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

Little direct evidence was found to assess the effects of increased temperature on piddocks, and the assessment is based on distribution records and evidence for spawning in response to temperature changes. The American piddock Petricolaria pholadiformis has a wide distribution and is found north as far as the Skaggerak, Kattegat and Limfjord (Jensen, 2010) and is also present in the Mediterranean, Gulf of Mexico and Caribbean (Huber & Gofas, 2015). Pholas dactylus occurs in the Mediterranean and the East Atlantic, from Norway to Cape Verde Islands (Micu, 2007). Barnea candida is distributed from Norway to the Mediterranean and West Africa (Gofas, 2015). Species distribution models show that the distribution of Pholas dactylus could expand northward in the next century due to ocean warming (Schultz et al., 2024).

Temperature influences the timing of reproduction in Pholas dactylus, which usually spawns between July and August. Increased summer temperatures in 1982 induced spawning in July on the south coast of England (Knight, 1984). Spawning of the piddock Petricolaria pholadiformis is initiated by increasing water temperature (>18 °C) (Duval, 1963a), so elevated temperatures outside of usual seasons may disrupt normal spawning periods. The spawning of Barnea candida was also reported to be disrupted by changes in temperature. Barnea candida normally spawns in September when temperatures are dropping (El-Maghraby, 1955). However, a rise in temperature in late June of 1956, induced spawning in some specimens of Barnea candida (Duval, 1963b). Disruption from established spawning periods, caused by temperature changes, may be detrimental to the survival of recruits as other factors influencing their survival may not be optimal, and some mortality may result. Established populations may otherwise remain unaffected by elevated temperatures.

Lüning (1990) reported that Ceramium virgatum (as Ceramium rubrum) survived temperatures from 0 to 25 °C with optimal growth at about 15 °C. According to OBIS (2025), Ceramium virgatum has been recorded in temperature ranges from 5 to 20 °C, with the most records occurring in the 10 to 15 °C range. The species is therefore likely to be tolerant of higher temperatures than it experiences in the seas around Britain and Ireland.

Gallon et al. (2014) recorded a change in red algae assemblages in Brittany, western France from 1992 to 2012, where sea surface temperatures range between 5 and 22 °C. The algal assemblage changes were linked to an increase of 0.7 °C in average sea surface temperature in that period. Among the many changes to these assemblages was a significant reduction in the frequency of Ceramium spp. observations. At a water-cooling discharge site in the Baltic Sea, the percentage cover of Ceramium tenuicorne overall decreased along a water temperature gradient from 18 to 26 °C except for in the winter where it increased slightly (Kibria, 2024). In the Galapagos Islands, Carr et al. (2018) found a significant negative relationship between temperature and the percentage cover of the red algae assemblage which included Ceramium spp. The temperature range observed was 18 to 30 °C.

Sensitivity assessment. The global distribution of the piddock species, Petricolaria pholadiformis, Pholas dactylus and Barnea candida, suggest that these species can tolerate warmer waters than currently experienced in the UK and may therefore be tolerant of a chronic increase in temperature. Short-term acute increases may, (depending on timing) interfere with spawning cues which appear to be temperature driven. The effects will depend on seasonality of occurrence and the species affected. Adult populations may be unaffected, and in such long-lived species, an unfavourable recruitment may be compensated for in a following year. The available evidence shows that in most cases, the macroalgae found within this biotope respond negatively to elevated temperatures. However, Ceramium rubrum survives a wide temperature range and is also distributed north and south of UK waters. Hence, resistance to a change in temperature is assessed as ‘High’ and recovery as ‘High’, and the biotope is considered ‘Not Sensitive’. For all characterizing species it should be noted that the timing of acute changes may lead to greater impacts. Temperature increases in the warmest months may exceed thermal tolerances whilst changes in colder periods may stress individuals acclimated to the lower temperatures.

Temperature decrease (local)

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

Evidence

Little empirical evidence was found to assess the effects of decreased temperature on piddocks and the assessment is based on distribution records and evidence for spawning in response to temperature changes. 

The American piddock Petricolaria pholadiformis has a wide distribution and is found north as far as the Skaggerak, Kattegat and Limfjord (Jensen, 2010) (Huber & Gofas, 2015). Pholas dactylus occurs in the Mediterranean and the East Atlantic, from Norway to Cape Verde Islands (Micu, 2007).   Barnea candida is distributed from Norway to the Mediterranean and West Africa (Gofas, 2015).

Temperature changes have been observed to initiate spawning by Pholas dactylus, which usually spawns between July and August. Increased summer temperatures in 1982 induced spawning in July on the south coast of England (Knight, 1984). Spawning of Petricolaria pholadiformis is initiated by increasing water temperature (>18 °C) (Duval, 1963a), so decreased temperatures may disrupt normal spawning periods where this coincides with the reproductive season. The spawning of Barnea candida was also reported to be disrupted by changes in temperature. Barnea candida normally spawns in September when temperatures are dropping (El-Maghraby, 1955). Disruption from established spawning periods, caused by decreased temperatures may be detrimental to the survival of recruits as other factors influencing their survival may not be optimal, and some mortality may result. Established populations may otherwise remain unaffected by decreased temperatures.

Lüning (1990) reported that Ceramium virgatum (as Ceramium rubrum) survived temperatures from 0 to 25 °C with optimal growth at about 15 °C. The species is therefore likely to be tolerant of lower temperatures than it experiences in the seas around Britain and Ireland. Sub-optimal temperatures may delay or slow reproduction.

Sensitivity assessment. Based on the wide range of temperature tolerance of Ceramium virgatum, it is concluded that the acute and chronic changes described by the benchmark would have limited effect.  The global distribution of the piddock species also suggest that these species can tolerate cooler waters than currently experienced in the UK and may therefore be tolerant of a chronic decrease in temperature at the benchmark level. Decreased temperatures may, depending on timing, interfere with spawning cues which appear to be temperature driven.  The effects will depend on seasonality of occurrence and the species affected. Adult populations may be unaffected and, in such long-lived species, an unfavourable recruitment may be compensated for in a following year. Based on the characterizing species, resistance to an acute and chronic decrease in temperature at the pressure benchmark is therefore assessed as ‘High’ and recovery as ‘High’ (within two years) and the biotope is considered ‘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

The biotope has only been recorded from conditions of full salinity (Connor et al., 2004; JNCC, 2015, 2022). However, intertidal biotopes will naturally experience fluctuations in salinity where evaporation increases salinity and inputs of reduces salinity. Species found in the intertidal are therefore likely to have some form of behavioural or physiological adaptations to changes in salinity. No direct empirical evidence was found to assess this pressure, and the assessment is based on the reported distribution of characterizing species. 

Pholas dactylus has been recorded in salinities ranging from 30 to 40 PSU, with most records being in the 30 to 35 PSU range (OBIS, 2025). Barnea candida is reported to extend into estuarine environments in salinities down to 20 PSU (Fish & Fish, 1996). Petricolaria pholadiformis is particularly common off the Essex and Thames estuary, e.g. the River Medway (Bamber, 1985) suggesting tolerance of brackish waters. Zenetos et al. (2009) suggest that at all sites where Petricolaria pholadiformis has been found has some freshwater inflow into the sea. According to the literature, the species in its native range inhabits environments with salinities between 29 and 35 PSU, while in the Baltic Sea it is reported from salinities 10 to 30 PSU (Gollasch & Mecke, 1996, cited from Zenetos et al. 2009). According to Castagna & Chanley (1973, cited from Zenetos et al. 2009) the lower salinity tolerance of Petricolaria pholadiformis is 7.5 to 10 PSU. It thus appears that reduced salinity facilitates its establishment (Zenetos et al., 2009).

The characteristic piddock species may be found in estuarine, reduced salinity conditions (see below) but no evidence of hypersaline resistance was found. However, the characteristic macroalgae may respond negatively to increased salinity. A study on macroalgae in saltwork ponds showed that Ceramium spp. abundance decreased with increasing salinity from 37.3 to 52.31 PSU (de Melo Soares et al., 2023). According to OBIS (2025), Ceramium virgatum occurs in salinity levels from 5 to 40 PSU, with most records being in the 25 to 30 PSU range.

Sensitivity assessment. This biotope occurs in waters with full salinity (30 TO 40 PSU) and in the intertidal zone, where tide pools form and can reach salinity levels up to 60 PSU due to evaporation (Terry et al., 2024). The evidence shows that the characteristic macroalgae of this biotope are affected by salinity levels similar to those of tide pools. There is no evidence of resistance to the pressure at the benchmark level for the characteristic piddock species. Therefore, there is Insufficient Evidence to assess the sensitivity of this biotope to this pressure at the benchmark level.

Salinity decrease (local)

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

Evidence

The biotope has only been recorded from conditions of full salinity (Connor et al., 2004; JNCC, 2015, 2022). However, intertidal biotopes will naturally experience fluctuations in salinity where evaporation increases salinity and inputs of rainwater reduces salinity. Species found in the intertidal are therefore likely to have some form of behavioural or physiological adaptations to changes in salinity. No direct empirical evidence was found to assess this pressure, and the assessment is based on the reported distribution of characterizing species. 

Barnea candida is reported to extend into estuarine environments in salinities down to 20 PSU (Fish & Fish, 1996). Petricolaria pholadiformis is particularly common off the Essex and Thames estuary, e.g. the River Medway (Bamber, 1985) suggesting tolerance of brackish waters. Zenetos et al. (2009) suggest that at all sites where Petricolaria pholadiformis has been found has some freshwater inflow into the sea. According to the literature, the species in its native range inhabits environments with salinities between 29 and 35 PSU, while in the Baltic Sea it is reported from salinities 10 to 30 PSU (Gollasch & Mecke, 1996, cited from Zenetos et al. 2009). According to Castagna & Chanley (1973, cited from Zenetos et al. 2009) the lower salinity tolerance of Petricolaria pholadiformis is 7.5 to 10 PSU. It thus appears that reduced salinity facilitates its establishment (Zenetos et al., 2009). No information was found for the salinity tolerance of Pholas dactylus.

The characterizing species Ceramium virgatum occurs over a very wide range of salinities. The species penetrates almost to the innermost part of Hardanger Fjord in Norway where it experiences very low salinity values and large salinity fluctuations due to the influence of snowmelt in spring (Jorde & Klavestad, 1963). According to OBIS (2025), Ceramium virgatum occurs in salinity levels from 5 to 40 PSU, with most records being in the 25 to 30 PSU range.

Sensitivity assessment. Based on the evidence and reported distributions of Ceramium sp. and piddocks it is considered that the benchmark decrease in salinity (from full to reduced) may not cause significant changes in the abundance of the characterising species. In areas experiencing prolonged decreases in salinity, the ratio of Petricolaria pholadiformis to other species may change due to its greater tolerance to reduced salinities, but this would not lead to re-classification of biotope. Therefore, resistance is assessed as 'High', resilience as 'High', and the biotope as 'Not Sensitive'. 

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

Established adult piddocks are, to a large extent, protected from direct effects of increased water flow, owing to their environmental position within the substratum. Increases or decreases in flow rates may affect suspension feeding by altering the delivery of suspended particles or the efficiency of filter feeding.  However, no evidence was found to inform the sensitivity assessment although other biotopes characterized by piddocks (IR.MIR.KR.Ldig.Pid and CR.MCR.SfR.Pid) have been found in areas where tidal flows vary between 0.5 -1.5 m/s (Connor et al., 2004), suggesting that changes in  flow rates within this range will not negatively impact piddocks.  Adult piddocks may become exposed should physical erosion occur at a greater rate than burrowing, and lost from the substratum. At higher densities bioerosion by piddocks may destabilise the substratum increasing vulnerability to erosion.  Increased scour, as a consequence of increased water flow could also inhibit settlement of juveniles and seaweed spores. The fronds of adults and germlings may also be damaged. Where the algal mat is dense some mitigation of water flow through friction and protection of the surface may occur although this effect will be lower than in biotopes characterized by robust species such as bivalves and fucoids and kelps.

The most damaging effect of increased flow rate would be the erosion of the peat substratum as this could eventually lead to loss of the habitat. Increased erosion would lead to the loss of habitat and removal of piddocks and the algal mat. No evidence was found to assess the water velocities at which erosion of peat occurs. Some erosion will occur naturally and storm events and wave action may be more significant in loss and damage of peats that surface water flow. Periodically peats are removed in storms to expose preserved submerged prehistoric landscapes around the UK.

Sensitivity assessment. No direct evidence was found to assess this pressure at the benchmark. Based on the exposure of piddocks in other biotopes to water flows between 0.5 and 1.5 m/s, the piddocks and algal mat are considered to be not sensitive to changes within this range as long as these do not lead to increased erosion of the substratum. Resistance is therefore assessed as 'High' and resilience as 'High' (based on no impact to recover from).

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

Adult piddocks and the algae that characterize this biotope have no mobility and cannot, therefore, migrate up or down the shore to adapt to changes in emergence. Within the peat substratum, adult piddocks will be afforded some protection from desiccation and temperature increases following increased emergence by their burrows, which will retain some moisture. The shells of piddocks do not completely enclose the animals, however , and, therefore, cannot be closed to prevent water loss. The tolerance of piddocks to increased and decreased emergence varies between species.  Pholas dactylus inhabits the shallow sub-tidal and lower shore and  Barnea candida and Petricolaria pholadiformis live slightly higher up the shore than Pholas dactylus (Duval, 1977) .  During extended periods of exposure, Pholas dactylus squirt some water from their inhalant siphon and extend their gaping siphons into the air (Knight, 1984). This may result in increased detection and predation by birds.  Changes in emergence may, therefore, alter species abundances and ratios although the biotope will remain recognisable as a piddock biotope. The algal mat covering the substratum, predominantly of the red seaweed Ceramium virgatum, may be more intolerant of an increase in desiccation. Ceramium virgatum occurs profusely in rockpools, on the lower shore and in the subtidal but not on the open shore away from damp places suggesting that it is intolerant of desiccation. As a consequence of an increase in emergence, the algal cover may become diminished.

A decrease in emergence will reduce exposure to desiccation and extremes of temperature and allow the resident  Pholas dactylus, Barnea candida, and Petricolaria pholadiformis to feed for longer periods and hence grow faster.  No information was found on factors controlling the lower limit of piddock populations and it is possible, for example, that predation (predominantly siphon nipping by gobies, and other species, Micu, 2007) may increase at the lower edge of the biotope. Competition for space with species better adapted to the changed conditions may also alter habitat suitability for this biotope.

Sensitivity assessment. The biotope occurs in the eulittoral zone, where it experiences regular immersion and emersion. Species present are therefore tolerant of periods of emergence to some extent, however, changes in emergence regime may alter habitat suitability and increase levels of predation and competition. Based on these considerations, resistance to changes in emergence is assessed as ‘Medium’ as changes may alter the upper or lower margins of the biotope. Resilience is assessed as ‘Medium’ for piddocks so that sensitivity is assessed as ‘Medium’.

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

No direct evidence was found to assess sensitivity to this pressure. The biotope typically occurs in moderately wave exposed locations (Connor et al, 2004). The piddocks are unlikely to be directly affected by changes in wave exposure, owing to their environmental position within the peat substratum, which protects them. At higher densities, bioerosion by piddocks may destabilise the substratum and increase vulnerability to erosion.

Ceramium virgatum occurs in extremely wave sheltered conditions (Connor et al., 2004) and is recorded in some of the most sheltered parts of Hardangerfjord in Norway (Jorde & Klavestad, 1963). Strong wave action could potentially cause some damage to fronds resulting in reduced photosynthesis and compromised growth. Furthermore, individuals may be damaged or dislodged by scour from sand and gravel mobilized by increased wave action (Hiscock, 1983).

However, an increase in wave height may facilitate upward expansion of biotope margins where wave splash ameliorates effects of emergence and desiccation, but this is not considered significant at the pressure benchmark.

The most damaging effect of increased wave height could be the erosion of the peat substratum as this could eventually lead to loss of the habitat and the removal of piddocks and the algal mat. No evidence was found to link significant wave height to erosion. Some erosion will occur naturally and storm events may be more significant in loss and damage of peats than changes in wave height at the pressure benchmark.

Sensitivity assessment. No direct evidence was found to assess this pressure at the benchmark. Based on the occurrence of this biotope in moderately wave exposed habitats the piddocks and algal mat are considered to have 'High' resistance to changes at the pressure benchmark where these do not lead to increased erosion of the substratum. Resilience is therefore assessed as 'High' and the biotope is considered to be 'Not sensitive', at the pressure benchmark.

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

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.

Synthetic compounds have been shown to cause oxidative stress in Ceramium virgatum (as C. rubrum), Ceramium virgatum was the most sensitive to detergent pollution, showing the highest level of antioxidant activity, up to 1.623 nmol Trolox equivalents/sample, indicating a strong stress response. In all three species, chlorophyll content also decreased with higher detergent concentrations, further suggesting physiological stress (Biris-Dorhoi et al., 2018).

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

Specific information concerning oxygen consumption and reduced oxygen tolerances were not found for important characterizing species within the biotope. Cole et al. (1999) suggested possible adverse effects on marine species below 4 mg O₂/l  and probable adverse effects below 2mg O₂/l . Duval (1963a) observed that conditions within the borings of Petricolaria pholadiformis were anaerobic and lined with a loose blue/black sludge, suggesting that the species may be relatively tolerant to conditions of reduced oxygen. However, insufficient information has been recorded. As this biotope occurs in the intertidal, emergence will mitigate the effects of hypoxic surface waters as will the exposure to wave action and water flows and this pressure is considered to be 'Not relevant'.

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. No evidence was found to assess the sensitivity of piddocks to this pressure. Hily et al. (1992) found that, in conditions of high nutrients, Ceramium virgatum (as Ceramium rubrum) and Ulva sp. dominated substrata in the Bay of Brest, France. Ceramium spp. are also mentioned by Holt et al. (1995) as likely to smother other species of macroalgae in nutrient enriched waters. Fletcher (1996) quoted Ceramium virgatum (as Ceramium rubrum) to be associated with nutrient enriched waters. It therefore seems that algal stands of Ceramium virgatum are likely to benefit from elevated levels of nutrients. Furthermore, nutrient enrichment that enhances productivity of phytoplankton may indirectly benefit the suspension feeding piddocks by increasing food supply.

Sensitivity assessment. No evidence of the effects of nutrient enrichment on the characteristic piddock species was found. Although, Ceramium spp. may increase in abundance in nutrient rich waters. Nutrient enrichment is unlikely to adversely affect the piddock community, and there is no other evidence of the effects of nutrients on the community as whole. Therefore, there is Insufficient Evidence to support an assessment.

Organic enrichment

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

Evidence

No evidence was found to assess this pressure. 

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 fossilised peat substratum, which supports populations of burrowing piddocks. A change to a sedimentary, rock or artificial substratum results in the loss of piddocks, 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 fossilised peat. If 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 on peat substratum. 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

This biotope is characterized by the fossilised peat substratum which supports populations of burrowing piddocks. A change to a sedimentary substratum would result in the loss of piddocks 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 fossilised peat; 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 on peat substratum. Although no specific evidence is described, confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.

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 30cm depth will remove the entire surface algal mat, associated biological assemblage and the piddocks, 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 fossilised peat, 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 on peat 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

Within this biotope the dense algal mat of red and green seaweed could be damaged and removed by surface abrasion.  Some species protruding from the surface, e.g. Lanice conchilega, Sabella pavonina may also be removed.  Although the piddocks are afforded some protection from surface abrasion by living in their burrows, the peat is soft which leaves many individuals, especially those near the surface of the clay, vulnerable to damage and death through exposure, sediment damage and compaction.   Micu (2007) for example observed that after storms in the Romanian Black Sea, the round goby, Neogobius melanostomus, removed clay from damaged or exposed burrows to be able to remove and eat piddocks.

The most significant impact from abrasion may be the habitat effects of removal and damage to the peat substratum. Natural erosion processes are, however, likely to be on-going within this habitat type. Where abundant the boring activities of piddocks contribute significantly to bioerosion, which can make the substratum habitat more unstable and can result in increased rates of coastal erosion (Evans 1968, Trudgill 1983, Trudgill & Crabtree, 1987).  Pinn et al. (2005) estimated that over the lifespan of a piddock (12 years), up to 41% of the shore could be eroded to a depth of 8.5 mm. The burrowing activities of piddocks may therefore weaken the substratum increasing the potential damage from substratum abrasion.

Sensitivity assessment. Surface abrasion may remove the algal mat and surface infauna and result in the loss of some piddocks and damge to habitat. Resistance is therefore assessed as ‘Low’ for the algal mat and ‘Medium’ for piddocks and substratum.  The algal mat and surface infauan are predicted to remover within 2 years, so that resilience is considered to be ‘High’ and sensitivity is ‘Low’. As the substratum  cannot recover, resilience is assessed as ‘Very Low’ and sensitivity of the overall biotope is considered to be  ‘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 will damage and remove the dense algal mat and surface fauna and could damage and expose piddocks. Piddocks 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 (Micu, 2007). 

The most significant impact may be the damage to, and removal of, the peat substratum. Where abundant the boring activities of piddocks can make the substratum habitat more unstable and can exacerbate erosion (Evans 1968, Trudgill 1983,Trudgill & Crabtree, 1987).  Pinn et al. (2005) estimated that over the lifespan of a piddock (12 years), up to 41% of the shore could be eroded to a depth of 8.5 mm. The piddock burrowing activities may therefore weaken the substratum so that it is more vulnerable to damage and erosion.

Sensitivity assessment. Sub-surface penetration and disturbance will remove and damage the algal mat and surface infauna and result in the loss of piddocks and damage to the habitat. Resistance is therefore assessed as ‘Low’ for the algal mat piddocks and substratum.  The algal mat and surface infauana are predicted to remover relatively rapidly and the piddocks within 2-10 years so that resilience of the biological assemblage 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 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

No direct evidence was found to assess this pressure. 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. Increased levels of particles may increase scour and deposition in the biotope depending on local hydrodynamic conditions. The piddocks are protected from scour within burrows and increased organic particles will provide a food subsidy.  Pholas dactylus occurs in habitats such as soft chalks where turbidity may be high and is therefore unlikely to be affected by an increase in suspended sediments at the pressure benchmark. Piddocks, in common with other suspension feeding bivalves, have efficient mechanisms to remove inorganic particles via pseudofaeces. Experimental work on Pholas dactylus showed that large particles can either be rejected immediately in the pseudofaeces or passed very quickly through the gut (Knight, 1984). Similarly Petricolaria pholadiformis is able to tolerate high-levels of suspended solids through the production of pseudofaeces (Purchon, 1955). Increased suspended sediments may impose sub-lethal  energetic costs on piddocks by reducing feeding efficiency and requiring the production of pseudofaeces with impacts on growth and reproduction.

Macroalgae within the biotope may be sensitive to decreased light penetration, however Hily et al. (1992) found that, in conditions of high turbidity, the characterizing species Ceramium virgatum (as Ceramium rubrum) (and Ulva sp.) dominated sediments in the Bay of Brest, France. It is most likely that Ceramium virgatum thrived because other species of algae could not. Whilst the field observations in the Bay of Brest suggested that an increase in abundance of Ceramium virgatum might be expected in conditions of increased turbidity, populations where light becomes limiting will be adversely affected. However, in shallow depths and the intertidal , photosynthesis can occur during low tides (as long as sediments are not deposited) and Ceramium virgatum may benefit from increased turbidity through decreased competition.

A significant decrease in suspended organic particles may reduce food input to the biotope resulting in reduced growth and fecundity of piddocks. However, local primary productivity may be enhanced where suspended sediments decrease, increasing food supply.  Decreased suspended sediment may increase macroalgal competition enhancing diversity but is considered unlikely to significantly change the character of the biotope.

Sensitivity assessment. No direct evidence was found to assess sensitivity to this pressure however, based on the tolerance of Ceramium sp. to increased turbidity and the occurrence of Pholas dactylus in turbid areas and evidence for the production of pseudofaeces by piddocks, resistance is assessed as ‘High’ and resilience as High (no impact to recover from). The biotope is therefore considered to be ‘Not sensitive’.  

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

The burrowing mechanisms of the piddocks Pholas dactylus and Barnea candida and other Pholads, mean that the burrows have a narrow entrance excavated by the juvenile. As the individual grows and excavates deeper the burrow widens resulting in a conical burrow from which the adult cannot emerge. Petricolaria pholadiformis excavates a cylindrical burrow (Ansell, 1970) and hence may be able to relocate in sandy sediments. Burrowing mechanisms have been studied but no evidence was found to suggest this species can re-emerge through sediments and re-bury. Piddocks cannot therefore emerge from layers of deposited silt as other more mobile bivalves can.

No examples of direct empirical evidence or experiments on mortality rates in response to siltation have been found for piddocks. Sometimes the substratum in which piddocks reside is covered by a thin layer of loose sandy material, through which the piddocks maintain contact with the surface via their siphons. It is likely that the piddocks would be able to extend their siphons through loose material, particularly where tidal movements shift the sand around. Pholas dactylus have been found living under layers of sand in Aberystwyth, Wales, (Knight, 1984) and in Eastbourne, with their siphons protruding at the surface (Pinn et al., 2008). Barnea candida has also been found to survive being covered by shallow layers of sand in Merseyside (Wallace & Wallace, 1983). Wallace & Wallace (1983) were unsure as to how long the Barnea candida could survive smothering but noted that, on the coast of the Wirral, the piddocks have survived smothering after periods of rough weather. Where smothering is constant, survival may be more difficult. The redistribution of loose material following storms off Whitstable Street, in the Thames Estuary, is thought to be responsible for the suffocation of many Petricolaria pholadiformis and it is possible that this species may be the most intolerant of the three piddock species associated with this biotope. However, it was not known how deep the layer of loose material was, nor how long it lasted for or what type of material it was made of.

Indirect indications for the impacts of siltation are provided by studies of Witt et al., (2004) on the impacts of harbour dredge disposal. Petricolaria (syn. Petricola) pholadiformis was absent from the disposal area, and Witt et al., (2004) cite reports by Essink (1996, not seen) that smothering of Petricolaria (as Petricola) pholadiformis from siltation could lead to mortality within a few hours. Hebda (2011) also identified that sedimentation may be one of the key threats to Barnea truncata populations. In Agigea, Romania, the smothering of clay beds by sand and finer sediments removed populations of Pholas dactylus. In this area, sand banks up to 1m thick frequently shift position driven by storm events and currents (Micu, 2007). Similar smothering was described in the case of Barnea candida populations boring into clay beds (Gomoiu & Muller 1962, cited from Micu, 2007).

Species comprising and living within the dense algal mat are likely to be intolerant of smothering. Sporelings would certainly be adversely affected as Vadas et al. (1992) stated that algal spores and propagules are adversely affected by a layer of sediment, which can exclude up to 98% of light.

Sensitivity assessment. As piddocks are essentially sedentary and as siphons are relatively short, siltation from fine sediments rather than sands, even at low levels for short periods could be lethal. Resistance to siltation is assessed as ‘Low’ for piddocks and the algal mat although effects would be mitigated where water currents and wave exposure rapidly removed the overburden and this will depend on shore height and local hydrodynamic conditions. Resilience is assessed as ‘Medium’ (2 to 10 years) for piddocks 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

The burrowing mechanisms of the piddocks Pholas dactylus and Barnea candida and other Pholads, mean that the burrows have a narrow entrance excavated by the juvenile. As the individual grows and excavates deeper the burrow widens resulting in a conical burrow from which the adult cannot emerge. Petricolaria pholadiformis excavates a cylindrical burrow (Ansell, 1970) and hence may be able to relocate in sandy sediments. Burrowing mechanisms have been studied but no evidence was found to suggest this species can re-emerge through sediments and re-bury. Piddocks cannot therefore emerge from layers of deposited silt as other more mobile bivalves can.

No examples of direct empirical evidence or experiments on mortality rates in response to siltation have been found for piddocks. Sometimes the substratum in which piddocks reside is covered by a thin layer of loose sandy material, through which the piddocks maintain contact with the surface via their siphons. It is likely that the piddocks would be able to extend their siphons through loose material, particularly where tidal movements shift the sand around. 

Indirect indications for the impacts of siltation are provided by studies of Witt et al., (2004) on the impacts of harbour dredge disposal. Petricolaria (syn. Petricola) pholadiformis was absent from the disposal area, and Witt et al., (2004) cite reports by Essink (1996, not seen) that smothering of Petricola pholadiformis from siltation could lead to mortality within a few hours. Hebda (2011) also identified that sedimentation may be one of the key threats to Barnea truncata populations. At Agigea (Micu, 2007) reported that smothering of clay beds by sand and finer sediments had removed populations of Pholas dactylus. In this area sand banks up to 1m thick frequently shift position driven by storm events and currents (Micu, 2007). Similar smothering was described in the case of Barnea candida populations boring into clay beds (Gomoiu & Muller 1962, cited from Micu, 2007).

Species comprising and living within the dense algal mat are likely to be intolerant of smothering. Sporelings would certainly be adversely affected as Vadas et al. (1992) stated that algal spores and propagules are adversely affected by a layer of sediment, which can exclude up to 98% of light.

Sensitivity assessment. Siltation at the pressure benchmark is considered to remove most or all of the piddocks and the surface algae and fauna. Resistance to siltation is therefore assessed as ‘None' although effects could be mitigated where water currents and wave exposure rapidly removed the overburden and this will depend on shore height and local hydrodynamic conditions. Resilience is assessed as ‘Medium’ (2 to 10 years) for piddocks and sensitivity is therefore assessed as ‘Medium’. 

Litter

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

Evidence

Not assessed.

Electromagnetic changes

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

Evidence

Evidence on the effect of electromagnetic fields (EMFs) on benthic organisms is still severely lacking. No studies examining the effect of EMFs on macroalgae were found. 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. No studies investigating the effect of EMFs at the population or community level for benthic organisms were found.

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

Underwater noise changes

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

Evidence

Not relevant.

Introduction of light or shading

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

Evidence

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

Sensitivity assessment. Given the rapid expansion of the evidence base but the continuing lack of data at the level of individual biotopes, resistance and resilience cannot be robustly assessed. Sensitivity is therefore recorded as Insufficient Evidence.

Barrier to species movement

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

Evidence

Not relevant.

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

Pholas dactylus reacts quickly to changes in light intensity, after a couple of seconds, by withdrawing its siphon (Knight, 1984). This reaction is ultimately an adaptation to reduce the risk of predation by, for example, approaching birds (Knight, 1984). However, its visual acuity is probably very limited and it is unlikely to be sensitive to visual disturbance. Birds are highly intolerant of visual presence and are likely to be scared away by increased human activity, therefore reducing the predation pressure on piddocks. Therefore, visual disturbance may be of indirect benefit to piddock populations and the biotope is considered to be ‘Not sensitive’. Resistance and resilience are therefore assessed as ‘High’ by default.

Genetic modification & translocation of indigenous species

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

Evidence

The species characterizing this biotop are not farmed or translocated and therefore this pressure is 'Not relevant' to this biotope.

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

Piddocks may be removed as bait and across Europe they have traditionally been harvested for food, however high levels of habitat damage are associated with the removal of boring molluscs (Fanelli et al., 1994) and this practice has largely been banned. The most sensitive component of this biotope to targeted harvesting is the peat substratum which may be damaged and removed if piddocks were excavated from their burrows, this effect is considered through the physical damage pressures, abrasion and penetration and sub-surface damage.

Sensitivity assessment. Removal of piddocks will result in loss of targeted individuals and damage to the habitat. Resistance is  assessed as ‘Low’ as piddocks are sedentary and burrow openings are readily detected.  Piddocks are predicted to recover within 2-10 years so that resilience is considered to ‘Medium’ and sensitivity is ‘Medium’. Resistance of the habitat to removal of substratum to extract sediments is assessed as 'Low' and resilience as 'Very Low' based on no recovery, so that sensitivity is 'High'. The more sensitive habitat assessment and associated confidence is presented in the table.

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

Surface algal mats and infauna 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 algal mat may alter the character of the biotope but it is unlikely that targeted harvesting of other species would remove all of the mat or unintentionally remove piddocks. Resistance of surface fauna and flora is  assessed as 'Medium' and resilience as 'High' so that sensitivity is assessed as 'Low' (based on the algal mat). 

The American slipper limpet, Crepidula fornicata

Evidence

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and the Essex coast in 1887-1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 1999, 2018; Hinz et al., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). It ranges from the Baltic Sea, the Kattegat and Skagerrak, the North Sea coasts of the UK, Germany, and Belgium, through the English Channels and into the Irish sea coasts of Ireland and south Wales with records in east and west Scotland, Northern Ireland, northwest France, Spain and south into the Mediterranean (NBN, 2024; OBIS, 2025).

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

Its density and ability to spread within and between sites (e.g., Bays) depends on the availability of suitable habitat, completion with other species, larval retention with the site, human activity (e.g., dredging) and summer and winter temperatures (especially in the intertidal). For example, the Crepidula fornicata population in the Bay of Mont-Saint-Michel grew by 50% between 1996 and 2004 and covered 25% at a high density (51 to 100% cover) aided by local oyster farming and shellfish dredging (Blanchard, 2009). However, in Arcachon Bay, France, Crepidula fornicata was limited to only 155 tonnes in 1999 and 312 tonnes in 2011 (De Montaudouin et al., 2001, 2018). Crepidula was limited to muddy sediments that were only ~8% of the bay and were colonized by Zostera beds and represented only 0.4% of suspension feeder biomass of the oysters Magallana gigas in the bay and did not show signs of increasing biomass at a 12-year scale. In addition, benthic trawling was prohibited in the bay (De Montaudouin et al., 2001, 2018). As a result, De Montaudouin et al. (2018) concluded that Crepidula was not invasive in the Bay of Arcachon.

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

For example, 75% to 98% of Crepidula larvae settled on dead Crepidula shells, in the eastern Solent harbours of Portsmouth, Langstone, and Chichester, while ~4% settled on stone, 2.5% on live Crepidula, 0.3% oyster shell, 0.6% cockle shell, 0.3% winkle shell and 0.1% perwinkle shell (Preston et al., 2020). However, in the Milford Harbour Waterway, the highest densities of Crepidula were found in areas of sediment with hard substrata (e.g., mixed fine sediment with shell, gravel, or both). While Crepidula density increased with increasing gravel cover in the subtidal zone, the opposite pattern was observed in the intertidal zone (Bohn et al., 2015). Gravel formed the base of most stacks of Crepidula in the intertidal, which suggested that initial colonization occurred on available hard substrata (i.e., gravel) in the absence of adult shells of Crepidula. The availability of hard substrata (e.g., gravel) may only restrict initial colonization as higher densities of Crepidula functions as substrata for subsequent colonization (Thieltges et al., 2004; Blanchard, 2009). Bohn et al. (2015) also noted that Crepidula density was low in areas of homogenous fine sediment and absent in areas dominated by boulders. Bohn et al. (2015) suggested that wave action (exposure) probably prevented the establishment of large numbers of Crepidula in high-energy areas. However, Hinz et al. (2011) recorded Crepidula off the Isle of Wight in the English Channel, at ~60 m on rough ground in areas of high tidal flow. Tillin et al. (2020) suggested that the effect of oscillatory wave meditated flow might have a greater effect on Crepidula than tidal flow, presumably due to mobilization of the substratum. Similarly, Crepidula was absent from sandy substrata in Swansea Bay but was most abundant in the shelter of the breakwater at Swansea east site (Powell-Jennings & Calloway, 2018).

It is recorded from the lower intertidal to ~160 m in depth but it most common in the shallow subtidal and low water springs (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 2025; Tillin et al., 2020). Bohn et al. (2012, 2013a, 2013b, 2015) suggested that extreme conditions in intertidal limited its upward distribution due to early post-settlement mortality. It reached its highest densities in the lower shore (below ~0.7 m) and was absent from high tidal level (~1.8 m) in the Milford Harbour Waterway (Bohn et al., 2015). Bohn et al. (2013b) noted that Crepidula spat in their experimental intertidal panels suffered high mortality 78 to 100% during emersion by low water spring tides. Thieltges et al. (2003) noted that Crepidula abundance at the intertidal to subtidal transition zone (~21 /m²) was significantly higher than in the upper, mid, and lower intertidal <3 /m²). Similarly, Diederich & Pechenik (2013) noted that Crepidula densities were not significantly different in the low intertidal (+0.2 m) and shallow subtidal (-1 m) but became lower at +0.4 and were absent above +0.6 m in Bissel Cove, Rhode Island where the mean high water was +1.38 m. They reported that intertidal adults experienced temperatures of ~42 °C, which were 15 °C higher than subtidal adults. However, there was no significant difference in the tolerance of subtidal and intertidal adults with a lethal range of 33 to 37 °C after 3 hours in the laboratory. Diederich & Pechenik, (2013) suggested that adult Crepidula were living close to their upper thermal limit in Rhode Island and would be driven into the subtidal due to climate change. Diederich et al. (2015) reported that most juvenile Crepidula died after aerial exposure under laboratory conditions (20 °C, 75% relative humidity), while adults from the intertidal and subtidal survived (26 °C, 75% relative humidity). Franklin et al. (2023) noted that the body mass index of adult Crepidula did not decrease significantly in winter months in New Hampshire, USA, but did decrease in spring and summer, probably due to its investment in reproduction.

The density of Crepidula populations in the northern Europe (Germany, Denmark, and Norway) are significantly lower (<100 /m²) than in southern waters. Thieltges et al. (2004) reported that the population of Crepidula was affected strongly by cold winters in the Wadden Sea. The winters of 2001 and 2003 resulted in ~56 to 64% mortality of intertidal Crepidula and up to 97% on one mussel bed, compared to only 11 to 14% in southern areas without frost. Crepidula almost vanished from the Wadden Sea after the 1978/79 winter and took ten years to recover due to moderate winters which regularly affected the population. Similarly, 25% mortality was observed in Crepidula populations on the south coast of the UK after the extreme 1962/63 winter (Crisp, 1964, Bohn et al., 2012). Thieltges et al. (2003) suggested that global warming may allow Crepidula populations become more abundant in northern Europe. Valdizan et al. (2011) noted higher water temperatures between 2000 to 2001 and 2006 to 2007 together with elevated chlorophyll-a corresponded to an increase in gametogenesis and the duration of broods in Crepidula population in Bournerf Bay, France. They suggested that rising temperatures in northern Europe could increase its reproductive success due favourable breeding temperatures and increased phytoplankton (Valdizan et al., 2011). Nehls et al. (2006) noted that the decline in mussel (Mytilus edulis) beds in the Wadden Sea was due to mild winters that favoured non-native oysters (Magellana gigas) and slipper limpets, which co-existed with the mussels.

Crepidula fornicata has one or two reproductive periods per year (depending on location), is highly fecund, and has long-lived pelagic larvae. Hence, dispersal is potentially high. However, Bohn et al. (2012, 2013a, 2013b, 2015) suggested that lack of suitable habitat rather than larval supply, together with local hydrography may limit the northward spread of Crepidula from Milford Harbour Waterway, and that post-settlement mortality is particularly important in the intertidal. Dupont et al. (2007) reported genetic isolation with distance along the English Channel but a high degree of genetic connectivity between the bays of northern France, which were consistent with hydrographic models of larval transport. They noted marked genetic isolation of the population in the semi-enclosed Bay of Brest. Dupont et al. (2007) suggested that Crepidula populations were isolated by hydrographic barriers over distances of ~100 km. Riel et al. (2009) noted that larval supply was low in the Bay of Mont Saint-Michel partly due to larval mortality and larval export out of the bay, although recruitment was still adequate to maintain the population. Bohn et al. (2012) suggested that homogenous sediments and boulders at the entrance to the Milford Harbour Waterway formed a barrier to dispersal and, together with high larval export probably explained the slow of northward expansion of Crepidula along the Welsh coast. Nevertheless, the initial spread of Crepidula was facilitated by human activities such as shipping, shellfish culture (e.g. oysters and mussels), ballast water (Blanchard, 1997) and fisheries (e.g., dredging) (Blanchard, 1997, 2009; De Montaudouin et al., 2018; Kostecki et al., 2011; McNeill et al., 2010; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015).

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

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

Sensitivity Assessment. No evidence was found in the literature to suggest that invasive non-indigenous species are present in UK marine and coastal peat habitats. According to Tillin et al. (2020), peat and clay exposures are unsuitable for Crepidula fornicata settlement, although this is stated with low confidence. Crepidula fornicata requires hard substrate for colonisation. The features of this biotope, i.e. fossilised peat with a Ceramium sp. turf are most likely unsuitable for Crepidula fornicata colonisation. Therefore, resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’.

The carpet sea squirt, Didemnum vexillum

Evidence

The Pacific oyster, Magallana gigas

Evidence

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

Since introduction, the species has invaded and established self-sustaining natural populations throughout Europe from the North Sea, Wadden Sea and Scandinavian coastlines to the Atlantic coastlines of Spain and Portugal, as well as the Mediterranean and Adriatic Sea (Wrange et al., 2010; GBNNSIP, 2011, 2012; Ezgeta-Balic et al., 2019; Spagnolo et al., 2019; Bergstrom et al., 2021; Hansen et al., 2023). In the UK, the species predominantly occurs around the southern and western coastlines (OBIS, 2025; NBN, 2024).

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

Magallana gigas has a high fecundity, a long-lived pelagic larval phase (2 to 4 weeks) and can produce up to 200 million eggs during spawning (Herbert et al., 2012, 2016; Alves et al., 2021; Wood et al., 2021; Hansen et al., 2023). Hence, as a broadcast spawner, it has a high dispersal potential of more than 1000 km (Padilla, 2010; Wood et al., 2021). Larval mortality can be as large as 99%, as larvae are sensitive to environmental conditions (Alves et al., 2021). But adults are long-lived so that populations can survive with infrequent recruitment (Padilla, 2010). Larval dispersal and mass spawning events have facilitated the settlement and establishment of Pacific oysters, as seen in the Oosterschelde estuary, Netherlands (Hansen et al., 2023). It has been suggested that the spread of the Pacific oyster in Scandinavia is due to northward larval drift on tidal and wind-driven currents (Hansen et al., 2023). Wood et al. (2021) suggested that larval dispersal of the Pacific oyster from populations within and outside the UK was possible via unaided (passive) transport by currents, but that aquaculture and offshore structures (e.g. windfarms) increased the risk of the invasive species spreading and the geographical extent of spread.

Magallana gigas is an ecosystem engineer and can dramatically change habitat structure when it invades. Once successfully settled, groups of Pacific oysters may form dense aggregations, potentially forming a reef, which in some regions can reach densities of 700 individuals/m² (Herbert et al., 2012, 2016). Once, the density of live or dead Pacific oysters reaches or exceeds 200 ind./m², little of the underlying substratum remains visible (Herbert et al., 2016). These reefs can stabilize the sediment surface locally (Troost, 2010). When such reefs are formed or, particularly when the species colonizes soft sediments such as mud or sand, it can change and affect local communities, by creating hard substrata for mobile species, which might not otherwise be present before the invasion (Padilla, 2010). However, Hansen et al. (2023) suggested that no immediate ecosystem risk is observed where the Pacific oyster occurs sporadically.

Magallana gigas also colonizes littoral intertidal biogenic reefs formed by the blue mussel Mytilus edulis or honeycomb worm Sabellaria alveolata (GBNNSIP, 2011, 2012; Kochmann, 2012; Kochmann et al., 2013; Herbert et al., 2016; Tillin et al., 2020). Evidence suggests the Pacific oyster can out-compete Mytilus edulis, particularly for food and space, as the faster growth rates of the oyster make it more competitive when food or space is limiting (Nehls et al., 2006; Padilla, 2010; Tillin et al., 2020; Joyce et al., 2021). The invasion of Magallana gigas may alter the structure and function of these intertidal reefs but can create a multi-layered structure of a mixture of oysters and blue mussels that is more resilient and accumulates a higher biodiversity of flora and fauna and supports the densities of other native species such as Littorina littorea (Reise et al., 2017; Andriana et al., 2020; Cornelius & Buschbaum, 2020; Hansen et al., 2023).

Magallana gigas requires hard substrata for successful settlement and establishment, including littoral rock, bedrock, chalk, bare boulders, cobbles and pebbles and shells (Kochmann, 2012; Kochmann et al., 2013; McKinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). It also prefers mudflats with mixed sediment composed of shingle and sand, attaching to whatever hard substrata are available within otherwise unsuitable fine muddy sediment (Spencer et al., 1994; McKinstry & Jensen, 2013; Tillin et al., 2020). Invasive populations of Magallana gigas has been found wave-exposed rocky shores to wave-sheltered soft sediment environments and it has been described as a habitat generalist (Troost, 2010; Kochmann, 2012; Kochmann et al., 2013). For example, in Scotland, wild Magallana gigas are mainly located in the lower intertidal on bedrock, bedrock encrusted with barnacles, within bedrock crevices, and large and small boulders (Cook et al., 2014). They are unlikely to occur under boulders as they require access to the water column (Tillin et al., 2020). Patches of Pacific oyster reefs have been recorded on littoral rock in Kent, southern England and on littoral sediments in southern England, the North Sea, and the English Channel (Herbert et al., 2012, 2016; Morgan et al., 2021).

Magallana gigas has been reported from estuaries growing on intertidal mudflats and sandflats, and other soft sediments (Padilla, 2010; Herbert et al., 2016; Cabral et al., 2020). The settlement of spat on hard substrata within sediments has been observed in the estuaries of the River Dart, Exe, Fal, Fowey, Tamar, Teign, and Yealm in Devon and Cornwall, the Menai Straits, Wales and large estuaries of Lough Swilly, Lough Foyle and the Shannon in Ireland, and the Tagus Estuary in Portugal (Spencer et al., 1994; Kochmann, 2012; Kochmann et al., 2013; Cabral et al., 2020). In Lough Swilly, Lough Foyle and the Shannon, the Pacific oyster was often associated with intertidal mud or sandflats (Kochmann et al., 2013). In contrast, the Pacific oysters were absent from sandflat areas in Poole Harbour (McKinstry & Jensens, 2013).

Although shorelines comprised of mainly mud were suggested to be unsuitable for spat settlement (Spencer et al., 1994), the presence of smaller hard substrata, such as shells or pebbles, can enable larvae to settle (Tillin et al., 2020). For example, in the River Teign estuary, Pacific oyster settlement was observed on shell-covered ground mainly attached to mussel shells, and occasionally attached to cockles, stones and common periwinkle (Littorina littorea) shells on a mud flat in the estuarine intertidal zone otherwise mainly comprised of sand and mud (Spencer et al., 1994). In addition, the Blue Lagoon on the north shore of Poole Harbour had the highest abundance of oysters on mud mixed with shingle and shell (McKinstry & Jensen, 2013). Outside of the Blue Lagoon, oysters were also recorded on mixed substrata composed of mud, gravel, and shell (McKinstry & Jensen, 2013). Tillin et al. (2020) concluded that while successful invasions occurred on mudflats, Magallana gigas prefers mixed substrata. Fine mud sediments without hard substrata (such as small stones, gravel, and shell) are unlikely to be suitable (Tillin et al., 2020). The speed of Magallana gigas reef formation on soft substrata seems to be dependent on the amount of hard substrata present, developing quicker once there is a sufficient amount (Troost, 2010). Bergstrom et al. (2021) reported that the presence of Magallana gigas was partially dependent on increasing gravel content up to 15% but remained stable with increasing percentages (measured up to 80%).

Pacific oysters prefer shallow gradients and, hence, extensive intertidal shores. In Poole Harbour, the species was recorded on shallow gradient shores in a relatively expansive intertidal area but absent from steep gradients (McKinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). Similarly, in some Irish estuaries, oysters had a greater presence on intertidal shores wider than 50 m and in the Solway Firth all Pacific oysters were found in areas with intertidal widths greater than 40 m (Kochmann et al., 2013; Cook et al., 2014; Tillin et al., 2020).

Dense macroalgal cover is unsuitable for the Magallana gigas (Herbert et al., 2012, 2016; Tillin et al., 2020), being rarely found under macroalgal cover in Northern Ireland, absent from exposed bedrock or large boulders with macroalgae cover in the Solway Firth, Scotland, and absent in Poole Harbour where there was competition with macroalgae (Kochmann, 2012; Kochmann et al., 2013; McKinstry & Jensen, 2013; Cook et al., 2014; Tillin et al., 2020). Fucus cover significantly reduced larval recruitment of the Pacific oyster in the Wadden Sea (Diederich, 2005). Hence, the Pacific oyster is more likely to colonize bare rock, boulders, or mussel beds without macroalgae (Diederich, 2005; Cook et al., 2014). Kochmann et al. (2013) suggested that macrophyte canopies prevent larvae from settling on the rocks underneath and macroalgal fronds inhibit settlement and recruitment by exuding metabolites.

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

On littoral rock in Brittany, the Pacific oyster colonizes all intertidal levels from Mean High Water to Mean Low Water on sheltered (low energy), moderately exposed (moderate energy) and high energy rock shores (Herbert et al., 2012). However, in the northwest Pacific, Magallana gigas is commonly found on sheltered low energy littoral rock and has less than 10% cover on exposed high energy littoral rock shores (Herbert et al., 2012, 2016). Magallana gigas was recorded in the mid intertidal on hard rock and artificial harbour structures in the northern and eastern Adriatic but absent from subtidal beam trawl surveys off the coast of Istria (Ezgeta-Balic et al., 2019). Ezgeta-Balic et al. (2019) noted that the only oyster found in their trawls were large Ostrea edulis, often confused for Magallana gigas by local fishers. Magallana gigas has not been found at extreme low water levels or subtidally beneath rocky habitats, as it has been in soft sediment areas (Herbert et al., 2012).

The oyster reefs, in the Wadden Sea and Brittany, on littoral muddy and sandy habitats formed predominantly at lower tidal levels from Mean Low Water levels to the shallow subtidal (Herbert et al., 2012, 2016). Pacific oyster spatfall was recorded in the estuarine intertidal zone on areas with hard substrata of stone and shell, particularly between the low water of spring tides and high-water of neap tides, such as in the Menai Strait (Spencer et al., 1994). In Lim Bay, Adriatic Sea, Magallana gigas is only found in the intertidal and on the sublittoral edge (at a depth of 1 m) and not at 3 m or 6 m depth (Stagličić et al., 2020; Tillin et al., 2020). It coexists here with Ostrea edulis which is abundant in the subtidal (Stagličić et al., 2020). Bergstrom et al. (2021) found that depth was one of the most important predictors of occurrence of Magallana gigas in the Skagerrak and suggested the optimal depth of the species was 0.5 m in the shallow subtidal, although it occurred down to 5 m.

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

Temperature and salinity affect spawning and recruitment of Magallana gigas populations. While Pacific oyster larvae are vulnerable to environmental change and less adaptable, it has been suggested that Magallana gigas adults and established populations are more resilient (GBNNSIP, 2011, 2012; Hansen et al., 2023). The broad geographical spread of Magallana gigas indicates the invasive species has a wide environmental tolerance.

The Pacific oyster can withstand a wide range of salinities (from 11 to 34 PSU) but no oysters were observed in areas which had salinities less than 20 PSU and most abundant populations occur in salinities above 20 PSU on the Swedish west coastline (Wrange et al., 2010; Kochmann, 2012; Chu et al., 1996 cited in Tillin et al., 2020). Bergstrom et al. (2021) noted that in the Skagerrak, Sweden native and Pacific oyster densities increased with rising salinity above 15 to 21 PSU up to the full range measured (27 PSU). Larvae can survive salinities between 19 to 35 PSU (Troost, 2010; Tillin et al., 2020). Kochmann (2012) reported 11 to 35 PSU as the optimal salinity range for Magallana gigas (cited in Wood et al. (2021). Growth of Pacific oysters can occur between 10 to 30 PSU (Troost, 2010).

Carrasco & Baron (2010) suggested that Magallana gigas has successfully adapted to colonize a range of thermal niches. Temperature is important for the life cycle of the Pacific oyster and influences the establishment of feral and wild populations (Alves et al., 2021). Within its native range, Magallana gigas occurs in areas where the sea surface temperature ranges from 14.0 °C and 28.6 °C in the warmest month of the year, and between -1.9 °C and 19.8 °C in the coldest month (Carrasco & Baron, 2010).

Magallana gigas has a seasonal reproductive cycle (Alves et al., 2021). Spawning occurs in the summer months, when temperatures are 16 to 34 °C and larvae require a water temperature of 18 °C or above for successful development (Mann 1979; Troost, 2010; Kochmann, 2012; Ezgeta-Balic et al., 2020; Alves & Tidbury, 2022). In Poole, UK, spawning temperatures were estimated at 19.7 °C (Alves & Tidbury, 2022). Ezgeta-Balic et al.‘s (2020) study indicated that temperatures in the Mediterranean and the Adriatic were favourable for Pacific oyster larval development, with gametogenesis initiated at temperatures from around 10 to 15 °C and spawning initiated at around 24 °C. However, the lower thermal limit for spawning was recognized as 16 °C (Carrasco & Baron, 2010) and once settled, larvae are unable to survive in temperatures below 3 °C (Alves & Tidbury, 2022).

Adults can survive in water temperatures up to 40 °C and at low tide, freezing air temperatures as low as -17 °C, depending on the salinity of the water in their shells (Troost, 2010; Tillin et al., 2020; Hansen et al., 2023). Growth of Pacific oysters occurs between 3 to 40 °C (Troost, 2010; Kochmann, 2012).

Increasing temperatures are associated with the spread of Pacific oysters in Europe (Diederich et al. 2005, Kochmann et al., 2013; Herbert et al., 2016; Pack et al., 2021; Alves & Tidbury, 2022). The decision to introduce Magallana gigas in Europe initially was based on the prediction that the lower seawater temperatures in Europe would reduce the risk of spreading by the Pacific oyster to natural neighbouring habitats, as predicted temperatures were lower than required for successful reproduction. However, an increase in mean seawater temperature allowed successful reproduction and increased the frequency of spawning events that led to the established populations in the Wadden Sea, margins of the Skagerrak and the Atlantic coast off Norway (Wrange et al., 2010; Carrasco & Baron, 2010, Alves et al., 2021).

The evidence suggests that invasion risks of Magallana gigas are likely to increase due to temperature increases associated with climate change (Alves & Tidbury, 2022; Glamuzina et al., 2024). Glamuzina et al. (2024) identified a high risk of invasion by Magallana gigas in the Mediterranean Sea, under IPCC climate change predicted scenarios. Reproduction and larval success are improved at warmer summer temperatures, so recent warming trends due to climate change may increase spawning frequency, recruitment, and settlement, furthering the spread of this invasive species, particularly to more northern colder regions such as Scotland, Denmark, and Norway (King et al., 2021; Alves & Tidbury, 2022). King et al. (2021) predicted a progressive northward expansion of Magallana gigas within the northwest European shelf by the end of the century under IPCC RCP 8.5 scenario, as the majority of the coastlines would be within the species’ thermal recruitment niche.

In the Bay of Brest, Pacific oyster reefs on rock had a greater diversity, species richness and biomass than the surrounding bare rock habitats (Lejart & Hily, 2011). There was an increase in macrograzers, suspension feeders, carnivores, deposit feeders and detritivores in the present on oyster reefs on rock compared with the surrounding bare rock. Lejart & Hily (2011) found that 15% of species present in the oyster reefs on rock were characteristic of mud habitats. In addition, they the surface available for epibenthic species in the Bay of Brest, increased 4-fold when oysters were present on rock, for every 1 m² of colonized substrata the oyster reef added 3.97 m² of surface area on rock. An increase in available settlement substrata, which is clean free of epibiota, could be the reason oyster reefs cause an increase in the macrofaunal abundance. Zwerschke et al. (2018) found at intertidal rocky sites and sites with gravel around the UK, Ireland and northern France, densities of Pacific oysters more than 10 m² had a different macrofaunal assemblage structure than sites with low density or no Magallana gigas. Their results showed a greater abundance of species such as barnacles in mud, rock, and gravel sites when Pacific oysters were superabundant (oyster density more than 99 /m²). However, a decrease in abundance of kelp, Fucus vesiculosus and periwinkle Littorina sp. was observed on the rocky shore sites colonized by the oysters (Zwerschke et al., 2018). In addition, settlement of Magallana gigas in the barnacle zone of exposed rocky shores in the Strait of Georgia, Canada provided a greater surface area for settlement while neighbouring species at the rocky sites facilitated the survival of the Pacific oyster, by reducing predation and physical stress (Ruesink et al., 2005; Herbert et al., 2012).

Similarly, in rocky habitats, in Argentina, four epifaunal species (crabs Cyrtograpsus angulatus, Chasmagnathus granulatus, isopod Melita palmata and snail Helebia australis) showed higher densities and abundance within Magallana gigas beds than outside these areas (Escapa et al., 2004; Herbert et al., 2012).

Magallana gigas is a trophic competitor of other bivalves and other filter feeders (Decottignies et al., 2007 cited in Tillin et al., 2020), likely to compete with native species including native oyster and filter feeders such as Sabellaria alveolata (Cognie et al., 2006; Tillin et al., 2020). However, evidence has suggested Magallana gigas and some native species coexist, often forming more diverse reefs and habitats (e.g. Mytilus edulis and Ostrea edulis). For example, all sites studied in the Skagerrak area, Sweden colonized by Magallana gigas contained thriving populations of native oyster Ostrea edulis (Bergstrom et al., 2021) and there is no spatial competition identified between native Ostrea edulis and the Pacific oyster in the Northern Adriatic Sea, although densities of the Pacific oyster were significantly higher (Stagličić et al., 2020). In Balgzand, Wadden Sea the impact on the food web and the biomass of Magallana gigas remained low (Jung et al., 2020).

The global spread of the Pacific oyster has facilitated the introduction of macrospecies, microparasites associated with oysters, including harmful algae and disease agents (Padilla, 2010). It is recognised that copepod parasites of Magallana gigas, Mytilicola orientalis and Myicola ostreae were introduced with imports of the oyster from France to Ireland (Tillin et al., 2020). Mytilicola orientalis was introduced into the Wadden Sea by Magallana gigas and infected blue mussels (Goedknegt et al., 2020). Predator avoidance by blue mussels in biogenic oyster reefs can indirectly affect parasite-host interactions. For example, in the Wadden Sea, one mixed mussel and oyster reef had significantly higher abundance of parasitic Mytilicola spp. in mussels at the top of the reef compared to at the bottom (Goedknegt et al., 2020). In contrast, with increasing oyster density, an increase in the presence of the trematode Renicola roscovita was seen in mussels (Goedknegt et al., 2019). Magallana gigas is also the predominant host of the shell-boring parasites Polydora ciliata and Polydora websteri in the Wadden Sea, with relatively higher densities of Polydora ciliata found in the Pacific oyster compared to the blue mussels (Waser et al., 2021).

Sensitivity Assessment. No evidence was found in the literature to suggest that invasive non-indigenous species are present in UK marine and coastal peat habitats. Tillin et al. (2020) believe that peat and clay exposures are potentially suitable for Magallana gigas, although this is stated with medium confidence. However, this biotope is found on moderately wave exposed shores, while Magallana gigas settlement is more successful on wave sheltered shores. In addition, the macroalgal feature of this biotope makes the Magallana gigas invasion less likely (Herbert et al., 2012, 2016; Tillin et al., 2020). Based on the evidence, resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’ with medium confidence.

Wireweed, Sargassum muticum

Evidence

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 is established currently on 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). In Plymouth, UK, De Leij et al. (2017) found that natural habitats with dense native macroalgal canopies, such as Laminaria hyperborea, Laminaria ochroleuca, Laminaria digitata and Saccharina latissima had more resistance to Undaria pinnatifida invasion than disturbed or sparse canopies, due to limited space and light availability for Undaria pinnatifida recruits. However, the dense canopies did not always prevent the invasion of Undaria pinnatifida as sporophytes were still recorded within dense Laminaria canopies, so canopy disturbance was not always required (De Leij et al., 2017; Epstein & Smale, 2018).

Undaria behaves as a winter annual, and recruitment occurs in winter followed by rapid growth through spring, maturity and then senescence through summer, with only the microscopic life stages persisting 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). 

Sensitivity Assessment. It is unlikely that fossilised peat would provide the same attachment stability as rocky or artificial substrata which favour Undaria pinnatifida. In addition, the Ceramium sp. turf would most likely inhibit the settlement of Undaria pinnatifida. Lastly, Undaria pinnatifida prefers low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018; Epstein et al., 2019a). It is therefore unlikely that Undaria pinnatifida poses a threat to this biotope. Resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’.

Other INIS

Evidence

The surface of the peat is friable and subject to on-going erosion and therefore this habitat does not support large, long-lived, attached species. The algal mat characterizing this biotope consists of small, ephemeral species, replacement of these species by invasive non-indigenous algal species could occur and alter the character of the biotope. However, no evidence was found in the literature to suggest that invasive non-indigenous species are present in UK marine and coastal peat habitats.

The presence of 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 shores. In Belgium and The Netherlands Petricolaria pholadiformis  almost completely displaced the native piddock, Barnea candida (ICES, 1972). However, this has not been observed elsewhere, and later studies have found that Barnea candida is now more common than Petricolaria pholadiformis in Belgium (Wouters, 1993) and there is no documentary evidence to suggest that Barnea candida has been displaced in the British Isles (J. Light & I. Kileen pers. comm. to Eno et al., 1997). Petricolaria pholadiformis is considered unlikely to displace Pholas dactylus which is more likely to occur subtidally. Should Petricolaria pholadiformis is be present in this biotope it is not considered to alter the character or ecological function of the biotope.

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).

Sensitivity assessment. Based on the lack of records of invasive non-indigenous species in this biotope, and the unsuitability of the habitat for algae and other attached epifauna this biotope is considered to have ‘High’ resistance to this pressure and, by default ‘High’ resilience, this biotope is therefore considered to be ‘Not sensitive’. This assessment may need revising in light of future invasions, e.g. the introduction of the whelk Rapana venosa.