The Marine Life Information Network

Temperature increase (local)

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

Evidence

Gordillo & Aitken (2000) in a review of environmental factors relevant to re-interpreting Late Quaternary environments from fossil collections suggest that Hiatella arctica is eurythermal, based on Aitken (1990) and Peacock (1993).  The current distribution of Hiatella arctica is predominantly arctic and boreal (Sejr et al., 2004; Gordillo, 2001) and palaeecological reviews describe the genus as ‘consistently linked to cool temperate and polar regions’ (Gordillo, 2001). However, populations of Hiatella arctica occur in the Mediterrannean and have clearly acclimated to the warmer temperatures (Oberlechner, 2008). Laboratory experiments on filtration rates of Hiatella arctica found that activity was strongly linked to temperature (Ali, 1970). Activity rates rose steadily between 0 ^oC to a maximum between 15 ^oC and 17 ^oC and fell sharply to almost no activity at 25 ^oC  (Ali, 1970).  Although activity may be reduced Hiatella arctica have very low metabolic rates and may be able to sustain a  period of reduced activity. Regression models developed by Bourget et al. (2003) found that temperature and water transparency (measured in metres and indicating the level of inorganic suspended solids) explained only 40% of the variation in biomass of Hiatella arctica fouling navigation buoys in the Gulf of St Lawrence system (Canada). These findings suggest that other variables play a more significant role in determining settlement, survival and growth over a year in this system. However the models did indicate that biomass is higher where temperatures were greater (around 14 ^oC) although a causal link was not identified (Bourget et al., (2003).

Asterias rubens is abundant throughout the north-east Atlantic, from Arctic Norway, along Atlantic coasts to Senegal, and only found occasionally in the Mediterranean (Mortensen, 1927).  The geographic range of Asterias rubens illustrates that the species is tolerant of a range of temperatures and probably becomes locally adapted. Asterias rubens was reported to be unaffected by the severe winter of 1962-1963 in Britain when anomalously low temperatures persisted for two months (Crisp, 1964). Ophiothrix fragilis also has a geographically wide distribution, ranging from northern Norway, south to the Cape of Good Hope.  Consequently, this species is exposed to temperatures both above and below those found in the UK. Temperature is also a critical factor stimulating or inhibiting reproduction in hydroids, most of which have an optimum temperature range for reproduction (Gili & Hughes, 1995). Most of the hydroid and bryozoan species within the biotope are recorded to the north or south of the UK and are unlikely to be adversely affected by long-term increases in temperature at the benchmark level.

Sensitivity assessment. No direct evidence was found to assess sensitivity to this pressure however, the experiments by (Ali, 1970) suggest that Hiatella arctica would be able to tolerate an acute or chronic increase in temperature at the pressure benchmark although an acute or chronic increase may result in sub-lethal effects on feeding and hence a reduction in growth and potentially reproduction. Based on the geographic range of Hiatella arctica and other associated species, the biotope would be able to tolerate either an acute or chronic change in temperature at the pressure benchmark. Resistance is therefore assessed as ‘High’ and resilience as ‘High’ by default, so that the biotope is considered to be ‘Not sensitive’. 

Temperature decrease (local)

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

Evidence

Gordillo & Aitken (2000) in a review of environmental factors relevant to re-interpreting Late Quaternary environments from fossil collections suggest that Hiatella arctica is eurythermal, based on Aitken (1990) and Peacock (1993).  The current distribution of Hiatella arctica is predominantly arctic and boreal (Sejr, et al., 2004; Gordillo, 2001) and palaeecological reviews describe the genus as ‘consistently linked to cool temperate and polar regions’ (Gordillo, 2001) suggesting that within temperate regions this species would not be sensitive to a decrease in temperature at the pressure benchmark. Regression models developed by Bourget et al. (2003) found that temperature and water transparency (measured in metres and indicating the level of inorganic suspended solids) explained only 40% of the variation in biomass of Hiatella arctica fouling navigation buoys in the Gulf of St Lawrence system (Canada). These findings suggest that other variables play a more significant role in determining settlement, survival and growth over a year in this system. However the models did indicate that biomass is higher where temperatures were greater (around 14 ^oC) although a causal link was not identified (Bourget et al., (2003). 

Asterias rubens is abundant throughout the north-east Atlantic, from Arctic Norway, along Atlantic coasts to Senegal, and only found occasionally in the Mediterranean (Mortensen, 1927).  The geographic range of Asterias rubens illustrates that the species is tolerant of a range of temperatures and probably becomes locally adapted.  Asterias rubens was reported to be unaffected by the severe winter of 1962-1963 in Britain when anomalously low temperatures persisted for two months (Crisp, 1964). Brittlestar populations have experienced mass mortalities when exposed to very low water temperatures in winter.  Populations of Ophiothrix fragilis inhabiting shallow subtidal habitats (5-7m depth) in the Dutch Oosterschelde Estuary were greatly reduced (to less than 10% spatial coverage) following cold winters in 1978-79, 1984-85 and 1985-86 (Leewis et al., 1994).  However, these decreases in temperature exceed the pressure benchmark. Temperature is also a critical factor stimulating or inhibiting reproduction in hydroids, most of which have an optimum temperature range for reproduction (Gili & Hughes, 1995). Most of the hydroid and bryozoan species within the biotope are recorded to the north or south of the UK and are unlikely to be adversely affected by acute or short-term decreases in temperature at the benchmark level.

Sensitivity assessment. Based on distribution of the key characterizing species Hiatella arctica and other associated species, the biotope is considered to have ‘High’ resistance to an acute or chronic decrease in temperature at the pressure benchmark and ‘High’ resilience (by default). The biotope is therefore considered to be ‘Not sensitive’.  

Salinity increase (local)

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

Evidence

Filipov et al., (2003, abstract only) tested the salinity tolerances of Hiatella arctica obtained from the White Sea. The salinity tolerance of individuals kept at 25 ppt was 17-36 ppt. Acclimation of Hiatella arctica allowed them to adapt to higher or lower salinities with the potential tolerance range of acclimated individuals assessed as 13-42 ppt. 

Sensitivity assessment. Caution should be used when extrapolating results from short-term, laboratory experiments on individuals collected from other geographic ranges. However, the results from Filippov et al., (2003) suggest that Hiatella arctica is relatively euryhaline and may acclimate to increases in salinity > 40 ppt. Resistance is therefore assessed as ‘High’ and resilience as ‘High’ (by default). The biotope is therefore classed as ‘Not sensitive’. Some reduction in species richness may occur as less tolerant associated species either move away or perish but this is not considered to significantly impact the character of the biotope. 

Salinity decrease (local)

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

Evidence

This biotope is reported to occur in full salinity (30-35 ppt) Connor et al., (2004). A change in salinity at the pressure benchmark therefore refers to a change to reduced (18-30 ppt) or variable salinity (18-35 ppt). Filipovv et al., (2003, abstract only) tested the salinity tolerances of Hiatella arctica obtained from the White Sea. The salinity tolerance of individuals kept at 25 ppt was 17-36 ppt. Acclimation of Hiatella arctica allowed them to adapt to higher or lower salinities with the potential tolerance range of acclimated individuals assessed as 13-42 ppt. Gordillo & Aitken (2000) in a review of environmental factors relevant to re-interpreting Late Quaternary environments from fossil collections suggest that the normal minimum salinity tolerance of Hiatella arctica is 20 ppt, based on Aitken (1990) and Peacock (1993).

 Echinoderms are stenohaline owing to the lack of an excretory organ and a poor ability to osmo- and ion-regulate (Stickle and Diehl, 1987, Russell 2013).  This means that they are unable to tolerate wide fluctuations in salinity and are considered sensitive to a decrease in salinity at the pressure benchmark. However, there are examples where brittlestars have been recorded to persist in low salinity habitats.  For example, dense Ophiothrix aggregations  have been recorded in areas where normal salinity is only 16.5 ppt (Wolff, 1968, cited from Hughes, 1998), with the species was found to persist down to 10 ppt; Asterias rubens has been reported from areas of reduced salinity, e.g. Loch Etive, Scotland (16 ‰) and the Baltic Sea (8 ‰), and (reported as Asteria vulgaris) the east coast of N. America (18 ‰), the Netherlands (18 ‰) and Maine (27.4 ‰) (Russell , 2013).  Binyon (1961) demonstrated all specimens exposed to 18‰ for one week died, while those exposed to 25‰ for the same period all survived.  Binyon (1961) determined that their LD50 was between 22-24‰.  He also noted that the Baltic specimens tolerated 8‰ and were probably a ‘physiological’ race; that is, adapted to low salinity.  Russell (2013) reviewed additional experimental studies in which Asterias rubens was reported to experience mortality at 26‰, 22‰ or 12‰, and tolerate 27.4‰ and 14‰.  The results suggest local or regional variation in tolerance.  Echinoderm larvae have a narrow range of salinity tolerance and will develop abnormally and die if exposed to reduced or increased salinity. Similarly Ryland (1970) stated that, with a few exceptions, bryozoans the Gymnolaemata were fairly stenohaline and restricted to full salinity (ca 35 psu) and noted that reduced salinities result in an impoverished bryozoan fauna.

Sensitivity assessment. Caution should be used when extrapolating results from short-term, laboratory experiments on individuals collected from other geographic ranges. However, the results from Filippov et al., (2003) suggest that Hiatella arctica is relatively euryhaline and may acclimate to decreases in salinity from full to reduced (18-30ppt) or variable (18-35 ppt) . The impact will be mediated by the length of exposure to lower salinities, with the evidence suggesting that long-term exposure to salinities < 20 ppt harmful. Reductions in salinity at the lower end of the pressure benchmark are likely to result in a reduction in species abundance and richness as less tolerant species either move away or perish. Resistance is assessed as ‘Medium’ based on the resistance of Hiatella arctica and resilience as ‘High’. The biotope is therefore classed as having 'Low' sensitivity. 

Water flow (tidal current) changes (local)

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

Evidence

The key characterizing species, Hiatella arctica are protected from water flows within burrows, although they and other associated species may be indirectly affected by changes in water movement where these impact the supply of food or larvae or other processes. Connor et al. (2004) report that this biotope is found in a range of water flows from moderately strong (1-3 knots- 0/5- 1.5 m/s) to areas where the flow is negligible Connor et al., (2004).

Most species are likely to be tolerant of changes in waterflow at the pressure benchmark. The hydroid Nemertesia antennina, which occurs in clumps in this biotope, is found in areas where water flows range from very weak to strong (negligible -3m/s); Dense brittlestar beds are found in a range of water flows from sea lochs with restricted water flows to higher-energy environments on open coastlines.  In the Dover Strait, Ophiothrix beds experience current speeds of up to 1.5m/s during average spring tides (Davoult & Gounin, 1995).  Similarly strong tidal streams (1.0 -1.2m/s) were also recorded over beds in the Isle of Man (Brun, 1969).  Davoult & Gounin (1995) found that current speeds below 0.2m/s were optimal for suspension feeding by Ophiothrix fragilis; if velocity exceeded 0.3 m/s the animals cease feeding, flatten themselves against the substratum and link arms, so increasing their collective stability in the current. 

Sensitivity assessment.  The range of flow rates experienced by the biotope is considered to indicate, by proxy, that the biotope would have ‘High’ resistance and by default ‘High’ resilience to a change in water flow at the pressure benchmark. The biotope is therefore classed as ‘Not sensitive’. This assessment is supported by evidence for the range of flow speeds in which associated species are found.

Emergence regime changes

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

Evidence

Changes in emergence are not relevant to this biotope (group) which is restricted to fully subtidal habitats. It should be noted that Hiatella arctica occur within the intertidal and subtidally and that the presence of suitable substratum rather than emergence regime is a more significant factor determining the distribution

Wave exposure changes (local)

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

Evidence

This biotope occurs in moderately exposed shores (JNCC, 2022). The key characterizing species, Hiatella arctica are protected within burrows from the oscillatory water flows at the seabed, although they and other associated species may be indirectly affected by changes in water movement where these impact the supply of food or larvae or other processes. Trudgill & Crabtree (1987) observed Hiatella arctica at both sheltered and wave exposed sites, suggesting that substratum, rather than wave action is a more significant factor determining distribution.

Sensitivity assessment. The range of wave exposures experienced by the similar biotope IR.MIR.KR.HiaSw is considered to indicate, by proxy, that this biotope would have ‘High’ resistance and by default ‘High’ resilience to a significant change in wave height (3 to 5%). The biotope is therefore classed as ‘Not Sensitive’.

Transition elements & organo-metal contamination

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

Evidence

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

Hydrocarbon & PAH contamination

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

Evidence

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

Synthetic compound contamination

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

Evidence

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

Radionuclide contamination

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

Evidence

No evidence.

Introduction of other substances

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

Evidence

This pressure is Not assessed.

De-oxygenation

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

Evidence

No evidence.

Nutrient enrichment

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

Evidence

This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations.  The benchmark is set at compliance with WFD criteria for good status, based on nitrogen concentration (UKTAG, 2014).  No evidence was found to assess the sensitivity of piddocks to this pressure.  Nutrient enrichment that enhances productivity of phytoplankton may indirectly benefit the suspension feeding piddocks by increasing food supply. No direct evidence was found to assess this pressure. Hiatella arctica is a fouling species present at fish farms suggesting that it is tolerant of the increased nutrient levels associated with fish aquaculture. Moderate increases in nutrient levels may benefit suspension feeding members of the associated species assemblage by increasing macroalgal and phytoplankton productivity, increasing the proportion of particulate and dissolved organic matter and hence increasing the food supply. 

Sensitivity assessment. The benchmark is relatively protective and the presence of the key characterizing species Hiatella arctica within fish farms suggests that biotope resistance to this pressure is 'High', resilience is 'High' (by default) and the biotope is considered to be 'Not sensitive'.

Organic enrichment

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

Evidence

No evidence was found for the key characterizing species Hiatella arctica.This biotope is found on vertical rock surface that may limit the deposition of organic matter on the surface. In addition the biotope is often found in areas with high levels of water flow that support assemblages of suspension feeders. Currents will remove organic matter limiting exposure to the pressure and over the course of the year low levels of input may be consumed by the brittle star Ophiothrix fragilis, sponges and ascidians and other suspension feeders within the biotope. Borja et al., (2000) and Gittenberger & van Loon (2011) in the development of the AZTI Marine Biotic Index (AMBI), a biotic index to assess disturbance (including organic enrichment), both assigned Asterias rubens to their ecological group III of species that are ‘tolerant to excess organic matter enrichment’.  Hall-Spencer et al., (2006) observed a much higher abundance (10-100 times higher abundance) of Asterias rubens beneath salmon farms where organic enrichment had led to a visible build-up of wastes compared to reference areas. 

Sensitivity assessment. This biotope occurs in areas where vertical rock is present, often with strong water flows, both these factors would reduce deposition of organic matter within the biotope. As suspension feeders within the biotope could capture and cycle organic matter, resistance to this pressure at the benchmark is assessed as 'High', resilience is assessed as 'High' (by default) 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. Further detail

Evidence

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

Physical change (to another seabed type)

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

Evidence

This biotope is characterized by the soft rock substratum which supports populations of burrowing Hiatella arctica. A change to a sedimentary, hard rock or artificial substratum result in the loss of burrowing Hiatella arctica significantly altering the character of the biotope. The biotope is therefore considered to have 'No' resistance to this pressure, recovery of the biological assemblage (following habitat restoration) is considered to be 'Medium' (2-10 years) but see caveats in the recovery notes. The biotope is dependent on the presence of soft rock, as the change at the pressure benchmark is considered to be permanent recovery is categorised as 'Very low'. Sensitivity is therefore assessed as '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 sediment type)

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

Evidence

Not relevant to biotopes occurring on bedrock.

Habitat structure changes - removal of substratum (extraction)

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

Evidence

The removal of substratum to 30 cm depth will remove the attached epiflora and epifauna and burrowing Hiatella arctica in the impact footprint.  Resistance is therefore assessed as ‘None’, recovery of the biological assemblage (following habitat restoration) is considered to be 'Medium' (2-10 years) but see caveats in the recovery notes. The biotope is dependent on the presence of soft rock to support populations of the characterizing Hiatella arctica, when lost restoration would not be feasible and recovery is therefore categorised as 'Very low'. Sensitivity is therefore assessed as 'High', based on the lack of recovery of the substratum. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

Abrasion / disturbance of the surface of the substratum or seabed

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

Evidence

Hiatella arctica burrow depths were approximately 2 cm (mean length of Hiatella arctica individuals was 1-1.2 cm) with a maximum depth of 4 cm on limestone shores off the coast of Ireland (Trudgill & Crabtree,1987). The burrowing life habit provides some protection from abrasion at the surface but the presence of burrows will weaken the mechanical strength of the rock. The surface epifauna and flora are more susceptible to damage and removal by surface abrasion.

The available evidence indicates that attached epifauna, such as members of this ecological group, can be entangled and removed by abrasion.  Drop down video surveys of Scottish reefs exposed to trawling showed that visual evidence of damage to bryozoans and hydroids on rock surfaces was generally limited and restricted to scrape scars on boulders (Boulcott & Howell, 2011).  The study showed that damage is incremental with damage increasing with frequency of trawls rather than a blanket effect occurring on the pass of the first trawls.  The level of impact may be mediated by the rugosity of the attachment, surfaces with greater damage occurring over smooth terrains where the fishing gear can move unimpeded across a flat surface.  Veale et al. (2000) reported that the abundance, biomass and production of epifaunal assemblages decreased with increasing fishing effort.  Re-sampling of grounds that were historically studied (from the 1930s) indicates that some upright species have increased in areas subject to scallop fishing (Bradshaw et al. 2002).  This study also found increases in the tough stemmed hydroids including Nemertesia spp., whose morphology may prevent excessive damage.  Bradshaw et al. (2002) suggested that as well as having high resistance to abrasion pressures, Nemertesia spp. have benthic larvae that could rapidly colonise disturbed areas with newly exposed substrata close to the adult. 

Re-sampling of grounds that were historically studied (from the 1930s) indicates that Ophiothrix fragilis has declined in areas subject to scallop fishing (Bradshaw et al., 2002).  Examination of historical and recent samples suggest that the spatial presence of Ophiothrix fragilis and Amphiura spp. in the North Sea has more than halved in comparison with the number of ICES rectangles in which they were sampled at the beginning of the century, apparently in response to fishing effort (Callaway et al., 2007). 

Sensitivity assessment. Erect epifauna are directly exposed to abrasion and sub-surface penetration which would displace, damage and remove individuals (de Groot 1984; Veale et al., 2000; Boulcott & Howell, 2011).  Abrasion may also damage the substratum resulting in loss of habitat and exposure of Hiatella arctica. Resilience of some associated species will be ‘High’ with recovery  occurring through repair, asexual reproduction and larval settlement. However, resilience of the biotope is assessed as ‘Medium ‘as some slower growing species may require longer to re-establish. Sensitivity is therefore assessed as ‘Medium’.

Penetration or disturbance of the substratum subsurface

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

Evidence

Penetration and disturbance below the surface of the substratum may damage and remove the surface dwelling fauna and could damage and expose the Hiatella arctica depnding on depth of penetration and burrow depth. Burrow depths were approximately 2 cm (mean length of Hiatella arctica individuals was 1-1.2 cm) with a maximum depth of 4 cm on limestone shores off the coast of Ireland (Trudgill & Crabtree (1997).  Suggesting that even shallow damage would impact individuals. Hiatella arctica in damaged burrows or those that are removed from the substratum are unlikely to be able to rebury and will be predated by fish and other mobile species. 

The evidence presented in the abrasion pressure is also relevant to sub-surface damage and penetration which will include a level of abrasion. Activities resulting in penetration and disturbance can directly affect epifauna by crushing or removal, Sub-surface disturbance may also remove the habitat by breaking up and removing the substratum. Natural erosion processes are, however, likely to be on-going within this habitat type. Where abundant the boring activities of Hiatella arctica contribute significantly to bioerosion, which can make the substratum habitat more unstable and can result in increased rates of erosion (Trudgill & Crabtree, 1987). 

Sensitivity assessment. Sub-surface penetration and disturbance could result in damage and removal of the surface epifauna and result in the damage, exposure and loss of Hiatella arctica and damage to the habitat. Resistance is therefore assessed as ‘Low’.  The associated surface dwelling fauana are predicted to recover relatively rapidly via regrowth, larval recolonisation and migration of adults in mobile species. Recovery of the key characterizing species, Hiatella arctica is predicted to require 2-10 years so that resilience is considered to be ‘Medium’ and sensitivity is ‘Medium’. As the substratum cannot recover, resilience is assessed as ‘Very Low’ and sensitivity of the overall biotope, based on the sedimentary habitat, is considered to be ‘High’.  

Changes in suspended solids (water clarity)

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

Evidence

Increased suspended particles will decrease light penetration, may enhance food supply (where these are organic in origin), or decrease feeding efficiency (where the particles are inorganic and require greater filtration efforts). Very high levels of silt may clog respiratory and feeding organs of some suspension feeders. Erosion of soft rock mean that this biotope, in common with other chalk biotopes, is susceptible to high turbidity and low light penetration (Connor et al., 2004). Hiatella arctica is a filter feeding bivalve, many other species of this type have efficient mechanisms to remove inorganic particles via pseudofaeces. For example, Petricolaria pholadiformis can cope in water laden with much suspended material by binding the material in mucus and using the palps to reject it (Purchon, 1955). Hiatella arctica is protected from scour within burrows and increased organic particles may provide a food subsidy. Increased suspended sediments may impose sub-lethal energetic costs on bivalves by reducing feeding efficiency and requiring the production of pseudofaeces with impacts on growth and reproduction.

Local increases in turbidity in waters previously within the photic zone, may alter local abundances of phytoplankton and surface diatoms and the zooplankton and other small invertebrates that feed on them. An increase in suspended solids may therefore indirectly reduce feeding efficiency. However, where the pressure results from an increase in suspended organic matter, this would be beneficial to some suspension feeders by providing increased food material (and perhaps local stimulation of phytoplankton abundance where nutrients are recycled back to the water column).

Regression models developed by Bourget et al. (2003) found that temperature and water transparency (measured in metres and indicating the level of inorganic suspended solids) explained only 40% of the variation in biomass of Hiatella arctica fouling navigation buoys in the Gulf of St Lawrence system (Canada). These findings suggest that other variables play a more significant role in determining settlement, survival and growth over a year in this system. However, the models did indicate that biomass is higher where water transparency was greater (around 15 m) and declined at higher levels of suspended solids (transparency 5 m) although a causal link was not identified (Bourget et al., 2003).

Sensitivity assessment. No direct evidence was found to assess sensitivity to this pressure. A decrease in turbidity increasing light penetration may allow some algae to colonize and could lead to the development of an assemblage resembling that of the infralittoral limestone biotope IR.MIR.KR.HiaSw. However, a year is not considered long enough to lead to the development of this community, particularly as space is occupied by attached epifauna. Resistance is therefore assessed as ‘High’, resilience as ‘High’ (by default) and the biotope is considered to be ‘Not Sensitive’. As the biotope occurs in turbid waters a further increase in turbidity may exceed the tolerances of Hiatella arctica and other associated species. Resistance is therefore assessed as ‘Medium’ and resilience as ‘High’ so that sensitivity is assessed as ‘Low’, albeit with ‘low’ confidence due to a lack of direct evidence.

Smothering and siltation rate changes (light)

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

Evidence

Exposure to siltation pressures will be mediated by site-specific topography and hydrodynamics as silts may not accumulate on vertical surfaces, especially where these are smooth, although some deposits may be trapped by epifauna and epiflora (where these occur). Water currents or wave action may also be sufficient to rapidly remove fine particles although this will depend partially on the scale of the impact. As Hiatella arctica are essentially sedentary with relatively short siphons, siltation from fine sediments rather than sands, even at low levels for short periods could be lethal. Siltation by fine sediments would also prevent larval settlement for species which require hard substratum (Berghahn & Offermann, 1999).  In general it appears that hydroids are sensitive to silting (Boero, 1984; Gili & Hughes, 1995) and decline in beds in the Wadden Sea has 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.

Sensitivity assessment. As this biotope occurs on vertical surfaces siltation may be limited. However, in general, resistance to this pressure is assessed as ‘Low’ as siltation may smother Hiatella arctica and other associated species. Resilience is assessed as ‘Medium’ and sensitivity is therefore assessed as ‘Medium’.

Smothering and siltation rate changes (heavy)

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

Evidence

Sensitivity to this pressure will be mediated by site-specific hydrodynamic conditions and the footprint of the impact. Where a large area is covered sediments may be shifted by wave and tides rather than removed. As Hiatella arctica are essentially sedentary with relatively short siphons, siltation from fine sediments rather than sands, even at low levels for short periods could be lethal.  In general it appears that hydroids are sensitive to silting (Boero, 1984; Gili & Hughes, 1995) and decline in beds in the Wadden Sea has 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.

Sensitivity assessment. As this biotope occurs on vertical surfaces siltation may be limited.  However, in general resistance to siltation is assessed as ‘Low’ as siltation may smother Hiatella arctica and other associated species. Resilience is assessed as ‘Medium’ (2-10 years) and sensitivity is therefore assessed as ‘Medium’.  Survival will be higher in winter months when temperatures are lower and physiological demands are decreased.  However, mortality will depend on the duration of smothering. Mortality is likely to be more significant in wave sheltered areas where the smothering sediment remains for prolonged periods and reduced where the smothering sediment is rapidly removed by wave action or currents.

Litter

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

Evidence

Not assessed.

Electromagnetic changes

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

Evidence

Underwater noise changes

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

Evidence

Not relevant.

Introduction of light or shading

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

Evidence

This biotope occurs on vertical surfaces where shading will prevent direct light. Where levels of turbidity are high, light penetration will be further reduced. Hiatella arctica and other species present in this biotope occur in the intertidal and shallow subtidal where light levels are high, as well as deeper water where light penetration is limited. Increases in light may allow a more diverse and abundant algal community to develop, however the soft rock surfaces and turbidity that characterize this biotope are generally unsuitable for large attached macroalgae and a shift to an algal dominated biotope is considered unlikely. 

Sensitivity assessment. This biotope is considered to have 'High' resistance to changes in light levels and 'High' resilience (by default), so the biotope is considered to be 'Not Sensitive' to this pressure.

Barrier to species movement

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

Evidence

Barriers that reduce the degree of tidal excursion may alter larval supply to suitable habitats from source populations. Conversely the presence of barriers may enhance local population supply by preventing the loss of larvae from enclosed habitats.  Hiatella arctica and species associated with the biotope are widely distributed and produce large numbers of larvae capable of long distance transport and survival, resistance to this pressure is assessed as 'High' and resilience as 'High' by default. This biotope is therefore considered to be 'Not sensitive'. 

Death or injury by collision

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

Evidence

Not relevant’ to seabed habitats.  NB. Collision by grounding vessels is addressed under ‘surface abrasion.

Visual disturbance

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

Evidence

Not relevant.

Genetic modification & translocation of indigenous species

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

Evidence

Key characterizing species within this biotope are not cultivated or translocated. This pressure is therefore considered ‘Not relevant’ to this biotope group.

Introduction of microbial pathogens

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

Evidence

No evidence was found for the impact of microbial pathogens on characterizing species, based on the lack of evidence for outbreaks of disease or significant mortality this biotope was considered to have 'High' resistance to this pressure and 'High' resilience (by default), and is therefore assessed as 'Not sensitive'.

Removal of target species

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

Evidence

No species within the biotope description (Connor et al., 2004) are targeted commercially. This pressure is therefore considered to be 'Not relevant' to this biotope. The effects of removal of non-target species (by-catch) are assessed separately.

Removal of non-target species

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

Evidence

The epifauna present in this biotope may be removed or damaged by activities targeting other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. Removal of the epifauana as by-catch would alter the character of the biotope and result in the loss of the ecosystem functions and habitat structure created by these species. In general the attached species present are found in low densities due to the nature of the substratum which is too soft for epifauna and flora to attach to or to maintain attachment (Connor et al., 2004). It is unlikely that targeted harvesting of other species would remove all of the species present or unintentionally remove the key characterizing Hiatella arctica species which are protected within burrows. Resistance of the biotope to this pressure is assessed as 'Medium' as the effects are through removal of the associated assemblage rather than the key characterizing Hiatella arctica species and resilience is assessed as 'High' so that sensitivity is assessed as 'Low'. This biotope is found on vertical surfaces which may limit exposure to activities that result in this pressure.

The American slipper limpet, Crepidula fornicata

Evidence

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).

A lack of published descriptions and an incomplete historical record, has led to the widespread misidentification of Didemnum vexillum and it is often recorded as Didemnum spp. Hence, the native range of the species is not known conclusively (Lambert, 2009; Stefaniak et al., 2012; Mckenzie et al., 2017; Holt, 2024). However, molecular data and limited historical evidence have suggested that the species may be native to Japan with its native range possibly extending into continental Asia and north-western Pacific (Stefaniak et al., 2012; Tillin et al., 2020; Holt, 2024). Previously unrecorded populations of a colonial ascidian have been recently identified as Didemnum vexillum (Tillin et al., 2020).

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

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

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

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

Human-meditated transport via aquaculture facilities, boat hulls, commercial fishing vessels, ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijstra et al., 2007; Griffiths et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009). 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 human-mediated dispersal, between marinas and shellfish culture sites, is the most likely pathway for connectivity of Didemnum vexillum populations throughout Ireland and Britain (Prentice et al., 2021; Holt, 2024). Didemnum vexillum can disperse away from artificial substrata, invading and colonizing natural substrata in surrounding areas (Tillin et al., 2020). Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. The current evidence of Didemnum vexillum’s ability to spread on natural habitats in this area is sparse and often conflicting, complicated by genetics and its apparent variable habitat preferences and tolerances and its variable ability to adapt to ‘new’ conditions (Holt 2024).

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

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

In addition, Didemnum vexillum is commonly found associated with artificial hard substrata, being mostly found in harbours and marinas where it covers a variety of maritime structures such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ships hulls and at aquaculture facilities (Valentine et al., 2007 a&b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012). Didemnum vexillum was abundant in the marinas at Terschelling, Texel, and Vlieland, in the Wadden Sea, (Gittenberger et al., 2015). In the UK, Didemnum vexillum was initially recorded in marinas and adjacent shallow man-made structures (Tillin et al., 2020). In Wales, it was first recorded in Holyhead marina, then subsequently reported in Plymouth marina and other marinas around the UK (Griffith et al., 2009; Minchin & Nunn, 2013; Bishop et al., 2015).

Didemnum sp. can colonize both horizontal and vertical surfaces of fouling and benthic communities, commonly occurring on upper horizontal surfaces in benthic habitats (Dijkstra et al., 2007; Tillin et al., 2020). It has been recorded on overhangs or the underside of boulders (Hitchin, 2012) or on the underside of docks, boat hulls, pontoons (Griffiths et al., 2009; Minchin & Nunn, 2013). In sheltered areas, colonies are lobed and beard-like, forming long tendrils that drop down from the underside of docks or other artificial substrata to establish new colonies if there are suitable substrata available (Valentine et al., 2007a). In areas of stronger current, colonies are low undulating mats (Valentine et al., 2007 a&b).

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

In aquaculture, Didemnum sp. has been recorded on fish farm gear, cages, nets and other equipment (Tillin et al., 2020). In British Columbia, oyster and mussel culture facilities are known to be heavily fouled by Didemnum sp. (Bullard et al., 2007). Didemnum sp. has been recorded densely covering nets of bluefin tuna cages in the Aegean Sea, almost clogging the opening on some parts of the nets (Cinar & Ozgul, 2023). These examples demonstrate economic impacts for oyster and mussel aquaculture operations through decline in farmed species and removal of biofouling on equipment (Tillin et al., 2020).

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

However, some studies on Georges Bank, USA, and in Sandwich, Massachusetts, observed that colonies were able to survive partial burial by sand (Bullard et al., 2007; Valentine et al., 2007a). Gittenberger et al. (2015) reported that Didemnum vexillum was able to overgrow sandy bottoms (citing Gittenberger, 2007). In the Netherlands, the coastal zone is composed primarily of mud and sand, with shells providing the only hard substrata. Didemnum sp. remained rare until 1996, when populations rapidly expanded and the species became a dominant invader. This expansion was attributed to an increase in available hard substrata for colonization following a cold winter between 1995 and 1996 that reduced the abundance of many marine animals (Gittenberger, 2007). Thus, Didemnum vexillum was able to colonize and establish in mud and sand habitats where hard substrata were present.

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

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

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

In northern Kent, Didemnum vexillum has been recorded covering London clay boulders on Whitstable Flats, West Beach; tabulate sandstone boulders (0.5 to 2 m across) on the mid shore; and sediment mounds on the low shore, characterized by larger areas of sand, mud, and low-lying sediment at Reculver and Bishopstone (Hitchin, 2012). It was also recorded in muddy substrata at that site. Hitchin (2012) noted that the site was exposed to enough waves and currents to cause sedimentation. However, Didemnum vexillum grew hanging from the underside of sandstone boulders nestled on sediment, on consolidated sediment mounds and firm clays, hence burial may prevent colonization and its survival rather than sedimentation alone.

An experiment in the Thames River estuary, Connecticut, found a significant difference between Didemnum vexillum growth rates at different depths, with faster growth rates seen in shallow water (1.0 m) compared to deeper depths (4.0 m) (Bullard & Whitlatch, 2009). It was also found that although Didemnum vexillum grew faster in shallow depths during the experiment, it grew well at all depths examined (1.0 m, 2.5 m and 4.0 m) and there was no significant difference in survival between the depths.

The Sandwich tide pools were subject to air exposure at low tide, daily changes in water depth and temperature (Valentine et al., 2007a). Didemnum vexillum colonies can survive exposure to air at low tides during rapid colony growth in summer months July to September (Valentine et al., 2007a). However, parts of the large established colonies, which were artificially exposed to air for two to three hours in October, were observed desiccated or predated on by grazing periwinkles 30 days later, in the winter month November (Valentine et al., 2007a). They suggested that the invasive tunicates’ ability to tolerate exposure to air varies with the seasonal growth cycle. Didemnum vexillum also tolerated emersion in Kent, as colonies on the mid-shore at Reculver flourish and survive in air exposure for up to three hours per cycle during spring tides (Hitchin, 2012). Hitchin (2012) suggested the porous nature of the sandstone boulders the species colonized retained water. The Kent shore was sheltered but held water due to its shallow slope and flats, which may allow Didemnum sp. to survive in the low to mid-shore. There is evidence that Didemnum vexillum died when exposed to air for more than 6 hours (Laing et al., 2010).

Zhang et al. (2020) suggested that in the current climate conditions (based on depth, current, temperature and salinity) Didemnum vexillum had not yet occupied their predicted suitable habitats, and suggested many suitable habitats around the world, are at risk due. Zhang et al. (2020) predicted that the Northern Atlantic coast is susceptible to invasion by Didemnum vexillum and that climate change will cause a poleward expansion of Didemnum vexillum.

Didemnum vexillum tolerates a wide range of environmental conditions including temperature and salinity (Herborg et al., 2009; Tillin et al., 2020). Didemnum vexillum can withstand a wide range of salinities from 20 to 44 ppt, is commonly found in marine waters around 33 ppt but is unable to survive in salinities below 20 ppt (Bullard & Whitlatch, 2009; Groner et al., 2011; Tillin et al., 2020). It has been recorded in estuarine conditions and tidal lagoons (Dijkstra et al., 2007; Tillin et al., 2020). In the Lagoon of Venice, Didemnum vexillum is found in waters at 30 PSU. It was absent in low salinity, such as the estuary and around the saltmarshes, but well established in the euhaline and tidally well flushed zones of the Lagoon of Venice (Tagliapietra et al., 2012). Similar results were found in Connecticut and Rhode Island where Didemnum vexillum was not found in environments with salinity less than 20 ppt (Bullard & Whitlatch, 2009). However, in the Wadden Sea, colonies of Didemnum vexillum were abundant in salinities between 17.91 to 25.97 ppt (Gittenberger, 2007; Gittenberger et al., 2015).

Salinity can influence the growth rates of Didemnum vexillum. For example, in an experiment in the Thames River estuary, Connecticut, Bullard & Whitlatch (2009) found growth rates were significantly higher in high salinity areas (26 to 30 ppt) and although survival at different salinities was not significantly different, the Didemnum vexillum colonies in low (10 to 26 ppt) and medium (15 to 28 ppt) salinities were bloated, discoloured and appeared to be dying. In unpublished data from Bullard & Whitlatch (2009), similar results were found in the laboratory as most colonies appeared to be dying after one week in 20 ppt and healthy in 30 ppt.

A study on Didemnum vexillum colonies from Holyhead Marina, Isle of Anglesey, found colony growth within a week was significantly impaired and reduced by two thirds at lower salinities (27 PSU and 20 PSU), while in ambient Holyhead Marina salinity (34 PSU) the growth increased and surface area doubled (Groner et al., 2011). Mortality was described as negligible in colonies of Didemnum vexillum in ambient salinity (34 PSU) after two weeks. However, mortality increased as salinity decreased. At the end of the two-week experiment, 72% of invasive colonies survived in 27 PSU and 55% of colonies survived in 20 ppt (Groner et al., 2011). When exposed to severe low salinity of 10 PSU for two hours, Didemnum vexillum showed no mortality, which suggested the duration of exposure influences mortality, not the stress intensity (Groner et al., 2011). Colonies of Didemnum vexillum collected from Angelsey, Wales, experienced more mortality under severe hypo-salinity (20 PSU, 38% colonies survived) compared to moderate hypo-salinity (27 PSU, 82% colonies survived) after two weeks, showing severe hypo-salinity creates more stressful conditions for Didemnum vexillum (Lenz et al., 2011). Therefore, Didemnum vexillum can tolerate a short-term severe decline in salinity but prolonged exposure over two weeks caused chronic stress and increases in mortality.

Didemnum vexillum is a temperate species that can survive a broad temperature range of -2 to 24 °C, with an upper survival limit suggested to be 25 °C (Bullard et al., 2007; Valentine et al., 2007a; Herborg et al., 2009; Kleeman, 2009; Mckenzie et al., 2017; Holt, 2024). It thrives best at 14 to 20 °C, with optimal growth temperature between 14 to 18 °C during summer months (May, June, September, October) (Gittenberger, 2007; Kleeman, 2009; Mckenzie et al., 2017). Didemnum vexillum has been recorded surviving in 4 to 15 °C in Georges Bank and 5 to 22 °C in Holyhead (Bullard et al., 2007; Valentine et al., 2007b; Griffith et al., 2009).

In New England, colonies tolerate temperatures as low as -2 °C (Bullard et al., 2007), but reports from the Netherlands show colonies “die-off” when temperatures drop below 5 °C during winter months from November to April (Gittenberger, 2007; Herborg et al., 2009). Cold winter months cause colonies to regress and reduce in size, yet they often regenerate as temperatures warm (Griffith et al., 2009; Kleeman, 2009, Mercer et al., 2009), although some populations may not survive winter at all (Dijkstra et al., 2007). Temperature changes are an important factor influencing the seasonal growth cycle and reproduction of Didemnum vexillum (Valentine et al., 2007a).

Didemnum sp. is known to affect aquaculture habitats by overgrowing shellfish and infrastructure but little is known of its effects on seabed habitats (Valentine et al., 2007b; Fletcher et al., 2013b).

Didemnum vexillum requires suitable hard substrata for successful settlement and establishment of invasive populations. It grows quickly and can establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems due to its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata, gravel, pebble, cobble or boulders (Tillin et al., 2020). The extensive mats formed by the invasive species over cobble-pebble substrata can bind or ‘glue’ small pebbles and cobbles together by filling spaces between the sediment particles, which alters the habitat complexity of the seafloor turning it into a more homogenous two-dimensional habitat rather than heterogeneous three-dimensional one (Griffith et al., 2009; Mercer et al., 2009; Lengyel et al., 2009).

Once established, Didemnum vexillum can expand rapidly, taking over most available hard substrata. Studies have hypothesized that this may alter species diversity and community composition and may decrease species abundance and biodiversity. In the Oosterschelde, Netherlands, populations of the brittlestar (Ophiotrix fragilis) declined from hundreds per m² to almost no specimens after a cold winter of 1995 to 1996, which resulted in a decrease in the abundance of many marine populations. This created large amounts of available space for Didemnum sp. colonies to expand, taking over newly available hard substrata (Gittenberger, 2007). Gittenberger (2007) stated that at this site, Didemnum sp. could cover around 95% of hard substrata, leaving little space for recruitment and growth of other species.

On Georges Bank, USA, Didemnum vexillum has altered the benthic community (Lengyel et al., 2009; Tillin et al., 2020). The pebble gravel substrata on Georges Bank is important to the success and survival of haddock (Melanogrammus aeglefinus) and Atlantic cod (Gadus morhua), and the settlement of sea scallop larvae (Placopecten magellanicus). Therefore, the invasion of Didemnum vexillum and its ability to change the habitat complexity of the seafloor, may in turn negatively impact the benthic community (Lengyel et al., 2009). In Georges Bank Lengyel et al. (2009)’s analysed photographs of the seabed and suggested that Didemnum vexillum outcompeted other epifaunal and macrofaunal species. Changes were seen in hydroids, the second most abundant epifaunal species at the location, which were overgrown by the invasive tunicate and negatively correlated with the percentage cover of Didemnum vexillum (Lengyel et al., 2009). The number of non-colonial macrofauna was also negatively related to the percentage cover of Didemnum vexillum (Lengyel et al., 2009). Dredge samples revealed clear differences in benthic species composition and revealed a significant difference in the species abundance before and after the colonization of Didemnum vexillum (Lengyel et al., 2009). Invasion of Didemnum vexillum also provided a new habitat for species not normally present, such as two polychaete species Nereis zonata and Harmothoe extenuata, changing the species composition. The increase in abundance of polychaetes Nereis zonata and Harmothoe extenuata were also seen in dredge samples collected from Georges Bank (Valentine et al., 2007b).

In contrast, some studies have suggested that potentially the overgrowth of Didemnum vexillum has little impact to benthic communities. In Long Island Sound, USA, Mercer et al. (2009) found the total abundance and richness of native epifaunal and infaunal species were either not different or significantly higher in samples taken inside Didemnum vexillum mats compared with samples collected outside the mats. While the mats did lead to subtle changes in community structure and shifts in species dominance, the authors suggested that benthic species may use Didemnum vexillum mats as a novel habitat and species living beneath the mats may use it for shelter and protection from epibenthic predators (Mercer et al., 2009). The predator protection could explain the high abundance of infaunal invertebrates found under the mats as well as the reduced abundance of crabs and demersal fish predators in areas dominated by Didemnum vexillum compared to uncolonized areas (Mercer et al., 2009). In addition, dredge samples taken from Georges Bank found 15 polychaete species and seven bivalve species living beneath the Didemnum vexillum mat (Valentine et al., 2007b). The comparisons of 85 benthic megafauna collected from dredge samples before and after Didemnum sp. became abundant in Georges Bank fishing ground showed slight changes in abundance, but changes to the invertebrate species composition were statistically marginally insignificant (Valentine et al., 2007b).

Some species have shown to tolerate overgrowth by Didemnum vexillum. Such as anemones (did not specify species name) which were observed in high densities of 10 to 339 individuals in transects with high percentage cover of Didemnum vexillum (Lengyel et al., 2009). In the Netherlands, the sea anemone Sagartia elegans and Sabella pavonia tubes were not overgrown by Didemnum sp. (Gittenberger, 2007). Botrylloides violaceus can overgrow Didemnum sp. (Gittenberger, 2007) although it was noted to be overgrown in other studies (Valentine et al., 2007a). In addition, Styela clava and Ascidiella aspera survived overgrowth by Didemnum vexillum as long as their siphons remained free (Gittenberger, 2007). However, Gittenberger (2007) stated that the boring sponge Clione celata, the sea anemone Diadumene cincta, Mytilus edulis, Magallana (syn. Crassostrea) gigas, Ostrea edulis, a variety of hydroids, the colonial ascidians Aplidium (Fig. 4) and Diplosoma listerianum and the solitary ascidians Ciona intestinalis start to die on contact with Didemnum sp.

A shift in species dominance was also seen in a long-term experiment comparing species diversity using deployed panels in New Hampshire, USA. No Didemnum vexillum was recorded between 1979 to 1982, but after invasion it became one of the most common and dominant species on the deployed panels and displaced native Mytilus edulis (Dijkstra & Harris, 2009). Coexistence was maintained as seasonal populations changed and Didemnum vexillum and other invasive sea squirts would die off, which would open up space for other species to move in (Dijkstra & Harris, 2009). The author concluded that the increase in space was facilitated by the regression of seasonally dominant Didemnum vexillum and other invasive ascidians (Dijkstra & Harris, 2009).

Didemnum vexillum mats may alter the flux of materials by creating a barrier from the water column to the sediment column, influencing the biogeochemical cycling of many nutrients. This barrier can prevent light and food from reaching the sessile community underneath it, prevents predators from feeding on the bottom and hinders larvae settlement (Mercer et al., 2009; Dijkstra, 2009 cited in Tillin et al., 2020). This has been seen in Zostera marina (Carman & Grunden, 2010; Long & Grosholz, 2015). The barrier may also influence the dissolved oxygen exchange between sediments and overlaying water, creating hypoxic conditions (Mercer et al., 2009).

Even though knowledge on the ecology of Didemnum vexillum is limited, information is available for other Didemnum species (Bullard et al., 2007). This evidence suggests that many Didemnum species have chemical defences and a highly acidic tunic (Bullard et al., 2007; Mercer et al., 2009). Toxic organic compounds found in congeneric species can affect invertebrate and vertebrate predators (e.g. fish, crabs, sea stars) that forage on or near the seafloor (Mercer et al., 2009). Therefore, the toxic characteristics of Didemnum vexillum may reduce natural predation.

Didemnum vexillum can overgrow bivalve species, such as oysters, scallops and mussels, as the hard shells can provide suitable hard substrata for settlement. It has been described as a ‘shellfish pest’ by the aquaculture industry because it is likely to completely encapsulate bivalves and smother them resulting in death or partially encapsulate and partially smother them resulting in reduced bivalve growth (Auker, 2010; Bullard et al., 2007; Coutts & Forrest, 2007, Valentine et al., 2007a; Carman et al., 2009; Kleeman, 2009; Fletcher et al., 2013b; Tillin et al., 2020). Didemnum vexillum has been recorded overgrowing mussels in Strangford Lough, Northern Ireland (Minchin & Nunn, 2013) and recorded forming large mats over Blue Mussel beds in the Gulf of Maine, completely covering individuals (Auker et al., 2014).

Sensitivity Assessment. There is no evidence of Didemnum vexillum colonization on chalk. However, it has been recorded on other soft rock substrata such as clay and limestone boulders (Hitchin, 2012). According to Tillin et al. (2020), clay exposures are potentially suitable substrata for Didemnum vexillum colonization, although this is stated with low confidence. The limestone component of this biotope is also suitable for Didemnum vexillum colonization. Didemnum vexillum prefer sheltered conditions so the moderately wave exposed conditions that characterize this biotope may mitigate its abundance. Therefore, a precautionary resistance of 'Medium' (some mortality, <25%) is suggested. Resilience is likely to be 'Very Low' as Didemnum vexillum would need to be physically removed to allow recovery. Hence, sensitivity to invasion by Didemnum vexillum is assessed as 'Medium'. However, confidence in the assessment is ‘Low’ due to the lack of direct evidence of damage macroalgae.

The Pacific oyster, Magallana gigas

Evidence

Wireweed, Sargassum muticum

Evidence

Sargassum muticum is a circumglobal invasive species (Engelen et al., 2015). It is recorded from Norway to Morocco and into the Mediterranean in the eastern Atlantic and from Alaska to Baja California in the eastern Pacific and from southern Russia to southern China in the western Pacific (Engelen et al., 2015). It colonizes a variety of habitats, can tolerate temperatures from -1° C to 30 °C, and survive salinities below 10 ppt. Although fertilization does not occur below 15 ppt and growth of germlings is limited below 10 °C, it can complete its life cycle as long as temperatures are over 8 °C for at least four months of the year (Engelen et al., 2015). Its distribution is limited by the availability of hard substratum (e.g., stones >10 cm) and light (Staehr et al., 2000; Strong & Dring 2011; Engelen et al., 2015). It is most abundant between 1 and 3 m below mean water, but it has been recorded at 18 m or 30 m in the clear waters of California. However, it is a poor competitor under low light and only develops dense canopies in shallow areas (Engelen et al., 2015). 

Sargassum muticum was shown to replace and out-compete leathery, canopy-forming macroalgae such as Saccharina latissima, Halidrys siliquosa, and Fucus spp. and, to a lesser degree, understorey species such as Codium fragile, Chondrus crispus and Dictyota dichotoma in Limfjorden, Denmark between 1984 and 1997 (Staehr et al., 2000; Engelen et al., 2015; de Bettignies et al., 2021). The invasion in Limfjorden had stabilized by 2005 although many of the native macroalgal species continued to decline (Engelen et al., 2015). In Limfjorden, the distribution of Sargassum muticum was limited to areas with hard substratum, in particular stones >10 cm in diameter, while smaller stones, gravel and sand were unsuitable. It was most abundant between 1 and 4 m in depth but had low cover at 0 to 0.5 m and 4 to 6 m, in the turbid waters of the Limfjorden. Limfjorden is wave sheltered but wave exposure has been reported to restrict the growth and survival of Sargassum muticum (Staehr et al., 2000). Viejo et al. (1995) reported that Sargassum muticum transplanted to wave exposed shores in Spain experienced >80% breakages within a month and that the growth of undamaged plants was significantly lower than that of plants on sheltered shores. Similarly, Andrew & Viejo (1998) noted that Sargassum muticum was restricted to intertidal rockpools in wave exposed sites in the Bay of Biscay. 

Strong & Dring (2011) used canopy removal experiments to investigate inter- and intraspecific competition between Sargassum muticum and Saccharina latissima in the Dorn, Strangford Lough, N. Ireland. The Dorn consists of tidal pools, very sheltered from wave action but with moderately strong tidal streams (1 to 2 knots). Sargassum muticum grew better in mixed stands with Saccharina latissima than in the highest density monospecific stands examined. However, the growth of Saccharina was not affected by the proportion of Sargassum in mixed stands. They concluded that Saccharina was not impacted significantly by the alien species while Sargassum benefited from growth in mixed stands. Experimental manipulation of subtidal algal canopies in the San Juan Islands, Washington State, USA, showed that Sargassum muticum reduced the abundance of native macroalgae, including the kelp Laminaria bongardiana due to shading. However, the experimental removal of Sargassum resulted in the recovery of native species within one year (Britton-Simmons, 2004; Engelen et al., 2015). The negative effects of Sargassum muticum on native macroalgae are mainly due to competition for light, rather than changes in nutrient availability, sedimentation or water flow (Britton-Simmons, 2004; Engelen et al., 2015).

Sensitivity Assessment. This circalittoral biotope is found is found in areas of high turbidity, low light penetration and moderate wave exposure shores. The low levels of light in this biotope make it unsuitable for macroalgae. In addition, Sargassum muticum prefers wave sheltered, shallow sites in the sublittoral fringe. No evidence was found for the effects of Sargassum muticum on the characterizing features of this biotope. Therefore, resistance is assessed as ‘High’, resilience as 'High', and sensitivity is assessed as ‘Not sensitive’.

Wakame, Undaria pinnatifida

Evidence

Undaria pinnatifida (Wakame or Asian kelp) is a large brown seaweed and an Invasive Non-Indigenous Species (INIS) that could out-compete native UK kelp species (see Farrell & Fletcher, 2006; Thompson & Schiel, 2012; Brodie et al., 2014; Heiser et al., 2014; Arnold et al., 2016; Epstein & Smale, 2017; Epstein & Smale, 2018; Kraan, 2017; Epstein et al., 2019a, b; Tidbury, 2020). Undaria pinnatifida originates from Japan but has spread to the coastlines of New Zealand, Australia, Northern France, Spain, Italy, the UK, Portugal, Belgium, Holland, Argentina, Mexico, and the USA (De Leij et al., 2017). Undaria pinnatifida was first recorded in the UK in the Hamble Estuary in 1994 (Macleod et al., 2016) and has since proliferated along UK coastlines. One year after its discovery at the Queen Anne Battery marina, Plymouth, it became a major fouling plant on pontoons (Minchin & Nunn, 2014). Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound. 

Undaria pinnatifida seems to settle better on artificial substrata (e.g., floats, marinas or piers) than on natural rocky shores among local kelps (Vaz-Pinto et al., 2014). It is found predominantly in low intertidal to shallow subtidal habitats (Epstein et al., 2019b) and is significantly more abundant on artificial substrata compared to natural rocky substrata (Heiser et al., 2014; Epstein & Smale, 2018). James (2017) suggested that Undaria pinnatifida could out-compete native species on artificial substrata (such as marinas and wharf structures).

Undaria pinnatifida behaves as a winter annual. Recruitment occurs in winter, followed by rapid spring growth, maturation in summer, and senescence by late summer, leaving only microscopic stages to persist through autumn. It exhibits multiple dispersal strategies, such as short-range spore dispersal, and long-range dispersal as whole drift plants or fragments. Undaria pinnatifida has spread rapidly across the UK and Europe, resulting in community-wide responses and impacts (Vaz-Pinto et al., 2014; Epstein & Smale, 2017). Its impacts are complex and context-specific, depending on space, time, and taxa present in the introduced location (Epstein & Smale, 2017; Teagle et al., 2017; Tidbury, 2020). 

Undaria pinnatifida has a wide physiological niche meaning it can occur in both coastal and estuarine environments showing tolerance for varying salinities, turbidity and siltation (Heiser et al., 2014; Epstein & Smale, 2018). Undaria pinnatifida has a greater preference for sites sheltered with low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018). In natural habitats, Undaria pinnatifida was not recorded if the wave fetch was greater than 642 km but increased in abundance and cover in very sheltered sites (Epstein & Smale, 2018). 

In St Malo, France, there was evidence that Undaria pinnatifida co-existed with Laminaria hyperborea under certain conditions (Castric-Fey et al., 1993). Epstein & Smale (2018) also observed that Undaria pinnatifida was relatively common (abundance of >70 individuals per 25 m transect) at three sites in Devon, UK (Jennycliff, Bovisand and Beacon Cove) where Laminaria spp. were abundant (40 to 79%) or superabundant (>80%), which suggested that Undaria pinnatifida could co-exist within refugia amongst areas with dense Laminaria spp.. 

In Plymouth Sound, UK, Heiser et al. (2014) observed that Laminaria hyperborea was significantly less abundant at sites with the presence of Undaria pinnatifida, with only ~0.5 Laminaria hyperborea individuals per m2 present compared to ~8 individuals per m2 at sites without the presence of Undaria pinnatifida. However, the results from their correlation study only showed that the species were not found together (pers. comm., Epstein, 2021). Whereas exclusion and succession experiments on reefs tell us that Laminaria spp. exclude Undaria pinnatifida, not the other way around. Epstein & Smale (2018) reported that in Devon, UK, persistent, dense, and intact Laminaria spp. canopies in rocky reef habitats exerted a strong influence over the presence/absence, abundance, and percentage cover of Undaria pinnatifida. A dense canopy of native kelp restricts the proliferation of Undaria pinnatifida and disturbance of the canopy is often the key to the recruitment of Undaria pinnatifida. Epstein et al. (2019b) reported that Undaria pinnatifida density and biomass were significantly negatively correlated with the sum of all Laminaria spp. in Plymouth, UK. The evidence indicated that native Laminaria spp. canopies in the UK inhibited Undaria pinnatifida and implied that Undaria pinnatifida was opportunistic but competitively inferior (Farrell & Fletcher, 2006; Heiser et al., 2014; Minchin & Nunn, 2014; De Leij et al., 2017; Epstein & Smale, 2018; Epstein et al., 2019b). However, Epstein et al. (2019b) also noted that Laminaria hyperborea had a non-significant positive relationship with Undaria pinnatifida due to low densities of Laminaria hyperborea across the study area, resulting in insufficient data. 

In Plymouth Sound (UK), Epstein et al. (2019b) found that within its depth range (+1 to –4 m), Undaria pinnatifida co-existed with seven species of canopy-forming brown macroalgae, including Laminaria hyperborea. De Leij et al. (2017) found that natural habitats with dense native macroalgal canopies had more resistance to Undaria pinnatifida invasion than disturbed or sparse canopies, due to limited space and light availability for Undaria recruits. However, the dense canopies will not prevent the invasion of Undaria, as sporophytes were still recorded within dense Laminaria canopies, and this suggests that canopy disturbance is not always required (De Leij et al., 2017; Epstein & Smale, 2018). 

Undaria pinnatifida was successfully eradicated on a sunken ship in Chatham Islands, New Zealand, by applying a heat treatment of 70 °C (Wotton et al., 2004). However, numerous other eradication attempts have failed and as noted by Fletcher & Farrell (1998), once established Undaria pinnatifida resists most attempts at long-term removal.

Sensitivity Assessment. This circalittoral biotope is found is found in areas of high turbidity, low light penetration and moderate wave exposure shores. The low levels of light in this biotope make it unsuitable for macroalgae. Undaria pinnatifida is found in low intertidal to shallow subtidal habitats and prefers sheltered sites with low tidal streams and low wave exposure. No evidence was found to suggest that Undaria pinnatifida poses any threat to this biotope. Therefore, resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’.

Other INIS

Evidence

The friable nature of the substratum which is subject to ongoing erosion means this biotope supports only a sparse epifauna and flora. This biotope is therefore unlikely to be invaded by sessile invasive non-indigenous species that require hard substratum. As the biotope occurs subtidally and turbidity levels are often high limiting light penetration this biotope is unlikely to provide suitable habitats for many species of invasive non-indigenous algae.  

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

The American piddock, Petricolaria pholadiformis is a non-native, boring piddock that was unintentionally introduced from America with the American oyster, Crassostrea virginica, not later than 1890 (Naylor, 1957). Rosenthal (1980) suggested that from the British Isles, the species has colonized several northern European countries by means of its pelagic larva and may also spread via driftwood, although it usually bores into clay, peat or soft rock intertidal habitats. This species is unlikely to displace Hiatella arctica in this biotope which occurs subtidally and on harder substrata.

Although not currently established in UK waters, the whelk Rapana venosa, may spread to habitats. This species has been observed predating on Pholas dactylus in the Romanian Black Sea by Micu (2007) and may pose a threat to other burrowing bivalves including Hiatella arctica.

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