented to decrease as temperatures dropped below 20°C (Pomponi & Meritt, 1985 cited in Stubler et al., 2024). Duckworth & Peterson (2013) demonstrated that an increase in water temperature to 26 and 31°C had no detectable effect on the boring activity of Cliona celata. Whereas cold winter temperatures (<5°C) can cause boring activity to cease, the incurrent papillae to withdraw and the excurrent papillae constrict (Fell et al., 1984; Carver et al., 2010). Indicating cold temperatures are more limiting to Cliona celata than hot temperatures.

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). In particular, 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 (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. However, 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., 1992).

Leucosolenia spp. are Boreal and Arctic calcareous sponges, and populations in the White Sea, Russia, experience seawater temperatures between 6 and 10°C from June to September (Lavrov et al., 2018; 2024). However, Leucosolenia sp. have been recorded in bottom water temperatures as low as 1.91 to -1.38°C in Prydz Bay, East Antarctica, where it was, along with Halichondria sp., the most prominent sponges present (Liu et al., 2020).

Crisp et al. (1964) observed 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 et al. (1964) 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 (1964b) general observations on all marine life stated that damage decreased the deeper the habitat and that the extremely cold temperatures (sea temperatures between 4 to 6°C colder than the five-year mean over a period of two months) were more extreme than the benchmark level for assessment. 

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, and 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. 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).

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. 

Nemertesia ramosa is found inshore to deeper water, being common throughout the British Isles and is distributed from Iceland to north-west Africa (Hayward & Ryland, 1994). The hydroid Plumularia setacea is often described as cosmopolitan and has been frequently reported along the Atlantic coasts (cf. Ansín Agís et al., 2001 cited in Moura et al., 2025). The hydroid genus Eudendrium is a monophyletic taxon with a worldwide distribution that includes more than 70 nominal species, and in the Mediterranean, a total of 15 species of Eudendrium are listed (González-Duarte, Megina & López-González, 2023). Eudendriums spp. are also found in cooler waters such as the Bay of Fundy, Canada (Calder, 2017). Palerud et al. (2004) described the presence of the characterizing hydroids Halecium Halecinum and Nemertesia sp. in Svalbard, suggesting that these hydroids are probably tolerant of the lowest temperatures they are likely to encounter in Britain and Ireland of ca 4°C (Beszczynska-Möller & Dye, 2013). Dendrodoa grossularia has been recorded as an abundant component of benthic fauna in Nottinghambukta, Svalbard (Różycki & Gruszczyński, 1991).

Clavelina lepadiformis is native to the Mediterranean Sea, Atlantic subarctic regions (i.e., Norwegian Sea and Greenland), and the Bay of Biscay, and has successfully spread to the coastline of South Africa, the coasts of North and Central America, and Brazil (Van Name, 1945, Turon et al., 2003, Pyo & Shin, 2011, and Reinhardt et al., 2010 cited in Hiebert et al., 2022). Reproduction of Clavelina lepadiformis is temperature-dependent (Berrill, 1975; Millar, 1970). A change in temperature could affect the time and duration of spawning, however, the distribution of Clavelina lepadiformis extends to the north and south of the British Isles from Norway to the Adriatic (Hayward & Ryland, 1996).

Clavelina lepadiformis undergoes stolonic asexual budding. At the end of the sexual breeding season, towards the end of the summer, zooids disappear or are resorbed. Over winter, the colony survives as 'winter buds' from which new zooids develop in spring (Berrill, 1950; Fish & Fish, 1996; Picton, 1997). In the winter months, when the zooids undergo de-differentiation, the resulting cylindrical bodies of many species of Clavelinidae are often found on rocky shores (Millar, 1970). During the severe winter of 1962-63, although no significant mortality of Clavelina lepadiformis was noted, Crisp (1964) found that many compound ascidians were retarded in renewal of the colony after 'winter budding', and some individuals may have been killed.

In lab experiments involving Clavelina lepadiformis collected from the harbour of La Spezia, Italy, Hiebert et al. (2022) noted that the production of their dormant form, the ‘winter buds’, occurred at 18°C. Adult zooids of Clavelina lepadiformis were also exposed to abrupt shifts in temperature (−20°C to 37°C) and salinity (10 to 44 ppt) for 24 hours, then transferred to 24°C, and the viability of zooids and the germination of dormant forms were observed after one week. Clavelina lepadiformis zooids were more resistant to lower temperatures (between 0 and 28°C) and higher salinity (30 to 45 ppt), and winter buds resisted a wider range of conditions compared to zooids (temperatures between 0 and 32°C and salinity between 15 and 45 ppt) (Hiebert et al., 2022). Clavelina lepadiformis have also been found to exist in the absence of zooids, and winter buds without zooids have been documented in the North Sea in the winter months (Orton, 1914 and Orton, 1921 and cited in Hiebert et al., 2022), in the summer months in the eastern Mediterranean, and Clavelina regress during summer months (called aestivation) in the western Mediterranean (Caralt et al., 2002; Mukai, 1977 cited in Hiebert et al., 2022). Hiebert et al. (2022) concludes that their results show that the dormant stages of Clavelina lepadiformis is resistant to extremes in temperature and salinity, thus, this linked capacity for dormancy and regeneration is likely an important adaptation that allows these species to survive drastic seasonal changes.

Casso et al. (2018) observed Clavelina lepadiformis growing in Fangar Bay, at the northern side of the Ebro Delta in the North East Iberian Coast and noted how Clavelina lepadiformis showed a slightly higher occurrence during colder seasons but was present throughout the year. Reproduction of Clavelina lepadiformis is highly seasonal, with larval occurrence restricted to two to three months during winter and spring (Caralt et al., 2002). Colonies disappear during summer (aestivation), and only survive as dormant stolons, with re-growth occurring in spring (Caralt et al., 2002; Picton & Morrow, 2004c).

Characterizing bryozoans, including Crisularia plumosa, Bugulina turbinata and Bowerbankia pustulosa, have been recorded across the British Isles, from the Channel Isles to the northern coast of Scotland (NBN, 2015).

Little information exists for Bugulina turbinata, however, evidence is available for the similar, yet invasive Bryozoa, Bugula neritina. Bugula spp. grows and reproduces in the summer months. However, day length and/or the phytoplankton bloom characteristic of temperate waters are probably more important than temperature (Ryland, 1967; 1970; Tyler-Walters, 2005c). Recruitment for Bugula spp. is known to occur between 19 and 24°C, and it can survive temperatures between 6 and 30°C (Lord, 2017; Jurgens et al., 2018; Kocak, Kucuksezgin & Bakal, 2019; Kitamura & Hirayama, 1984 cited in Fortic et al., 2025). For example, at Zikim beach, in the eastern Mediterranean, Bugula neritina and other Bugula spp. were observed in mean sea water surface temperatures of 20.27 ± 1.81°C (Ros, Guerra-García & Hoffman, 2016).

In a follow-up study, Marshall (2021) studied how temperature selected offspring size in Bugula neritina over four years. Temperature was kept at 18°C (±0.5) during settlement but then varied between 10.5 and 22.5°C during the experimental period. At temperatures cooler than 17°C, selection on both mothers and offspring favoured larger offspring, and at the hottest temperature (24°C), selection on both mothers and offspring favoured smaller offspring (Marshall, 2021). However, between these two temperatures (18 to 22°C), selection on offspring favoured larger offspring sizes, but selection on mothers favoured smaller offspring sizes (Marshall, 2021). Marshall (2021) concluded that offspring size closely tracked the local environmental temperature across cohorts, with this offspring size-temperature covariance appearing to be adaptive, at least from the perspective of mothers. When temperatures were warmer, the relationship between offspring size and performance was weak; when temperatures were cooler, the relationship was strongly positive.

Sensitivity assessment

There is evidence of sponge mortality at extremely low temperatures in the British Isles. Given this evidence, it is likely that a cooling of 5°C for a month could potentially affect the characterizing sponges, and resistance has been assessed as ‘Medium’ with a resilience of ‘High’. Therefore, sensitivity is assessed as ‘Low’.

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

Kazanidis et al. (2019) reported how water temperature and salinity explained a significant amount of variation in sponge density, and 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.  

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).

Cliona celata is tolerant of higher salinities and is found in more polyhaline (e.g. >20 psu) parts of waterbodies, such as the Chesapeake Bay, USA, where it was differentially distributed along the Bay’s salinity gradient, peaking in the mid-salinity sites (Anchondo et al., 2024). However, when Choptank in Chesapeake Bay experienced a high-salinity event, pre-2017, clionid populations apparently thrived, but then subsequently decreased over the next two years until 2019 due to a freshet (the flood of a river from heavy rain or melted snow), which reduced salinity (Anchondo et al., 2024).

Castric-Fey & Chassé (1991) conducted a factorial analysis of the subtidal rocky ecology near Brest, France and rated the distribution of species from estuarine to offshore conditions. Raspailia ramosa was rated as indifferent to this range, and Cliona celata 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 this study do not reach the benchmark level. Some of the sponges occur in harbours and estuaries, including Hymeniacidon perlevis (Ackers et al., 1992). 

Little evidence for the characterizing hydroids could be found. Studies on hydroids in general have found that prey capture rates may be affected by salinity and temperature (Gili & Hughes, 1995), although no evidence was found for species that characterize this biotope. 

In lab experiments involving Clavelina lepadiformis collected from the harbour of La Spezia, Italy, Hiebert et al. (2022) noted that the production of their dormant form, the ‘winter buds’, occurred at 18°C. Adult zooids of Clavelina lepadiformis were also exposed to abrupt shifts in temperature (−20°C to 37°C) and salinity (10 to 44 ppt) for 24 hours, then transferred to 24°C, and the viability of zooids and the germination of dormant forms were observed after one week. Clavelina lepadiformis zooids were more resistant to lower temperatures (between 0 and 28°C) and higher salinity (30 to 45 ppt), and winter buds resisted a wider range of conditions compared to zooids (temperatures between 0 and 32°C and salinity between 15 and 45 ppt) (Hiebert et al., 2022). Clavelina lepadiformis have also been found to exist in the absence of zooids, and winter buds without zooids have been documented in the North Sea in the winter months (Orton, 1914 and Orton, 1921 and cited in Hiebert et al., 2022), in the summer months in the eastern Mediterranean, and Clavelina regress during summer months (called aestivation) in the western Mediterranean (Caralt et al., 2002; Mukai, 1977 cited in Hiebert et al., 2022). Hiebert et al. (2022) concludes that their results show that the dormant stages of Clavelina lepadiformis are resistant to extremes in temperature and salinity, thus, this linked capacity for dormancy and regeneration is likely an important adaptation that allows these species to survive drastic seasonal changes.

Soule & Soule (1979) cite Hastings (1927), who described the presence of five bryozoans in hypersaline conditions in the Suez Canal. Little information exists for Bugulina turbinata, however, evidence is available for the similar, yet invasive Bryozoa, Bugula neritina. Bugula neritina is known to occur and grow between 18 and 30 psu (Fofonoff et al., 2025 cited in Fortic et al., 2025). However, Kocak, Kucuksezgin & Bakal (2019) documented Bugula neritina growing in salinity ranging from 37.05 to 39.45 ppt in the Eastern Aegean Sea, similar to Ros, Guerra-García & Hoffman (2016) who observed the same species in the eastern Mediterranean in waters of 38.19 ± 0.37 psu.

Sensitivity assessment

The CR.MCR.CFaVS.CuSpH biotope (and its sub-biotopes) occur in full (30 to 35) and variable (18 to 40) salinity. An increase in salinity (for example, due to local hypersaline effluents) at the benchmark level would result in hypersaline (>40) conditions. The evidence suggests that Clavelina lepadiformis and fouling Bugula spp. could survive at >40 ppt, but the evidence for the diverse range of species that are recorded in the biotope, and especially the dominant cushion sponges, is limited. Therefore, the evidence is ‘Insufficient’ to allow an assessment to be presented.

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, and 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. 

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).

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, following 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 species from estuarine to offshore conditions. Raspailia ramosa was rated as indifferent to this range, and Cliona celata 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 this study do not reach the lower benchmark level. Some of the sponges occur in harbours and estuaries, including Hymeniacidon perlevis (Ackers et al., 1992). 

Carver et al. (2010) suggested Cliona celata can function efficiently in salinities as low as 20‰ and can withstand exposure to 15‰ for brief periods, but prevailing salinities less than 10 to 15‰ are likely to be lethal (Hartman 1958; Hopkins 1962). Cliona celata is found in more polyhaline (e.g. >20 psu) parts of waterbodies, such as the Chesapeake Bay, USA, where it was differentially distributed along the Bay’s salinity gradient, peaking in the mid-salinity sites (Anchondo et al., 2024). However, when Choptank in Chesapeake Bay experienced a high-salinity event, pre-2017, clionid populations apparently thrived, but then subsequently decreased over the next two years until 2019 due to a freshet (the flood of a river from heavy rain or melted snow) (Anchondo et al., 2024).

Little evidence for the characterizing hydroids could be found. Studies on hydroids in general have found that prey capture rates may be affected by salinity and temperature (Gili & Hughes, 1995), although no evidence was found for species that characterize this biotope. Stebbing (1981) found that, for the hydroid Campanularia flexuosa, growth was inhibited in 70% seawater (ca 25‰) and that exposure to below 30% seawater (ca 10‰) was lethal after three days. Nemertesia spp. and Halecium halecinum were recorded as occurring in variable to full salinity biotopes (18 to 35‰). Dendrodoa grossularia has been recorded in biotopes occurring from full to low salinity regimes (<18 to 35‰). The species of the hydroid genus Eudendrium can be found in coastal lagoons, such as Bahía de Chengue, near the city of Santa Marta, Colombia (Alvarez-Leon et al., 1995 cited in Wedler, 2017). As of February 2016, Eudendrium was still present in the region where they grow on stilt roots of the red mangrove, Rhizophora mangle, in the lagoon channel, and near the opening of Cienaga Grande de Santa Marta estuary (Wedler, 2017).

In lab experiments involving Clavelina lepadiformis collected from the harbour of La Spezia, Italy, Hiebert et al. (2022) noted that the production of their dormant form, the ‘winter buds’, occurred at 18°C. Adult zooids of Clavelina lepadiformis were also exposed to abrupt shifts in temperature (−20°C to 37°C) and salinity (10 to 44 ppt) for 24 hours, then transferred to 24°C, and the viability of zooids and the germination of dormant forms were observed after one week. Clavelina lepadiformis zooids were more resistant to lower temperatures (between 0 and 28°C) and higher salinity (30 to 45 ppt), and winter buds resisted a wider range of conditions compared to zooids (temperatures between 0 and 32°C and salinity between 15 and 45 ppt) (Hiebert et al., 2022). Clavelina lepadiformis have also been found to exist in the absence of zooids, and winter buds without zooids have been documented in the North Sea in the winter months (Orton, 1914 and Orton, 1921 and cited in Hiebert et al., 2022), in the summer months in the eastern Mediterranean, and Clavelina regress during summer months (called aestivation) in the western Mediterranean (Caralt et al., 2002; Mukai, 1977 cited in Hiebert et al., 2022). Hiebert et al. (2022) concluded that their results showed that the dormant stages of Clavelina lepadiformis was resistant to extremes in temperature and salinity, thus, this linked capacity for dormancy and regeneration is likely an important adaptation that allows these species to survive drastic seasonal changes.

De Castro et al. (2018) undertook a study to see if low salinity could be used as a biosecurity tool for minimizing biofouling on ship sea chests. Settlement panels were suspended at a depth of 1.5 m in a Plymouth marina for 24 months by which time they had developed mature biofouling assemblages. Panels were exposed to three different salinities (7, 20 and 33) for two hours using a model sea chest placed in the marina and flushed with freshwater. Fouling organism diversity and abundance were assessed before panels were treated, immediately after treatment, and then one week and one month later. Some native ascidian Dendrodoa grossularia survived, but all other macrobenthos were killed by the salinity seven treatment after one week (De Castro et al., 2018). The salinity 20 treatment was not effective at killing the majority of fouling organisms (de Castro et al., 2018).

Little information exists for Bugulina turbinata, however, evidence is available for the similar, yet invasive Bryozoa, Bugula neritina. Bugula neritina is known to occur and grow between 18 and 30 psu (Fofonoff et al., 2025 cited in Fortic et al., 2025). However, Kocak, Kucuksezgin & Bakal (2019) documented Bugula neritina growing in salinity ranging from 37.05 to 39.45 ppt in the Eastern Aegean Sea, similar to Ros, Guerra-García & Hoffman (2016) who observed the same species in the eastern Mediterranean in waters of 38.19 ± 0.37 psu.

Sensitivity assessment

This biotope (CR.MCR.CFaVS.CuSpH) occurs in 'full' or 'variable' salinity. Many of the characterizing species also appear in similar biotopes with lower salinities (Connor et al., 2004). However, the biotope CFaVS.CuSpH.VS occurs in areas of less stable and lower salinities (but variable salinities) and has a lower species diversity than the CFaVS.CuSpH.As or CFaVS.CuSpH. Therefore, a reduction in salinity at the benchmark level (i.e. from full to reduced) is likely to favour the less species-rich CFaVS.CuSpH.VS biotope but CFaVS.CuSpH will remain. Therefore, resistance is assessed as ‘High’, resilience as ‘High’, and the biotope is assessed as ‘Not sensitive’ at the benchmark level.

Water flow (tidal current) changes (local)

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

Evidence

Sponges and corals are suspension feeders, relying on water currents to supply food (Hiscock, 1983; O’Reilly et al., 2022). 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 that generate water currents in sponges using flagella (De Vos et al., 1991). Some studies have 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 sponge removal from the substrata (Kazanidis et al., 2019). However, pedunculate, papillate, and arborescent morphotypes were more abundant at the low flow areas, as these morphotypes may prevent sedimentation on sponges (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). Bell (2002) documented Cliona celata regeneration, following artificial damage, at two sites within Lough Hyne. The study demonstrated 80% of Cliona celata were fully regenerated at the site which received high water flow (2 m/sec) within 100 days, whereas no Cliona celata in slower water velocities (0.1 m/sec) had fully regenerated. The author suggested an increase in water flow may increase food availability and therefore be of benefit to Cliona celata, whereas slow water flow (0.1 m/sec) may limit food supply and therefore slow regeneration if damaged.

Whilst little evidence for the characterizing sponges could be found, the important characterizing hydroids are typically found in places of low to moderate water movement, although Hayward & Ryland (1995b) note that abundant communities occur in narrow straits and headlands which may experience high levels of water flow. Hydroids can bend passively with water flow to reduce drag forces to prevent detachment and enhance feeding (Gili & Hughes, 1995). The hydroid growth form also varies to adapt to prevailing conditions, allowing species to occur in a variety of habitats (Gili & Hughes, 1995). Hiscock (1979) assessed the feeding behaviour of the hydroid Tubularia indivisa in response to different flow rates. At flow rates <0.05 m/s, polyps actively moved their tentacles. Increasing the flow rate to 0.2 m/s increased capture rates, but at higher flow rates from 0.5 to 0.9 m/s, the tentacles were extended down current and pushed together, and feeding efficiency was reduced.

In general, flow rates are an important factor for feeding in hydroids, and prey capture appears to be higher in more turbulent conditions that prevent self-shading by the colony (Gili & Hughes, 1995). The capture rate of zooplankton by hydroids is correlated with prey abundance (Gili & Hughes, 1995), thus prey availability can compensate for sub-optimal flow rates. Water movements are also important to hydroids to prevent siltation, which can cause death (Round et al., 1961). Tillin & Tyler-Walters (2014) suggested that the range of flow speeds experienced by biotopes in which hydroids are found indicates that a change (increase or decrease) in the maximum water flow experienced by mid-range populations for the short periods of peak spring tide flow would not have negative effects on this ecological group.

As sessile filter feeders, ascidians generally require a reasonable water flow rate in order to ensure sufficient food availability. It was shown that in stagnant water, phytoplankton density became reduced in a 20 to 30 cm layer immediately above a dense colony of Ciona intestinalis (Riisgård et al., 1996). If dislodged, juvenile and adult ascidians have a limited capability to re-attach, given calm conditions and prolonged contact with the new substrata (Carver et al., 2006; Millar, 1971).

Clavelina lepadiformis thrives in areas where there is little water movement (Hiscock & Hoare, 1975; De Caralt et al., 2002). Naranjo et al. (1996) found that the species was dominant in a low rate of water renewal, excess silting and high suspended solid concentrations. High water flow rates may be detrimental to feeding ability and posture, but are unlikely to cause detachment. Clavelina lepadiformis, observed growing in the shallow coastal lagoon of Ria de Aveiro, Portugal, experiences tidal current velocities of 1 m/s, which progressively get weaker in the many innermost lagoon canals (Martins et al., 2010 cited in Marques et al., 2022).

Water flow has been shown to be important for the development of bryozoan communities and the provision of suitable hard substrata for colonization (Eggleston, 1972b; Ryland, 1976). In addition, areas subject to high mass transport of water, such as the Menai Strait and tidal rapids, generally support large numbers of bryozoan species (Moore, 1977a). Although active suspension feeders, their feeding currents are probably fairly localized, and they are dependent on water flow to bring adequate food supplies within reach (McKinney, 1986). A substantial decrease in water flow will probably result in impaired growth due to a reduction in food availability, and an increased risk of siltation (Tyler-Walters, 2005c).

Little information exists for Bugulina turbinata, however, evidence is available for the similar, yet invasive Bryozoa, Bugula neritina. Bugula neritina is a common biofouling Bryozoa, and while studying the en route survivorship and post-voyage growth of the species, Schimanski et al. (2016) noted that survival (in one-week-old recruits) was not influenced by vessel speed, and speeds ranged from 6 to 18 knots (equivalent to very strong tidal streams). Okamura (1984) reported that an increase in water flow from ‘slow’ flow (1 to 2 cm/s) to ‘fast’ flow (2 to 10 cm/s) (equivalent to weak tidal streams) reduced feeding efficiency in small colonies but not in large colonies of Bugula stolonifera. 

Sensitivity assessment

The CR.MCR.CFaVS complex biotopes occur in moderate energy habitats (moderate water flow and wave sheltered habitats) (Conor et al., 2004). However, CFaVS.CuSpH, CFaVS.CuSpH.As, and CR.MCR.CFaVS.CuSpH.VS are recorded from strong to weak tidal streams and in a range of wave exposures from moderately exposed to extremely sheltered. Therefore, water flow is probably the most important contributor to water movement in many but not all examples of the biotopes. A significant decrease in water flow could result in a change to the CR.LCR.BrAs group of biotopes, whereas a significant increase may result in a change to CR.HCR.FaT. However, change at the benchmark level of 0.1 to 0.2 m/s is unlikely to be significant, and resistance is, therefore, assessed as ‘High’, resilience as ‘High’, and the biotope is assessed as ‘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

Changes in emergence are not relevant to this biotope as it is restricted to fully subtidal/circalittoral conditions. The pressure benchmark is relevant only to littoral and shallow sublittoral fringe biotopes.

Wave exposure changes (local)

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

Evidence

Roberts et al. (2006) studied deep sponge reef communities (18 to 20 m) in sheltered and exposed locations in Australia. They reported greater diversity and cover (>40% cover) of sponges in wave-sheltered areas compared with a sparser and more temporal cover in exposed sites (25% cover). Erect sponges dominated the sheltered sites, while encrusting sponges dominated in exposed locations. However, in the Faroe-Shetland Channel, sponge abundance was noted to be highest in the region of internal wave activity at the seabed (Eerkes-Medrano et al., 2020). High sponge diversity and aggregations are likely found in areas of high-wave activity, such as near shelf breaks, due to the wave activity providing an abundant and stable food supply to them, with diversity and densities of sponges decreasing away from such areas (Santín et al., 2018).

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).

Faucci et al. (2000) recorded hydroid communities at two sites of different wave exposures and recorded the presence of Obelia dichotoma and Halecium spp. in both the exposed and sheltered sites, but only found Kirchenpaueria sp. in the sheltered site. High-energy wave action can be detrimental to ascidian populations. This is mainly through physical damage to the sea squirts and through the abrasive action of suspended sediment (Jackson, 2008). Jackson (2004) suggested that Nemertesia ramosa was intolerant of high wave exposure because it was only found in sheltered areas.

Clavelina lepadiformis is tolerant of a wide range of exposure but is most abundant in moderately exposed sites (Picton, 1997). Ciona intestinalis is often dominant in highly sheltered areas such as harbours (Carver et al., 2006). Decreases in wave exposure are unlikely to have any effect. If dislodged, juvenile and adult Ciona intestinalis have a limited capability to re-attach, given calm conditions and prolonged contact with the new substratum (Carver et al., 2006; Jackson, 2008; Millar, 1971) but increases in wave exposure above moderately exposed are likely to cause a proportion of the population to die, especially in the shallower examples of the biotope if the cobbles and pebbles on which the biotope occurs are mobilized by wave action. Ascidia mentula has rarely been recorded at depths shallower than 15 m (Svane, 1984), it is possible that damage could occur if subjected to increased wave exposure.

Bugula spp. produce flexible erect tufts, which are likely to move with the oscillatory flow created by wave action. Bugula turbinata has been recorded from very wave-exposed to very wave-sheltered habitats (Tyler-Walters, 2005c).

Sensitivity Assessment

The CR.MCR.CFaVS complex biotopes occur in moderate energy habitats (moderate water flow and wave-sheltered habitats) (Conor et al., 2004). However, CFaVS.CuSpH, CFaVS.CuSpH.As, and CR.MCR.CFaVS.CuSpH.VS are recorded from strong to weak tidal streams and in a range of wave exposures from moderately exposed to extremely sheltered. Therefore, water flow is probably the most important contributor to water movement in many, but not all, examples of the biotopes. A decrease in wave exposure is probably 'not relevant', because the biotopes are recorded in wave-sheltered conditions. However, an increase in wave exposure by one MNCR category (e.g. from wave sheltered to moderately exposed is likely to result in a fundamental change to mixed faunal turf (XFa) or echinoderm and crustose communities (EcCr) in examples of the biotopes exposed to wave sheltered rather than very or more sheltered conditions, Therefore, resistance is assessed as ‘Low’’, resilience as ‘Medium’ and sensitivity is assessed as ‘Medium’ at the benchmark level. However, the assessment is based on expert judgement and confidence assessed as ‘Low’.

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 organometal pollutants on the characterizing sponges could be found.

Although no information on the effects of heavy metals on the assessed hydroids was found, evidence suggests that hydroids may suffer at least sub-lethal effects and possibly morphological changes and reduced growth due to heavy metal contamination. Various heavy metals have been shown to have sublethal effects on growth in the few hydroids studied experimentally (Bryan, 1984). Stebbing (1981a) reported that Cu, Cd, and tributyl-tin fluoride affected growth regulators in Laomedea (as Campanularia) flexuosa, resulting in increased growth. Stebbing (1976) reported that 1 µg/l Hg²⁺ was stimulatory, although the effect was transitory, exposure resulting in reduced growth towards the end of his 11-day experiments. Cadmium (Cd) was reported to cause irreversible retraction of 50% of hydranths in Laomedea loveni after seven days of exposure at concentrations between 3 µg/l (at 17.5°C and 10 ppt salinity) and 80 µg/l (at 7.5°C and 25 ppt salinity) (Theede et al., 1979). Laomedea loveni was more tolerant of Cd exposure at low temperatures and low salinities. Karbe (1972, summary only) examined the effects of heavy metals on the hydroid Eirene viridula (Campanulidae). He noted that Cd and Hg caused cumulative effects and morphological changes. Mercury (Hg) caused irreversible damage at concentrations as low as 0.02 ppm. He reported threshold levels of heavy metals for acute effects in Eirene viridula of 1.5-3 ppm Zn, 1-3 ppm Pb, 0.1-0.3 ppm Cd, 0.03-0.06 ppm Cu and 0.001-0.003 ppm Hg. Karbe (1972, summary only) suggested that Eirene viridula was a sensitive test organism when compared to other organisms.

Although no information on the effects of heavy metals on assessed hydroid species was found, the above evidence suggests that hydroids may suffer at least sub-lethal effects and possibly morphological changes and reduced growth due to heavy metal contamination.

Trace metals (particularly mercury and copper) have been found to affect embryogenesis and larval settlement in Ciona intestinalis (Bellas et al., 2004). Whilst there are extensive studies of larval intolerance to TBT (Mansueto et al., 1993; Pellerito et al., 1996; Bellas, 2005) and zinc pyrithione (Bellas, 2005), data appears non-existent for the adult stage.

Chesher (1971) found that Ascidia niagra was surprisingly intolerant of desalination effluent (50% mortality in 5.8% effluent solution after 96 hours), far less tolerant than the other species included in the study (echinoids, crabs and gorgonians). Whilst the presence of copper was considered the most deleterious factor across the study, the increased sensitivity of the ascidians was attributed to synergistic copper and temperature effects, although the presence of other contaminants (e.g. nickel) could not be ruled out (Chesher, 1971).

Bryozoans are common members of fouling communities and amongst those organisms most resistant to antifouling measures, such as copper-containing anti-fouling paints. Bryozoans were also shown to bioaccumulate heavy metals to a certain extent (Soule & Soule, 1979; Holt et al., 1995). Bryan & Gibbs (1991) reported that there was little evidence regarding TBT toxicity in bryozoa with the exception of the encrusting Schizoporella errata, which suffered 50% mortality when exposed for 63 days to 100 ng/l TBT. Rees et al. (2001) reported that the abundance of epifauna (including bryozoans) had increased in the Crouch Estuary in the five years since TBT was banned from use on small vessels. This last report suggests that bryozoans may be at least inhibited by the presence of TBT. Hoare & Hiscock (1974) suggested that Polyzoa (Bryozoa) were amongst the most intolerant species to acidified halogenated effluents in Amlwch Bay, Anglesey and reported that Flustra foliacea did not occur less than 165 m from the effluent source. The evidence, therefore, suggests that Securiflustra securifrons would be sensitive to synthetic compounds.

For the invasive bryozoan Bugula neritina, maternal copper exposure has been documented to have a strong negative carryover effect on adult offspring growth and survival in populations without historical exposure, especially when larvae themselves were exposed to copper (Neylan, Sih & Stachowicz, 2022). Conversely, little to no maternal or offspring treatment effect on adult growth and survival in the population with a history of copper exposure (Neylan, Sih & Stachowicz, 2022). In addition, parents with a history of copper exposure produced larger larvae on average and were able to increase the size of their larvae in response to copper exposure, providing a potential mechanism for maintaining fitness and suggesting transgenerational plasticity through maternal provisioning (Neylan, Sih & Stachowicz, 2022). Kocak & Kucuksezgin (2025) studied the effect of metal pollution (primarily from urbanisation and antifouling paints) in Bugula neritina at eastern Aegean ports and marinas and noted it as one of the most abundant species present. In addition, the levels of lead (Pb), cadmium (Cd), copper (Cu), vanadium (V), and zinc (Zn) in bryozoans did not show significant differences among species or sampling seasons, except for Pb (Kocak & Kucuksezgin, 2025).

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 circalittoral turf communities is likely to be limited. However, as in the case of the Prestige oil spill off the coast of France, high swell and winds can cause oil pollutants to mix with the seawater and potentially negatively affect 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.

Little information on the effects of hydrocarbons on hydroids was found, although hydroid species adapted to a wide variation in environmental factors and with cosmopolitan distributions tend to be more tolerant of polluted waters (Boero, 1984; Gili & Hughes, 1995).

Ignatiades & Becacos-Kontos (1970) found that Ciona intestinalis resisted the toxicity of oil-polluted water, and ascidians 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).

Little information on the effects of hydrocarbons on bryozoans was found. Ryland & De Putron (1998) did not detect adverse effects of oil contamination on the bryozoan Alcyonidium spp. or other sessile fauna in Milford Haven or St. Catherine's Island, south Pembrokeshire. Houghton et al. (1996) reported a reduction in the abundance of intertidal encrusting Bryozoa (no species given) at oiled sites after the Exxon Valdez oil spill. Soule & Soule (1979) reported that the encrusting bryozoan Membranipora villosa was not found in the impacted area for seven months after the December 1976 Bunker C oil spill in Los Angeles Harbour. In addition, Soule & Soule (1979) reported that Bugula neritina was lost from breakwater rocks in the vicinity (in December 1979) of the Bunker C oil spill and had not recovered within a year. However, Bugula neritina had returned to a nearby area within five months (May 1977) even though the area was still affected by sheens of oil. Furthermore, only three of the eight recorded species were present two weeks after the incident in April within the affected breakwater area. All the species had been replaced by dense growths of the erect bryozoan Scrupocellaria diegensis by June. Banks & Brown (2002) found that exposure to crude oil significantly impacted recruitment in the bryozoan Membranipora savartii.

Mohammad (1974) reported that Bugula spp. and Membranipora spp. were excluded from settlement panels near an oil terminal in Kuwait, subject to minor but frequent oil spills. Encrusting bryozoans are also probably intolerant of the smothering effects of acute hydrocarbon contamination and pollution, resulting in suffocation of colonies and communities may be lost or damaged. Circalittoral communities are likely to be protected from the direct effects of oil spills by their depth. However, the biotope may be exposed to emulsified oil treated with dispersants, especially in areas of turbulence, or may be exposed to water-soluble fractions of oils, PAHs or oil adsorbed onto particulates (Tyler-Walters, 2002).

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.

The species richness of hydroid communities decreases with increasing pollution, but hydroid species are adapted to a wide variation in environmental factors and, with cosmopolitan distributions, tend to be more tolerant of polluted waters (Boero, 1984; Gili & Hughes, 1995). Stebbing (1981a) reported that Cu, Cd, and tributyl-tin fluoride affected growth regulators in Laomedea (as Campanularia) flexuosa, resulting in increased growth. Stebbing (1981) cited reports of growth stimulation in Obelia geniculata caused by methylcholanthrene and dibenzanthrene. Bryan & Gibbs (1991) reported that virtually no hydroids were present on hard bottom communities in TBT contaminated sites and suggested that some hydroids were intolerant of TBT levels between 100 and 500 ng/l. No information concerning the resistance of the hydroids assessed was found. However, the above evidence suggests that several species of hydroid exhibit sublethal effects due to synthetic chemical contamination and lethal effects due to TBT contamination.

Bryozoans are common members of the fouling community, and amongst those organisms most resistant to antifouling measures, such as copper-containing anti-fouling paints (Soule & Soule, 1979; Holt et al., 1995). Hoare & Hiscock (1974) suggested that the Bryozoa (as Polyzoa) were amongst the most intolerant species to acidified halogenated effluents in Amlwch Bay, Anglesey, e.g. Electra pilosa occurred at lower abundance on laminarian holdfasts within the bay, compared to sites outside the affected area. They also reported that Flustra foliacea did not occur less than 165 m from the effluent source. Xu et al. (2020) studied the effect of the anti-fouling compound di(1H-indol-3-yl)methane (DIM) on a marine biofouling species, the invasive bryozoan Bugula neritina. They noted how 3-hour DIM treatments at 4 μg/ml or higher concentration, and 12-hour DIM treatments at 2 μg/ml or higher concentration, induced significant larval mortality and metamorphic abnormality in Bugula neritina. Overall, larval mortality increased with elevating DIM concentration and prolonged exposure time (Xu et al., 2020).

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

In general, respiration in most marine invertebrates does not appear to be significantly affected until extremely low concentrations are reached. For many benthic invertebrates, this concentration is about 2 ml/l (Herreid, 1980; Rosenberg et al., 1991; Diaz & Rosenberg, 1995). Cole et al. (1999) suggested possible adverse effects on marine species below 4 mg/l and probable adverse effects below 2 mg/l. Most epifauna are suspension feeders adapted to water flow, although their tolerance or requirement for water flow varies with species or taxonomic group. For example, Gray et al. (2002) concluded that fish were more sensitive to hypoxia than crustaceans and echinoderms, which in turn were more sensitive than annelids, with molluscs the most tolerant. Riedel et al. (2012) noted that infauna were generally more tolerant than epifauna. They also noted that decapods, echinoderms, and polychaetes showed lower tolerance while ascidians and anthozoans showed higher tolerance to hypoxia and anoxia (Riedel et al. 2012). Mobile species, such as fish, would probably move away from the affected area.

Hiscock & Hoare (1975) reported an oxycline forming in the summer months (Jun-Sep) in a quarry lake (Abereiddy, Pembrokeshire) from close to full oxygen saturation at the surface to <5% saturation (ca 0.5 mg/l) below ca 10 m. Despite the presence of Tethya citrina, Kirchenpaueria pinnata, Hymeniacidon pereleve, Polymastia boletiformis or Ascidia mentula in shallower water, no sponges or ascidians were recorded at depths below the oxycline at 10 to 11 m. Demosponges maintained under laboratory conditions can tolerate hypoxic conditions for brief periods. (Gunda & Janapala, 2009) investigated the effects of variable dissolved oxygen (DO) levels on the survival of the marine sponge, Haliclona pigmentifera. Under hypoxic conditions (1.5-2.0 ppm DO), Haliclona pigmentifera with intact ectodermal layers and subtle oscula survived for 42 ± 3 days. Sponges with prominent oscula, foreign material, and damaged pinacoderm exhibited poor survival (of 1-9 days) under similar conditions. Complete mortality of the sponges occurred within 2 days under anoxic conditions of <0.3 ppm DO (ca 0.3 mg/l).

Riisgård (2024) studied the oxygen extraction efficiency and tolerance to hypoxia in sponges. In terms of the demosponge Halichondria panicea, respiration rate was constant down to about 1.5 mL O2/L, which shows that the extraction efficiency increases with a 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.

Micaroni et al. (2022) exposed sponges to a total of five hypoxic treatments, with increasing severity (3.3, 1.6, 0.5, 0.4 and 0.13 mg O2/L, over 7 to 12 days). They found that sponges are generally very tolerant of hypoxia for short periods, with all sponges (Cliona celata, Suberites carnosus, and Suberites australiensis) surviving in the experimental conditions, except Polymastia crocea, which showed significant mortality at the lowest oxygen concentration (0.13 mg O₂/l), with a lethal median time of 286 hours) (Micaroni et al., 2022).

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).

Hydroids mainly inhabit environments in which the oxygen concentration exceeds 5 ml/l (ca 7 mg/l) (Gili & Hughes, 1995). Although no information was found on oxygen consumption for the characterizing hydroids, Sagasti et al. (2000) reported that epifaunal species, including several hydroids and Obelia bidentata (as bicuspidata) in the York River, Chesapeake Bay, tolerated summer hypoxic episodes of between 0.5 and 2 mg O₂/l (0.36 and 1.4 ml/l) for 5 to 7 days at a time, with few changes in abundance or species composition.

The ability of solitary ascidians to withstand decreasing oxygen levels has not been well documented. Mazouni et al. (2001) noted that whilst oysters (Magallana gigas) can survive short-term exposure to periods of anoxia (Thau Lagoon, France), the associated biofouling community dominated by Ciona intestinalis suffered heavy mortality.  It should be noted, however, that Ciona intestinalis is frequently found in areas with restricted water renewal where oxygen concentrations may drop (Carver et al., 2006). Whilst adverse conditions could affect health, feeding, reproductive capability and could eventually lead to mortality, recovery should be rapid.

Little information on the effects of oxygenation on bryozoans was found. Sagasti et al. (2000) reported that epifauna communities, including dominant species such as the bryozoans, were unaffected by periods of moderate hypoxia (ca 0.35 to 1.4 ml/l which corresponds to ca 0.5 to 2 mg/l) and short periods of hypoxia (<0.35 ml/l which corresponds to <0.5 mg/l) in the York River, Chesapeake Bay, although bryozoans were more abundant in the area with generally higher oxygen. However, estuarine species are likely to be better adapted to periodic changes in oxygenation. 

Little information exists for Bugula spp., however, evidence is available for the similar, yet invasive bryozoan, Bugula neritina. Kocak, Kucuksezgin & Bakal (2019) documented Bugula neritina growing in waters in the Eastern Aegean Sea where dissolved oxygen levels ranged from 5.21 to 6.67 mg/l. Lagos, Albarrán-Melzer & Gaitán-Espitia (2025) studied the breeding zone for the adult and larval stages of Bugula neritina by measuring their critical sensitivity to low oxygen for upper and lower zones. Individuals were collected from Lamma Island, Hong Kong, where Bugula neritina displays seasonal phenology with presence between November and March (winter months) in waters characterized by dissolved oxygen levels of 6.9 ± 0.9 (mg/l; mean ± s.d.) and 91.2 ± 10 (% saturation; mean ± s.d.) (Lagos, Albarrán-Melzer & Gaitán-Espitia, 2025). Although the results show similar intra-colony tolerances in the adults, differences were found in their larvae. While the lower zones of the colonies showed higher tolerant larvae, the upper zones had larvae with higher sensitivity and a tendency to avoid low-oxygen microhabitats, and these larvae settled more quickly and in greater numbers compared with their lower-zone counterparts (Lagos, Albarrán-Melzer & Gaitán-Espitia, 2025).

In a previous study, Lagos, White & Marshall (2016) studied how oxygen availability (the influence of normoxic and hypoxic water) affected habitat selection for Bugula neritina. They noted how larvae used cues in a hierarchical way; the oxygen levels in the water prime larvae to respond, and the response to different biofilms is contingent on oxygen levels in the water (Lagos, White & Marshall, 2016). When oxygen levels varied throughout biofilm formation, larvae responded differently depending on the history of the biofilm, with Bugula neritina larvae using cues about current and historical oxygen levels to select the appropriate microhabitat and maximize their fitness (Lagos, White & Marshall, 2016).

Sensitivity assessment

The CFaVS complex of biotopes occurs in moderate energy environments, so that low oxygen episodes may be short-lived. The above evidence suggests that sponges are probably resistant to hypoxia for short periods (albeit based on a small number of species), while some members of ascidian, bryozoan and hydroid communities may also tolerate hypoxia. However, prolonged hypoxia, for example, due to seasonal oxythermoclines would probably significantly affect the community. Therefore, a reduction in oxygen levels to <=2 mg/l for one week may adversely affect a proportion of the community but significantly affect the biotope. Hence, resistance is assessed as ‘Medium’ as a precaution. Resilience is likely to be ‘High’, so sensitivity is assessed as ‘Low’ at the benchmark level. Confidence in the assessment is ‘Low’ due to the possible variation in response between the component species in the biotope.

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

Wood et al. (2025) exposed marine sponges to nitrogenous fertilizer for 13 days and found that sponges (Cliona celata, Stelligera stuposa, Axinella dissimilis, and Suberites carnosus) showed high survival rates (>95%) and only one species, Cliona celata, showed evidence of health effects. Cliona celata exhibited both a significant change in respiration rates, coupled with visible changes in surface colouration, but only at the highest fertilizer concentration. High survival rates for all sponges were seen at 4 to 6 mg/l NO₃-N, and low survival was seen at 16 to 19 mg/l NO₃-N (Wood et al., 2025). Suberites carnosus, although less affected, also showed a decline in abundance, including necrosis in the control treatment (ambient seawater, no fertilizer) after 8 days, and two additional sponges exposed to 16 to 19 mg/l NO₃- N concentration, which showed necrosis after 13days of exposure (13 % affected) (Wood et al., 2025). Statistically, there was a significant effect of fertilizer treatment on respiration rates in Suberites carnosus (F(2,33) = 3.45; p = 0.04) with pairwise comparisons revealing a significantly lower respiration rate in the 16 to 19 mg/l NO₃- N concentration treatment compared to the control on day 8 (p < 0.01) (Wood et al., 2025). The reduction in respiration rate (ranging from 1.8 to 17.8 mg O₂ gAFDW/h) due to high fertilizer concentration after eight days was similarly seen in Micaroni et al. (2022), whereby a 50% reduction in the respiration rate of Suberites carnosus was observed after seven days of exposure to low oxygen levels, with respiration rates ranging from 0.09 to 2.6 mg O₂ gAFDW/h. The North East Atlantic portion of this study was conducted in Lough Hyne, Ireland. Lough Hyne has experienced major shifts in intertidal and subtidal communities in the last two decades, and in particular, a large decline in the abundance of subtidal sponges at some sites in the lough (Micaroni et al., 2021 cited in Micaroni et al., 2025). Excess nitrogen has been proposed as a possible cause of these changes in the subtidal sponges (Micaroni et al., 2021 cited in Micaroni et al., 2025).

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). 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 to 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 a 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.

Gochfeld et al. (2012) studied the effect of nutrient enrichment (≤0.05 to 0.07 μM for nitrate and ≤0.5 μM for phosphate) as a potential stressor in Aplysina caulifornis and its bacterial symbionts and found that nutrient enrichment had no effects on sponge or symbiont physiology when compared to control conditions. This study does contradict findings in Gochfeld et al. (2007) in which Aplysina spp. were virtually absent from a site of anthropogenic stress in Bocas del Toro, Panama, which experienced high rainfall and terrestrial runoff. The author suggested that whilst this site did include elevated nutrient concentrations, other pressures and stresses could be contributing.

In contrast, Stubler et al. (2024) evaluated the impacts of nitrate and phosphate addition on the bioerosion of Cliona celata inhabiting carbonate substrata in the subtropical southeastern USA. Overall, there were no differences in loss of calcium carbonate substratum among treatments in any of the experiments, though very high rates of bioerosion (up to 0.11 g CaCO₃/day) were observed in the field experiments. Cliona celata is known to thrive under nutrient-replete conditions and presents high adaptive plasticity to environmental variables, and it is likely that the sponges were generally unimpacted by the nutrient treatments (Stubler et al., 2024). However, previous studies have shown a positive correlative relationship between organic nutrient loading and abundance of clionid boring sponges in the Caribbean and Mediterranean (Rose & Risk, 1985; Ward-Paige et al., 2005; Chaves-Fonnegra et al., 2007 and Muricy, 1991 cited in Stubler et al., 2024). Rose & Risk (1985) described an increase in abundance of Cliona delitrix in an organically polluted section of the Grand fringing reef affected by the discharge of untreated faecal sewage. Ward-Paige et al. (2005) described that the greatest size and biomass of clionids corresponded with the highest nitrogen, ammonia and δ¹⁵N levels. 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).

Witt et al. (2004) found that the hydroid Obelia spp. was more abundant in a sewage disposal area in the Weser estuary (Germany), which experienced sedimentation of 1 cm for more than 25 days. It should be noted that another hydroid (Sertularia cupressina) was reduced in abundance when compared with control reference areas. As suspension feeders, an increase in organic content at the benchmark is likely to be of benefit to the characterizing hydroids.

Clavelina lepadiformis was found to dominate Spanish harbours and nearby zones with highly transformed substrata, low rate of water renewal and excess silting and suspended matter and was described as biofouling and opportunist (Naranjo et al., 1996). Ascidia mentula has been reported in Iberian bays subject to both nutrient-rich upwelling events and anthropogenic pollution (Aneiros et al., 2015). There is some suggestion that there are possible benefits to ascidians from the increased organic content of water; ascidian ‘richness’ in Algeciras Bay was found to increase in higher concentrations of suspended organic matter (Naranjo et al. 1996).

Hartikainen et al. (2009) reported that increased nutrient concentrations resulted in freshwater bryozoans achieving higher biomass. O’Dea & Okamura (2000) found that the annual growth of Flustra foliacea in western Europe has substantially increased since 1970. They suggest that this could be due to eutrophication in coastal regions due to organic pollution, leading to increased phytoplankton biomass (see Allen et al., 1998).

Little information exists for Bugula spp., however, evidence is available for the similar, yet invasive bryozoan, Bugula neritina. Kocak, Kucuksezgin & Bakal (2019) documented Bugula neritina growing in waters in the Eastern Aegean Sea where ammonium (NH₄-N) ranged from 1.63 to 15.14 μM and total phosphate (TPO₄P) ranged from 0.74 to 5.67 μM. Although there was no significant relationship between Bugula neritina and nutrient concentrations, it was emphasized that eutrophic conditions in Abu Qir port, Egypt, where average values of ammonia and total phosphate were measured as 6.25 μM and 1.10 μM, respectively, could support the occurrence of Bugula neritina (Ramadan et al., 2006 and Abdel-Salam and Ramadan, 2008 cited in Kocak, Kucuksezgin & Bakal, 2019).

Gauff et al. (2022) studied how pollution gradients lead to local adaptation and small-scale spatial variability of communities in the Marina du Château, Brest, France. From observing the inner, middle, and entrance of the marina (whereby the entrance was the least polluted area), the results suggest that microgeographic adaptations exist at a <100 m scale, where populations develop location-specific tolerance to contamination gradients (Gauff et al., 2022). Specifically, Bugula neritina experienced local adaptation in community structure, respiration, metabolomes, and pollutant uptake, as individuals were impacted by the transplant to different marina locations, with a disadvantage for individuals transplanted from the entrance to the inner location (Gauff et al., 2022).

Sensitivity assessment

The evidence above suggests that several of the characteristic species are resistant to increased nutrients and may thrive in nutrient-enriched conditions. However, some species, e.g. Cliona celata were reduced in abundance at the highest nutrient level tested (Wood et al., 2025). Therefore, resistance is assessed as ‘Medium’ but with ‘Low’ confidence due to the variation in response to nutrients between species or species groups. Hence, resilience is assessed as ‘High’ and sensitivity as ‘Low’.

Organic enrichment

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

Evidence

Rose & Risk (1985) described an increase in abundance of the sponge Cliona delitrix in an organically polluted section of Grand Cayman fringing reef affected by the discharge of untreated faecal sewage. De Goeij et al. (2008) used ¹³C to trace the fate of dissolved organic matter in the coral reef sponge Halisarca caerulea. Biomarkers revealed that the sponge incorporated dissolved organic matter through both bacteria-mediated and direct pathways, suggesting that it feeds, directly and indirectly, on dissolved organic matter. 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).

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. and Hymeniacidon perleve (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 a 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.

Witt et al. (2004) found that the hydroid Obelia spp. was more abundant in a sewage disposal area in the Weser estuary (Germany), which experienced sedimentation of 1 cm for more than 25 days. However, another hydroid (Sertularia cupressina) was reduced in abundance when compared with unimpacted reference areas. As suspension feeders, an increase in organic content at the benchmark is likely to be of benefit to the characterizing hydroids.

There is some suggestion that there are possible benefits to the ascidians from the increased organic content of water; Ascidian ‘richness’ in Algeciras Bay was found to increase in higher concentrations of suspended organic matter (Naranjo et al. 1996). Kocak & Kucuksezgin (2000) noted that Ciona intestinalis was one of the rapid breeding opportunistic species that tended to be dominant in Turkish harbours enriched by organic pollutants and was frequently found in polluted environments (Carver et al., 2006). Ascidia mentula has been reported in Iberian bays subject to both nutrient-rich upwelling events and anthropogenic organic pollution (Aneiros et al., 2015).

Sensitivity assessment

The above evidence suggests that resistance to this pressure is 'High'. Therefore, resilience is assessed as 'High', and the biotope is assessed as 'Not sensitive'.

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

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). In addition, Halichondria bowerbanki has been observed growing on cave walls (Ereskovsky et al., 2018).

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). 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.

The hydroid Plumularia setacea has been observed overgrowing artificial structures in ports (Moura et al., 2025). The hydroid genus Eudendrium is a common inhabitant of harbours and marinas, and Eudendrium racemosum, for example, is widely found in both natural and artificial Mediterranean environments (Guerra-García et al., 2025). Interestingly, native Eudendrium spp. have a preference for natural rock habitats, while invasive Eudendrium spp. have a preference for artificial structures such as floating pontoons (Megina, González-Duarte & López-González, 2016).

Clavelina lepadiformis was observed growing in the shallow coastal lagoon of Ria de Aveiro, Portugal, on oyster shells and was the dominant species present (Marques et al., 2022), as well as being abundant on PVC plates in Fangar Bay, at the northern side of the Ebro Delta in the NE Iberian Coast (Casso et al., 2018).

Bugula spp. require a hard substratum to settle, and in lab experiments, both Bugula flabellata and Bugula neritina larvae swam less and explored more when exposed to plastic surfaces, suggesting a preference for this substrate and resulting in lower energy expenditures associated with searching for habitat (Pinochet, Urbina & Lagos, 2020). Larvae actively chose to settle on plastics rather than on wood or concrete substrates, and suggests that for Bugula larvae, the likelihood of colonizing plastic surfaces is higher than other materials commonly found in urbanized coastal areas (Pinochet, Urbina & Lagos, 2020). The strong preference of invertebrate larvae for plastics can potentially extend the distribution range of many marine species as they are able to travel long distances attached to floating debris (Pinochet, Urbina & Lagos, 2020).

Sensitivity assessment

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

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

The characterizing species are likely to be affected by physical disturbances. 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 increase when fishing activity is removed (Jennings & Kaiser, 1998; Sheehan et al., 2017; Kaiser et al., 2018; Long et al., 2021; Langton, Stirling & Boulcott, 2023). Also, heavy mobile gears could also result in the movement of boulders (Bullimore, 1985; Jennings & Kaiser, 1998). 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).

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.

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 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). However, studies have shown Ascidia spp. to become more abundant following disturbance events (Bradshaw et al., 2000). 

Resampling of grounds that were historically studied (from the 1930s) indicates that some species have increased in areas subject to scallop fishing (Bradshaw et al., 2002).  This study also found an (unquantified) increase in abundance of tough-stemmed hydroids, including Nemertesia spp. Bradshaw et al. (2002) suggested that, as well as having high resistance to abrasion pressures, Nemertesia spp. have benthic larvae that could rapidly colonize disturbed areas with newly exposed substrata close to the adult. Hydroids may also recover rapidly, as the surface covering of hydrorhizae may remain largely intact, from which new uprights are likely to grow. In addition, the resultant fragments of colonies may be able to develop into new colonies. Hydroid colonies were still present in the heavily fished area, albeit at lower densities than in the closed area. This may largely be because the Isle of Man scallop fishery is closed from 1^st June to 31^st October (Andrews et al., 2011), so at the time the samples were taken for the study in question, the seabed had been undredged for at least 3.5 months. The summer period is also the peak growing/breeding season for many marine species. (Bradshaw et al., 2003).

Clavelina lepadiformis is permanently attached to the substratum and is unable to move out of the way of abrasive objects. The body of the species is soft and delicate, so abrasion is likely to cause physical damage and possibly death.

Sensitivity assessment

Whilst a large proportion of the sponge community is likely to be affected by abrasion events, 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). Hydroids have rapid growth rates, and potentially high recruitment and can recover quickly from fragments or dormant resting stages. Ascidians are also likely to be significantly affected, although, given their high resilience, they are likely to recover quickly. The physiology of the bryozoans affords some protection in the event of abrasion events, and recovery is likely to be rapid if stolons remain undamaged. The assessment is therefore based on the sponge component of the biotope. Hence, resistance is assessed as ‘Low’, resilience as ‘Medium’ and sensitivity is assessed as ‘Medium’.

Please note, Boulcott & Howell (2011) did not mention the abrasion caused by fully loaded collection bags on the Newhaven dredges. A fully loaded Newhaven dredge may cause higher damage to the community than indicated in their study.

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 sediment may have a deleterious effect on the suspension-feeding community. It is likely to clog their feeding apparatus to some degree, resulting in a reduced ingestion over the benchmark period and, subsequently, a decrease in growth rate (Jackson, 2004).

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

Bell et al. (2015) noted that upright forms intercepted a smaller amount of settling sediment than encrusting forms. For example, Bell & Barnes (2002; cited in Bell et al., 2015) reported considerable variation in the branching characteristics of Raspaillia ramosa and Stelligera stuposa across a sediment gradient in Lough Hyne, Northern Ireland, although the patterns were due to the interaction between sedimentation and water flow. Raspailia ramosa and Stelligera stuposa have a reduced maximum size in areas of high sedimentation (Bell et al., 2002).

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. 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.

Some of the characterizing sponges occur in harbours and estuaries, including Halichondria spp. and Hymeniacidon perleve (Ackers et al., 1992). Storr (1976) observed the sponge Sphecispongia vesparium ‘backwashing’ to eject sediment and noted that other sponges (such as Condrilla nucula) use secretions to remove settled material. Tjensvoll (2013) found that Geodia barretti physiologically shuts down (86% reduction in respiration) when exposed to sediment concentrations of 100 mg/l. Rapid recovery to initial respiration levels directly after the exposure indicated that Geodia barretti can cope with a single short exposure to elevated sediment concentrations. However, it should be noted that a laboratory study on the impact of elevated sedimentation rates on deep-water sponges found that sediment load of 30 mg sed./l resulted in significantly higher sponge mortality compared with sponges exposed to 5 and 10 mg sed./l, although no additional information was provided (Hoffman & Tore Rapp, pers comm. cited in Lancaster et al., 2014).

Pineda et al. (2017a) examined the effect of suspended sediments in three species of sponge from New Zealand; two phototrophic (due to symbiotic algae) (Cliona orientalis and Carteriospongia foliascens) and one heterotrophic (Ianthella basta) under laboratory conditions. All sponges exhibited a short-term response to suspended sediment, e.g. closed oscula, mucus production, and tissue regression. Most survived low to medium turbidity (≤33 mg/l) for up to 28 days, but at high turbidity (≤76 mg/l), Cliona orientalis and Carteriospongia foliascens experienced 20-90% mortality, and Ianthella basta showed tissue regression. Pineda et al. (2017a) suggested that suspended sediment combined with low light due to turbidity increased mortality in the phototrophic species, but noted that there was considerable interspecies variation in their response. 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 water depth, the average concentration of suspended POM is about four times higher than between 150 and 4,000 m water depth.

It should be noted that many of the characterizing sponges have been recorded in the turbid biotopes CFaVS.CuSpH.VS and CFaVS.CuSpH.As (Connor et al., 2004). For example, CR.MCR.CFaVS.CuSpH.VS occurs in estuarine waters and, therefore, the species in the biotope are likely to be well adapted to turbid conditions. Wass et al. (1999) described suspended sediment maxima for ‘medium-sized’ rivers as rarely exceeding 500 mg/l, with a few rivers (including the Don and the Swale) experiencing concentrations in excess of 1,000 mg/l. Langston et al. (2003) described annual mean suspended sediment concentrations in the Tamar as varying from 61 mg/l to 1039 mg/l in the upper estuary, 6 to 18 mg/l in the outer estuary and 2 to 9 mg/l beyond. It should be noted that the values quoted are mean annual concentrations, and the same report states that conditions could be ‘very turbid’ in the outer estuary. The estuarine turbidity maximum (the point at which the highest turbidity is experienced) can be highly variable and has been reported to move by ca 12 km down-estuary during the transition from neap to spring tides in the Humber estuary (Uncles et al., 2005). Estuarine environments are likely to experience variable turbidity, and the species present are probably tolerant of significant short-term changes in suspended solid concentrations.

Nemertesia ramosa is a passive suspension feeder, extracting seston from the water column. Increased siltation may clog up the feeding apparatus, requiring energetic expenditure to clear. Recovery is likely to take only a few days. (Jackson, 2004)  

Increased siltation can cause clogging of ascidian respiratory organs (Bakus, 1968). Clavelina lepadiformis has relatively wide apertures that help prevent clogging from particles (Naranjo et al., 1996). The simplistic structure of its branchial sac (Fiala-Medioni, 1978) may be less efficient in expelling particles, and more likely to suffer from clogging of feeding apparatus than other forms of sea squirts, such as Ciona intestinalis. However, Clavelina lepadiformis was found to dominate Spanish harbours and nearby zones with a low rate of water renewal, excess silting and suspended matter and the species was described as biofouling and opportunistic (Naranjo et al., 1996).

Bryozoan turfs form preferentially on steep surfaces and under overhangs, and larvae preferentially settle under overhangs, presumably to avoid smothering and siltation (Ryland, 1977; Hartnoll, 1983). In addition, bryozoans are suspension feeders that may be adversely affected by increases in suspended sediment, due to clogging of their feeding apparatus. Wendt (1998) noted that Bugula neritina grew faster on downward-facing surfaces than upward-facing surfaces, presumably due to siltation and reduced feeding efficiency on upward-facing surfaces. But where water flow is sufficient to prevent siltation, Bugulina turbinata may colonize upward-facing surfaces (Hiscock & Mitchell, 1980).

Long-term increase in turbidity may affect primary production in the water column and, therefore, reduce the availability of diatom food, both for suspension feeders and deposit feeders. In addition, primary production by the microphytobenthos on the sediment surface may be reduced, further decreasing food availability for deposit feeders. However, primary production is probably not a major source of nutrient input into the system, and, furthermore, phytoplankton will also immigrate from distant areas, so the effect may be decreased.

Sensitivity assessment

CR.MCR.CFaVS.CuSpH.VS and CuSpH.As are considered to be turbid (unquantified) (Connor et al., 2004). Mortality at the benchmark level is, therefore, considered unlikely, and resistance is ‘High’, resilience is ‘High’, and the biotope is assessed as ‘Not sensitive’. The lack of evidence for characterizing species, as well as a lack of consensus in the literature, results in a ‘Low’ quality confidence score.

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

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). 

Pineda et al. (2017b) exposed three phototrophic (due to symbiotic algae) and two heterotrophic sponges from New Zealand to repeated deposition events and sediment cover over 80 to 100% of sponge surface to a depth of ca 0.5 mm for up to 30 days in laboratory conditions. All five species survived with minimal physiological effects. However, Wulff (2006) described mortality in three sponge groups following four weeks of complete burial under sediment; 16% of Amphimedon biomass died compared with 40% and 47% in Iotrochota and Aplysina, respectively.

In Lough Hyne, a small (approx. 0.5 km²) lough on the southwest coast of Ireland, sponge communities experience varying levels of sedimentation, from negligible levels (3 ± 0.2 mm) to higher rates (18 to 34 g /m/ day), and sponges have been continually recorded in these areas under these levels of sedimentation (Micaroni et al., 2025).

Micaroni et al. (2025) examined the changes in sponge-dominated benthic rock communities from 2018 to 2021, after a mass sponge mortality in 2010 in Lough Hyne. In doing so, they recorded the change in sponge communities with water flow and sedimentation. The entrance (Whirlpool cliff) experienced flow rates up to 2.5 m/s and negligible sedimentation (3 ± 0.2 mm) and was dominated by cnidarians (mostly jewel anemones and soft corals) and sponges. Three internal sites, Glannafeen, Labhra Cliff, and Goleen, had similar communities, of which Glannafeen had the highest flow, 0.09 to 0.18 m/s and 7 ± 0.5 mm sedimentation, while Goleen had 0 to 0.2 m/s flow and the highest sedimentation (11 ± 0.3 mm). The internal sites were dominated by sponge and faunal turfs. The innermost site (West cliff) had the lowest flow (0.15 cm/s, but only during spring tides), and the highest sedimentation rate of all the sites. It was dominated by fewer sponges, sea squirts, and hydroids. Encrusting sponges were abundant at all sites. However, papillate (Polymastia spp.) and massive (Dysidea spp., Suberites carnosus) sponges were most abundant at Glannafeen, while erect (e.g. Axinella dissimilis, A. damicornis, Raspailia ramosa, Stelligera spp.) sponges were most abundant at Glannafeen and Labhra Cliff. Raspailid, axinellid and tethyid (e.g. Tethya aurantia) sponge populations were the most stable over the four years of the study, while Polymastiid, Stelligerid, and Mycalid sponges and Hymeniacidon kitchingi, H. perlevis and Suberites carnosus showed some temporal variation, while Haliclona spp., Amphilectus fucorum showed marked temporal variation. However, Caryophyllia spp. and Corynactis viridis populations were very stable (Micaroni et al., 2025).

The complete disappearance of the ‘sponges associated’ with the sea squirt Ascidiella aspersa in the Black Sea near the Kerch Strait was attributed to siltation (Terent'ev, 2008 cited in Tillin & Tyler-Walters, 2014).

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). Cliona celata can either have a boring life form (it bores into rock) or a massive form (grows on top of rock). With the boring form, only inhalant and exhalant papillae are visible just above the rock surface (Wood, 2007). The massive form has raised, rounded ridges with large specimens growing up to 1 m across and 50 cm high (Snowden, 2007). 

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

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

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 com cited in Lancaster et al., 2014).

In general, it appears that hydroids are sensitive to silting (Boero, 1984; Gili & Hughes, 1995) and the decline of hydroid beds in the Wadden Sea have been linked to environmental changes, including siltation. Round et al. (1961) reported that the hydroid Sertularia (now Amphisbetia) operculata died when covered with a layer of silt after being transplanted to sheltered conditions. Boero (1984) suggested that deep water hydroid species develop upright, thin colonies that accumulate little sediment, while species in turbulent water movement were adequately cleaned of silt by water movement. Hughes (1977) found that maturing hydroids that had been smothered with detritus and silt lost most of the hydrocladia and hydranths. After one month, the hydroids were seen to have recovered, but although neither the growth rate nor the reproductive potential appeared to have been affected, the viability of the planulae may have been affected. Therefore, if the deposition is removed fairly rapidly, the impact may be limited. Nemertesia ramosa is an upright hydroid with a height of up to 15 cm. The colony structure is fairly tough and flexible. Halecium halecinum can grow up to 25 cm, and Kirchenpaueria pinnata can grow to ca 10 cm (Hayward & Ryland, 1994). Some of the community is, therefore, likely to survive smothering by 5 cm. Smothering with 5 cm of sediment may cover over some individuals, others may just have the lower section of the main stem covered (Hayward & Ryland, 1994). 

The solitary ascidians considered in this report are permanently attached to the substratum and are active suspension feeders. Dendrodoa grossularia is a small ascidian, capable of reaching a size of approx. 8.5 mm (Millar, 1954) and is, therefore, likely to be inundated by deposition of 5 or 30 cm of sediment.  Clavelina lepadiformis reaches up to 2 cm in height and often colonizes vertical surfaces and overhangs (Fish & Fish, 1992). Smothering by 5 cm depth of sediment would completely cover the majority of the population, with only those colonizing overhangs and vertical surfaces protected. The complete disappearance of the sea squirt Ascidiella aspera biocoenosis in the Black Sea near the Kerch Strait was attributed to siltation (Terent'ev, 2008 cited in Tillin & Tyler-Walters, 2014). 

Sensitivity assessment

Smothering by 5 cm of sediment is likely to impact hydroids, bryozoans, ascidians, and smaller cushion sponge species. The evidence suggests that hydroids, bryozoans and ascidians are likely to be adversely affected by smothering. While the sponge species vary in their ability to tolerate smothering, it may lead to mortality depending on the depth of cover and duration. However, it is likely that the moderate energy and, in particular, tidal flow experienced by this biotope would remove deposited spoil within a few tidal cycles, while vertical faces would probably be unaffected. Therefore, resistance has been assessed as ‘Medium’ to represent some loss of abundance in some members of the community resilience as ‘High’, and sensitivity has been assessed as ‘Low’ at the benchmark level.

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

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). 

Pineda et al. (2017b) exposed three phototrophic (due to symbiotic algae) and two heterotrophic sponges from New Zealand to repeated deposition events and sediment cover over 80 to 100% of sponge surface to a depth of ca 0.5 mm for up to 30 days in laboratory conditions. All five species survived with minimal physiological effects. However, Wulff (2006) described mortality in three sponge groups following four weeks of complete burial under sediment; 16% of Amphimedon biomass died compared with 40% and 47% in Iotrochota and Aplysina, respectively.

In Lough Hyne, a small (approx. 0.5 km²) lough on the southwest coast of Ireland, sponge communities experience varying levels of sedimentation, from negligible levels (3 ± 0.2 mm) to higher rates (18 to 34 g /m/ day), and sponges have been continually recorded in these areas under these levels of sedimentation (Micaroni et al., 2025).

Micaroni et al. (2025) examined the changes in sponge-dominated benthic rock communities from 2018 to 2021, after a mass sponge mortality in 2010 in Lough Hyne. In doing so, they recorded the change in sponge communities with water flow and sedimentation. The entrance (Whirlpool cliff) experienced flow rates up to 2.5 m/s and negligible sedimentation (3 ± 0.2 mm) and was dominated by cnidarians (mostly jewel anemones and soft corals) and sponges. Three internal sites, Glannafeen, Labhra Cliff, and Goleen, had similar communities, of which Glannafeen had the highest flow, 0.09 to 0.18 m/s and 7 ± 0.5 mm sedimentation, while Goleen had 0 to 0.2 m/s flow and the highest sedimentation (11 ± 0.3 mm). The internal sites were dominated by sponge and faunal turfs. The innermost site (West cliff) had the lowest flow (0.15 cm/s, but only during spring tides), and the highest sedimentation rate of all the sites. It was dominated by fewer sponges, sea squirts, and hydroids. Encrusting sponges were abundant at all sites. However, papillate (Polymastia spp.) and massive (Dysidea spp., Suberites carnosus) sponges were most abundant at Glannafeen, while erect (e.g. Axinella dissimilis, A. damicornis, Raspailia ramosa, Stelligera spp.) sponges were most abundant at Glannafeen and Labhra Cliff. Raspailid, axinellid and tethyid (e.g. Tethya aurantia) sponge populations were the most stable over the four years of the study, while Polymastiid, Stelligerid, and Mycalid sponges and Hymeniacidon kitchingi, H. perlevis and Suberites carnosus showed some temporal variation, while Haliclona spp., Amphilectus fucorum showed marked temporal variation. However, Caryophyllia spp. and Corynactis viridis populations were very stable (Micaroni et al., 2025).

The complete disappearance of the ‘sponges associated’ with the sea squirt Ascidiella aspersa in the Black Sea near the Kerch Strait was attributed to siltation (Terent'ev, 2008 cited in Tillin & Tyler-Walters, 2014).

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). Cliona celata can either have a boring life form (it bores into rock) or a massive form (grows on top of rock). With the boring form, only inhalant and exhalant papillae are visible just above the rock surface (Wood, 2007). The massive form has raised, rounded ridges with large specimens growing up to 1 m across and 50 cm high (Snowden, 2007). 

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

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

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 com cited in Lancaster et al., 2014).

In general, it appears that hydroids are sensitive to silting (Boero, 1984; Gili & Hughes, 1995) and the decline of hydroid beds in the Wadden Sea have been linked to environmental changes, including siltation. Round et al. (1961) reported that the hydroid Sertularia (now Amphisbetia) operculata died when covered with a layer of silt after being transplanted to sheltered conditions. Boero (1984) suggested that deep water hydroid species develop upright, thin colonies that accumulate little sediment, while species in turbulent water movement were adequately cleaned of silt by water movement. Hughes (1977) found that maturing hydroids that had been smothered with detritus and silt lost most of the hydrocladia and hydranths. After one month, the hydroids were seen to have recovered, but although neither the growth rate nor the reproductive potential appeared to have been affected, the viability of the planulae may have been affected. Therefore, if the deposition is removed fairly rapidly, the impact may be limited. Nemertesia ramosa is an upright hydroid with a height of up to 15 cm. The colony structure is fairly tough and flexible. Halecium halecinum can grow up to 25 cm, and Kirchenpaueria pinnata can grow to ca 10 cm (Hayward & Ryland, 1994). Some of the community is, therefore, likely to survive smothering by 5 cm. Smothering with 5 cm of sediment may cover over some individuals, others may just have the lower section of the main stem covered (Hayward & Ryland, 1994). 

The solitary ascidians considered in this report are permanently attached to the substratum and are active suspension feeders. Dendrodoa grossularia is a small ascidian, capable of reaching a size of approx. 8.5 mm (Millar, 1954) and is, therefore, likely to be inundated by deposition of 5 or 30 cm of sediment.  Clavelina lepadiformis reaches up to 2 cm in height and often colonizes vertical surfaces and overhangs (Fish & Fish, 1992). Smothering by 5 cm depth of sediment would completely cover the majority of the population, with only those colonizing overhangs and vertical surfaces protected. The complete disappearance of the sea squirt Ascidiella aspera biocoenosis in the Black Sea near the Kerch Strait was attributed to siltation (Terent'ev, 2008 cited in Tillin & Tyler-Walters, 2014). 

Sensitivity assessment

Smothering by 30 cm of sediment is likely to impact hydroids, bryozoans, ascidians, and smaller cushion sponge species. The evidence suggests that hydroids, bryozoans and ascidians are likely to be adversely affected by smothering. While the sponge species vary in their ability to tolerate smothering, it may lead to mortality depending on the depth of cover and duration. However, it is likely that the moderate energy and, in particular, tidal flow experienced by this biotope would remove deposited spoil within a few tidal cycles, while vertical faces would probably be unaffected. Therefore, resistance has been assessed as ‘Medium’ to represent some loss of abundance in some members of the community resilience as ‘High’, and sensitivity has been assessed as ‘Low’ at the benchmark level.

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 or noise vibrations. Therefore, the pressure is probably 'Not relevant' in this biotope and its sub-biotopes.

Introduction of light or shading

Benchmark. A change in incident light via anthropogenic means (Introduced light or shade pressure definition).

Evidence

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).

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.

Gili & Hughes (1995) reviewed the effect of light on a number of hydroids and found that there is a general tendency for most hydroids to be less abundant in well-lit situations. Whilst hydroid larvae can be positively or negatively phototactic, the planulae of Nemertesia antennina show no response to light (Hughes, 1977).

In vitro studies of solitary ascidians indicate that both spawning and settlement are controlled by light. However, Ciona intestinalis in vivo has been observed to spawn and settle at any time of the day (Svane & Havenhand, 1993).

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 could be found for the effect of light on the characterizing species of these biotopes, it is unlikely that these species would be impacted. As a circalittoral biotope, 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). Increased shading (e.g. by the construction of a pontoon, pier, etc.) could benefit the characterizing species of this biotope. A significant increase in light (e.g. from underwater installations), however unlikely, may be detrimental if it increases competition from macroalgae. Nevertheless, there is currently ‘Insufficient evidence’ of the likely effects of artificial light on circalittoral communities to form the basis of an assessment.

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion (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). Specimens of Cliona spp. exhibited blackened damage since 2013 in Skomer. Preliminary results have shown that clean, fouled and blackened Cliona all have very different bacterial communities. The blackened Cliona are effectively dead and have a bacterial community similar to that of marine sediments. The fouled Cliona have a very distinct bacterial community that may suggest a specific pathogen caused the effect (Burton, pers. comm; Preston & Burton, 2015). 

However, Longo et al. (2016) studied the ability of the 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.

Hydroids exhibit astonishing regeneration and rapid recovery from injury (Sparks, 1972), and the only inflammatory response is active phagocytosis (Tokin & Yaricheva, 1959;1961, as cited in Sparks, 1972). No record of diseases in the characterizing hydroids could be found. 

There appears to be little research into ascidian diseases, particularly in the Atlantic. The parasite Lankesteria ascidiae targets the digestive tubes and can cause ‘long faeces syndrome’ in Ciona intestinalis (although it has also been recorded in other species). Mortality occurs in severely affected individuals within about a week following the first symptoms. (Mita et al., 2012).

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. Resistance is, therefore, assessed as 'Medium', resilience as 'High' and sensitivity as 'Low’. Given the on-going work described above, this pressure should be revisited in light of any new 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

Spongia officinalis (a Mediterranean species) has been targeted as a commercial species for use as bath sponges, although this species does not occur in the British Isles, and no record of commercial exploitation of sponges in the British Isles could be found.  Hiscock (2003) stated that the greatest loss of Axinella dissimilis at Lundy might have been due to collecting during scientific studies in the 1970s. No indication of recovery was evident.  Axinella damicornis was harvested in Lough Hyne during the 1980s (for molecular investigations), and the populations were reduced to very low densities, which subsequently recovered very slowly, although they are now considered to be back to their original densities (Bell, 2007).  No evidence of targeted removal of the characterizing species could be found.  Despite historic harvesting of the hydroid Sertularia cupressina in the Wadden Sea (Wagler et al., 2009), no evidence for the harvesting of the characterizing hydroids could be found and targeted extraction is highly unlikely.  Despite novel proposals to farm Ciona intestinalis as a potential feedstock for aquaculture in Sweden (Laupsa, 2015), it is very unlikely that solitary ascidians would be targeted for extraction. 

Sensitivity assessment

None of the characterizing species is harvested, and targeted removal is therefore ‘Not relevant’ to this biotope.

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, removal of the characteristic epifauna due to bycatch 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 in this biotope.  Whilst a large proportion of the sponge community is likely to be affected by abrasion events, there is some debate as it the level of effects depending on the size of the sponge and the type of abrasion effect (see Coleman et al., 2013).  The majority of the literature agrees that damage would fall within the ‘Low’ bracket of 25-75% reduction.  Ascidians are also likely to be significantly affected, although given their high resilience, they are likely to recover quickly (Bradshaw et al., 2000). 

Sensitivity assessment. Based on the broad agreement  of trawl impacts on sponge communities and the likely disturbance to the sessile epifaunal ascidians and hydroids, resistance is assessed as ‘Low’, resilience as ‘Medium’ and sensitivity is assessed as ‘Medium’.

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 (0.1-0.9/m²) in one faunal turf biotope (CR.MCR.CFaVS.CuSpH.As) (JNCC, 2015). Faunal turfs are dominated by suspension feeders, so larval predation is probably high, which may prevent colonization by Crepidula. Also, faunal turf species actively compete for space, and many are fast-growing and opportunistic, so they may out-compete Crepidula for space even if it gained a foothold in the community. 

Sensitivity assessment

The circalittoral rock characterizing this biotope could be suitable for colonization by Crepidula fornicata due to the sheltered to extremely sheltered wave conditions, 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 circalittoral habitats. At present, there is 'Insufficient evidence' to suggest that the circalittoral biotopes are sensitive to colonization by Crepidula fornicata or other invasive species; further evidence is required.

The carpet sea squirt, Didemnum vexillum

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009). Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024). In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Minchin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

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

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

Human-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 circalittoral 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 moderately strong to strong water flow (1 to 6 m/s) and sheltered to extremely sheltered wave exposure. Didemnum vexillum regresses as temperatures decline in winter, so shallow examples may be able to recover their condition in winter (Gittenberger, 2007; Valentine et al., 2007a; Herborg et al., 2009). However, deeper examples may not experience enough temperature change to trigger the decline in Didemnum vexillum (Valentine et al., 2007a). If Didemnum sp. could gain a 'foothold', it might overgrow, smother or cause mortality of epifauna. Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. Therefore, a resistance of 'Medium' (some, <25% mortality) is suggested as a precaution in case Didemnum vexillum could colonize the biotope, but with 'Low' confidence due to the lack of direct evidence. Resilience is assessed as 'Very low' as recovery would require the physical removal of Didemnum sp., so sensitivity is assessed as 'Medium'. 

The Pacific oyster, Magallana gigas

Evidence

The majority of the evidence indicates that circalittoral 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 depth and sedimentation probably exclude macroalgae from this biotope. Hence, it is unlikely to be colonized by Sargassum. Therefore, this biotope is probably 'Not sensitive to this INIS.

Wakame, Undaria pinnatifida

Evidence

The depth and sedimentation probably exclude macroalgae from this biotope. Hence, it is unlikely to be colonized by Undaria. Therefore, this biotope is probably 'Not sensitive’ to this INIS.

Other INIS

Evidence

This biotope is classified as circalittoral and therefore no algal species have been considered. Several invasive bryozoans are of concern, including Bugula neritina (Porter et al., 2017), Schizoporella japonica (Ryland et al., 2014) and Tricellaria inopinata (Dyrynda et al., 2000; Cook et al., 2013b), however, evidence of potential effects is sparse. Limited evidence exists for Bugula neritina in the UK; however, it was found in Northern Ireland for the first time in 2006, then seen again in both 2008 and 2013 in abundance and attached to buoys and kelp stipes throughout Carrickfergus Marina (Porter et al., 2017). Tricellaria inopinata has been reported to colonize the byssal threads of the mussel Mytilus galloprovincialis, Hymeniacidon perleve and the ascidian Styela plicata (Dyrynda et al., 2000). Tricellaria inopinata dominated the fouling community in the Lagoon of Venice, within seven years of being introduced (Ambrogi, 2000).

The Orange striped anemone, Diadumene lineata, is a possible invasive species to British waters. Diadumene lineata is a ‘cosmopolitan invader’ with extreme tolerance to environmental stress, especially salinity (Podbielski et al., 2016). At salinities of 24 to 34, Diadumene lineata has been reported as growing quickly, exhibiting a fivefold population growth through asexual reproduction within four weeks (Podbielski et al., 2016). The impacts from Diadumene lineata are not fully clear, however, as the species has a high potential for rapid population expansion, even from a single individual, with up to 939 individuals per m² (Podbielski et al., 2016), it may smother or out-compete native species and alter benthic community compositions, as has been the case with blue mussel, Mytilus edulis, beds (Podbielski et al., 2016).

At present, there is 'Insufficient evidence' to suggest that the circalittoral biotopes are sensitive to colonization by algal or other invasive species; further evidence is required.