The Marine Life Information Network

Global warming (extreme)

Extreme emission scenario (by the end of this century 2081-2100) benchmark of:

• A 5°C rise in SST and NBT (coastal to the shelf seas),

• A 6°C rise in surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf, and

• A 5°C rise in surface air temperature in intertidal habitats exclusive to Scotland (Global warming pressure definitions).

Evidence

Little empirical 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. More extensive evidence on thermal tolerance and physiological effects was found for Mytilus edulis.

The American piddock Petricolaria pholadiformis is a cold-temperate species, which originates from the east coast of America, distributed from the Gulf of St Lawrence to the Caribbean (www.obis.org). From there it was unintentionally introduced into southern England with Crassostrea oysters, and from the UK, this species has colonized several northern European countries (Zenetos et al., 2009). It has since established a small population in the Saranikonos Gulf, in the Eastern Mediterranean (Zenetos et al., 2009). Pholis dactylus inhabits the mid-littoral and shallow sublittoral the Mediterranean and the East Atlantic, from Norway to the 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 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).

Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas that frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992). In the north Atlantic, this species occurs from Norway to the coast of Spain. In the western Atlantic, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits have contracted approximately 350 km north of its previous southern limits in Cape Hatteras, North Carolina, due to increases in water temperature beyond the lethal limit (Jones et al., 2009).

Wells & Gray (1960) suggested that the mean summer water temperatures of 26.6 °C set the southern range limit. Gonzalez & Yevich (1976) found that Mytilus edulis could not tolerate sustained temperatures of 27°C, and feeding stopped after 25°C. Pearce (1969) found that whilst some populations of Mytilus edulis could survive 27°C for over a month, none of these populations could survive temperatures of 28°C for more than four days. Read & Cummings (1967) estimated the upper tolerance limit to be 27°C, and Chapple et al. (1998) found that Mytilus edulis could not acclimate to temperatures above 28.5°C. Almada-Villela et al. (1982) found that growth significantly declined in juvenile Mytilus edulis as temperatures increased above 20°C. Similarly, Hiebenthal et al. (2013) found that the growth rate decreased by 60% as temperatures increased from 20 - 25°C and resulted in 25% mortality under experimental conditions. Incze et al. (1980) found that Mytilus edulis growth decreased at 20°C and mortality occurred at 25°C, although mortality occurred at lower temperatures when phytoplankton abundance was low, suggesting that mortality occurred through a combination of reducing food source at a time of metabolic stress. Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) and it appears that tolerance varies, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgard, 2005). After acclimation of individuals of Mytilus edulis to 18°C, Kittner & Riisgaard (2005) observed that the filtrations rates were at their maximum between 8.3 and 20°C and below this at 6°C the mussels closed their valves. However, after acclimation at 11°C for five days, the mussels maintained the high filtration rates down to 4°C.  Hence, given time, mussels can acclimatise, shifting their temperature tolerance.

Rising air temperatures can also lead to significant mortality in Mytilus edulis. Intertidal ecosystems are likely to be more negatively impacted than subtidal ecosystems, due to their increased daily and seasonal variations in temperatures (Jones et al., 2009). Tsuchiya (1983) documented the mass mortality of Mytilus edulis in August 1981 due to air temperatures of 34°C that resulted in mussel tissue temperatures over 40°C. In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died. It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two. Under experimental conditions exposure to air temperatures greater than 30°C led to significant mortality (Jones et al., 2009), suggesting this may be an upper temperature threshold for this species.

At the upper range of a mussels tolerance limit, heat shock proteins are produced, indicating high stress levels (Jones et al., 2010). After a single day at 30°C, heat shock proteins were still present over 14 days later, although at a reduced level. Increased temperatures can also affect reproduction in Mytilus edulis (Myrand et al., 2000). In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20°C. These mussels had a low energetic content post-spawning and had stopped shell growth.  The high temperatures likely caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

Temperature changes may also lead to indirect effects. For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004).  Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact upon the health of the mussels if they become infected. There is evidence that increases in temperature will also give a competitive advantage to invasive species. For example, in the Dutch Wadden Sea mild winters favour Magallana gigas recruitment while cold winters favour Mytilus edulis (Deiderich, 2005).

Sensitivity assessment. Sea surface temperatures around the UK are currently between 6-19°C (Huthnance, 2010). Under the three scenarios (middle and high emission and extreme), summer sea temperatures in the south of the UK may rise to temperatures of 22, 23, and 24°C respectively. The global distribution of the piddock species suggests that these species can tolerate warmer waters than currently experienced in the UK. Mytilus edulis is a eurythermal species, and the maximum upper thermal limit of this species appears to generally be somewhere between 25-28°C, above which this species experiences mortality, with tolerance related to exposure. As ocean warming will occur gradually, across the course of this century, it is expected that both piddocks and Mytilus edulis will be able to withstand these increases in temperature.

As these species occur in the intertidal, they will also have to cope with increasing air temperatures. In July, temperatures can reach up to an average of 25°C in the south of the UK, although the highest temperature recorded 1961-2010 was 38.5°C (Perry & Golding, 2011). If air temperatures rise by 3, 4, and 6°C by the end of the century (middle and high and extreme emission scenarios, respectively), this could lead to temperatures reaching average summer high temperatures of between 28 - 31°C.

Under the middle and high emission scenario with seawater temperatures reaching up to 23°C and air temperatures reaching 29°C, both piddocks and Mytilus edulis may be able to adapt to global warming. Most studies place Mytilus edulis upper thermal limit at between 25-28°C for seawater temperatures and 30°C for air temperature, although these temperatures may lead to a decrease in growth. Under these scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, this biotope is assessed as ‘Not sensitive’ to ocean warming under the middle and high emission scenarios.

Piddocks are likely to be able to withstand the temperatures expected under the extreme scenario (based on their biogeography), although Mytilus edulis is likely to be impacted. Whilst seawater temperatures are expected to remain below the threshold upper temperature limit for this species, air temperatures are likely to rise to 31°C, which exceeds potential upper air temperature limits, and is likely to lead to some mortality in the south of the UK. As such, resistance has been assessed as ‘Medium’, whilst resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ to ocean warming under the extreme scenario.

Global warming (high)

High emission scenario (by the end of this century 2081-2100) benchmark of:

• A 4°C rise in SST, NBT (coastal to the shelf seas) and surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf, and

• A 3°C rise in surface air temperature in intertidal habitats exclusive to Scotland. 

Evidence

Little empirical 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. More extensive evidence on thermal tolerance and physiological effects was found for Mytilus edulis.

The American piddock Petricolaria pholadiformis is a cold-temperate species, which originates from the east coast of America, distributed from the Gulf of St Lawrence to the Caribbean (www.obis.org). From there it was unintentionally introduced into southern England with Crassostrea oysters, and from the UK, this species has colonized several northern European countries (Zenetos et al., 2009). It has since established a small population in the Saranikonos Gulf, in the Eastern Mediterranean (Zenetos et al., 2009). Pholis dactylus inhabits the mid-littoral and shallow sublittoral the Mediterranean and the East Atlantic, from Norway to the 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 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).

Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas that frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992). In the north Atlantic, this species occurs from Norway to the coast of Spain. In the western Atlantic, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits have contracted approximately 350 km north of its previous southern limits in Cape Hatteras, North Carolina, due to increases in water temperature beyond the lethal limit (Jones et al., 2009).

Wells & Gray (1960) suggested that the mean summer water temperatures of 26.6 °C set the southern range limit. Gonzalez & Yevich (1976) found that Mytilus edulis could not tolerate sustained temperatures of 27°C, and feeding stopped after 25°C. Pearce (1969) found that whilst some populations of Mytilus edulis could survive 27°C for over a month, none of these populations could survive temperatures of 28°C for more than four days. Read & Cummings (1967) estimated the upper tolerance limit to be 27°C, and Chapple et al. (1998) found that Mytilus edulis could not acclimate to temperatures above 28.5°C. Almada-Villela et al. (1982) found that growth significantly declined in juvenile Mytilus edulis as temperatures increased above 20°C. Similarly, Hiebenthal et al. (2013) found that the growth rate decreased by 60% as temperatures increased from 20 - 25°C and resulted in 25% mortality under experimental conditions. Incze et al. (1980) found that Mytilus edulis growth decreased at 20°C and mortality occurred at 25°C, although mortality occurred at lower temperatures when phytoplankton abundance was low, suggesting that mortality occurred through a combination of reducing food source at a time of metabolic stress. Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) and it appears that tolerance varies, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgard, 2005). After acclimation of individuals of Mytilus edulis to 18°C, Kittner & Riisgaard (2005) observed that the filtrations rates were at their maximum between 8.3 and 20°C and below this at 6°C the mussels closed their valves. However, after acclimation at 11°C for five days, the mussels maintained the high filtration rates down to 4°C.  Hence, given time, mussels can acclimatise, shifting their temperature tolerance.

Rising air temperatures can also lead to significant mortality in Mytilus edulis. Intertidal ecosystems are likely to be more negatively impacted than subtidal ecosystems, due to their increased daily and seasonal variations in temperatures (Jones et al., 2009). Tsuchiya (1983) documented the mass mortality of Mytilus edulis in August 1981 due to air temperatures of 34°C that resulted in mussel tissue temperatures over 40°C. In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died. It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two. Under experimental conditions exposure to air temperatures greater than 30°C led to significant mortality (Jones et al., 2009), suggesting this may be an upper temperature threshold for this species.

At the upper range of a mussels tolerance limit, heat shock proteins are produced, indicating high stress levels (Jones et al., 2010). After a single day at 30°C, heat shock proteins were still present over 14 days later, although at a reduced level. Increased temperatures can also affect reproduction in Mytilus edulis (Myrand et al., 2000). In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20°C. These mussels had a low energetic content post-spawning and had stopped shell growth.  The high temperatures likely caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

Temperature changes may also lead to indirect effects. For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004).  Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact upon the health of the mussels if they become infected. There is evidence that increases in temperature will also give a competitive advantage to invasive species. For example, in the Dutch Wadden Sea mild winters favour Magallana gigas recruitment while cold winters favour Mytilus edulis (Deiderich, 2005).

Sensitivity assessment. Sea surface temperatures around the UK are currently between 6-19°C (Huthnance, 2010). Under the three scenarios (middle and high emission and extreme), summer sea temperatures in the south of the UK may rise to temperatures of 22, 23, and 24°C respectively. The global distribution of the piddock species suggests that these species can tolerate warmer waters than currently experienced in the UK. Mytilus edulis is a eurythermal species, and the maximum upper thermal limit of this species appears to generally be somewhere between 25-28°C, above which this species experiences mortality, with tolerance related to exposure. As ocean warming will occur gradually, across the course of this century, it is expected that both piddocks and Mytilus edulis will be able to withstand these increases in temperature.

As these species occur in the intertidal, they will also have to cope with increasing air temperatures. In July, temperatures can reach up to an average of 25°C in the south of the UK, although the highest temperature recorded 1961-2010 was 38.5°C (Perry & Golding, 2011). If air temperatures rise by 3, 4, and 6°C by the end of the century (middle and high and extreme emission scenarios, respectively), this could lead to temperatures reaching average summer high temperatures of between 28 - 31°C.

Under the middle and high emission scenario with seawater temperatures reaching up to 23°C and air temperatures reaching 29°C, both piddocks and Mytilus edulis may be able to adapt to global warming. Most studies place Mytilus edulis upper thermal limit at between 25-28°C for seawater temperatures and 30°C for air temperature, although these temperatures may lead to a decrease in growth. Under these scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, this biotope is assessed as ‘Not sensitive’ to ocean warming under the middle and high emission scenarios.

Piddocks are likely to be able to withstand the temperatures expected under the extreme scenario (based on their biogeography), although Mytilus edulis is likely to be impacted. Whilst seawater temperatures are expected to remain below the threshold upper temperature limit for this species, air temperatures are likely to rise to 31°C, which exceeds potential upper air temperature limits, and is likely to lead to some mortality in the south of the UK. As such, resistance has been assessed as ‘Medium’, whilst resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ to ocean warming under the extreme scenario.

Global warming (middle)

Middle emission scenario (by the end of this century 2081-2100) benchmark of:

• A 3°C rise in SST, NBT (coastal to the shelf seas) and surface air temperature (in eulittoral and supralittoral habitats).

• A 1°C rise in Deep-sea habitats (>200 m) off the continental shelf.

• A 2°C rise in surface air temperature in intertidal habitats exclusive to Scotland. 

Evidence

Little empirical 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. More extensive evidence on thermal tolerance and physiological effects was found for Mytilus edulis.

The American piddock Petricolaria pholadiformis is a cold-temperate species, which originates from the east coast of America, distributed from the Gulf of St Lawrence to the Caribbean (www.obis.org). From there it was unintentionally introduced into southern England with Crassostrea oysters, and from the UK, this species has colonized several northern European countries (Zenetos et al., 2009). It has since established a small population in the Saranikonos Gulf, in the Eastern Mediterranean (Zenetos et al., 2009). Pholis dactylus inhabits the mid-littoral and shallow sublittoral the Mediterranean and the East Atlantic, from Norway to the 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 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).

Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas that frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992). In the north Atlantic, this species occurs from Norway to the coast of Spain. In the western Atlantic, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits have contracted approximately 350 km north of its previous southern limits in Cape Hatteras, North Carolina, due to increases in water temperature beyond the lethal limit (Jones et al., 2009).

Wells & Gray (1960) suggested that the mean summer water temperatures of 26.6 °C set the southern range limit. Gonzalez & Yevich (1976) found that Mytilus edulis could not tolerate sustained temperatures of 27°C, and feeding stopped after 25°C. Pearce (1969) found that whilst some populations of Mytilus edulis could survive 27°C for over a month, none of these populations could survive temperatures of 28°C for more than four days. Read & Cummings (1967) estimated the upper tolerance limit to be 27°C, and Chapple et al. (1998) found that Mytilus edulis could not acclimate to temperatures above 28.5°C. Almada-Villela et al. (1982) found that growth significantly declined in juvenile Mytilus edulis as temperatures increased above 20°C. Similarly, Hiebenthal et al. (2013) found that the growth rate decreased by 60% as temperatures increased from 20 - 25°C and resulted in 25% mortality under experimental conditions. Incze et al. (1980) found that Mytilus edulis growth decreased at 20°C and mortality occurred at 25°C, although mortality occurred at lower temperatures when phytoplankton abundance was low, suggesting that mortality occurred through a combination of reducing food source at a time of metabolic stress. Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) and it appears that tolerance varies, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgard, 2005). After acclimation of individuals of Mytilus edulis to 18°C, Kittner & Riisgaard (2005) observed that the filtrations rates were at their maximum between 8.3 and 20°C and below this at 6°C the mussels closed their valves. However, after acclimation at 11°C for five days, the mussels maintained the high filtration rates down to 4°C.  Hence, given time, mussels can acclimatise, shifting their temperature tolerance.

Rising air temperatures can also lead to significant mortality in Mytilus edulis. Intertidal ecosystems are likely to be more negatively impacted than subtidal ecosystems, due to their increased daily and seasonal variations in temperatures (Jones et al., 2009). Tsuchiya (1983) documented the mass mortality of Mytilus edulis in August 1981 due to air temperatures of 34°C that resulted in mussel tissue temperatures over 40°C. In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died. It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two. Under experimental conditions exposure to air temperatures greater than 30°C led to significant mortality (Jones et al., 2009), suggesting this may be an upper temperature threshold for this species.

At the upper range of a mussels tolerance limit, heat shock proteins are produced, indicating high stress levels (Jones et al., 2010). After a single day at 30°C, heat shock proteins were still present over 14 days later, although at a reduced level. Increased temperatures can also affect reproduction in Mytilus edulis (Myrand et al., 2000). In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20°C. These mussels had a low energetic content post-spawning and had stopped shell growth.  The high temperatures likely caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

Temperature changes may also lead to indirect effects. For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004).  Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact upon the health of the mussels if they become infected. There is evidence that increases in temperature will also give a competitive advantage to invasive species. For example, in the Dutch Wadden Sea mild winters favour Magallana gigas recruitment while cold winters favour Mytilus edulis (Deiderich, 2005).

Sensitivity assessment. Sea surface temperatures around the UK are currently between 6-19°C (Huthnance, 2010). Under the three scenarios (middle and high emission and extreme), summer sea temperatures in the south of the UK may rise to temperatures of 22, 23, and 24°C respectively. The global distribution of the piddock species suggests that these species can tolerate warmer waters than currently experienced in the UK. Mytilus edulis is a eurythermal species, and the maximum upper thermal limit of this species appears to generally be somewhere between 25-28°C, above which this species experiences mortality, with tolerance related to exposure. As ocean warming will occur gradually, across the course of this century, it is expected that both piddocks and Mytilus edulis will be able to withstand these increases in temperature.

As these species occur in the intertidal, they will also have to cope with increasing air temperatures. In July, temperatures can reach up to an average of 25°C in the south of the UK, although the highest temperature recorded 1961-2010 was 38.5°C (Perry & Golding, 2011). If air temperatures rise by 3, 4, and 6°C by the end of the century (middle and high and extreme emission scenarios, respectively), this could lead to temperatures reaching average summer high temperatures of between 28 - 31°C.

Under the middle and high emission scenario with seawater temperatures reaching up to 23°C and air temperatures reaching 29°C, both piddocks and Mytilus edulis may be able to adapt to global warming. Most studies place Mytilus edulis upper thermal limit at between 25-28°C for seawater temperatures and 30°C for air temperature, although these temperatures may lead to a decrease in growth. Under these scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. Therefore, this biotope is assessed as ‘Not sensitive’ to ocean warming under the middle and high emission scenarios.

Piddocks are likely to be able to withstand the temperatures expected under the extreme scenario (based on their biogeography), although Mytilus edulis is likely to be impacted. Whilst seawater temperatures are expected to remain below the threshold upper temperature limit for this species, air temperatures are likely to rise to 31°C, which exceeds potential upper air temperature limits, and is likely to lead to some mortality in the south of the UK. As such, resistance has been assessed as ‘Medium’, whilst resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ to ocean warming under the extreme scenario.

Marine heatwaves (high)

High emission scenario benchmark: A marine heatwave occurring every two years, with a mean duration of 120 days, and a maximum intensity of 3.5°C (Marine heatwave pressure definitions).

Evidence

Marine heatwaves (middle)

Middle emission scenario benchmark:  A marine heatwave occurring every three years, with a mean duration of 80 days, with a maximum intensity of 2°C. 

Evidence

Ocean acidification (high)

High emission scenario benchmark: a further decrease in pH of 0.35 (annual mean) and corresponding 120% increase in H+ ions, seasonal aragonite saturation of 20% of UK coastal waters and North Sea bottom waters, and the aragonite saturation horizon in the NE Atlantic, off the continental shelf, occurring at a depth of 400 m by the end of this century 2081-2100 (Ocean acidification pressure definitions).

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop by a further 0.35 units by the end of this century, dependent on emission scenario. In general, it is thought that calcifying invertebrates will be more sensitive to ocean acidification than non-calcifying invertebrates, which appear to have a more mixed response (Hofmann et al., 2010), although bivalves generally appear to be tolerant to a decrease in pH (Kroeker et al., 2011, Garrard et al., 2014).

Mytilus edulis is a calcified organism but it is unlikely this species will be significantly negatively impacted by ocean acidification, because acidification does not appear to lead to mortality, even at levels which far exceed levels of ocean acidification expected for the end of this century (e.g. Berge et al., 2006, Melzner et al., 2011). For example, levels of growth in Mytilus edulis were maintained at pH 7.6 -7.7, although growth does decrease under pH levels < 7.4 (Berge et al., 2006, Melzner et al., 2011).

The calcified shell of Mytilus edulis consists of an outer calcite layer and an inner aragonite layer (Fitzer et al., 2015). When cultured at levels of acidification expected for the end of this century under both the middle (550 ppm) and high (1000 ppm) emission scenario, results showed that Mytilus edulis shells became more brittle (Fitzer et al., 2015). There was no impact of ocean acidification on production or strength of the byssal threads (Dickey et al., 2018). Beesley et al. (2008) found that the health of Mytilus edulis decreased as a result of 60 days exposure to increased CO₂, which they suggested was due to the elevated concentration of calcium ions in the haemolymph.  Sun et al. (2017) found that ocean acidification damaged the ultrastructure of haemocytes and led to a reduction in phagocytosis.

The Baltic Sea is naturally low in carbonates, and exhibits seasonal aragonite undersaturation and borderline calcite undersaturation, even at levels of low pCO₂ (Thomsen & Melzner, 2010), yet Mytilus edulis is one of the most conspicuous animals present on the rocky sublittoral of the northern Baltic (Westerbom et al., 2002). In Kiels Fjord in the Baltic Sea, pH can reach levels of <7.5 in the summer, and yet Mytilus edulis can be found there in high densities, and juvenile settlement occurs in the summer when pH values are at their lowest (Thomsen et al., 2010). This suggests that this species will be able to tolerate pH levels expected for the end of this century around the UK. The piddock Barnea candida is also known to be present in the Baltic Sea (Richter & Sarnthein, 1977), suggesting that this species is tolerant of a decrease in pH.

Sensitivity Assessment. Whilst levels of ocean acidification expected for the end of this century appear to decrease organism health through alterations to immune response, and an increase in shell brittleness, it is not certain how these impacts will lead to population level responses. In situ data show that Mytilus edulis can survive, and is abundant in Kiel Fjord, where pH can fluctuate from <7.5 - > 8.2 (Thomsen et al., 2010), whilst the piddock Barnea candida is also present in the Baltic, where pH is highly variable. Therefore, under both the middle and high emission scenario resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a score of ‘Not sensitive’.

Ocean acidification (middle)

Middle emission scenario benchmark: a further decrease in pH of 0.15 (annual mean) and a corresponding 35% increase in H+ ions with no coastal aragonite undersaturation and the aragonite saturation horizon in the NE Atlantic, off the continental shelf, at a depth of 800 m by the end of this century, 2081-2100. 

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop by a further 0.35 units by the end of this century, dependent on emission scenario. In general, it is thought that calcifying invertebrates will be more sensitive to ocean acidification than non-calcifying invertebrates, which appear to have a more mixed response (Hofmann et al., 2010), although bivalves generally appear to be tolerant to a decrease in pH (Kroeker et al., 2011, Garrard et al., 2014).

Mytilus edulis is a calcified organism but it is unlikely this species will be significantly negatively impacted by ocean acidification, because acidification does not appear to lead to mortality, even at levels which far exceed levels of ocean acidification expected for the end of this century (e.g. Berge et al., 2006, Melzner et al., 2011). For example, levels of growth in Mytilus edulis were maintained at pH 7.6 -7.7, although growth does decrease under pH levels < 7.4 (Berge et al., 2006, Melzner et al., 2011).

The calcified shell of Mytilus edulis consists of an outer calcite layer and an inner aragonite layer (Fitzer et al., 2015). When cultured at levels of acidification expected for the end of this century under both the middle (550 ppm) and high (1000 ppm) emission scenario, results showed that Mytilus edulis shells became more brittle (Fitzer et al., 2015). There was no impact of ocean acidification on production or strength of the byssal threads (Dickey et al., 2018). Beesley et al. (2008) found that the health of Mytilus edulis decreased as a result of 60 days exposure to increased CO₂, which they suggested was due to the elevated concentration of calcium ions in the haemolymph.  Sun et al. (2017) found that ocean acidification damaged the ultrastructure of haemocytes and led to a reduction in phagocytosis.

The Baltic Sea is naturally low in carbonates, and exhibits seasonal aragonite undersaturation and borderline calcite undersaturation, even at levels of low pCO₂ (Thomsen & Melzner, 2010), yet Mytilus edulis is one of the most conspicuous animals present on the rocky sublittoral of the northern Baltic (Westerbom et al., 2002). In Kiels Fjord in the Baltic Sea, pH can reach levels of <7.5 in the summer, and yet Mytilus edulis can be found there in high densities, and juvenile settlement occurs in the summer when pH values are at their lowest (Thomsen et al., 2010). This suggests that this species will be able to tolerate pH levels expected for the end of this century around the UK. The piddock Barnea candida is also known to be present in the Baltic Sea (Richter & Sarnthein, 1977), suggesting that this species is tolerant of a decrease in pH.

Sensitivity Assessment. Whilst levels of ocean acidification expected for the end of this century appear to decrease organism health through alterations to immune response, and an increase in shell brittleness, it is not certain how these impacts will lead to population level responses. In situ data show that Mytilus edulis can survive, and is abundant in Kiel Fjord, where pH can fluctuate from <7.5 - > 8.2 (Thomsen et al., 2010), whilst the piddock Barnea candida is also present in the Baltic, where pH is highly variable. Therefore, under both the middle and high emission scenario resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a score of ‘Not sensitive’.

Sea level rise (extreme)

Extreme scenario benchmark: a 107 cm rise in average UK sea-level by the end of this century (2018-2100) (Sea-level rise pressure definitions).

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 1996, Crooks, 2004, Fujii & Raffaelli, 2008).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position with the coastal energy gradient (Crooks, 2004, Fujii & Raffaelli, 2008).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (Pethick, 1996) suggested that a sea-level rise rate of 6 mm/yr. could result in landward movement of estuaries by 10 m/yr., long-shore migration of open coast landforms of 50 m/yr. and ebb-tidal deltas to extend laterally by 300 m/yr. 

The effects of sea-level rise and increased wave action may be increased further due to storms and storms surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Lowe et al., 2018, Palmer et al., 2018). 

This biotope occurs on the lower a shore of the intertidal zone and therefore an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. It may be able to expand its range and migrate upwards to compensate for sea-level rise, if not constrained by lack of suitable clay habitat. Where landward migration is not possible, it is expected that depth distribution of Mytilus edulis and piddocks on eulittoral firm clay will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery.  In this assessment, as the ability to migrate inshore will be site-specific, we will assess on a worst-case-scenario basis, assuming that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and  Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), this biotope may be completely lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small portion of this biotope may be lost.  Therefore, under the medium and high emission scenarios, resistance has been assessed as ‘Low’, as more than 25% of this biotope may be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise, and sensitivity is assessed as ‘High’. Under the extreme scenario resistance has been assessed as ‘None’, as it is likely that more than 75% of this biotope could be lost. Hence, resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise, and sensitivity is assessed as ‘High’.

Sea level rise (high)

High emission scenario benchmark: a 70 cm rise in average UK sea-level by the end of this century (2018-2100). 

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 1996, Crooks, 2004, Fujii & Raffaelli, 2008).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position with the coastal energy gradient (Crooks, 2004, Fujii & Raffaelli, 2008).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (Pethick, 1996) suggested that a sea-level rise rate of 6 mm/yr. could result in landward movement of estuaries by 10 m/yr., long-shore migration of open coast landforms of 50 m/yr. and ebb-tidal deltas to extend laterally by 300 m/yr. 

The effects of sea-level rise and increased wave action may be increased further due to storms and storms surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Lowe et al., 2018, Palmer et al., 2018). 

This biotope occurs on the lower a shore of the intertidal zone and therefore an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. It may be able to expand its range and migrate upwards to compensate for sea-level rise, if not constrained by lack of suitable clay habitat. Where landward migration is not possible, it is expected that depth distribution of Mytilus edulis and piddocks on eulittoral firm clay will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery.  In this assessment, as the ability to migrate inshore will be site-specific, we will assess on a worst-case-scenario basis, assuming that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and  Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), this biotope may be completely lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small portion of this biotope may be lost.  Therefore, under the medium and high emission scenarios, resistance has been assessed as ‘Low’, as more than 25% of this biotope may be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise, and sensitivity is assessed as ‘High’. Under the extreme scenario resistance has been assessed as ‘None’, as it is likely that more than 75% of this biotope could be lost. Hence, resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise, and sensitivity is assessed as ‘High’.

Sea level rise (middle)

Middle emission scenario benchmark: a 50 cm rise in average UK sea-level by the end of this century (2081-2100).

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 1996, Crooks, 2004, Fujii & Raffaelli, 2008).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position with the coastal energy gradient (Crooks, 2004, Fujii & Raffaelli, 2008).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (Pethick, 1996) suggested that a sea-level rise rate of 6 mm/yr. could result in landward movement of estuaries by 10 m/yr., long-shore migration of open coast landforms of 50 m/yr. and ebb-tidal deltas to extend laterally by 300 m/yr. 

The effects of sea-level rise and increased wave action may be increased further due to storms and storms surges.  IPCC (2019) note that the frequency of extreme sea-level events (e.g. due to storms) are predicted to increase as sea-level rises, however, there is no consensus on the future storm and, hence, wave climate around UK coasts (Lowe et al., 2018, Palmer et al., 2018). 

This biotope occurs on the lower a shore of the intertidal zone and therefore an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. It may be able to expand its range and migrate upwards to compensate for sea-level rise, if not constrained by lack of suitable clay habitat. Where landward migration is not possible, it is expected that depth distribution of Mytilus edulis and piddocks on eulittoral firm clay will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery.  In this assessment, as the ability to migrate inshore will be site-specific, we will assess on a worst-case-scenario basis, assuming that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and  Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), this biotope may be completely lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small portion of this biotope may be lost.  Therefore, under the medium and high emission scenarios, resistance has been assessed as ‘Low’, as more than 25% of this biotope may be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise, and sensitivity is assessed as ‘High’. Under the extreme scenario resistance has been assessed as ‘None’, as it is likely that more than 75% of this biotope could be lost. Hence, resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise, and sensitivity is assessed as ‘High’.

Temperature increase (local)

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

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.

Local Mytilus edulis populations may be acclimated to the prevailing temperature regime and may, therefore, exhibit different tolerances to other populations subject to different salinity conditions and therefore caution should be used when inferring tolerances from populations in different regions. Mytilus edulis is a eurytopic species found in a wide temperature range from mild, subtropical regions to areas that frequently experience freezing conditions and are vulnerable to ice scour (Seed & Suchanek, 1992). In recent years, Mytilus edulis has been observed to be expanding its range pole-wards and has reappeared in Svalbard, due to an increase in sea temperature in that region (Berge et al., 2005), whilst its equatorial limits are contracting due to increases in water temperature beyond the lethal limit (Jones et al., 2010). In British waters, 29 °C was recorded as the upper sustained thermal tolerance limit for Mytilus edulis (Read & Cumming, 1967; Almada-Villela, et al., 1982), although it is thought that European mussels will rarely experience temperatures above 25 °C (Seed & Suchanek, 1992). 

A growing body of experimental and field-based evidence supports the conclusion that adult Mytilus edulis exhibits substantial thermal plasticity and acclimation capacity across a wide geographic range. Long-term field observations in the White Sea detected no measurable change in heart rate despite an approximately 4 °C increase in summer seawater temperature, indicating effective physiological compensation under naturally variable thermal regimes (Bakhmet et al., 2019). Controlled laboratory heating experiments further demonstrate that Mytilus edulis has a higher thermal tolerance and lower, more stable heart rate responses than the closely related Mytilus trossulus, with significantly lower (negligible) mortality during progressive warming trials, despite temperature increases of 2 °C/hour from 1.8 to 28.8 °C (Bakhmet et al., 2022).

Transcriptomic and proteomic studies across multiple regions indicate that adult mussels can tolerate short-term warming to at least 27 to 30 °C with limited cellular stress responses. Greenland populations showed minimal transcriptional disruption even at 32 °C, indicating a broad capacity for acclimation across thermal histories (mussels were collected at 27, 19, and 3 °C; Clark et al., 2021). Similarly, mussels from the remote, sub-Antarctic Kerguelen Archipelago exposed to rapid warming from 7.5 to 20 °C exhibited a typical but controlled stress response consistent with conserved protein homeostasis mechanisms rather than acute physiological failure (Bultelle et al., 2021). Archival shell records from the Belgian coast further suggest long-term acclimation or adaptation to increasing thermal variability, with no clear reduction in shell growth despite increased energetic demands associated with warming (Telesca et al., 2021).

Tsuchiya (1983) documented the mass mortality of Mytilus edulis in in Mutsu Bay, northern Japan in August 1981 due to air temperatures of 34 °C that resulted in mussel tissue temperatures in excess of 40 °C. In one hour, 50% of the Mytilus edulis from the upper 75% of the shore had died. It could not be concluded from this study whether the mortality was due to high temperatures, desiccation or a combination of the two. Lethal water temperatures appear to vary between areas (Tsuchiya, 1983) although it appears that their tolerance at certain temperatures vary, depending on the temperature range to which the individuals are acclimatised (Kittner & Riisgaard, 2005). After acclimation of individuals of Mytilus edulis to 18 °C, Kittner & Riisgaard (2005) observed that the filtrations rates were at their maximum between 8.3 and 20 °C and below this at 6 °C the mussels closed their valves. However, after being acclimated at 11 °C for five days, the mussels maintained the high filtration rates down to 4 °C. Hence, given time, mussels can acclimatise and shift their temperature tolerance. Filtration in Mytilus edulis was observed to continue down to -1 °C, with high absorption efficiencies (53-81%) (Loo, 1992).

Elevated temperatures are associated with changes in feeding behaviour, metabolic demand and energy allocation. Experimental warming (+ 3.5 and + 6 °C) has been shown to increase clearance rates under moderate thermal stress but reduce condition index and deplete energy reserves as temperatures approach upper tolerance limits (Guinle et al., 2025). Lipidomic and gene expression analyses indicate substantial membrane remodelling and tissue-specific stress responses, reflecting significant metabolic costs even where short-term survival is maintained (Guinle et al., 2025).

Metabolic suppression and recovery experiments demonstrate that mussels progressively reduce feeding and respiration as temperatures exceed approximately 24 to 26 °C, with partial or full recovery occurring during subsequent cooling phases (Vajedsamiei, Melzner et al., 2021). These dynamics can improve growth performance under extremely high average temperatures when large daily thermal fluctuations occur, but can reduce growth at less extreme temperatures typical of current summer conditions (Vajedsamiei, Melzner et al., 2021). Increased temperatures have also been shown to delay valve reopening following predator cues, indicating altered behavioural trade-offs under thermal stress (Clements et al., 2021).

At the cellular and immunological level, Mytilus edulis shows a capacity to accommodate elevated temperatures, although responses indicate cumulative stress as additional pressures are added. Barrett et al. (2022) examined gene expression responses under combined warming (30 and 33 °C) and reduced salinity (15 vs 23 PSU) and found that similar physiological pathways were activated across all treatments, with progressively stronger upregulation as stressors accumulated. This pattern suggests that while Mytilus edulis is highly resilient to heat stress and can acclimate efficiently to reduced salinity, combined stressors increasingly push individuals towards physiological limits, particularly under more extreme hyposaline conditions (5 PSU), where strong expression of stress and osmoregulatory marker genes was observed.

Molecular responses to acute heat stress are strongly conditioned by prior thermal history. Péden et al. (2016) exposed mussels acclimated to present-day (16.9 to 21.2 °C) or future (18 to 26.2 °C) summer temperature regimes to an identical acute thermal stress and found that gill proteomes differed markedly between treatments. Mussels acclimated to higher temperatures showed improved cellular responses, including increased expression of heat shock proteins, maintenance of cell integrity and a reallocation of energy production towards anaerobic and alternative metabolic pathways. The importance of acclimation temperature was further demonstrated by Péden et al. (2018), who examined mussels from a heavily polluted site following acclimation to lower (16.9 to 21.1 °C) or higher (17.6 to 26.2 °C) temperatures before exposure to acute heat stress (up to 35.2 °C). Mortality was substantially higher in mussels acclimated to lower temperatures (51.7%) compared to those acclimated to warmer conditions (8.3%). While both groups showed activation of protein folding and degradation pathways, surviving mussels from the warmer acclimation treatment exhibited a more effective heat shock response, indicating a greater capacity to withstand combined pollution and thermal stress.

Temperature-related acclimation capacity has also been demonstrated for immune function. Beaudry et al. (2016) assessed haemocyte viability and phagocytic competence following cadmium exposure at 5, 10 and 20 °C. Haemocyte viability increased significantly after 28 days at 10 °C compared to 5 °C, declined slightly but not significantly at 20 °C, and stabilised after longer exposure. Mussels maintained at 5 °C were better able to cope with cumulative stress challenges, indicating that while immune parameters can adjust across a wide thermal range, lower temperatures may confer greater resilience to multiple stressors.

Long-term field datasets indicate that moderate warming can enhance growth under some conditions, but sustained temperature increases are associated with declining condition and biomass at broader spatial and temporal scales. Growth rates of Mytilus edulis in the Wadden Sea were positively correlated with temperature over a 40-year dataset, within a relatively narrow thermal range (11.2 to 14.4 °C), suggesting benefits of mild warming (Beukema et al., 2017). In contrast, evidence from Scottish rocky shore monitoring does not indicate temperature-driven declines in Mytilus edulis at recent levels of warming. As part of the MarClim project, the thermal preference of Mytilus edulis was assessed alongside population trends around Scotland, and analyses covering 2020 to 2022 found no consistent relationship between species’ thermal affinity, position within their geographic range, and whether populations increased or declined between survey periods, despite measurable warming of Scottish coastal waters since 2014 to 2015 (Burrows et al., 2020). In the shorter term, mussel body condition declined linearly with increasing temperature over a five-month experimental period spanning 4.8 to 8.2 °C, indicating energetic constraints under sustained warming (Melzner et al., 2020).

Regional declines in mussel abundance have been documented in rapidly warming areas such as the Gulf of Maine, where populations have declined by more than 60% over the past 40 years, coinciding with increasing temperatures (>0.2 °C/year) and shifts in community composition (Sorte et al., 2017; Matoo et al., 2021). This interpretation is consistent with a synthesis by Metcalf (2019), which identified sustained increases in sea surface temperature (>2 °C over approx. 20 years) as the primary driver of the widespread decline of blue mussel populations in the Gulf of Maine. In the North East Atlantic, population-scale modelling suggests that while individual growth potential may increase slightly under future climate scenarios (RCP 8.5 - the IPCC’s “worst case” scenario), overall biomass is likely to decline due to altered recruitment phenology and spatially heterogeneous responses to warming, including a total halt in recruitment during summer (Thomas & Bacher, 2018; Thomas et al., 2020).

Heat shock proteins are produced at the upper range of a mussels’ tolerance limit indicating high stress levels (Jones et al., 2010). After a single day at 30 °C, the heat shock proteins were still present over 14 days later, although at a reduced level. Increased temperatures can affect reproduction in Mytilus edulis (Myrand et al., 2000). In shallow lagoons, mortality began in late July at the end of a major spawning event when temperatures peaked at >20 °C. These mussels had a low energetic content post-spawning and had stopped shell growth. It is likely that the high temperatures caused mortality due to the reduced condition of the mussels post-spawning (Myrand et al., 2000). Gamete production does not appear to be affected by temperature (Suchanek, 1985).

More recent evidence indicates that reproductive responses to temperature are non-linear and context dependent. Oliveira et al. (2021) analysed fecundity patterns in Portuguese mussel populations and showed that relative fecundity was strongly influenced by the number of days with seawater temperatures exceeding 14 °C during the preceding four months. The highest positive effect occurred at approximately 80 warm days, while relative fecundity declined sharply when the number of warm days exceeded 100. Average sea surface temperature also influenced fecundity, but only when temperatures exceeded 16 °C. The authors cautioned that predictive models based on linear temperature-fecundity relationships may overestimate the ability of Mytilus edulis to cope with future warming scenarios.

A number of experimental studies indicate that survival declines rapidly once upper thermal thresholds are exceeded, particularly under sustained exposure. Complete mortality has been observed after 3 to 32 days at 30 °C, with faster mortality at higher temperatures and declining condition index prior to death (Kamermans & Saurel, 2022). Laboratory mortality experiments also demonstrate strong interactions between temperature and salinity, with low salinity substantially reducing upper thermal tolerance; exposure to air temperatures of 30 to 36 °C resulted in sharply increased mortality under hyposaline conditions compared to full salinity treatments (Nielsen et al., 2021).

Dynamic and constant heatwave experiments further indicate that survival outcomes depend strongly on thermal regime. Vajedsamiei et al. (2024) found that under constant heatwave conditions, survival dropped to zero within seven days at 29 °C and within 30 days at 28 °C, whereas fluctuating (“dynamic”) heatwave regimes allowed partial survival at similar mean temperatures, highlighting the mitigating role of thermal variability. These findings suggest that while Mytilus edulis can tolerate brief exposure to extreme temperatures, sustained or repeated exposure leads to rapid mortality. Furthermore, evidence from both laboratory simulations and real-world heatwave events indicates that repeated exposure to elevated temperatures can progressively erode thermal tolerance, even when individual exposure events are sub-lethal. Repeated daily heat stress experiments reproducing body temperature profiles recorded during the 2018 English Channel heatwave demonstrated that thermal tolerance consistently declined with successive exposures, leading to increased mortality under otherwise moderate climatic conditions (Seuront et al., 2019).

Experimental marine heatwave simulations involving repeated +5 °C warming events increased valve micro-closure frequency, indicating heightened stress, with some behavioural and molecular effects persisting beyond the recovery period (Grimmelpont et al., 2024). Field-based studies suggest that survival during extreme heat events is strongly influenced by microhabitat and biological context, with shell-associated endolithic microbial communities providing a measurable thermal buffering effect of up to 3.2 °C and enhancing survival during the 2022 English Channel heatwave, particularly on the high shore where thermal pressure was greatest (Zardi et al., 2024).

In contrast to adults, early life stages appear substantially more sensitive to elevated temperatures. Laboratory studies show increasing larval abnormalities and developmental arrest with rising temperatures, with complete failure of development at 24° C (Boukadida et al., 2021). Juvenile mussels exposed to short-term warming up to 24 °C exhibited 37% higher mortality than controls, although surviving individuals displayed compensatory increases in shell and tissue growth (Guillou et al., 2023).

Experimental selection studies indicate that extreme thermal events early in life can alter cohort genetic composition, favouring heat-tolerant genotypes but reducing overall performance and tissue mass (Nascimento-Schulze et al., 2025). Recruitment experiments simulating future thermal regimes demonstrate that recruitment success can be reduced by more than 96% under +4 °C warming, even where surviving recruits show enhanced thermal recovery capacity (Vajedsamiei, Wahl et al., 2021). These findings suggest that warming may disproportionately affect population renewal rather than adult persistence.

Temperature changes may also lead to indirect effects. For example, an increase in temperature increases the mussels’ susceptibility to pathogens (Vibrio tubiashii) in the presence of relatively low concentrations of copper (Parry & Pipe, 2004). Increased temperatures may also allow for range expansion of parasites or pathogens which will have a negative impact on the health of the mussels if they become infected.

Altered predator-prey interactions represent an additional indirect pathway by which warming may affect mussel populations. Lugo et al. (2020) conducted a two-month predation experiment and found that a +4 °C temperature increase reduced predation by the native starfish Asterias rubens, with energy intake decreasing by 86%, but approximately doubled predation rates by the invasive crab Hemigrapsus takanoi. Although crab growth rates did not increase, higher consumption was required to sustain existing growth levels, indicating that warming may increase per capita predation pressure by invasive crabs even as native predators become less effective.

Temperature effects on Mytilus edulis are frequently mediated by interactions with additional stressors. Elevated temperatures increase susceptibility to parasitic infection and reduce condition when combined with invasive copepod infection, although mortality responses remain strongly temperature-dependent (Jolma et al., 2025). Burial experiments demonstrate significantly higher mortality under elevated (+5 °C) temperature treatments, particularly in organically enriched and fine sediments, likely due to increased bacterial activity and metabolic demand under hypoxic conditions (Cottrell et al., 2016).

Combined warming and ocean acidification experiments show substantially reduced survival compared to warming alone (Voet et al., 2022), while nutrient enrichment and thermal stress together have produced variable mortality responses depending on exposure timing (Carrier-Belleau et al., 2024). Long-term hazard analyses from the Wadden Sea indicate increased failure rates of newly formed mussel beds under combinations of elevated temperature, low salinity and reduced oxygen, even where hypoxia alone showed no effect (Johansson et al., 2024).

Several studies indicate that elevated temperature can exacerbate the effects of other environmental pressures even where warming alone causes limited mortality. In a field-based experiment on the south coast of the UK, Greatorex & Knights (2023) recorded 6% mortality over eight weeks at both control (15 °C) and elevated (20 °C) temperatures. However, mortality increased to 23% when elevated temperature was combined with simulated ocean acidification, while no mortality occurred under acidification alone. Experimental work under controlled laboratory conditions supports this interaction. Li et al. (2015) exposed adult mussels to temperatures of 19, 22 and 25 °C for two months and found that elevated temperature intensified the effects of ocean acidification, resulting in reduced calcification, lower shell calcium content, altered Ca/Mg ratios and downregulation of biomineralization-related genes.

Transcriptomic responses to combined future stressors have also been documented under short-term exposure. Martino et al. (2019) exposed mussels from the Gulf of Maine to increased temperature and reduced pH alongside decreased food availability for two weeks and identified a shared “core stress response”, including increased expression of genes associated with aerobic metabolism, cellular stress and potential protein degradation. These responses indicate elevated energetic demand under combined stress conditions over short timescales.

Power stations have the potential to cause an increase in sea temperature of up to 15 °C (Cole et al., 1999), although this impact will be localised. However, as mussels are of the most damaging biofouling organisms on water outlets of power stations, they are clearly not adversely affected (Whitehouse et al., 1985; Thompson et al., 2000).

Evidence from recent extreme heat events in north-west Europe demonstrates that mussels in UK-relevant biotopes can experience body temperatures substantially exceeding ambient seawater temperatures during calm, sunny conditions. Real-world heatwave observations show that intertidal mussel body temperatures frequently exceed 30 °C for multiple consecutive days, with localised mass mortality recorded despite relatively moderate climatic heatwave conditions (Seuront et al., 2019; Talevi et al., 2023).

Although some individuals and populations exhibit remarkable tolerance and recovery capacity, including enhanced thermal limits following prior heatwave exposure (King et al., 2024), repeated or cumulative heat stress can reduce feeding, condition and survival, particularly in shallow intertidal settings (Dereuder et al., 2025; Fly et al., 2015). These impacts have broader ecological consequences, as declines in mussel abundance are associated with reduced habitat complexity and shifts towards algal-dominated assemblages (Sorte et al., 2017).

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. In UK waters, a temperature increase of 5 °C for one month or 2 °C for one year relative to current average seawater temperatures is unlikely to cause significant mortality of Mytilus edulis or result in loss of mussel bed extent or structure. Observed responses at these temperature increases are predominantly sub-lethal, including changes in metabolism, feeding behaviour and energy allocation, without clear evidence of compromised population viability of the mussel bed itself. Short-term acute increases may (depending on timing) interfere with spawning cues of piddocks and Mytilus edulis 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. Resistance is therefore assessed as ‘High’, Resilience as ‘High’, and the overall sensitivity of the biotope to temperature increase at the benchmark level is assessed as ‘Not sensitive’ at the benchmark level.

Temperature decrease (local)

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

Evidence

Salinity increase (local)

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

Evidence

The biotope 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. 

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).

Mytilus edulis is found in a wide range of salinities from variable salinity areas (18 to 35 PSU) such as estuaries and intertidal areas to areas of more constant salinity (30 to 35 PSU) in the sublittoral (Connor et al., 2004). Furthermore, mussels in rock pools are likely to experience hypersaline conditions on hot days (Terry et al., 2024). Newell (1979) recorded salinities as high as 42 PSU in intertidal rock pools, suggesting that Mytilus edulis can tolerate hypersaline conditions.

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). While Mytilus edulis appears to have high resistance to increased salinity, no evidence of resistance to the pressure was found 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 (Salinity regime change pressure definition detail).

Evidence

Water flow (tidal current) changes (local)

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

Evidence

Emergence regime changes

Benchmark.  1) A change in the time covered or not covered by the sea for a period of ≥1 year, or 2) an increase in relative sea level or decrease in high water level for ≥1 year. (Emergence regime change pressure definition).

Evidence

Adult piddocks and the clumps of Mytilus edulis that characterize this biotope have no mobility and cannot, therefore, migrate up or down the shore to adapt to changes in emergence. Within the clay substratum, adult piddocks will be afforded some protection by their burrows from desiccation and temperature increases, following increased emergence, by their burrows which will retain some moisture. 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 predation by birds. 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. 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). Changes in emergence may, therefore, alter species abundances and ratios within the piddock population although the biotope will remain recognisable as a piddock biotope.

Mytilus edulis beds are found at a wide range of shore heights from in the strandline down to the shallow sublittoral (Connor et al., 2004).  Their upper limits are controlled by temperature and desiccation (Suchanek, 1978; Seed & Suchanek 1992; Holt et al., 1998) while the lower limits are set by predation, competition (Suchanek, 1978) and sand burial (Daly & Mathieson 1977).  Mussels found higher up the shore display slower growth rates (Buschbaum & Saier, 2001) due to the decrease in time during which they can feed and also a decrease in food availability.  It has been estimated that the point of zero growth occurs at 55% emergence (Baird, 1966) although this figure will vary slightly depending on the conditions of the exposure of the shore (Baird, 1966; Holt et al., 1998). Increasing shore height does, however, increase the longevity of the mussels due to reduced predation pressures (Seed & Suchanek 1992; Holt et al., 1998), resulting in a wider age class of mussels found on the upper shore.

A decrease in emergence would reduce exposure to desiccation and extremes of temperature and allow the piddocks and Mytilus edulis to feed for longer periods and hence grow faster.  Piddocks and mussels are therefore likely to be tolerant of a decrease in emergence and as a result, the biotope may be able to colonize further up the shore, providing a suitable substrate was available. 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, see Micu, 2007, and other species) may increase at the lower edge of the biotope. The lower limit of Mytilus beds is mainly set by predation from Asterias rubens and Carcinus maenas which may increase with a decrease in emergence potentially reducing the lower limit or reducing the number of size classes and age of the mussels at the lower range of the bed (Saier, 2002).  Competition for space with species better adapted to the changed conditions may also alter habitat suitability for this biotope. The Therefore, in the short-term, a decrease in emergence is likely to change the population structure of the mussel bed and, possibly, the piddock populations at their lower limits, probably reducing the species richness of the biotope. Although the mussel patches and piddock populations will effectively survive, the lower limit of the biotope as described may be lost although this biotope will probably colonize further up the shore, if the profile and substratum are suitable.

Sensitivity assessment. This 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, recovery as ‘Medium’ (within 2-10 years) 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 (Wave action pressure definition). 

Evidence

Transition elements & organo-metal contamination

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

Evidence

The sensitivity of this biotope to this pressure is Not assessed. For a Rapid Evidence Assessment on the effects of 'Transitional elements & organometal' contaminants on Mytilus spp., see the full 'Mytilus evidence review'.

Hydrocarbon & PAH contamination

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

Evidence

The sensitivity of this biotope to this pressure is Not assessed. For a Rapid Evidence Assessment on the effects of 'Transitional elements & organometal' contaminants on Mytilus spp., see the full 'Mytilus evidence review'.

Synthetic compound contamination

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

Evidence

The sensitivity of this biotope to this pressure is Not assessed. For a Rapid Evidence Assessment on the effects of 'Transitional elements & organometal' contaminants on Mytilus spp., see the full 'Mytilus evidence review'.

Radionuclide contamination

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

Evidence

No evidence.

Introduction of other substances

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

Evidence

This pressure is Not assessed.

De-oxygenation

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

Evidence

Nutrient enrichment

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

Evidence

Organic enrichment

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

Evidence

Organic enrichment can result from inputs of additional organic matter. Organic enrichment may lead to eutrophication with adverse environmental effects including deoxygenation, algal blooms and changes in community structure (see nutrient enrichment and de-oxygenation). No evidence was found for piddocks to support the assessment of sensitivity to this pressure. Mytilus edulis, however, has sometimes been found to be insensitive to increased organic matter resulting from human activities. 

Mytilus edulis has been recorded in areas around sewage outflows (Akaishi et al. 2007; Lindahl & Kollberg, 2008; Nenonen et al. 2008; Giltrap et al. 2013) suggesting that they are highly tolerant of the increase in organic material that would occur in these areas. A number of studies have also highlighted the ability of Mytilus edulis to utilise the increased volume of organic material available at locations around salmon farms. Reid et al. (2010) noted that Mytilus edulis could absorb organic waste products from a salmon farm with great efficiency. Increased shell length, wet meat weight, and condition index were shown at locations within 200 m from a farm in the Bay of Fundy allowing a reduced time to market (Lander et al., 2012). It has been shown that regardless of the concentration of organic matter Mytilus edulis will maintain its feeding rate by compensating with changes to filtration rate, clearance rates, production of pseudofaeces and absorption efficiencies (Tracey, 1988; Bayne et al., 1993; Hawkins et al., 1996).

Tolerance to organic enrichment is context-dependent and may be reduced under conditions where organic material accumulates within sediments and oxygen availability is restricted. Experimental burial of Mytilus edulis beneath organically enriched sediments resulted in significantly elevated mortality within days, particularly under fine sediments with high organic content and elevated temperatures, where mortality exceeded 50% within short burial durations and increased further with prolonged burial, up to 80% after 32 days (Cottrell et al., 2016). The authors suggested that pathogenic infection associated with microbial activity in organically enriched sediments, combined with burial-induced hypoxia and increased metabolic demand at higher temperatures, was a key driver of mortality. These findings indicate that while mussels tolerate elevated organic matter in well-flushed environments, organic enrichment that leads to sediment accumulation, hypoxia or microbial proliferation may have negative effects.

Evidence from biomonitoring studies indicates that chronic exposure to organically enriched and contaminated environments may be associated with sub-lethal health effects. Mussels sampled near a salmon aquaculture site in Loch Creran exhibited a high prevalence of tissue-level indicators of stress, with 90% of tissue samples showing one or more signs of pollutant-induced pathology, including immune cell infiltration, accumulation of cellular waste products associated with stress and ageing, parasitic infection and abnormal tissue growth (Nippard & Ciocan, 2019). While such effects do not necessarily translate into immediate mortality or population decline, they suggest potential longer-term consequences for condition, disease susceptibility and resilience under sustained organic loading.

Exposure to certain bloom-forming algae has been shown to elicit physiological responses in early life stages. Short-term exposure of larvae to Pseudo-nitzschia multiseries and Prorocentrum lima at elevated concentrations increased immune activity, indicating a stress response even over brief exposure periods (De Rijcke et al., 2015), for example. At the extreme end of organic enrichment, exceptionally high phytoplankton biomass has been associated with mass mortality events in mussels, attributed to hypoxia and physical clogging of the gills during intense bloom conditions (Richardson et al., 2021).

Sensitivity assessment. While Mytilus edulis are fairly resistant to organic enrichment, there is Insufficient evidence to make a sensitivity assessment for the whole biotope due to the lack of evidence regarding piddock sensitivity to this pressure.

Physical loss (to land or freshwater habitat)

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

Evidence

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

Physical change (to another seabed type)

Benchmark. Permanent change from sedimentary or soft rock substrata to hard rock or artificial substrata, or vice versa (Physical change in subtratum type pressure definition).

Evidence

This biotope is characterized by the clay substratum which supports populations of burrowing piddocks. A change to a sedimentary, rock or artificial substratum will result in the loss of piddocks significantly altering the character of the biotope. The biotope is therefore considered to have no resistance (resistance = 'None') 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 clay, when lost natural habitat restoration is unlikely and recovery is, therefore, assessed as 'Very low'. Hence, biotope sensitivity is assessed as 'High', based on the lack of recovery of clay 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) (Physical change in sediment type pressure definition). 

Evidence

This biotope is characterized by the clay substratum which supports populations of burrowing piddocks. A change in 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 (resistance = 'None') to this pressure, recovery of the biological assemblage (following habitat restoration) is considered to be 'Medium' (2-10 years). The biotope is dependent on the presence of clay when lost restoration would not be feasible and recovery is, therefore, assessed as 'Very low'. Sensitivity is therefore assessed as 'High', based on the lack of recovery on clay 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) (Removal of substratum pressure definition). 

Evidence

The removal of substratum down to 30 cm depth will remove the biological assemblage and the substratum.   Resistance is, therefore, assessed as ‘None’, recovery of the biological assemblage (following habitat restoration) is considered to be 'Medium' (2-10 years). However, the biotope is dependent on the presence of clays, when lost habitat restoration is unlikely and recovery is, therefore, assessed as 'Very low'. Hence, sensitivity is assessed as 'High', based on the lack of recovery of clay 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) (Surface abrasion/disturbance pressure definition).

Evidence

Penetration or disturbance of the substratum subsurface

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

Evidence

Penetration and disturbance below the surface of the substratum may damage and remove the Mytilus edulis clumps, surface-dwelling fauna and could damage and expose piddocks depending on the depth of penetration and their burrow depth. Duval (1977) found that the depth of the piddock burrow depended on the size of the animal. For example, an animal with a shell length of 1.2 cm could bore a 2.7 cm burrow whereas animals 4.8 cm long could bore burrows of 12 cm. Piddocks in damaged burrows or those that are removed from the substratum are unlikely to be able to rebury (Duval, 1963a; Barnes, 1980) and will be predated by fish and other mobile species (Micu, 2007). 

Mytilus edulis lives on the surface of the substratum held in place by byssus threads that either attach to the substratum or to other mussels in the bed.  Activities resulting in penetration and disturbance can either directly affect the mussel by crushing or removal, or indirectly affect them by the weakening or breaking of their byssus threads making them vulnerable to displacement (Denny, 1987) where they are unlikely to survive (Dare, 1976). Sub-surface disturbance may also remove mussels by breaking up and removing the substratum. Where mussels are removed attached species including macroalgae and barnacles will also be removed.  In addition, abrasion and sub-surface damage attract mobile scavengers and predators including fish, crabs, and starfish to feed on exposed, dead and damaged individuals and discards  (Kaiser & Spencer, 1994; Ramsay et al., 1998; Groenewold & Fonds, 2000; Bergmann et al., 2002).  This effect could increase predation pressure on surviving damaged and intact Mytilus edulis.

A significant impact resulting from this pressure may be removal and damage of the clay resulting in the clay being more vulnerable to erosion. 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.

Sensitivity assessment. Sub-surface penetration and disturbance could result in damage and removal of the surface infauna including clumps of Mytilus edulis and result in the damage, exposure and loss of piddocks and damage to the habitat. Resistance is, therefore, assessed as ‘Low’ for piddocks, Mytilus edulis and their clay substratum.  The associated surface-dwelling fauna are predicted to recover relatively rapidly via larval recolonization and migration of adults in mobile species. Recovery of the key characterizing species, piddocks and Mytilus edulis are likely to require 2-10 years so that resilience is considered to ‘Medium’. However, as the substratum cannot recover, resilience is assessed as ‘Very Low’ and sensitivity of the overall biotope, based on the sedimentary habitat, is considered to be ‘High’.  

Changes in suspended solids (water clarity)

Benchmark. A change in one rank on the WFD (Water Framework Directive) scale, e.g. from clear to intermediate for one year (Suspended sediment pressure definition).

Evidence

In general, increased suspended particles may enhance food supply where these are organic in origin, or decrease food supply by limiting light availability and increasing the effort needed for filter feeders to obtain food. Food limitation may result in reduced growth and fecundity of the characterising species. Very high levels of silt may clog respiratory and feeding organs of the suspension-feeding piddocks and Mytilus edulis. Increased levels of particles may increase scour and deposition in the biotope depending on local hydrodynamic conditions, although changes in substratum are assessed through the physical change (to another seabed type) pressure.

The piddocks are protected from scour within burrows and increased organic particles would 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). Petricolaria (syn. Petricola) pholadiformis is able to cope in water laden with much suspended material by binding the material in mucus and using the palps to reject it (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.

Mytilus edulis is often found in areas with high levels of turbidity. For example, the average suspended particulate matter (SPM) concentration at Hastings Shingle Bank was 15 to 20 mg/l in June 2005, reaching 50 mg/l in windier (force 4) conditions, although a concentration of 200 mg/l was recorded at this site during gales (Last et al., 2011). Winter (1972, cited by Moore, 1977) recorded 75% mortality of Mytilus edulis in concentrations of 1.84-7.36 mg/l when food was also available. However, a relatively small increase in SPM concentration e.g. from 10 mg/l to 90 mg/l was found to increase growth rates (Hawkins et al., 1996). Concentrations above 250 mg/l have been shown to impair the growth of filter-feeding organisms (Essink, 1999). But Purchon (1937) found that concentrations of particulates as high a 440 mg/l did not affect Mytilus edulis and that mortality only occurred when mud was added to the experiment bringing the concentrations up to 1220 mg/l. The reason for some of the discrepancy between studies may be due to the volume of water used in the experiment. Loosanoff (1962) found that in small quantities of turbid water (due to particulates) the mussel can filter out all of the particulates within a few minutes whereas in volumes >50 gallons per individual the mussel becomes exhausted before the turbidity has been significantly lowered, causing it to close its shell and die.

Based on a comprehensive literature review, Moore (1977) concluded that Mytilus edulis displayed a higher tolerance to high SPM concentrations than many other bivalves although the upper limit of this tolerance was not certain. He also hypothesised that the ability of the mussel to clean its shell in such conditions played a vital role in its success along with its pseudofaecal expulsion.

Recent evidence further supports the importance of suspended particulate material as a food resource. Field and experimental studies show that detrital and particulate organic matter can constitute a substantial proportion of the diet of Mytilus edulis, contributing at least 16% of ingested material (Both et al., 2020), and in estuarine systems more than 50% of annual food intake (Jung et al., 2019). Growth responses are often positive. In Australia, mussel size and dry mass increased with increasing particulate organic matter concentrations, particularly in shallow environments where detritus availability was high (Bearham et al., 2020). 

It may be possible for Mytilus edulis to adapt to a permanent increase in SPM by decreasing their gill size and increasing their palp size in areas of high turbidity (Theisen, 1982; Essink, 1999). In areas of variable SPM, it is likely that the gill size would remain the same but the palp would adapt (Essink, 1999). In addition to morphological adjustment, feeding performance itself appears to be flexible. Steeves et al. (2020) demonstrated that capture efficiency, pumping rate and overall ingestion in Mytilus edulis varied both within and between populations along a fjord gradient (reflecting changes in suspended matter and water clarity). Reciprocal transplant experiments showed that these traits can adjust in response to local environmental conditions, indicating short- to medium-term physiological plasticity rather than fixed population-level differences. Whilst the ability to adapt may prevent immediate declines in health, the energetic costs of these adaptations may result in reduced fitness; the extent of which is still to be established.

Mytilus edulis uses the circadian clock to determine the opening of the shell gape in nocturnal gape cycles (Ameyaw-Akumfi & Naylor, 1987). Last et al. (2011) investigated the effects on increased SPM concentrations on both the gape pattern and mortality in order to establish the effect that aggregate dredging will have on Mytilus edulis and other benthic invertebrates. Therefore, they tested concentrations similar to those expected within a few hundred meters of an aggregate extraction site. The highest concentration tested using a pVORT (paddle Vortex Resuspension Tanks) was approx. 71 mg/l. They reported a significant reduction of the strength of the nocturnal gape cycle at high suspended sediment loads as well as a change in the gape period. The effects of these changes are not fully known but as it is likely that the gape pattern is a strategy to avoid diurnal predators the change may result in an increased risk of predation. On the other hand, the increased turbidity may reduce predation by visual predators such as fish and birds (Essink, 1999). After continued measurements of the gape cycle for four days post-treatment, Last et al. (2011) observed that the cycle took longer than this to recover from the cycle disruption. Further study is required to determine the length of time required for recovery of this behavioural response (Last et al., 2011). Importantly, suspended material associated with turbidity is not inherently detrimental and often represents a significant energetic subsidy. Across a range of coastal and estuarine systems, particulate organic matter supports growth and condition rather than causing energetic limitation (Bearham et al., 2020; Both et al., 2020; Jung et al., 2019). As such, short-term behavioural disruption under elevated suspended sediment loads does not necessarily translate into reduced growth or population-level impacts.

Mytilus edulis may be more sensitive to decreased turbidity where this reflects a decrease in the availability of organic matter and seston. Winter (1972) (cited by Moore, 1977) recorded 75% mortality of Mytilus edulis in concentrations of 1.84-7.36 mg/l when food was also available. However, a relatively small increase in SPM concentration e.g. from 10 mg/l to 90 mg/l was found to increase growth rates (Hawkins et al., 1996). 

Sensitivity assessment. Based on the occurrence of Pholas dactylus in turbid areas and evidence for the production of pseudofaeces by piddocks, they are considered to have high resistance to this pressure. Evidence indicates that Mytilus edulis can tolerate a broad range of suspended solids. The benchmark for this pressure refers to a change in turbidity of one rank on the Water Framework Directive (WFD) scale. Mussel beds form in relatively clear waters of open coasts and wave exposed shores and on sediments in sheltered coast (where turbulent water flow over the mussel beds could resuspend sediments locally) and in turbid bays and estuaries. Therefore, is unlikely that a change in turbidity by of one rank (e.g. from 300 to 100 mg/l or <10 to 100 mg/l) will significantly affect the Mytilus edulis or piddocks. Resistance to this pressure is therefore assessed as ‘High’. Recovery is assessed ‘High’ (no impact to recover from), and sensitivity is, therefore '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 (Smothering pressure definition).

Evidence

In general, increased suspended particles may enhance food supply where these are organic in origin, or decrease food supply by limiting light availability and increasing the effort needed for filter feeders to obtain food. Food limitation may result in reduced growth and fecundity of the characterising species. Very high levels of silt may clog respiratory and feeding organs of the suspension-feeding piddocks and Mytilus edulis. Increased levels of particles may increase scour and deposition in the biotope depending on local hydrodynamic conditions, although changes in substratum are assessed through the physical change (to another seabed type) pressure.

The piddocks are protected from scour within burrows and increased organic particles would 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). Petricolaria (syn. Petricola) pholadiformis is able to cope in water laden with much suspended material by binding the material in mucus and using the palps to reject it (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.

Mytilus edulis is often found in areas with high levels of turbidity. For example, the average suspended particulate matter (SPM) concentration at Hastings Shingle Bank was 15 to 20 mg/l in June 2005, reaching 50 mg/l in windier (force 4) conditions, although a concentration of 200 mg/l was recorded at this site during gales (Last et al., 2011). Winter (1972, cited by Moore, 1977) recorded 75% mortality of Mytilus edulis in concentrations of 1.84-7.36 mg/l when food was also available. However, a relatively small increase in SPM concentration e.g. from 10 mg/l to 90 mg/l was found to increase growth rates (Hawkins et al., 1996). Concentrations above 250 mg/l have been shown to impair the growth of filter-feeding organisms (Essink, 1999). But Purchon (1937) found that concentrations of particulates as high a 440 mg/l did not affect Mytilus edulis and that mortality only occurred when mud was added to the experiment bringing the concentrations up to 1220 mg/l. The reason for some of the discrepancy between studies may be due to the volume of water used in the experiment. Loosanoff (1962) found that in small quantities of turbid water (due to particulates) the mussel can filter out all of the particulates within a few minutes whereas in volumes >50 gallons per individual the mussel becomes exhausted before the turbidity has been significantly lowered, causing it to close its shell and die.

Based on a comprehensive literature review, Moore (1977) concluded that Mytilus edulis displayed a higher tolerance to high SPM concentrations than many other bivalves although the upper limit of this tolerance was not certain. He also hypothesised that the ability of the mussel to clean its shell in such conditions played a vital role in its success along with its pseudofaecal expulsion.

Recent evidence further supports the importance of suspended particulate material as a food resource. Field and experimental studies show that detrital and particulate organic matter can constitute a substantial proportion of the diet of Mytilus edulis, contributing at least 16% of ingested material (Both et al., 2020), and in estuarine systems more than 50% of annual food intake (Jung et al., 2019). Growth responses are often positive. In Australia, mussel size and dry mass increased with increasing particulate organic matter concentrations, particularly in shallow environments where detritus availability was high (Bearham et al., 2020). 

It may be possible for Mytilus edulis to adapt to a permanent increase in SPM by decreasing their gill size and increasing their palp size in areas of high turbidity (Theisen, 1982; Essink, 1999). In areas of variable SPM, it is likely that the gill size would remain the same but the palp would adapt (Essink, 1999). In addition to morphological adjustment, feeding performance itself appears to be flexible. Steeves et al. (2020) demonstrated that capture efficiency, pumping rate and overall ingestion in Mytilus edulis varied both within and between populations along a fjord gradient (reflecting changes in suspended matter and water clarity). Reciprocal transplant experiments showed that these traits can adjust in response to local environmental conditions, indicating short- to medium-term physiological plasticity rather than fixed population-level differences. Whilst the ability to adapt may prevent immediate declines in health, the energetic costs of these adaptations may result in reduced fitness; the extent of which is still to be established.

Mytilus edulis uses the circadian clock to determine the opening of the shell gape in nocturnal gape cycles (Ameyaw-Akumfi & Naylor, 1987). Last et al. (2011) investigated the effects on increased SPM concentrations on both the gape pattern and mortality in order to establish the effect that aggregate dredging will have on Mytilus edulis and other benthic invertebrates. Therefore, they tested concentrations similar to those expected within a few hundred meters of an aggregate extraction site. The highest concentration tested using a pVORT (paddle Vortex Resuspension Tanks) was approx. 71 mg/l. They reported a significant reduction of the strength of the nocturnal gape cycle at high suspended sediment loads as well as a change in the gape period. The effects of these changes are not fully known but as it is likely that the gape pattern is a strategy to avoid diurnal predators the change may result in an increased risk of predation. On the other hand, the increased turbidity may reduce predation by visual predators such as fish and birds (Essink, 1999). After continued measurements of the gape cycle for four days post-treatment, Last et al. (2011) observed that the cycle took longer than this to recover from the cycle disruption. Further study is required to determine the length of time required for recovery of this behavioural response (Last et al., 2011). Importantly, suspended material associated with turbidity is not inherently detrimental and often represents a significant energetic subsidy. Across a range of coastal and estuarine systems, particulate organic matter supports growth and condition rather than causing energetic limitation (Bearham et al., 2020; Both et al., 2020; Jung et al., 2019). As such, short-term behavioural disruption under elevated suspended sediment loads does not necessarily translate into reduced growth or population-level impacts.

Mytilus edulis may be more sensitive to decreased turbidity where this reflects a decrease in the availability of organic matter and seston. Winter (1972) (cited by Moore, 1977) recorded 75% mortality of Mytilus edulis in concentrations of 1.84-7.36 mg/l when food was also available. However, a relatively small increase in SPM concentration e.g. from 10 mg/l to 90 mg/l was found to increase growth rates (Hawkins et al., 1996). 

Sensitivity assessment. Based on the occurrence of Pholas dactylus in turbid areas and evidence for the production of pseudofaeces by piddocks, they are considered to have high resistance to this pressure. Evidence indicates that Mytilus edulis can tolerate a broad range of suspended solids. The benchmark for this pressure refers to a change in turbidity of one rank on the Water Framework Directive (WFD) scale. Mussel beds form in relatively clear waters of open coasts and wave exposed shores and on sediments in sheltered coast (where turbulent water flow over the mussel beds could resuspend sediments locally) and in turbid bays and estuaries. Therefore, is unlikely that a change in turbidity by of one rank (e.g. from 300 to 100 mg/l or <10 to 100 mg/l) will significantly affect the Mytilus edulis or piddocks. Resistance to this pressure is therefore assessed as ‘High’. Recovery is assessed ‘High’ (no impact to recover from), and sensitivity is, therefore 'Not sensitive'. 

Smothering and siltation rate changes (heavy)

Benchmark. ‘Heavy’ deposition of up to 30 cm of fine material added to the seabed in a single discrete event (Smothering pressure definition).

Evidence

A deposit of 30 cm of fine material would lead to smothering of the key characterizing species and the associated biological assemblage. No examples of direct empirical evidence for the response to siltation was found for piddocks, although smothering has been cited as a key threat to piddocks. Piddocks cannot emerge from layers of deposited silt. 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).

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.

Sand burial has been shown to determine the lower limit of Mytilus edulis beds (Daly & Mathieson, 1977a). Burial of Mytilus edulis beds by large scale movements of sand, and resultant mortalities have been reported from Morecambe Bay, the Cumbrian coast and Solway Firth (Holt et al., 1998). Essink (1999) recorded fatal burial depths of 1 to 2 cm for Mytilus edulis and suggested that Mytilus edulis a low tolerance of sedimentation based on investigations by R.Bijkerk (cited by Essink, 1999). However, Widdows et al. (2002) noted that mussels buried by 6 cm of sandy sediment (caused by resuspension of sediment due to turbulent flow across the bed) were able to move to the surface within one day. 

Mytilus edulis occurs in areas of high suspended particulate matter (SPM) and therefore a level of siltation is expected from the settling of SPM. In addition, the high rate of faecal and pseudofaecal matter production by the mussels naturally results in siltation of the seabed, often resulting in the formation of large mounds beneath the mussel bed. For example, at Morecambe Bay, an accumulation of mussel-mud (faeces, pseudofaeces and washed sand) of 0.4-0.5 m between May 1968 and September 1971 resulted in the mortality of young mussels (Daly & Mathieson, 1977). In order to survive the mussels needed to keep moving upwards to stay on the surface. Many individuals did not make it to the surface and were smothered by the accumulation of mussel-mud (Daly & Mathieson, 1977), so that whilst Mytilus edulis does have the capacity to vertically migrate through sediment some individuals will not survive.

Experimental evidence supports the conclusion that burial to depths relevant to the light smothering benchmark can be tolerated for short periods, but that mortality increases with burial duration, sediment fineness and temperature. Last et al. (2011) exposed Mytilus edulis to sudden burial at depths of 2, 5 and 7 cm using coarse, medium and fine sediments for up to 32 days. Overall mortality across all treatments was relatively low (13%), but increased strongly with duration (from 4% after two days to 44% after 32 days) and was substantially higher in fine sediments (28%) than in coarse sediments (2%), reflecting reduced emergence success. Only 16% of buried mussels died after 16 days compared to almost 50% mortality at 32 days. Mortality also increased sharply with a decrease in particle size and with increases in temperature from 8.0 and 14.5 to 20 °C. The ability of a proportion of individuals to emerge from burial was again demonstrated with approximately one quarter of the individuals buried at 2 cm resurfacing. However, at depths of 5 cm and 7 cm no emergence was recorded. The lower mortality when buried in coarse sands may be related to the greater number of individuals who were able to emerge in these conditions and emergence was to be significant for survival (Last et al., 2011). 

The capacity for vertical migration through accumulating sediment has also been demonstrated under gradual burial scenarios. Hutchison et al. (2020) showed that mussels were able to migrate upwards through coarse, medium and fine sediments deposited at rates of 0.5 to 1.5 cm per day over periods of up to 16 days. However, the proportion of buried individuals increased with both burial rate and duration, with approximately 30% buried at 0.5 cm/day and nearly 95% buried at 1.5 cm/day after 16 days, indicating that sustained or rapid sedimentation can overwhelm this compensatory response.

It is unclear whether the same results would be recorded when mussels are joined by byssal threads or whether this would have an impact on survival (Last et al., 2011), although Daly & Mathieson (1977) recorded loose attachments between juvenile mussels during a burial event and some of these were able to surface. It was not clear whether the same ability would be shown by adult mussels in a more densely packed bed.

Burial-associated mortality is further influenced by the organic content of the smothering material. Cottrell et al. (2016) demonstrated significantly higher mortality of Mytilus edulis under 5 cm burial when sediments contained elevated organic matter, particularly in fine sediments. Mortality increased rapidly within two days and rose with both organic content and temperature, reaching over 50% in fine sediments with 1% organic matter under summer temperature conditions (20 °C). The authors suggested that enhanced microbial activity and pathogenic processes within organically enriched sediments contributed to mortality, particularly under low-oxygen conditions.

Sensitivity assessment. Sensitivity to this pressure will be mediated by site-specific hydrodynamic conditions and the footprint of the impact. Where a large area is covered sediments may be shifted by wave and tides rather than removed. The inability of Mytilus edulis to emerge from sediment deeper than 2 cm (Last et al., 2011, Essink, 1999, Daly & Matthieson, 1977) and the increased mortality with depth and reduced particle size observed by Last et al. (2011) indicates that there may be significant mortality of mussels. As piddocks are essentially sedentary with relatively short siphons, siltation from fine sediments rather than sands, even at low levels for short periods could be lethal. Therefore, resistance to siltation is assessed as ‘Low’ for piddocks and Mytilus edulis and resilience is assessed as ‘Medium’ (2 to 10 years) so that sensitivity is assessed as 'Medium'. Survival will be higher in winter months when temperatures are lower and physiological demands are decreased. However, mortality will depend on the duration of smothering. Mortality is likely to be more significant in wave sheltered areas where the smothering sediment remains for prolonged periods and reduced where the smothering sediment is rapidly removed by wave action or currents.

Litter

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

Evidence

Mytilus edulis ingest microplastics. A laboratory experiment using microbeads of polystyrene demonstrated uptake of particles by Mytilus edulis within 12 hours (Browne et al., 2008). After three days some of the beads were translocated to the circulatory system. Microplastics were excreted in faecal pellets but were still present in hemolymph 48 days later. No toxicological effects were observed and there were no changes in filter-feeding activity (Browne et al., 2008). As exposure was short-term it is not clear whether lethal or sub-lethal effects would occur in wild populations over extended periods. There is currently no evidence to assess the level of impact and this pressure is 'Not assessed'.

Electromagnetic changes

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

Evidence

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

Albert et al. (2022) experimentally investigated the effects of magnetic fields comparable to those generated by buried subsea power cables on adult Mytilus edulis. Mussels were exposed to a direct current magnetic field of 300 µT, substantially above the ambient geomagnetic field (approx. 47 µT), reflecting field strengths measured in close proximity to power cables. No significant differences were observed in valve gaping behaviour or filtration rates between exposed and control individuals, indicating no detectable impairment of feeding behaviour under the conditions tested.

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

No evidence was found for the effects of underwater noise on the piddock species which characterise this biotope.

Mytilus edulis are sessile as adults but have planktonic larvae and post-larval stages that are potentially more sensitive to environmental stressors, including anthropogenic underwater noise. Evidence indicates that both behavioural and physiological responses can be elicited by exposure to noise across a range of frequencies, intensities, and durations.

Studies on larval and early life stages show that exposure to vessel noise at environmentally relevant levels can induce physiological and behavioural effects. Jolivet et al. (2016) reported that low-frequency vessel noise (127 ± 3 dB re 1 µPa between 100 and 1,000 Hz) interacted with food availability to significantly increase larval settlement, while Veillard, Beauclercq, Ghafari et al. (2025) and Veillard, Beauclercq, Palacios et al. (2025) demonstrated that shipping noise during embryogenesis and early post-larval development provoked stress-related metabolic disruption, delayed metamorphosis, and altered energy pathways, potentially reducing larval fitness and affecting subsequent recruitment. In contrast, Aspirault et al. (2023) found no impact on larval feeding behaviour at similar noise levels, while Haque & Kwon (2018) reported near-total larval mortality only under specialized high-frequency ultrasound treatments used for antifouling, which are unlikely to occur under typical shipping conditions.

Adult Mytilus edulis respond to underwater noise through partial valve closure, changes in clearance rates, and sublethal oxidative and DNA damage (Hubert et al., 2022; Roberts et al., 2015; Spiga et al., 2016; Wale et al., 2019), although behavioural responses often decay over repeated exposures, indicating some capacity for habituation. Continuous low-frequency noise did not impair byssal thread production in adults (Wang et al., 2024), which may partly explain their persistence in noisy environments. Noise exposure has also been associated with suppression of immune function in adults when combined with bacterial challenge, although no direct mortality was observed (Chapuis et al., 2025). Overall, these studies indicate that underwater noise is unlikely to cause widespread mortality in adult mussels but can induce sublethal physiological stress, behavioural modifications, and, for early life stages, potential delays in development that may influence recruitment success.

Sensitivity assessment. Based on the available evidence, adult Mytilus edulis show a high tolerance to typical underwater noise at the benchmark level, with no significant mortality or reduction in population viability. Sub-lethal effects, including valve closure, altered clearance rates, oxidative stress, and immune modulation, have been observed, but these do not appear to threaten overall population persistence. Larval and post-larval stages may experience developmental delays, metabolic stress, or altered settlement patterns when exposed to sustained or high-intensity noise, which could influence recruitment. However, these effects are unlikely to result in population-level declines under the benchmark exposure, and subsequent reproductive cycles may compensate for temporary recruitment reductions. Despite this, there is Insufficient evidence for a sensitivity assessment for the whole biotope due to the lack of evidence for piddock sensitivity to this pressure.

Introduction of light or shading

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

Evidence

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion (Barrier to species movement pressure definition).

Evidence

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 (Death for collision pressure definition).

Evidence

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

Visual disturbance

Benchmark. The daily duration of transient visual cues exceeds 10% of the period of site occupancy by the feature (Visual disturbance pressure definition). 

Evidence

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’.

Genetic modification & translocation of indigenous species

Benchmark. Translocation of indigenous species or the introduction of genetically modified or genetically different populations of indigenous species may result in changes in the genetic structure of local populations, hybridization, or a change in community structure (Translocation pressure definition).

Evidence

This pressure is only relevant to Mytilus edulis as other species within the biotope are not subject to translocation or cultivation. Commercial cultivation of Mytilus edulis involves the collection of juvenile mussel ‘seed’ or spat (newly settled juveniles ca 1-2cm in length) from wild populations, with subsequent transportation around the UK for re-laying in suitable habitats. As the seed is harvested from wild populations from various locations the gene pool will not necessarily be decreased by translocations.  Movement of mussel seed has the potential to transport pathogens and non-native species (see relevant pressure sections). This pressure assessment is based on Mainwaring et al. (2014) and considers the potential impacts on natural mussel beds of genetic flow between translocated stocks and wild mussel beds.

Two species of Mytilus occur in the UK, Mytilus edulis and Mytilus galloprovincialis.  Mytilus edulis appears to maintain genetic homogeneity throughout its range whereas Mytilus galloprovincialis can be genetically subdivided into a Mediterranean group and an Atlantic group (Beaumont et al., 2007).  Mytilus edulis and Mytilus galloprovincialis have the ability to hybridise in areas where their distribution overlaps e.g. around the Atlantic and European coast (Gardner, 1996; Daguin et al., 2001; Bierne et al., 2002; Beaumont et al., 2004).  In the UK overlaps occur on the North East coast, North East Scotland, South West England and in the North, West and South of Ireland (Beaumont et al., 2007).  It is difficult to identify Mytilus edulis, Mytilus galloprovincialis or hybrids based on shell shape because of the extreme plasticity of shape exhibited by mussels under environmental variation, and a genetic test is required (Beaumont et al., 2007).  There is some discussion questioning the distinction between the two species as the hybrids are fertile (Beaumont et al., 2007).  Hybrids reproduce and spawn at a similar time to both Mytilus edulis and Mytilus galloprovincialis which supports genetic flow between the taxa (Doherty et al., 2009).

There is some evidence that hybrid larvae have a faster growth rate to metamorphosis than pure individuals which may leave pure individuals more vulnerable to predation (Beaumont et al., 1993).  As the physiology of both the hybrid and pure Mytilus edulis is so similar there is likely to be very little impact on the tolerance of the bed to pressures nor a change in the associated fauna.

A review by Svåsand et al. (2007) concluded that there was a lack of evidence distinguishing between different populations to accurately assess the impacts of hybridisation and in particular how the gene flow may be affected by aquaculture.  Therefore, it cannot be confirmed whether farming will have an impact on the genetics of this species beyond the potential for increased hybridization.

Sensitivity assessment. No direct evidence was found regarding the potential for negative impacts of translocated mussel seed on adjacent natural beds.  While it is possible that translocation of mussel seed could lead to gene flow between cultivated beds and local wild populations, there is currently no evidence to assess the impact (Svåsand et al., 2007).  Hybrid beds perform the same ecological functions as Mytilus edulis so that any impact relates to the genetic integrity of a bed alone.  This impact is considered to apply to all mussel biotopes equally, as the main habitat-forming species Mytilus edulis is translocated.  Also, given the uncertainty in the identification of the species, habitats or biotopes that are considered to be characterized by Mytilus edulis may, in fact, contain Mytilus galloprovincialis, their hybrids or a mosaic of the three. Presently, there is no evidence of impact resulting from genetic modification and translocation on Mytilus edulis beds in general or the clumps that characterize 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) (pathogen or disease pressure definition).

Evidence

No evidence was found for microbial pathogen impacts on piddocks.

No evidence was found for microbial pathogen impacts on piddocks. As a species with commercial significance, more research effort has been expended on Mytilus edulis and this assessment is based on a recent evidence review by Mainwaring et al. (2014) of the impacts of Marteilia refringens on Mytilus edulis populations. It should be noted that Mytilus edulis beds are host to a diverse array of disease organisms, parasites and commensals from many animal and plant groups including bacteria, blue-green algae, green algae, protozoa, boring sponges, boring polychaetes, boring lichen, the intermediary life stages of several trematodes, copepods and decapods (Bower, 1992; Gray et al., 1999; Bower, 2010). However, at usual levels of infestation, these are not considered to lead to high levels of mortality and these are not considered by the sensitivity assessment. Outbreaks of Bonamia may cause significant mortalities in some shellfish populations but this protozoan has been shown not to infect Mytilus edulis (Culloty et al., 1999).

Marteilia refringens can infect and have significant impacts on the health of Mytilus edulis. There is some debate as to whether there are two species of Marteilia, one which infects oysters (Marteilia refringens) and another that infects blue mussels (Marteilia maurini) (Le Roux et al., 2001) or whether they are just two strains of the same species (Lopez-Flores et al.,2004; Balseiro et al., 2007). Both species are present in southern parts of the United Kingdom. The infection of Marteilia results in Marteiliosis which disrupts the digestive glands of Mytilus edulis especially at times of spore release. Heavy infection can result in a reduced uptake of food, reduced absorption efficiency, lower carbohydrate levels in the haemolymph and inhibited gonad development particularly after the spring spawning resulting in an overall reduced condition of the individual (Robledo et al., 1995).

Recent evidence suggests that Marteilia is transferred to and from Mytilus edulis via the copepod Paracartia grani. This copepod is not currently prevalent in the UK waters, with only a few records in the English Channel and along the South coast. However, it is thought to be transferred by ballast water and so localised introductions of this vector may be possible in areas of mussel seed transfer. The mussel populations here are considered to be naive (i.e. not previously exposed) and therefore could be heavily affected, although the likelihood is slim due to the dependence on the introduction of a vector that is carrying Marteilia and then it being transferred to the mussels.

Experimental infection of Mytilus edulis with Marteilia pararefringens in Norway induced pronounced pathology, including degeneration of digestive tubules and approximately 50% mortality, although it could not be confirmed whether mortality was directly and exclusively attributable to the infection (Bøgwald et al., 2022).

Berthe et al. (2004) concluded that Mytilus edulis is rarely significantly affected by Marteilia sp. However, occasions have been recorded of nearly 100% mortality when British spat have been transferred from a ‘disease-free area’ to areas in France were Marteilia sp. are present. This suggests that there is a severe potential risk if naive spat are moved around the UK from northern waters into southern waters where the disease is resident (enzootic) or if increased temperatures allow the spread of Marteilia sp. northwards towards the naive northern populations. In addition, rising temperatures could allow increased densities of the Marteilia sp. resulting in heavier infections which can lead to mortality.

Vibrio spp., particularly Vibrio splendidus, have been linked to abnormal mortality events (up to 99% mortality) in wild and cultured mussel populations (Bechemin et al., 2014; Cheikh et al., 2016). Experimental exposure of adult Mytilus edulis to highly pathogenic strains has produced cumulative mortality ranging from 17 to 83% depending on family lineage, indicating both high susceptibility and moderate heritability of resistance traits (Ajithkumar et al., 2024, 2025). Challenge studies with Vibrio spp. suggest that some populations are resistant to naturally occurring doses, with effects typically only observed at high concentrations(approx. 108 CFU (colony-forming units, an estimate of viable microbial cells)/ml, vs approx. 105 CFU/ml found in the environment), and some populations retaining some resistance even to the higher doses (Charles et al., 2020). Vibrio infection can also disrupt the native microbiota, contributing to dysbiosis and mortality (Cheikh & Travers, 2022). Mussels experimentally exposed to Vibrio splendidus and Pseudomonas fluorescens showed a 4.9% increase in cell mortality after four hours, indicating potential acute effects (Gendre et al., 2023).

Larval stages are particularly vulnerable to microbial pathogens. Exposure to naturally occurring concentrations of Vibrio spp. has been shown to reduce larval viability and development (De Rijcke et al., 2016), with experimental challenge causing up to 98% mortality within five days (Eggermont et al., 2017; Wang et al., 2021). Similarly, toxic dinoflagellates such as Karlodinium armiger can cause substantial mortality in both embryos and larvae, with natural blooms affecting recruitment and survival in wild populations (Binzer et al., 2018).

Other microbial organisms have been detected in European Mytilus edulis populations, with potential but currently uncertain impacts on health. Francisella halioticida has been recorded in wild mussels in Normandy, France and in the Tamar estuary, UK, where mass mortality events have occurred, although no causal link has yet been proven (Bouras et al., 2023; Cano et al., 2022). Experimental infection of adults caused up to 36% mortality at very high bacterial doses, though lower, more ecologically relevant, doses produced no significant effects (Bouras et al., 2023). Protozoan parasites, including Cryptosporidium spp. and Toxoplasma gondii, can be bioaccumulated by mussels for at least 21 days (Bigot-Clivot et al., 2022), with haemocytes responding to Toxoplasma gondii exposure, although no measurable pathology was reported for Cryptosporidium spp. (Le Guernic et al., 2020). These findings indicate that while these pathogens can be taken up and retained by mussels, direct effects on survival and condition remain unclear, and further investigation is needed to assess their ecological significance.

Historical records indicate that subtidal Mytilus edulis beds in the Wadden Sea have suffered severe declines associated with disease outbreaks, particularly in combination with other pressures such as intensive fishing. Following such declines, populations have shown partial recovery over decades, suggesting that while adults can survive endemic pathogen presence, recruitment and long-term population stability are vulnerable to novel pathogen introduction (Reise & Buschbaum, 2017, cited in Ricklefs et al., 2020).

A further emerging microbial disease risk is transmissible disseminated neoplasia. Genetic evidence indicates that a clonally transmissible cancer lineage originating in Mytilus trossulus has crossed species boundaries and infected Mytilus edulis populations in Europe, where outbreaks have been associated with extremely high mortality (90 to 100%) affecting both juvenile and adult mussels (Benabdelmouna & Ledu, 2016; Yonemitsu et al., 2019). Although such outbreaks appear spatially restricted, they demonstrate the potential for severe population-level impacts following introduction of novel transmissible pathogens.

Sensitivity assessment. There is no evidence for impacts of microbial pathogens on piddocks, so this assessment solely considers the sensitivity of Mytilus edulis. Mytilus edulis is host to a range of microbial pathogens, including Marteilia spp., Vibrio spp., Francisella halioticida, protozoan parasites such as Cryptosporidium spp. and Toxoplasma gondii, and other bacterial and protozoan organisms. While Marteilia spp. infections can cause reduced condition and occasionally high mortality in naive populations, other pathogens such as Francisella halioticida or Vibrio spp. may cause moderate mortality under high exposure, with effects highly dependent on dose, environmental context, and host genetics. The impacts of protozoans are less clear, with some immune responses but no confirmed widespread mortality. Bower (2010) noted that although Marteilia was a potentially lethal pathogen of mussels, most populations were not adversely affected by marteilioisis, but that in some areas, mortality can be significant in mariculture (Berthe et al., 2004). The resultant population would be more sensitive to other pressures, even where the disease only resulted in a reduced condition. Given the variability in susceptibility and the possibility of high mortality in naive or stressed populations, a precautionary resistance of ‘Medium’ is suggested (<25% mortality), with a resilience of ‘Medium’ (2 to 10 years) resulting in a sensitivity of ‘Medium’.

Removal of target species

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

Evidence

Within this biotope, both Mytilus edulis and piddocks may be targeted as bait or food by fishers (Holt et al., 1998). Commercial harvesting of piddocks has been banned across Europe due to the high levels of habitat damage associated with the removal of boring molluscs (Fanelli et al., 1994). In addition, this biotope is intertidal, and Mytilus edulis occur in patches rather as opposed to a continuous bed. Therefore, it is highly unlikely that the mussels in this biotope would be targeted by dredging activities at high tide. However, hand-harvesting at low tide is possible, and evidence of hand-harvesting is presented here.

Even hand-picking for mussels as bait is likely to significantly deplete the biomass of mussels within this biotope, where they occur as clumps on the substratum (Smith & Murray 2005). Recreational fishermen will often collect moulting Carcinus maenas or whelks by hand from intertidal mussel beds for bait. The removal of predators may benefit Mytilus edulis although this effect is not considered in the sensitivity assessment.

Sensitivity assessment. Mytilus edulis and piddocks have no avoidance mechanisms to escape targeted harvesting. Removal of piddocks and Mytilus edulis will result in loss of targeted individuals and damage to the habitat. Resistance is assessed as ‘Low’ for the piddocks and Mytilus edulis as these sessile species are easily detected and removed. Piddocks and clumps of Mytilus edulis are predicted to recover within 2 to 10 years so that resilience is considered to ‘Medium’ and sensitivity is ‘Medium’.

Removal of non-target species

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

Evidence

This assessment is based on the ecological effects of removal, direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. As Mytilus edulis and piddocks are key characterizing species for this biotope their removal as by-catch would significantly alter the character of the biotope. Mytilus edulis clumps may be removed or damaged by activities targeting other species, this would alter the physical structure of the biotope, reducing habitat for attached and mobile species associated with the mussel clumps. It is unlikely that targeted harvesting of other species would unintentionally remove piddocks.

Sensitivity assessment.  Removal of Mytilus edulis will result in loss of individuals and consequently habitat structure. Resistance is assessed as ‘Low’ for Mytilus and resilience as ‘Medium’ (within 2-10 years) and biotope sensitivity is, therefore, assessed as ‘Medium’.

The American slipper limpet, Crepidula fornicata

Evidence

The carpet sea squirt, Didemnum vexillum

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009).

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

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

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

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

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

Human-mediated transport via aquaculture facilities, boat hulls, commercial fishing vessels, and ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijkstra et al., 2007; Griffith et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009). For example, in oyster farms in British Columbia, large fragments of Didemnum sp. come off oyster strings when they are pulled out of water, and other fragments can be pulled off oysters and mussels and thrown back into the water, which is likely to aid dispersal of the invasive species (Bullard et al., 2007). Dijkstra et al. (2007) hypothesised that Didemnum sp. was introduced to the Gulf of Maine with oyster aquaculture in the Damariscotta River and transported via Pacific oysters.

Didemnum vexillum was likely introduced into the UK from northern Europe or Ireland via poorly maintained or not antifouled vessels, movement of contaminated shellfish stock and aquaculture equipment, or via marine industries such as oil, gas, renewables and dredging (Holt, 2024). Recent evidence from genetic material suggests human-mediated dispersal, between marinas and shellfish culture sites, is the most likely pathway for connectivity of Didemnum vexillum populations throughout Ireland and Britain (Prentice et al., 2021; Holt, 2024). Didemnum vexillum can disperse away from artificial substrata, invading and colonizing natural substrata in surrounding areas (Tillin et al., 2020). Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. The current evidence of Didemnum vexillum’s ability to spread on natural habitats in this area is sparse and often conflicting, complicated by genetics, its apparent variable habitat preferences and tolerances and its variable ability to adapt to ‘new’ conditions (Holt 2024).

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

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

Once established, Didemnum vexillum can expand rapidly, taking over most available hard substrata. Studies have hypothesized that this may alter species diversity and community composition and may decrease species abundance and biodiversity. Gittenberger (2007) stated that at this site, Didemnum sp. could cover around 95% of hard substrata, leaving little space for recruitment and growth of other species.

Didemnum vexillum can overgrow bivalve species, such as oysters, scallops, and mussels, as the hard shells can provide suitable hard substrata for settlement. It has been described as a ‘shellfish pest’ by the aquaculture industry because it is likely to completely encapsulate bivalves and smother them resulting in death or partially encapsulate and partially smother them resulting in reduced bivalve growth (Auker, 2010; Bullard et al., 2007; Coutts & Forrest, 2007, Valentine et al., 2007a; Carman et al., 2009; Kleeman, 2009; Fletcher et al., 2013b; Tillin et al., 2020;). Didemnum vexillum has been recorded overgrowing mussels in Strangford Lough, Northern Ireland (Minchin & Nunn, 2013) and has been recorded forming large mats over blue mussel beds in the Gulf of Maine, completely covering individuals (Auker et al., 2014). Didemnum vexillum fouling on aquaculture equipment and bivalve species causes great economic impacts, as Didemnum vexillum removal methods are expensive, labour-intensive, and not always effective (Coutts & Forrest, 2007; Carman et al., 2009; Kleeman, 2009; Fletcher et al., 2013b; Tillin et al., 2020; Holt, 2024). The fouling on aquaculture nets and bags can restrict water flow and food availability for shellfish, and smothering on mussel farms may result in crop losses (Coutts & Forrest, 2007; Carver et al., 2003, cited by Carman et al., 2009; Fletcher et al., 2013b; Holt, 2024). The effects on mussels are likely to become more prominent as Didemnum vexillum becomes more abundant (Auker, 2010).

The epibiotic relationship between Didemnum vexillum and Mytilus edulis negatively impacts mussel growth (Auker, 2010). Clean control mussels with no Didemnum vexillum overgrowth had thicker shells, a significantly thicker lip, and a greater tissue index, compared to mussels overgrown by Didemnum vexillum (Auker, 2010). The clean mussels’ average length ranged from 3.2 cm to 5.37 cm, and had significantly greater shell lengths than overgrown mussels, which had an average length from 3.4 cm to 4.86 cm (Auker, 2010). Mortality of both control and overgrown mussels was relatively low over the one-year study period, but higher mortality was seen in overgrown mussels (6.7% died) compared to the clean control mussels (1.1% died) (Auker, 2010). Food is an important factor contributing to the decrease in mussel growth (Auker, 2010). Auker (2010) also found that Didemnum vexillum affected reproduction and recruitment of Mytilus edulis as the invasive species grew over gamete release points (siphons) or inhibited settlement of recruits, but this varied seasonally.

Didemnum vexillum has also been recorded forming large mats over blue mussel beds in the Gulf of Maine, completely covering individuals (Auker et al., 2014). However, the overgrowth of mussels by Didemnum vexillum reduced the predation risk on mussels (Auker, 2010; Auker et al., 2014; Lyu et al., 2020). The Didemnum vexillum mats act as refuges for blue mussels (Lyu et al., 2020). Evidence suggested that the relationship between Didemnum vexillum and Mytilus edulis reduced predation by the green crab as Didemnum vexillum deters predator attacks (Auker et al., 2014). It was suggested that the negative impacts of Didemnum vexillum overgrowth on mussel growth, resulting in smaller-sized blue mussels, may protect smaller blue mussels from predation as these are preferred over larger blue mussels by predators (Auker et al., 2014). In Auker’s (2010) study, Carcinus maenas consumed fewer mussels that were overgrown by Didemnum vexillum. The toxic chemical defences of Didemnum vexillum and the release of secondary metabolites and sulfuric acid may deter crab predators (Lyu et al., 2020). The protection from predators provided by Didemnum vexillum may vary seasonally because the invasive ascidians deteriorate during the winter months, potentially reducing predation protection for mussels during this time (Auker et al., 2014).

However, Fletcher et al. (2013b) reported that smaller-sized Perna canaliculus mussels (20-40 mm) were significantly affected by fouling of Didemnum vexillum on cultured mussel ropes. The cultured ropes included ambient fouling (ropes left to be naturally colonized by Didemnum vexillum and other species), enhanced fouling (ambient fouling with ropes that were artificially inoculated with Didemnum vexillum) and a control (small levels of fouling maintained by the removal of Didemnum vexillum). The average mussel density and average mussel weight of smaller mussels were higher in the control than in the treatments fouled by Didemnum vexillum. After 15 months, the smaller mussels were significantly smaller than the medium (40-60 mm) and large (60-70 mm)-sized mussels, which remained a similar size by the end of the experiment. They estimated there was a 40% reduction in small-sized mussel density per kilogram of Didemnum vexillum, indicating a negative relationship between small-sized mussel density and increasing Didemnum vexillum. Small-sized mussels had a significant difference in mussel loss from the larger mussels, with greater loss of the smaller mussels seen in the ambient and enhanced fouling treatments. The small mussels were displaced and overgrown by Didemnum vexillum. Displacement was also evident to a lesser extent in the medium mussels, but was less of a threat to larger mussels. However, the fouling treatments alone did not have a significant overall effect on mussel loss. It was suggested that high levels of fouling on the ropes may have resulted in small mussel loss as the mussels carry out a process of self-thinning, but high levels of fouling did not appear to affect individual mussel size or condition directly (Fletcher et al., 2013b). Fletcher et al. (2013b) also noted that Didemnum vexillum clogged cages and mesh used to house shellfish (e.g. mussels and oysters), which could reduce shellfish growth rates. Fletcher et al. (2013b) concluded that there were no direct effects of Didemnum vexillum fouling on mussel size and condition, but did indicate negative effects on small-sized mussels. However, in their study, Didemnum vexillum was only one of the fouling species contributing to fouling effects (Fletcher et al., 2013b).

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

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

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

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

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

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

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

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

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

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

Sensitivity Assessment. There is evidence of Didemnum vexillum colonization on clay boulders (Hitchin, 2012). According to Tillin et al. (2020), clay exposures are potentially suitable substrata for Didemnum vexillum colonization, although this is stated with low confidence. Therefore, if Didemnum were to colonize this biotope, their mats could cover the burrows from which piddocks need to extend their siphons to feed. In addition, mussels provide a suitable hard substratum for colonization by Didemnum vexillum. It can overgrow mussels and can survive in the lower intertidal zone where this biotope occurs. Resistance is therefore ‘Low’, resilience is ‘Very Low’ as Didemnum vexillum would need to be physically removed to allow recovery, and sensitivity is assessed as ‘High’, albeit with ‘Low’ confidence due to a lack of direct 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, 2011b, 2012a). 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, 2011b, 2012a; Humphreys et al., 2014 cited in Alves et al., 2021; Hansen et al., 2023).

Since its 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, 2011b, 2012a; 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, 2011b, 2012a; 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 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 have 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 mainly of 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 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). Nevertheless, 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.

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 the 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 the 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, 2011b, 2012a; 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 and 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 and 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 to 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 recognised 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 and 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.

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).

In the Wadden Sea and the North Sea, Magallana gigas overgrows mussel beds in the intertidal zone, on sedimentary and rocky habitats of low or high energy (Diederich, 2005, 2006; Nehls et al., 2006; Kochmann et al., 2008; Wrange et al., 2010; Padilla, 2010; GBNNSIP, 2011b, 2012a; Kochmann, 2012; Kochmann et al., 2013; Herbert et al., 2016; Tillin et al., 2020). 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). For example, in Sylt, Wadden Sea, mudflats and mussel beds have now been changed into Magallana gigas reefs (Tillin et al., 2020). In the northern Wadden Sea, this change is considered permanent (Tillin et al., 2020).

Experimental and field evidence indicate that replacement of mussel beds by Pacific oyster reefs can alter associated habitat structure and primary producer communities. In the Wadden Sea, Adriana et al. (2020) compared experimentally constructed mussel- and oyster-dominated reefs and found that oyster reefs promoted bloom-forming green algae and lower habitat complexity, whereas mussel reefs supported meadow-like algal assemblages dominated by fucoids. Field surveys further showed that invasion of mussel beds by Magallana gigas reduced the development of fucoid-dominated algal communities, indicating that oyster-driven shifts in reef structure can have cascading effects on broader reef-associated communities (Adriana et al., 2020).

Diederich (2005, 2006) examined the settlement, recruitment, and growth of Magallana gigas and Mytilus edulis in the northern Wadden Sea. Magallana gigas recruitment success was dependent on temperature, and in the northern Wadden Sea, it only occurred in six of the 18 years since Magallana gigas was first introduced. Survival of juveniles is higher in mild than in cold winters. Also, the survival of both juveniles and adults on mussel beds is higher than that of the mussels themselves. However, recruitment of Magallana gigas was significantly higher in the intertidal than in the shallow subtidal, although the survival of adult oysters or mussels in the subtidal is limited by predation. Deiderich (2005) concluded that hot summers could favour Magallana gigas reproduction, while cold winters could lead to high mussel recruitment the following summer. Diederich (2005, 2006) noted that the high survival rate of Magallana gigas adults and juveniles in the intertidal was likely to compensate for years of poor recruitment. Magallana gigas also prefers to settle on conspecifics, so that it can build massive oyster reefs, which themselves are more resistant to storms or ice scour than the mussel beds they replace, as oysters are cemented together, rather than dependent on byssus threads. Magallana gigas also grows faster than Mytilus edulis in the intertidal and reaches ca 2-3 times the length of mussels within one year. In addition, growth rates in Magallana gigas were independent of the tidal level (emergence regime, substratum, Fucus cover and barnacle epifauna (growing on both mussels and oysters), while the growth rate of Mytilus edulis was decreased by these factors. The faster growth rate could make Magallana gigas more competitive than Mytilus edulis, where space or food is limiting. Diederich (2006) concluded that the massive increase in Magallana gigas in the northern Wadden Sea was caused by high recruitment success, itself due to anomalously warm summer temperatures, the preference for settlement on conspecifics (and hence reef formation), and high survival rates of juveniles. As oyster reefs form on former mussel beds, the available habitat for Mytilus edulis could be restricted (Diederich, 2006). In addition, in the northern German Wadden Sea, the decrease in blue mussel beds and increase in Pacific oysters was linked to climatic conditions rather than caused by the invasion of the Pacific oyster (Nehls et al., 2006).

The strength of oyster impacts on mussel beds appears to be context-dependent and influenced by hydrodynamic conditions. Joyce et al. (2019) demonstrated that the relative impact potential of Magallana gigas declined under higher flow velocities, with lower-flow environments more susceptible to oyster-driven modification, suggesting that local water movement may mediate the extent to which Pacific oysters alter mussel bed structure and function, potentially contributing to spatial variability in invasion outcomes across otherwise similar habitats.

Kent and Essex Inshore Fisheries and Conservation Authority (IFCA) (cited in Herbert et al., 2012) reported that Magallana gigas had developed a significant stock on mussel beds on the Southend foreshore and that, by 2012, there were few mussels left in the affected area, but made no conclusions as to the reason for the decline in mussels. Herbert et al. (2016) reported that many Mytilus edulis beds have changed into mixed reefs dominated by 95% Magallana gigas in the Wadden Sea. Studies of mixed oyster-mussel reefs indicate that coexistence is often associated with fine-scale spatial structuring rather than uniform mixing. In the Wadden Sea, Buschbaum et al. (2016) reported that blue mussels were concentrated at the base of mixed reefs at approximately twice the density observed at the top, where mussels were more heavily colonized by epibiotic barnacles. Mussels located lower in the reef experienced substantially reduced barnacle cover, suggesting that the physical structure of oyster reefs can provide partial refuge from epibiont overgrowth. Long-term observations from Denmark indicate that mixed bivalve beds may persist for decades without strong vertical or temporal segregation of the two species, with mussel-dominated areas occurring alongside oyster-dominated zones within the same reef system (Holm et al., 2016). Population-level data from Danish mussel beds used as primary habitat by Magallana gigas further show that oyster abundance can remain stable over time while mussel densities increase, even under episodic oyster recruitment (Holm et al., 2015). Mixed reefs can also modify predator-prey interactions involving mussels. Waser et al. (2015) found that the presence of oysters significantly reduced mortality of juvenile mussels (<20 mm shell length) in the presence of small shore crabs (Carcinus maenas), although mortality in the presence of larger crabs was less affected. The authors concluded that oyster reefs can alter mussel population structure by disproportionately reducing mortality of early life stages, thereby influencing size distributions within mixed bivalve beds.

Despite concerns that the Pacific oyster can out-compete the Mytilus edulis, research indicates that mixed reefs can shift densities of resident species without suppressing native mussels, and the two species can coexist as mixed ‘oyssel’ beds (Reise et al., 2017; Cornelius & Buschbaum, 2020; Joyce et al., 2021). The invasion of Magallana gigas may alter the structure and function of intertidal reefs in the short term, but can sometimes create a multi-layered structure of a mixture of oysters and blue mussels in the long term 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 (Andriana et al., 2020; Cornelius & Buschbaum, 2020). Reise et al. (2017) noted that in the initial stage of colonization, oysters used mussels for settlement and smothered the bed. Ten years later, the oyster bed became the preferred substratum for settlement, and after 20 years, mussels were no longer the preferred substratum for oyster larvae and were able to use the oyster bed to shelter from predation and parasites (Reise et al., 2017). However, on the remaining hummocks of mussel mud, mussels dominate the top of the hummock and oysters on the sides (Reise et al., 2017). Native and invasive oysters are known to provide a refuge from predators within the biogenic reef they create (Troost, 2010; Goedknegt et al., 2020). The blue mussel Mytilus edulis can make use of the shelter provided by Pacific oysters to escape predators by migrating to the bottom of the Pacific oyster reef, reducing mussel predation by crabs and birds (Goedknegt et al., 2020). Therefore, the presence of Magallana gigas in mussel beds can adjust the mussel predator avoidance. Mixed oyster-mussel beds (‘oyssel’ beds) were reported to exhibit increased species richness, abundance, biomass, and number of deposit feeders compared to mussel beds in the German Wadden Sea (Markert et al., 2010; Herbert et al., 2016; Cornelius & Buschbaum, 2020). However, although mussels may persist within mixed reefs, several studies indicate that oysters often dominate total biomass and influence mussel morphology. In the Wadden Sea, Markert (2020) reported that reefs described as mussel beds were frequently dominated by Magallana gigas in terms of shell mass and total biomass, with oysters accounting for 80-90% of shell material depending on reef complexity. While oyster density did not directly reduce mussel density, mussels in mixed reefs exhibited more slender shell shapes, which may facilitate vertical movement through oyster-dominated structures (Markert, 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 Magallana gigas can colonize clay habitats, although Tillin et al. (2020) believe that clay exposures are potentially suitable for Magallana gigas, but this is stated with medium confidence. In addition, Mytilus edulis shells provide a secondary hard substratum, which is suitable for Magallana gigas, which in turn facilitates the settlement of more Magallana gigas. Therefore, as a precaution, resistance is assessed as ‘Low’, resilience as ‘Very low’ as Magallana gigas would need to be physically be removed to allow recovery, and sensitivity is assessed as ‘High’, albeit with ‘Low’ confidence due to a lack of direct evidence in this habitat.

Wireweed, Sargassum muticum

Evidence

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

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

Sensitivity Assessment. No evidence was found of Sargassum muticum presence in clay habitats. In addition, the level of wave exposure experienced by this biotope is generally unfavourable for Sargassum muticum which thrives better in sheltered sites. Based on the evidence, resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity is assessed as ‘Not Sensitive’, albeit with low confidence.

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. No evidence was found of Undaria pinnatifida presence in clay habitats. Undaria pinnatifida prefers low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018; Epstein et al., 2019a), while this biotope occurs in exposed and moderately exposed sites. 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