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

Ostrea edulis is native to the North-East Atlantic and can be found from the coast of Norway south through the North Sea down to the Iberian Peninsula and the Atlantic coast of Morocco and in the Mediterranean and Black Seas (UKBAP, 1999), suggesting tolerance to a wide range of temperatures.

Filtration rate, metabolic rate, assimilation efficiency and growth rates of adult Ostrea edulis increase with temperature (Newell et al., 1977; Mann, 1979; Haure et al., 1998).  Growth was predicted to be optimal at 17°C or, for short periods, at 25°C (Korringa, 1952; Yonge, 1960; Buxton et al., 1981; Hutchinson & Hawkins, 1992; Haure et al., 1998;), whilst maximum clearance efficiency occurs at 20°C (Newell et al., 1977). No upper lethal temperature was found. Kinne (1970) reported that gill tissue activity fell to zero between 40-42°C, although values derived from single tissue studies should be viewed with caution.  Buxton et al. 1981 reported that specimens survived short-term exposure to 30°C.

Spärck's (1951) data suggest that temperature is an important factor in the recruitment of Ostrea edulis, especially at the northern extremes of its range and Korringa (1952) reported that warm summers resulted in good recruitment.  Spawning around Europe was initiated once the temperature had risen to 13-16°C (Burke et al., 2008; Korringa, 1952; Yonge, 1960) although, in Canada, spawning appeared to occur at 18°C, showing local adaptation (Burke et al., 2008). Davis & Calabrese (1969) reported that larvae grew faster with increasing temperature and that survival was optimal between from 12.5 - 27.5°C but that survival was poor at 30°C. Prado et al. (2016) found that temperature did not affect the survival of spat, but that survival of umbonate and veliger larvae was maintained at temperatures up to 26°C but decreased by almost 50 % at 30°C. Pediveliger larval survival was low at all experimental temperatures but declined at temperatures ≥ 22°C.  As the adult stage appears tolerant to high temperatures, larval temperature tolerance may set the limit for thermal optimums.  Therefore, recruitment and the long-term survival of an oyster bed is probably affected by temperature and may benefit from an increase in temperature.

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, whilst in Scotland, summer sea surface temperatures may rise to 17, 18, and 19°C. In winter, minimum temperatures are expected to rise to 12, 13, and 14°C in the south and to 9, 10, and 11°C in the north. Ostrea edulis is a eurythermal species, and the maximum upper thermal limit of this species has not been defined. Spawning is induced when water temperatures hit 15°C and significant larval mortality has been shown at temperatures ≥ 22°C (Prado et al., 2016), although increasingly warm waters are likely to induce an earlier spawning season spawning so that larval stages avoid summer high temperatures. As ocean warming will occur gradually, and this species occurs in the Mediterranean, it is expected that Ostrea edulis will be able to withstand increases in temperature predicted for each of the three scenarios. Therefore, under the middle and high emission and extreme scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. This biotope is assessed as ‘Not sensitive’ to ocean warming.

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

Ostrea edulis is native to the North-East Atlantic and can be found from the coast of Norway south through the North Sea down to the Iberian Peninsula and the Atlantic coast of Morocco and in the Mediterranean and Black Seas (UKBAP, 1999), suggesting tolerance to a wide range of temperatures.

Filtration rate, metabolic rate, assimilation efficiency and growth rates of adult Ostrea edulis increase with temperature (Newell et al., 1977; Mann, 1979; Haure et al., 1998).  Growth was predicted to be optimal at 17°C or, for short periods, at 25°C (Korringa, 1952; Yonge, 1960; Buxton et al., 1981; Hutchinson & Hawkins, 1992; Haure et al., 1998;), whilst maximum clearance efficiency occurs at 20°C (Newell et al., 1977). No upper lethal temperature was found. Kinne (1970) reported that gill tissue activity fell to zero between 40-42°C, although values derived from single tissue studies should be viewed with caution.  Buxton et al. 1981 reported that specimens survived short-term exposure to 30°C.

Spärck's (1951) data suggest that temperature is an important factor in the recruitment of Ostrea edulis, especially at the northern extremes of its range and Korringa (1952) reported that warm summers resulted in good recruitment.  Spawning around Europe was initiated once the temperature had risen to 13-16°C (Burke et al., 2008; Korringa, 1952; Yonge, 1960) although, in Canada, spawning appeared to occur at 18°C, showing local adaptation (Burke et al., 2008). Davis & Calabrese (1969) reported that larvae grew faster with increasing temperature and that survival was optimal between from 12.5 - 27.5°C but that survival was poor at 30°C. Prado et al. (2016) found that temperature did not affect the survival of spat, but that survival of umbonate and veliger larvae was maintained at temperatures up to 26°C but decreased by almost 50 % at 30°C. Pediveliger larval survival was low at all experimental temperatures but declined at temperatures ≥ 22°C.  As the adult stage appears tolerant to high temperatures, larval temperature tolerance may set the limit for thermal optimums.  Therefore, recruitment and the long-term survival of an oyster bed is probably affected by temperature and may benefit from an increase in temperature.

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, whilst in Scotland, summer sea surface temperatures may rise to 17, 18, and 19°C. In winter, minimum temperatures are expected to rise to 12, 13, and 14°C in the south and to 9, 10, and 11°C in the north. Ostrea edulis is a eurythermal species, and the maximum upper thermal limit of this species has not been defined. Spawning is induced when water temperatures hit 15°C and significant larval mortality has been shown at temperatures ≥ 22°C (Prado et al., 2016), although increasingly warm waters are likely to induce an earlier spawning season spawning so that larval stages avoid summer high temperatures. As ocean warming will occur gradually, and this species occurs in the Mediterranean, it is expected that Ostrea edulis will be able to withstand increases in temperature predicted for each of the three scenarios. Therefore, under the middle and high emission and extreme scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. This biotope is assessed as ‘Not sensitive’ to ocean warming.

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

Ostrea edulis is native to the North-East Atlantic and can be found from the coast of Norway south through the North Sea down to the Iberian Peninsula and the Atlantic coast of Morocco and in the Mediterranean and Black Seas (UKBAP, 1999), suggesting tolerance to a wide range of temperatures.

Filtration rate, metabolic rate, assimilation efficiency and growth rates of adult Ostrea edulis increase with temperature (Newell et al., 1977; Mann, 1979; Haure et al., 1998).  Growth was predicted to be optimal at 17°C or, for short periods, at 25°C (Korringa, 1952; Yonge, 1960; Buxton et al., 1981; Hutchinson & Hawkins, 1992; Haure et al., 1998;), whilst maximum clearance efficiency occurs at 20°C (Newell et al., 1977). No upper lethal temperature was found. Kinne (1970) reported that gill tissue activity fell to zero between 40-42°C, although values derived from single tissue studies should be viewed with caution.  Buxton et al. 1981 reported that specimens survived short-term exposure to 30°C.

Spärck's (1951) data suggest that temperature is an important factor in the recruitment of Ostrea edulis, especially at the northern extremes of its range and Korringa (1952) reported that warm summers resulted in good recruitment.  Spawning around Europe was initiated once the temperature had risen to 13-16°C (Burke et al., 2008; Korringa, 1952; Yonge, 1960) although, in Canada, spawning appeared to occur at 18°C, showing local adaptation (Burke et al., 2008). Davis & Calabrese (1969) reported that larvae grew faster with increasing temperature and that survival was optimal between from 12.5 - 27.5°C but that survival was poor at 30°C. Prado et al. (2016) found that temperature did not affect the survival of spat, but that survival of umbonate and veliger larvae was maintained at temperatures up to 26°C but decreased by almost 50 % at 30°C. Pediveliger larval survival was low at all experimental temperatures but declined at temperatures ≥ 22°C.  As the adult stage appears tolerant to high temperatures, larval temperature tolerance may set the limit for thermal optimums.  Therefore, recruitment and the long-term survival of an oyster bed is probably affected by temperature and may benefit from an increase in temperature.

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, whilst in Scotland, summer sea surface temperatures may rise to 17, 18, and 19°C. In winter, minimum temperatures are expected to rise to 12, 13, and 14°C in the south and to 9, 10, and 11°C in the north. Ostrea edulis is a eurythermal species, and the maximum upper thermal limit of this species has not been defined. Spawning is induced when water temperatures hit 15°C and significant larval mortality has been shown at temperatures ≥ 22°C (Prado et al., 2016), although increasingly warm waters are likely to induce an earlier spawning season spawning so that larval stages avoid summer high temperatures. As ocean warming will occur gradually, and this species occurs in the Mediterranean, it is expected that Ostrea edulis will be able to withstand increases in temperature predicted for each of the three scenarios. Therefore, under the middle and high emission and extreme scenarios, resistance has been assessed as ‘High’, whilst resilience is assessed as ‘High’. This biotope is assessed as ‘Not sensitive’ to ocean warming.

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 due to increased air-sea heat flux are predicted to occur more frequently, last for longer and at increased intensity by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Whilst extreme cold winters have caused mass mortality of Ostrea edulis (Crisp, 1964), there is little available evidence of the impact of a marine heatwave on Ostrea edulis, although adult life stages appear to be able to withstand high temperatures and, in the laboratory, no mortality has been observed at temperatures of 30°C (Newell et al., 1977, Haure et al., 1998). A study comparing how Ostrea edulis responds to rapid and gradual warming found that oysters could tolerate higher temperatures when the heat increased quickly (Götze et al., 2025). Under rapid warming (2°C per hour), 50% mortality occurred by around 37°C, whereas, under gradual warming (2°C per day), 50% mortality occurred by around 35°C. During rapid warming, oysters shifted to anaerobic metabolism shortly after 30°C, and by 36°C cardiac activity had begun to fail, leading to mortality. Under gradual warming, oysters had more time to make metabolic adjustments and therefore delayed the onset of anaerobic metabolism, allowing them to survive for longer periods, but only to lower temperatures. Thus, rapid warming elevated the maximum temperature tolerated but reduced survival time, while gradual warming extended survival time but reduced maximum thermal limits (Götze et al., 2025).

In addition, a mesocosm experiment simulating marine heatwave conditions in tidal pools demonstrated that Ostrea edulis (and Magallana gigas) collected from intertidal sites in Strangford Lough, Northern Ireland, survived one week at 25.2°C (heatwave treatment). No mortality occurred two weeks after the treatment ended. The growth rate for Ostrea edulis was reduced under increased temperature (compared to an ambient temperature of 15.5°C). Ostrea edulis gained more biomass than Magallana gigas under heatwave conditions, providing that there was higher food availability. Respiration rate was increased during the heatwave exposure but decreased to below ambient levels two weeks after the experiment (Gilson et al., 2021), potentially suggesting a delayed physiological effect. These results suggest that Ostrea edulis has some tolerance to marine heatwaves up to 25.5°C and may have a competitive advantage over Magallana gigas in such conditions, if there is high enough food supply.

Ostrea edulis larvae also appear to be tolerant to increased temperatures up to 30°C. Prado et al. (2016) showed that, while late stage larvae suffered a significant decrease in survival ≥ 22°C, earlier larval stages survived until 26°C, with a steep increase in mortality only at 30°C. Subsequent studies also reported that larval growth rate and settlement success increased in higher temperatures of 25 to 30°C (Robert et al., 2017; Alter et al., 2024; Jung et al., 2025). Therefore, heatwaves may have a positive impact on recruitment, providing temperatures do not exceed 30°C. However, adult Ostrea edulis survive better in cooler conditions.

Sensitivity Assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, heatwaves could lead to summer sea temperatures reaching up to 24°C in southern England. Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, the heatwave could last the entire summer with temperatures reaching up to 26.5°C in the south of the UK. Ostrea edulis is thought to have an upper thermal limit of above 30°C, although maximum growth and clearance efficiency occur between 17 and 25°C (see Global Warming). As such, mature Ostrea edulis are likely to be able to tolerate a heatwave of this magnitude, and it may even facilitate recruitment success. Therefore, resistance has been assessed as ‘High'. Resilience as ‘High’, leading to a sensitivity assessment of ‘Not sensitive’ for this biotope under both the middle and high emission scenario.

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

Marine heatwaves due to increased air-sea heat flux are predicted to occur more frequently, last for longer and at increased intensity by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Whilst extreme cold winters have caused mass mortality of Ostrea edulis (Crisp, 1964), there is little available evidence of the impact of a marine heatwave on Ostrea edulis, although adult life stages appear to be able to withstand high temperatures and, in the laboratory, no mortality has been observed at temperatures of 30°C (Newell et al., 1977, Haure et al., 1998). A study comparing how Ostrea edulis responds to rapid and gradual warming found that oysters could tolerate higher temperatures when the heat increased quickly (Götze et al., 2025). Under rapid warming (2°C per hour), 50% mortality occurred by around 37°C, whereas, under gradual warming (2°C per day), 50% mortality occurred by around 35°C. During rapid warming, oysters shifted to anaerobic metabolism shortly after 30°C, and by 36°C cardiac activity had begun to fail, leading to mortality. Under gradual warming, oysters had more time to make metabolic adjustments and therefore delayed the onset of anaerobic metabolism, allowing them to survive for longer periods, but only to lower temperatures. Thus, rapid warming elevated the maximum temperature tolerated but reduced survival time, while gradual warming extended survival time but reduced maximum thermal limits (Götze et al., 2025).

In addition, a mesocosm experiment simulating marine heatwave conditions in tidal pools demonstrated that Ostrea edulis (and Magallana gigas) collected from intertidal sites in Strangford Lough, Northern Ireland, survived one week at 25.2°C (heatwave treatment). No mortality occurred two weeks after the treatment ended. The growth rate for Ostrea edulis was reduced under increased temperature (compared to an ambient temperature of 15.5°C). Ostrea edulis gained more biomass than Magallana gigas under heatwave conditions, providing that there was higher food availability. Respiration rate was increased during the heatwave exposure but decreased to below ambient levels two weeks after the experiment (Gilson et al., 2021), potentially suggesting a delayed physiological effect. These results suggest that Ostrea edulis has some tolerance to marine heatwaves up to 25.5°C and may have a competitive advantage over Magallana gigas in such conditions, if there is high enough food supply.

Ostrea edulis larvae also appear to be tolerant to increased temperatures up to 30°C. Prado et al. (2016) showed that, while late stage larvae suffered a significant decrease in survival ≥ 22°C, earlier larval stages survived until 26°C, with a steep increase in mortality only at 30°C. Subsequent studies also reported that larval growth rate and settlement success increased in higher temperatures of 25 to 30°C (Robert et al., 2017; Alter et al., 2024; Jung et al., 2025). Therefore, heatwaves may have a positive impact on recruitment, providing temperatures do not exceed 30°C. However, adult Ostrea edulis survive better in cooler conditions.

Sensitivity Assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, heatwaves could lead to summer sea temperatures reaching up to 24°C in southern England. Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, the heatwave could last the entire summer with temperatures reaching up to 26.5°C in the south of the UK. Ostrea edulis is thought to have an upper thermal limit of above 30°C, although maximum growth and clearance efficiency occur between 17 and 25°C (see Global Warming). As such, mature Ostrea edulis are likely to be able to tolerate a heatwave of this magnitude, and it may even facilitate recruitment success. Therefore, resistance has been assessed as ‘High'. Resilience as ‘High’, leading to a sensitivity assessment of ‘Not sensitive’ for this biotope under both the middle and high emission scenario.

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 1700’s to 8.14 in the 1990’s (Jacobson, 2005) and is 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). It must be noted that many species show variation in their response to pCO₂ independent of their taxonomic group or habitat preferences (Widdicombe & Spicer, 2008; Kroeker et al., 2013).

Whilst Ostrea edulis is a calcifying organism, it appears relatively robust to ocean acidification at levels expected for the end of this century (Prado et al., 2016, Lemasson et al., 2018), unlike other species of oyster such as Magallana gigas (formerly Crassostrea gigas) (Barton et al., 2012, Lemasson et al., 2018) and Ostrea lurida (Hettinger et al., 2013), which exhibit negative responses to ocean acidification. For example, Lemasson et al. (2018) showed that adult Ostrea edulis did not appear affected by an experimental decrease in pH (to 1000 ppm pCO₂), whilst Magallana gigas exhibited a decrease in clearance rate and condition index. However, Caillon et al., (2023) found the opposite to be true. They reported that Magallana gigas were able to withstand a decrease in pH from 7.7 to 6.4 for 39 days without any mortality, whereas Ostrea edulis were less tolerant of reduced pH, with < 40% survival after 50 days at a pH of 6.6. Furthermore, the physiological functioning of Ostrea edulis reduced linearly with decreasing pH, whereas for Magallana gigas, there appeared to be a tipping point, after which, certain functions declined rapidly. These results suggests that Magallana gigas has a better tolerance for decreased pH than Ostrea edulis, which may give them a competitive advantage where they are found co-occurring (Caillon et al., 2023).

A study investigating the uptake of certain pollutants (cadmium, cobalt, and caesium) showed that exposure to reduced pH (from 8.1 to 7.5) did not significantly alter uptake rate by Ostrea edulis, suggesting that increased ocean acidification may not alter contamination risk in this species (Sezer et al., 2018). Furthermore, lower pH reduced bacterial growth, particularly of Vibrio sp., a common pathogen of Ostrea edulis larvae, thus potentially reducing the likelihood of oysters getting infected with pathogens. Evidence also suggests that anti-predator defences (adductor muscle and shell strength) of Ostrea edulis under ocean acidification remain unaffected in elevated atmospheric carbon dioxide concentrations (up to 1000 ppm) (Lemasson & Knights, 2021).

Sezer et al. (2018) and Caillon et al. (2023) reported bleaching of Ostrea edulis shells in response reduced pH between 7.8 and 6.9. At pH 6.9, shells had fully bleached (Caillon et al., 2023). This bleaching was suggested to be caused by the dissolution of the periostracum and/or a change in the micro-community on the shell surface (Sezer et al., 2018).

Survival rates of the larvae of Ostrea edulis increased by up to 25% in response to ocean acidification, suggesting a positive benefit of increased hypercapnia (Prado et al., 2016). However, thermal stress of up to 30°C negated this positive effect of lower pH on the larvae, resulting in mortality (Prado et al., 2016).

Sensitivity assessment. Experimental evidence suggests that Ostrea edulis may become impacted by a reduced pH of 7.7 and below, after which, physiological functioning will decline, and it may be outcompeted by other more tolerant species. However, it appears robust to future levels of ocean acidification projected for both the middle emission and high emission scenarios. Therefore, under both the middle and high emission scenarios, resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a sensitivity assessment 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 1700’s to 8.14 in the 1990’s (Jacobson, 2005) and is 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). It must be noted that many species show variation in their response to pCO₂ independent of their taxonomic group or habitat preferences (Widdicombe & Spicer, 2008; Kroeker et al., 2013).

Whilst Ostrea edulis is a calcifying organism, it appears relatively robust to ocean acidification at levels expected for the end of this century (Prado et al., 2016, Lemasson et al., 2018), unlike other species of oyster such as Magallana gigas (formerly Crassostrea gigas) (Barton et al., 2012, Lemasson et al., 2018) and Ostrea lurida (Hettinger et al., 2013), which exhibit negative responses to ocean acidification. For example, Lemasson et al. (2018) showed that adult Ostrea edulis did not appear affected by an experimental decrease in pH (to 1000 ppm pCO₂), whilst Magallana gigas exhibited a decrease in clearance rate and condition index. However, Caillon et al., (2023) found the opposite to be true. They reported that Magallana gigas were able to withstand a decrease in pH from 7.7 to 6.4 for 39 days without any mortality, whereas Ostrea edulis were less tolerant of reduced pH, with < 40% survival after 50 days at a pH of 6.6. Furthermore, the physiological functioning of Ostrea edulis reduced linearly with decreasing pH, whereas for Magallana gigas, there appeared to be a tipping point, after which, certain functions declined rapidly. These results suggests that Magallana gigas has a better tolerance for decreased pH than Ostrea edulis, which may give them a competitive advantage where they are found co-occurring (Caillon et al., 2023).

A study investigating the uptake of certain pollutants (cadmium, cobalt, and caesium) showed that exposure to reduced pH (from 8.1 to 7.5) did not significantly alter uptake rate by Ostrea edulis, suggesting that increased ocean acidification may not alter contamination risk in this species (Sezer et al., 2018). Furthermore, lower pH reduced bacterial growth, particularly of Vibrio sp., a common pathogen of Ostrea edulis larvae, thus potentially reducing the likelihood of oysters getting infected with pathogens. Evidence also suggests that anti-predator defences (adductor muscle and shell strength) of Ostrea edulis under ocean acidification remain unaffected in elevated atmospheric carbon dioxide concentrations (up to 1000 ppm) (Lemasson & Knights, 2021).

Sezer et al. (2018) and Caillon et al. (2023) reported bleaching of Ostrea edulis shells in response reduced pH between 7.8 and 6.9. At pH 6.9, shells had fully bleached (Caillon et al., 2023). This bleaching was suggested to be caused by the dissolution of the periostracum and/or a change in the micro-community on the shell surface (Sezer et al., 2018).

Survival rates of the larvae of Ostrea edulis increased by up to 25% in response to ocean acidification, suggesting a positive benefit of increased hypercapnia (Prado et al., 2016). However, thermal stress of up to 30°C negated this positive effect of lower pH on the larvae, resulting in mortality (Prado et al., 2016).

Sensitivity assessment. Experimental evidence suggests that Ostrea edulis may become impacted by a reduced pH of 7.7 and below, after which, physiological functioning will decline, and it may be outcompeted by other more tolerant species. However, it appears robust to future levels of ocean acidification projected for both the middle emission and high emission scenarios. Therefore, under both the middle and high emission scenarios, resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a sensitivity assessment 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

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006).  This biotope is recorded between 0 – 20 m depth in the UK, although Ostrea edulis can be found at depths of up to 50 m (OSPAR, 2008).  Therefore, an increase in depth of between 50 – 107 cm is unlikely to have large implications for this species.

Ostrea edulis beds occur on shallow sublittoral muddy mixed sediment in sheltered environments with weak tidal streams (JNCC).  Understanding of how sea-level rise will affect exposure or tidal energy is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. The effects of sea-level rise and increased wave action may be increased further due to storms and storm 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 (Mossman et al., 2015, Lowe et al., 2018, Palmer et al., 2018).

Sensitivity assessment. This habitat occurs from 0 - 20 m depth, although Ostrea edulis beds can be found at depths of up to 50 m. Any change to the habitat in terms of its exposure or tidal currents is likely to negatively impact this biotope, although evidence suggests that changes to tidal currents and tidal amplitude with sea-level rise will be site-specific, and cannot be evaluated on a UK-wide basis. Therefore, under the available evidence, resistance to sea-level rise has been assessed as ‘High’ for both the middle (50 cm) and high (70 cm) emission scenario, and the extreme scenario (107 cm). As no recovery is deemed necessary, resilience has been assessed as ‘High’, and therefore this biotope has been classified as ‘Not sensitive’ to sea-level rise at each of the benchmarks albeit with ‘Low’ confidence.

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

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006).  This biotope is recorded between 0 – 20 m depth in the UK, although Ostrea edulis can be found at depths of up to 50 m (OSPAR, 2008).  Therefore, an increase in depth of between 50 – 107 cm is unlikely to have large implications for this species.

Ostrea edulis beds occur on shallow sublittoral muddy mixed sediment in sheltered environments with weak tidal streams (JNCC).  Understanding of how sea-level rise will affect exposure or tidal energy is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. The effects of sea-level rise and increased wave action may be increased further due to storms and storm 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 (Mossman et al., 2015, Lowe et al., 2018, Palmer et al., 2018).

Sensitivity assessment. This habitat occurs from 0 - 20 m depth, although Ostrea edulis beds can be found at depths of up to 50 m. Any change to the habitat in terms of its exposure or tidal currents is likely to negatively impact this biotope, although evidence suggests that changes to tidal currents and tidal amplitude with sea-level rise will be site-specific, and cannot be evaluated on a UK-wide basis. Therefore, under the available evidence, resistance to sea-level rise has been assessed as ‘High’ for both the middle (50 cm) and high (70 cm) emission scenario, and the extreme scenario (107 cm). As no recovery is deemed necessary, resilience has been assessed as ‘High’, and therefore this biotope has been classified as ‘Not sensitive’ to sea-level rise at each of the benchmarks albeit with ‘Low’ confidence.

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

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006).  This biotope is recorded between 0 – 20 m depth in the UK, although Ostrea edulis can be found at depths of up to 50 m (OSPAR, 2008).  Therefore, an increase in depth of between 50 – 107 cm is unlikely to have large implications for this species.

Ostrea edulis beds occur on shallow sublittoral muddy mixed sediment in sheltered environments with weak tidal streams (JNCC).  Understanding of how sea-level rise will affect exposure or tidal energy is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. The effects of sea-level rise and increased wave action may be increased further due to storms and storm 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 (Mossman et al., 2015, Lowe et al., 2018, Palmer et al., 2018).

Sensitivity assessment. This habitat occurs from 0 - 20 m depth, although Ostrea edulis beds can be found at depths of up to 50 m. Any change to the habitat in terms of its exposure or tidal currents is likely to negatively impact this biotope, although evidence suggests that changes to tidal currents and tidal amplitude with sea-level rise will be site-specific, and cannot be evaluated on a UK-wide basis. Therefore, under the available evidence, resistance to sea-level rise has been assessed as ‘High’ for both the middle (50 cm) and high (70 cm) emission scenario, and the extreme scenario (107 cm). As no recovery is deemed necessary, resilience has been assessed as ‘High’, and therefore this biotope has been classified as ‘Not sensitive’ to sea-level rise at each of the benchmarks albeit with ‘Low’ confidence.

Temperature increase (local)

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

Evidence

Filtration rate, metabolic rate, assimilation efficiency and growth rates of adult Ostrea edulis increase with temperature. Clearance rate was also shown to increase in large oysters was temperature increased (Fabra et al., 2025). Growth was predicted to be optimal at 17°C or, for short periods, at 25°C (Korringa, 1952; Yonge, 1960; Buxton et al., 1981; Hutchinson & Hawkins, 1992).  

Spärck's (1951) data suggest that temperature is an important factor in recruitment of Ostrea edulis, especially at the northern extremes of its range. With warm summers having resulted in good recruitment (Korringa, 1952). Water temperature has been shown to influence the timing of spawning as larval release is initiated once the temperature has risen to around 15 or 16°C, although local adaptation is likely (Korringa, 1952; Yonge, 1960). The number of degree-days (sum of degrees per day above 7°C) was the most significant predictor of peak larval abundance in Loch Ryan, Scotland, which occurred at 617 degree-days (Chapman et al., 2021), and at two locations in the south-west Netherlands after 576 degree-days and 1,100 degree-days (Maathuis et al., 2020). This is an important consideration for restoration efforts that intend to add shell cultch to the environment to provide additional settlement substrata, as it allows a predictable window of when highest larval settlement is likely. Davis & Calabrese (1969) reported that larvae grew faster with increasing temperature and that survival was optimal between from 12.5 and 27.5°C but that survival was poor at 30°C. Subsequent studies support this by demonstrating that Ostrea edulis larvae growth rate and settlement success increased in higher temperatures of 25 to 30°C (Robert et al., 2017; Alter et al., 2024; Jung et al., 2025). However, Prado et al. (2016) contradicted these findings by showing that larval stages appeared less tolerant of high temperatures and suffered a significant decrease in survival at ≥ 22°C. Increases in temperature from climatic changes may improve recruitment in this species, however, adults are known to perform better in cooler temperatures.

Huchinson & Hawkins (1992) noted that temperature and salinity were co-dependent, so that high temperatures and low salinity resulted in marked mortality, with no individuals surviving more than 7 days at 16 psu and 25°C, although these conditions rarely occur in nature. A 40-day experiment which exposed Ostrea edulis to various temperature and salinities further demonstrated that this species became less tolerant to lower salinity as temperatures increased (Pereiro et al., 2025). At 5°C, 100% mortality occurred for oysters in salinities of ≤ 11 after around 27 days. At 10°C, mortality occurred more quickly, with all oysters exposed to salinity of ≤ 11 dying by day 23. At 15°C and 20°C, mortality occurred by day three, with 90% of mortality in salinity of ≤ 11 occurring by day 14. At 25°C, oysters were no longer able to survive in salinities below 23, with over 80% mortality within 15 days (Pereiro et al., 2025). Therefore, this species might be less resistant to increases in water temperature in areas susceptible to freshwater inputs such as Galway Bay, west coast of Ireland (Pereiro et al., 2025).

Under normal salinity regimes, experimental studies have demonstrated that increased temperature can negatively impact the physiological functioning of Ostrea edulis. Oysters were shown to survive in 30°C for one week (Kamermans & Saurel, 2022), however, the condition of oysters (i.e. the ratio of tissue weight to shell weight) decreased as temperature increased from 3 to 30°C. Zapata-Restrepo et al. (2019) also showed that tissue biomass to shell ratio reduced with increasing temperature for this species (from 10 to 18°C). Year-old Ostrea edulis from the Mediterranean were exposed to 21°C (control), 25°C and 28°C for 26 days (Georgoulis et al., 2024). At 25°C and 28°C, the expression of heat shock proteins hsp70 and hsp90, and antioxidant genes were significantly higher compared to at 21°C, even after just four hours of exposure to that temperature. These levels remained significantly elevated throughout the 26-day experiment, indicating both thermal and oxidative stress as a result of temperature over 25°C. At 28°C, oysters moved to anaerobic metabolism to cope with physiological stress (Georgoulis et al., 2024). In addition, Eymann et al. (2020) showed that after 40 hours at 26°C, gill tissue of adult oysters became partly anaerobic and cardiac dysfunction occurred at 28°C. Lethal temperature was around 36°C when all oysters died within 18 hours.

A study comparing how Ostrea edulis responds to rapid and gradual warming found that oysters could tolerate higher temperatures when the heat increased quickly (Götze et al., 2025). Under rapid warming (2°C per hour), 50% mortality occurred by around 37°C, whereas, under gradual warming (2°C per day), 50% mortality occurred by around 35°C. During rapid warming, oysters shifted to anaerobic metabolism shortly after 30°C, and by 36°C cardiac activity had begun to fail, leading to mortality. Under gradual warming, oysters had more time to make metabolic adjustments and therefore delayed the onset of anaerobic metabolism, allowing them to survive for longer periods, but only to lower temperatures. Thus, rapid warming elevated the maximum temperature tolerated but reduced survival time, while gradual warming extended survival time but reduced maximum thermal limits (Götze et al., 2025). These studies, therefore, suggest that the lethal temperature for this species may be between 34°C and 36°C (Eymann et al., 2020; Götze et al., 2025) depending on whether the warming is rapid or gradual.  However, Kinne (1970) reported that gill tissue activity did not reach zero until between 40 and 42°C, although values derived from single tissue studies should be viewed with caution.

Ostrea edulis are protrandric hermaphrodites and switch from male to female, and back again to male within the reproductive season (Bayne, 2017). Warmer water temperature has been shown to influence the sex ratio of this species by increasing the proportion of males to females. Eagling et al., (2017) demonstrated this effect from populations at two sites: Loch Ryan, Scotland, where water temperature ranged between 4.5 to 18.1°C, and Chichester Harbour, England, where temperature remained greater than 10°C for most of the year and consistently above 16°C. They found that, generally, the ratio in the cooler Loch Ryan was up to 2:1 (males to females), however, after a temperature spike, sex ratios became significantly more skewed towards males. Whereas the warmer Chichester Harbour population had a more consistent sex ratio of 3:1 (males to females). They determined that the temperature threshold after which the proportion of males shifted significantly was around 16.5°C in Loch Ryan, and 17.5°C in Chichester Harbour (Eagling et al., 2017). Zapata-Restrepo et al. (2019) also investigated whether temperature affected sex determination as well as gametogenesis. Over four months, oysters were kept at either 10, 14 or 18°C. By the end of this time, there was a higher proportion of females (4:1) at 10°C compared to a higher proportion of males at 14°C (up to 2:1 males: females). Unusually, at 18°C, females dominated again. The authors suggested that, because they were fed frequently, this allowed the oysters to undergo energetically costly female gametogenesis even at higher temperature. Usually, species that undergo sex changes like Ostrea edulis would be expected to become male at higher temperatures given the reduced energetic demands (Pérez et al., 2013, Wright, 1988, cited in Zapata-Restrepo et al., 2019). In addition, gametogenesis occurred faster in Ostrea edulis with increasing temperature, demonstrating that temperature influenced gamete production as well as sex ratios in this species (Zapata-Restrepo et al., 2019). Should temperatures increase, it could skew the population structure significantly toward male dominance, potentially reducing fertilization efficiency within a reproductive season, and hampering population persistence or recovery.

Ostrea edulis and many of the other species in the biotope occur from the Mediterranean to the Norwegian coast and are unlikely to be adversely affected by long-term changes in temperatures in Britain and Ireland. Most of the other characterizing species within the biotope have a wide distribution in Europe suggesting that they are able tolerate a wider range of temperatures than those found in British waters. Delicate species may not be so tolerant, and mobile species may leave the biotope temporarily resulting in a decline in species richness.

Sensitivity assessment. When coupled with other stressors like lowered salinity, the survival of Ostrea edulis was significantly reduced in higher temperatures. However, this biotope is found in fully marine conditions and so increases in temperature may not be exacerbated by reduced salinity unless the biotope experienced localised increased freshwater runoff. In fully marine conditions typical of this biotope, Ostrea edulis survived at 25°C but with thermal and oxidative stress, loss of body conditioning, and a reduced ability to tolerate lower salinities. Studies identified 34 to 36°C as the lethal temperature for this species due to metabolic crash and total mortality within 18 hours. Increases in temperature alters sex ratios towards male dominance during reproduction, and increases at the benchmark level may result in a more highly skewed sex ratio, thus reducing fertilization success within a spawning period. Nevertheless, the evidence suggests that the native oyster would survive a short-term increase by 5°C or long-term increase of 2°C for a year. Hence, resistance is assessed as 'High', resilience as 'High' and sensitivity 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

Hutchinson & Hawkins (1992) suggested that Ostrea edulis switched to a reduced winter metabolic state below 10°C that enabled it to survive low temperatures and low salinities encountered in shallow coastal waters around Britain. Kamermans & Saurel (2022) showed that Ostrea edulis from the Netherlands survived at 3°C for up to 28 days (end of experiment), and Korringa (1952) reported that British, Dutch and Danish oysters could withstand even lower temperatures of 1.5°C for several weeks. However, heavy mortalities of native oyster were reported after the severe winters of 1939/40 (Orton, 1940) and 1962/63 (Waugh, 1964). The 1962/63 winter lasted from late December 1962 until late February 1963 during which temperatures were 5 to 6°C below seasonal averages. Mortality was attributed to relaxation of the adductor muscle so that the shell gaped, resulting in increased susceptibility to low salinities or to clogging with silt. Reduced temperatures may also lead to lower clearance and feeding rates (Fabra et al., 2025). In a laboratory study, the clearance rate of large oysters (> 70 mm) was significantly lower at 8°C (in March) compared to 10°C and 12°C (in April and May, respectively). Lowest average clearance rate was 0.27 litres per hour (8°C), compared to 0.77 and 0.61 L at 10°C and 12°C, respectively (Fabra et al., 2025).

Low temperatures and cold summers are correlated with poor recruitment in Ostrea edulis, presumably due to reduced food availability and longer larval developmental time, especially at the northern limits of its range. Davis & Calabrese (1969) noted that larval survival was poor at 10°C.

The severe winters of 1939/40 and 1962/63 (Orton, 1940; Waugh, 1964) also resulted in the death of associated fauna, e.g. Sabella pavonina and other polychaetes died in great numbers, Crepidula fornicata incurred about 25% mortality and Ocenebra erinacea died in large numbers, while only small Carcinus maenas remained on the beds (Orton, 1940; Waugh, 1964). However, starfish, crabs such as Hyas araneus and Urosalpinx cinerea, and Ascidiella aspersa were little affected (Orton, 1940; Waugh, 1964).

Mobile predatory species found within this biotope, such as fish and crabs, probably migrate further offshore in winter months, reducing predation pressure. Changes in the average summer temperature may have significant effects on recruitment. In addition, Spärck (1951) noted marked changes in the populations of Ostrea edulis in the Limfjord, Denmark, between 1852 and 1949. In periods of poor recruitment and the absence of fishing pressure, populations gradually declined, becoming restricted to the most favourable areas of the Limfjord. In some areas there was a 90% decrease in stock. Temperature was probably the most important controlling factor in recruitment in the Limfjord population (Spärck, 1951).

Sensitivity assessment. The decreases in temperature experienced in the 1962/63 severe winter were more extreme than the pressure benchmark (5°C for one month) but suggest that seasonally low temperatures could result in mortality. In addition, long-term decreases in temperature could potentially affect overall recruitment of Ostrea edulis. Other members of the community may be intolerant of short-term acute decreases in temperature. Resistance is assessed as ‘Medium’, and resilience has been assessed as ‘Medium’ resulting in the sensitivity of this biotope being ‘Medium’ to the pressure at the benchmark.

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

Ostrea edulis is usually found in salinity of 25 to 40 psu (OBIS, 2026), but also exists in the Mar Menor, Spain, a hypersaline lagoon with an average salinity between 42 and 45 psu (Hernandis et al., 2025).

To determine the extreme salinity tolerance of this species, oysters from the Mar Menor population were exposed to an acute (48 hours) or elongated (one month) period of salinity, of either 20, 30, 40, 50 or 60 psu (Hernandis et al., 2025). In the acute exposure, at the extreme salinities (20 and 60 psu), clearance rate was too low to be measured, and at 20 psu oxygen consumption was almost zero and very little feeding occurred. This species is euryhaline and had a positive scope for growth (i.e. enough energy to grow) in salinities of 20 to 50 psu in the long-term exposure treatment. However, after one month exposure to 60 psu, clearance rate was too low to be measured despite similar oxygen consumption as the control, which suggests that the oysters were prioritizing energy resources towards maintaining osmotic balance (Hernandis et al., 2025). Despite no mortalities, physiological functioning was severely reduced for this species at 60 psu, potentially indicating an upper salinity tolerance for Ostrea edulis (Hernandis et al., 2025).

Sensitivity assessment. Ostrea edulis is considered a euryhaline species as it is known to survive within a variety of salinities. Individuals from the Mar Menor population are able to withstand hyper-saline conditions up to 60 psu, but this may not be the case universally. Nevertheless, this biotope is found subtidally in full salinity waters (30 to 35 ppt) and is unlikely to experience hyper-saline effluent that would increase salinity to this extreme. Therefore, this biotope is considered to have ‘High’ resistance to increases in salinity, making its resilience ‘High’ by default and the biotope ‘Not Sensitive’, albeit with ‘Low’ confidence due to the limited number of studies available.

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

Ostrea edulis is euryhaline and colonizes estuaries and coastal waters exposed to freshwater influence (Yonge, 1960). It was previously thought that this species could not withstand salinities below 23 psu (Yonge, 1960). However, more recent evidence suggests they have some survival capacity, but with reduced physiological functioning. In addition, the tolerance of Ostrea edulis to decreases in salinity appears to depend on temperature. Hutchinson & Hawkins (1992) noted that scope for growth was severely affected below 22 psu, likely because the oyster's valves were closed, but that 19 to 16 psu could be tolerated if the temperature did not exceed 20°C. At 25°C animals did not survive more than seven days at 16 psu. Hutchinson & Hawkins (1992) noted that at low temperatures (10°C or less) the metabolic rate was minimal. This may help Ostrea edulis survive in low salinities associated with storm runoff. Pereiro et al. (2025) also demonstrated that Ostrea edulis were less tolerant to low salinities as temperature increased. At 5°C, 100% mortality occurred for oysters in salinities of ≤ 11 after around 27 days. At 10°C, mortality occurred more quickly, with all oysters exposed to salinity of ≤ 11 dying by day 23. At 20°C, mortality occurred by day three, with almost 100% mortality occurring by day 10 in salinity of ≤ 11, and all oysters dying in salinity of ≤ 17 occurring by day 30. At 25°C, oysters were no longer able to survive in salinities below 23, with over 80% mortality within 15 days (Pereiro et al., 2025). Therefore, this species might be less resistant to increases in water temperature in areas susceptible to a series of freshwater inputs, such as Galway Bay, west coast of Ireland (Pereiro et al., 2025).

To determine the extreme salinity tolerances of this species, oysters from the Mar Menor population were exposed to an acute (48 hours) or elongated (one month) period of salinity, of either 20, 30, 40, 50 or 60 psu (Hernandis et al., 2025). In the acute exposure, at the lowest measured salinity (20 psu), clearance rate was too low to be measured, oxygen consumption was almost zero, and very little feeding occurred. However, in the elongated exposure, the physiological function of the oysters in 20 psu was not different from those at 30 or 40 psu and there remained a positive scope for growth (i.e. enough energy to grow) (Hernandis et al., 2025). This study, therefore, suggested that Ostrea edulis has some ability to acclimate to 20 psu after an initial decline in physiological functioning within the first 48 hours (Hernandis et al., 2025).

The duration and extent to which Ostrea edulis open their valves determine the rate at which they breathe and feed, and in closing their gape, protect themselves from predators and poor-quality water. Decreasing salinity from 33.6 to 18.2 caused valve closure in oysters such that by a salinity of 20.5, full valve closure occurred (Bamber, 2023).

Several of the characterizing species in this biotope are commonly found in estuarine and full salinity waters and are probably tolerant of reduced salinity. For example, Lanice conchilega and Ascidiella aspersa will tolerate salinities as low as 18 psu (Fish & Fish, 1996). However, this biotope has only been recorded from full salinity habitats, therefore, a proportion of the epifauna and infauna may not tolerate a reduction in salinity and may be lost. Predatory starfish and other echinoderms are generally not able to tolerate low salinity and may be excluded.

Sensitivity assessment. Ostrea edulis may be able to tolerate short-term acute reductions in salinity due to freshwater input. However, a decrease in the salinity regime for an extended period is likely to have a negative impact on the biotope. Ostrea edulis can withstand lower salinities when temperatures are lower, however, mortality occurs faster and at higher salinities when temperatures increase. At 20°C and below, practically no mortality occurred at 29 and 23 psu, but there was marked mortality of oysters in ≤ 17 psu (Pereiro et al., 2025). This biotope is found only in fully marine conditions (30 to 40; Connor et al., 2004) which suggests that it would not survive in a reduced salinity regime (18 to 30). Therefore, resistance has been assessed as ‘Medium’ and resilience as ‘Medium’. Giving the biotopes a sensitivity of ‘Medium’ to the pressure at the benchmark. However, exposure to storm runoff and thermal stress due to hot weather may be more detrimental.

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

This biotope occurs in weak (< 0.5 m/s) to very weak (negligible) tidal streams. An increase in water flow above that of the pressure benchmark, for example from weak to strong, is likely to remove (erode) fine particulates, leaving coarser substrata and making more hard substratum available for settlement by oysters and other members of the community, e.g. Ascidiella sp. and epifauna. An increase in water flow rate could cause oysters to be swept away by strong tidal flow if the substratum to which they are attached is removed. Therefore, a proportion of the oyster bed may be lost, depending on the nature of the substratum.

Increased water flow can affect the ability of oysters to feed and reduce the time they are able to feed. Yet could improve the availability of suspended particles on which oysters feed. The former is thought to affect the biotope more significantly whilst the latter the individual species. With increased water flow rate the oyster filtration rate increases, up to a point where the oysters are unable to remove more particles from the passing water and thus individual species are likely to benefit from increased water flow rate. Maniero et al. (2020) demonstrated that mortality in broodstock oysters was 10, 20 and 50 to 60% in 1.0, 2.0 and 3.0 L/hour/oyster water flows, respectively, likely due to reduced clearance rate and feeding efficiency (Maniero et al., 2020). This study also reported that higher flow rates coincided with increases in Vibrio and heterotrophic bacteria in the tissue and intervalvular fluid of the oysters (Maniero et al., 2020).

Reproductive success and successful recruitment to an oyster bed may also be affected by a change in water flow. Recruitment is already known to be sporadic and dependent on the hydrographic regime and local environmental conditions but will be enhanced by the presence of adults and shell material (Cole, 1951). An increase in water flow rate may interfere with settlement of spat and it is thought that growth rates of Ostrea edulis are faster in sheltered sites than exposed locations, although this is thought to be attributed to the seston volume rather than flow speed or food availability (Valero, 2006).

Sensitivity assessment. A change in water flow at the benchmark level (a change of 0.1 to 0.2 m/s) is unlikely to cause any effect on this biotope. However, an increase above the benchmark of this pressure could have a negative impact. Both the resilience and resistance of this biotope are assessed as ‘High’, which results in the biotope being assessed as ‘Not sensitive’.

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

Beds of the native oyster Ostrea edulis may occur low on the shore and are exposed for a proportion of the tidal cycle.  Ostrea edulis is known to be able to survive aerial exposure at low temperatures during storage and are known to be capable of anaerobic respiration (Korringa, 1952; Yonge, 1960), which suggests that they can tolerate aerial exposure.  In addition, in the mariculture of oysters (native and introduced species) oyster trays are positioned in the low intertidal, and regularly exposed to the air.  Therefore, an increase in desiccation in this biotope, is unlikely to result in death of the oysters themselves at the level of the benchmark.  However, exposure to the air prevents feeding, and anaerobic respiration usually results in an oxygen debt, an energetic cost that the organism must make up on return to aerated water, resulting in reduced growth and reproductive capacity.

The associated epifauna may be more intolerant, such as ascidians (e.g. Ascidiella spp.) and Asterias rubens.  Burrowing infauna are likely to be protected from desiccation by their infaunal habit and species such as Lanice conchilega, Myxicola infundibulum and Chaetopterus variopedatus may be found at low water.  Mobile epifaunal species would probably move to deeper water while delicate hydroids and bryozoans may be damaged or killed by desiccation.

This biotope is subtidal so that an increase in emergence is unlikely to have an adverse effect on the community.  However, increased emergence may allow the oyster bed to spread further up the shore, although at a slow rate.  Therefore, the biotope may benefit from the factor.

Sensitivity assessment.  Oyster beds may resist an increase in desiccation at the benchmark level.  Both resistance and resilience are assessed as ‘High’, giving the biotope a ‘Not sensitive’ assessment at the level of the benchmark.

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

SS.SMx.IMx.Ost is found in sheltered to extremely sheltered conditions.  This biotope is found from 0 to 20 m in depth.  The shallow wave action in shallow water results in oscillatory water flow, the magnitude of which is greatest in shallow water and attenuated with depth.  While the oysters' attachment is permanent, increased wave action may result in erosion of its substratum and the oysters with it.  Areas where sufficient shell debris has accumulated may be less vulnerable to this disturbance.  However, a proportion of the bed is likely to be displaced by an increase in wave action.  Similarly, infaunal species, burrowing polychaetes and epifauna are characteristic of wave sheltered conditions and may be lost, e.g. Ascidiella sp.  The biotope may be replaced by communities characteristic of stronger wave action and coarser sediments. 

Sensitivity assessment.  At the benchmark of this pressure, it is highly unlikely that the change will cause any effect on this biotope. However, an increase above the benchmark of this pressure could have a negative impact.  Both the resilience and resistance of this biotope are assessed as ‘High’, which results in the biotope being classified as ‘Not sensitive’.

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 results of the Rapid Evidence Assessment on the effects of 'Transitional metal or organometal' contaminants on oyster species (Crassostrea spp., Ostrea spp., Saccostrea spp. and Magallana gigas). are summarized below. The full 'Oyster species evidence review' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Transitional metal or organometal' contaminant examined, together with an overall pressure assessment. 

Transitional metals. In adults and juveniles, 12% of the results reported ‘severe’ mortality and 33% reported ‘significant’ mortality but 43% reported sublethal effects.  However, ‘severe’ mortality was reported in 31% of the results from early life stages, and ‘significant’ mortality in 57% of the results.  There is considerable evidence to suggest that exposure to copper, cadmium, zinc, silver, mercury and lead could result in ‘severe’ or ‘significant’ mortality, although experimental designs and exposure concentrations vary.  Several other metals were only included in a few studies. Overall, all life stages were reported to experience mortality after exposure to transitional metals, with the exception of cobalt.  The evidence agrees with His et al. (2000) who ranked the toxicity of metals and organometals to bivalve larvae as follows: tributyltin >mercury >silver >copper >zinc >nickel >lead >cadmium.  Therefore, resistance to transitional metals exposure in adult, juvenile, and early life stages of oyster species is assessed as ‘None’, resilience as ‘Very low’ and sensitivity as ‘High’.

Curiously, CCA (copper-chrome-arsenic) affected the swimming of Crassostrea gigas larvae swimming but did not result in mortality when used as an antifoulant mixture (Prael et al., 2001).  Similarly, copper-oxide paint was reported to have only sublethal effects on adult C. gigas (His et al., 1987).

Ostrea edulis was the least studied oyster species.  The adults and juveniles of Ostrea edulis were reported to experience ‘significant’ mortality after exposure to nickel, mercury, and chromium.  The early life stages of Ostrea edulis were only studied in one article that only examined the effects of mercury, which in turn resulted in ‘significant’ mortality.  Bryan (1984) reported a 48-hour LC50 for Hg of 1-3.3 ppb in Ostrea edulis larvae compared with a 48-hour LC50 for Hg of 4,200 ppb in adults.

Ostrea edulis was able to survive in the lower reaches of Restronguet Creek, one of the most heavy metal polluted estuaries in the world, where metals from mining wastes reached concentrations several orders of magnitude above normal (Bryan et al., 1987).  Bryan et al. (1987) noted that O. edulis in the creek were highly polluted, that is, specimens were reported to be 'green' since the 1880s, in 1927, and in specimens collected in 1980 (Bryan et al., 1987).  Ostrea edulis from the Falmouth estuary were shown to be able to detoxify metals (Cu and Zn) in amoebocytes.  Bryan et al. (1987) noted that Cu and Zn were accumulated in the tissues of Ostrea edulis, with estimates ranging from ca 1,000 to ca 16,500 µg/g dry weight.  Bryan et al. (1987) concluded that this detoxification mechanism allowed O. edulis to survive in the lower reaches of the creek.

Ostrea edulis is, therefore, resistant to high levels of Cu and Zn and is able to survive in the lower reaches of Restronguet Creek, where other species are excluded by the heavy metal pollution.  Larval stages may be less resistant, but larval recruitment must be high enough for a population of oysters to survive for ca 123 years in the lower reaches of Restronguet Creek and the Falmouth estuaries.  Bryan et al. (1987) do not clarify their abundance/density in the creek.  However, it appears that Ostrea edulis is capable of localized adaption to transitional metal contamination, in particular, from copper and zinc. Although the adult Ostrea edulis may be tolerant of heavy metal pollution the larval effects suggest that recruitment may be impaired resulting in a reduction in the population over time, and hence a reduction in the associated fauna.

Therefore, the resistance of Ostrea edulis to transitional metals is assessed as ‘Low’, with the possible exception of Cu and Zn and the understanding that long-term exposure could result in localised adaption.  Hence, resilience is assessed as ‘Low’ and sensitivity as ‘High’, albeit with ‘Low’ confidence. 

Little information on the tolerance of ascidians or sponges was found. However, polychaetes are thought to be relatively tolerant of heavy metal pollution, even though some heavy metals may suppress reproduction (Bryan, 1984). Similarly, Bryan (1984) suggested that adult gastropod molluscs were also relatively tolerant of heavy metal pollution. Therefore, most other characteristic species in this biotope may be relatively tolerant of heavy metal pollution.

Organometals. Overall, 12% of the results of exposure to organotins reported ‘severe’ mortality, 49% ‘significant’ mortality, 6% ‘some’, 8% no mortality and 24.5% sublethal effects (Table 4.1).  In adults and juveniles, ‘severe’ mortality was reported in 6% of results, ‘significant’ in 31% and sublethal in 43% of results. However, in early life stages, 11% of the results of exposure to organotins reported ‘severe’ mortality, but 66.7% ‘significant’ mortality, 16.7% no mortality and 5.5% sublethal effects.  The evidence suggests that early life stages are more sensitive than adults.  In the five studies that examined Ostrea edulis, organotin exposure was reported to result in ‘significant’ or some ‘mortality’.  Therefore, the resistance of oyster species to organotins is assessed as ‘None’, resilience as ‘Very low’ and sensitivity as ‘High’. 

Rees et al. (2001) suggested that TBT contamination may have locally reduced population sizes of Ostrea edulis. In Ostrea edulis, TBT has been reported to cause reduced growth of new spat at 20 ng/l, a 50% reduction in growth at 60 ng/l. Although older spat grew normally at 240 ng/l for 7 days, larval production in adults was prevented by exposure to 240 and 2,620 ng/l for 74 days (Thain & Waldock, 1986; Bryan & Gibbs, 1991). Adults bioaccumulate TBT. Thain & Waldock (1986) and Thain et al. (1986) noted that TBT retarded normal sex change (male to female) in Ostrea edulis.

TBT also has marked effects on other marine organisms. For example, TBT causes imposex in prosobranch gastropods, especially the neogastropods such as Nucella lapillus, Ocenebra erinacea and Urosalpinx cinerea resulting in markedly reduced reproductive capacity and population decline. Ascidian larval stages were reported to be intolerant of TBT (Mansueto et al., 1993 cited in Rees et al., 2001). Beaumont et al. (1989) investigated the effects of tri-butyl tin (TBT) on benthic organisms. At concentrations of 1-3 µg/l there was no significant effect on the abundance of Hediste diversicolor or Cirratulus cirratus (family Cirratulidae) after 9 weeks in a microcosm. However, no juvenile polychaetes were retrieved from the substratum and hence there is some evidence that TBT had an effect on the larval and/or juvenile stages of these polychaetes. No information concerning the polychaetes characteristic of this biotope was found. Surveys of the Crouch estuary suggested that benthic epifauna were recovering since a reduction in TBT contamination suggesting that populations of several epifaunal species, including Ascidiella sp., had previously been reduced (Rees et al., 1999; 2001).

While the loss of predatory neogastropods (which are particularly intolerant of TBT) may be of benefit to Ostrea edulis populations, TBT has been shown to reduce reproduction and the growth of spat. Rees et al. (1999; 2001) reported that the epifauna of the inner Crouch estuary had largely recovered within 5 years (1987-1992) after the ban on the use of TBT on small boats in 1987. Increases in the abundance of Ascidiella sp. and Ciona intestinalis were especially noted. Ostrea edulis numbers increased between 1987 -1992 with a further increase by 1997. However, they noted that the continued increase in Ostrea edulis numbers and the continued absence of neogastropods suggested that recovery was still incomplete at the population level.

Nanoparticulate metals. Short-term (2-6 hours) exposure of Crassostrea virginica to titanium dioxide (TiO₂) did not result in negative effects. However, 48-hour exposure of C. virginica embryos to silver nanoparticles significantly impaired development (Ringwood et al., 2010). Therefore, the resistance of oyster species to TiO₂ is assessed as ‘High’ but exposure to silver nanoparticulates may be ‘Low’.  Hence, resilience is assessed as ‘Low’ and sensitivity as ‘High’ but with ‘Low’ confidence due to the lack of evidence.

Overall sensitivity assessment. The evidence suggests that oysters are highly sensitive to transitional metal (except cobalt) and organometal (organotin) exposure depending on the concentration and duration of exposure and the life stage of the oysters. The evidence on the effects of nanoparticulate metals is limited to two studies but reported larval sensitivity to nanoparticulate silver.  Therefore, the worst-case resistance of native oyster (Ostrea edulis) beds to transitional metals and organometals is assessed as 'Low'. Hence, resilience is assessed as 'Low' and sensitivity as 'High'. Ostrea spp. was the least studied species in the evidence review but as several studies provided direct evidence confidence in the assessment is 'Medium'. 

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 results of the Rapid Evidence Assessment on the effects of 'Hydrocarbons and PAH' contaminants on oyster species (Crassostrea spp., Ostrea spp., Saccostrea spp. and Magallana gigas). are summarized below. The full 'Oyster species evidence review' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Hydrocarbon' contaminant examined, together with an overall pressure assessment. 

Oil spills. Only two studies reported on the direct effect of oil spills on oyster beds.  Levings et al. (1994) reported that the Galeta oil spills resulted in a significant reduction in C. virginica beds along the mangrove fringe, which lasted at least five years.  Powers et al. (2017) reported a ‘severe’ reduction in the abundance of C. virginica in intertidal beds after the Deepwater Horizon spill.  However, any effect on subtidal beds was obscured by mass mortality caused by freshwater runoff.  Therefore, the evidence suggests that direct oiling of oyster beds could cause ‘severe’ or ‘significant’ mortality amongst the oysters.  Hence, resistance is assessed as ‘None’, resilience as ‘Very low’ and sensitivity as ‘High’ for oysters as a group.  However, evidence of the direct effect of oiling on Ostrea edulis was not found.  In the UK, Ostrea edulis beds occur in the shallow subtidal (0-20 m) and rarely in shallows exposed at low tide. But in sheltered areas, oil is likely to persist, and reach the shallow sea bed adsorbed to particulates or in solution. Therefore, the resistance of Ostrea edulis beds is assessed as ‘Medium’ to represent the chance that the most shallow extent of the biotopes might be exposed to an oil spill that coincided with the lowest tides.  Hence, resilience is assessed as ‘Medium’ and sensitivity to oil spills as ‘Medium’ but with ‘Low’ confidence. 

Petroleum hydrocarbons – oils and dispersed oils.  The effects of various petroleum hydrocarbons as oils (e.g. crude, fuel, and diesel) and dispersed oils (e.g. chemically enhanced water accommodated fractions, CEWAFs) were examined by 31 separate articles on adult, juveniles, early life stages and gametes of oyster species. There was considerable variation in the types of oil or oil and dispersant mixtures studied, experimental design and, hence, the results. For example, 3.6% of the results from the studies of the effects of complex hydrocarbons (crudes oils, WAF/WSF/HEWAF) on oysters reported ‘severe’ mortality, 47% reported ‘significant’ mortality, 4.8% ‘some’, 9.6% no mortality and 35% reported only sublethal effects.  However, if only early life stages are included, 5.4% report ‘severe’ mortality, but 65% report ‘significant’ mortality, 7% ‘some’, 14% no mortality and only 7% report only sublethal effects. The effects of dispersed oils (dispersant and oil mixtures) are similar. For example, 72% of the results of the effects of dispersed oils reported ‘severe’ or ‘significant’ mortality but 100% of the results from early life stages reported ‘severe’ or ‘significant’ mortality. However, only one article examined the effects of dispersed oil on adults and reported only sublethal effects. Similarly, only two articles examined the effects of oils on oysters, both of which reported sublethal effects.

His et al. (2000) also noted that oils, detergents, and their mixtures were usually toxic to the early life stages of bivalves at concentrations in the order of 1 ppm or higher.  His et al., 2000 also noted that refined oils were more toxic than crude oils but rarely at environmentally realistic concentrations. Oils and detergents also inhibited settlement in Crassostrea virginica while oil inhibited settlement in Ostrea edulis at 1-2 ppt (Renzoni, 1973b; Smith & Hackney, 1989; His et al., 2000).

Therefore, the resistance of the early life stages of oyster species to oils (WAF/WSF/HEWAFs) and dispersed oils is assessed as ‘None’.  Although limited evidence suggests that the adults may be resistant, we may assume that loss of larval stages would result in a decline in resident populations that are dependent on recruitment for their abundance (Bayne, 2017).  No evidence of the effects of oils and dispersed oils on Ostrea edulis was found.  Hence, the resistance of oyster beds is assessed as ‘Low’ to represent the resultant loss in annual recruitment and potential population decline.  Therefore, resilience is assessed as ‘Low’ and sensitivity as ‘High’ but with ‘Medium’ confidence due to the lack of evidence from adult populations.

Dispersants. The majority (72%) of the results of the effects of dispersants on oyster species reported ‘significant’ mortality, 5.5% ‘severe’ mortality, 5.5% no mortality and 16.7% reported only sublethal effects.  The proportion of ‘severe’ and ‘significant’ mortality results increase to 7.7% and 80.7% respectively in early life stages.  However, none of the results from adults and juveniles reported ‘severe’ mortality, but 66% reported ‘significant mortality and 33% reported sublethal effects.  Therefore, it appears that dispersants alone are more toxic to oyster species as adults, juveniles, or early life stages than complex hydrocarbons and dispersed oils.  This conclusion is consistent with the finding of Woelke (1972; cited in His et al., 2000).  Therefore, the resistance of oyster species to dispersants is assessed as ‘None’, resilience as ‘Very low’ and sensitivity as ‘High’.  However, Ostrea edulis may be an exception as only ‘significant’ mortality was reported in this species.  Hence, the resistance of Ostrea edulis and its beds to dispersants is assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’.

Polyaromatic hydrocarbons (PAHs). The results of exposure of oyster species to PAHs were split evenly between ‘severe’ and sublethal effects, albeit based on only 12 articles.  However, early life stages were more sensitive, with 64% of the results reporting ‘severe’ mortality’, 18% ‘significant and only 9% either no mortality or sublethal effects.  PAH exposure was reported to result in reduced scope for growth in adults, reduced sperm motility and reduced fertilization rate and abnormal larval development (His et al., 1997; Jeong & Cho, 2005; Choy et al., 2007; Kim et al., 2007; Wessel et al., 2007; Nogueira et al., 2017; Xie et al., 2017b).  The toxicity of PAHs also increased in light (UV exposure) (Lyons et al., 2002).  Therefore, the resistance of oyster species to PAHs is assessed as ‘None’, resilience as ‘Very low’ and sensitivity as ‘High’.  However, as no direct evidence of the effects of PAHs on Ostrea spp. was found, confidence is assessed as 'Low'. 

Nonylphenol.  Nice et al. (2000; 2003) examined the effect of nonylphenol on Crassostrea gigas larvae or gametogenesis (Nice, 2005).  Nonylphenol reduced sperm production, caused hermaphroditism in some specimens after 48-hour exposures, and had transgenerational effects in which offspring had reduced survival if one parent was exposed to nonylphenol during larval development.  However, 72-hour exposure of larvae to 1.0 mg/l nonyphenol resulted in 100% larval mortality.  Therefore, the resistance of oyster species to nonylphenol is assessed as ‘None’, resilience as ‘Very low’ and sensitivity as ‘High’.

Dioxin. Cooper et al. (2009) investigated the effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on the embryonic development of Crassostrea virginica.  They reported 97-99% mortality in larvae at 2-10 µg/l TCDD (‘severe’ mortality) but only sublethal effects on adults.  Cooper et al. (2009) suggested their results might explain the lack of self-sustaining populations of bivalves in estuaries contaminated by TCDD.  Therefore, the resistance of oyster species to TCDD is assessed as ‘None’ in early life stages but ‘High’ in adults.  However, if Cooper et al. (2009) suggestion is correct, and TCDD contamination might result in population decline, the resistance of oyster beds may be assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’ but with ‘Low confidence as the evidence is based on a single study.

Others. The results for ‘other’ petrochemicals are dominated by aromatic hydrocarbons and detailed in the ‘evidence summary spreadsheet’.  Exposure of Crassostrea virginica embryos for 48 hours to 25 mg/l 2,4-Dimethylphenol was reported to result in ‘severe’ mortality in one study.  Exposure to biphenyl was reported to result in sublethal effects in one study.  However, exposure to one or more of 19 aromatic petrochemicals (individually) was reported to result in significant mortality in five of the studies reviewed.  Therefore, the sensitivity of the early life stages of oyster species to 2,4-Dimethylphenol is probably ‘High’ (resistance is ‘None’ and ‘resilience ‘Very low’), albeit at high concentrations.  But oysters are probably ‘Not sensitive’ to biphenyl, albeit based on limited evidence.  Overall, the early life stages and hence populations of oyster species are probably of ‘High’ sensitivity (resistance and resilience are ‘Low’) to aromatic petrochemicals depending on the individual chemical, the exposure concentration, exposure duration and life stage exposed.  

Polychaetes, bivalves and amphipods are generally particularly affected by oil spills in infaunal habits, and echinoderms are also particularly intolerant of oil contamination (Suchanek, 1993).

Overall sensitivity.  The evidence suggests that oyster species, especially in their early life stages, are highly sensitive to a range of hydrocarbons and some PAHs. This conclusion agrees with the findings of His et al. (2000). Limited evidence suggests that the adults may be resistant to hydrocarbon contamination, but we may assume that loss of larval stages would result in a decline in resident populations that are dependent on recruitment for their abundance (Bayne, 2017). Therefore, the worst-case resistance of native oyster (Ostrea edulis) beds to hydrocarbon or PAH contamination is assessed as 'Low'. Beds may be protected from oil spills due to their depth depending on local conditions. Nevertheless, resilience is assessed as 'Low' and sensitivity as 'High'. However, Ostrea spp. was the least studied species in the evidence review so confidence in the assessment is 'Low'. 

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 results of the Rapid Evidence Assessment on the effects of 'Synthetic compound' contaminants on oyster species (Crassostrea spp., Ostrea spp., Saccostrea spp. and Magallana gigas). are summarized below. The full 'Oyster species evidence review' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'synthetic compound' contaminant examined, together with an overall pressure assessment. 

Pesticides/biocides. Adults and juveniles were more resistant to pesticide/biocide exposure than early life stages. For example, 5.8% of the results of exposure of adults and juveniles to pesticides/biocides reported ‘severe’ or ‘significant’ mortality compared to 88% that reported sublethal effects, whereas 89% reported ‘severe’ or ‘significant mortality after exposure of early life stages, compared to only 6% that reported sublethal effects.  Over 200 different pesticides/biocides, their metabolic or degradation products were catalogued, and divided amongst 22 different functional (e.g. herbicide, insecticide) or structural (e.g. organohalogen, organophosphate) groups. Therefore, it is not possible to discuss the sensitivity of each pesticide/biocide reviewed and the 'Oyster species evidence summary' should be referred to for details. 

Overall, there is considerable evidence to suggest that adults and juveniles are resistant to most pesticides, with the exception of some insecticides, organophosphates and organohalogens but that early life stages (e.g. larvae) are sensitive to a wide range of pesticides/biocides. Therefore, the resistance of the early life stages oyster species to pesticides/biocides is assessed as ‘None’. Hence, the resistance of oyster beds is assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’ on the assumption that loss of recruitment would lead to population decline.

Pharmaceuticals. ‘Severe’ mortality was reported in 16.6% of the results of exposure to pharmaceuticals, 53% reported ‘significant’, 2.8% ‘some’, 5.5% ‘no mortality, and the remaining 22% only reported sublethal effects.  However, the five articles that examined adult and juvenile oysters did not report any mortality (22% of results) or only sublethal effects (88% of results).  Conversely, the 11 articles that examined early life stages reported ‘severe’ mortality in 22% of the results, ‘significant’ in 70%, ‘some’ in 3.7%, ‘none’ in 3.7% and sublethal effects in 3.7% of the results. Overall, 28 separate pharmaceuticals were reported in the review but most of the chemicals were only tested in a single study (see ‘Oyster species evidence summary' spreadsheet).

The resistance of the early life stages in oyster species to pharmaceuticals is assessed as ‘None’.  However, the ‘worst-case’ resistance of adults and juveniles is probably ‘High’ (no mortality and/or sublethal effects).  Therefore, the resistance of oyster beds is assessed as ‘Low’ based on the assumption that loss of recruitment would lead to population decline.  Hence, resilience is assessed as ‘Low’ and sensitivity as ‘High’. 

Other synthetics. Overall, 13% of the results of exposure to ‘synthetics (other)’ reported ‘severe’ mortality, 59.4% reported ‘significant mortality, 4.3% ‘some’ mortality, 4.3% no mortality and 18.8% reported sublethal effects.  No mortality or only sublethal effects were reported for adults and juveniles.  However, the early life stages were less resistant to their effects.  For example, 17.6% of the results reported ‘severe’ mortality, 72.5% reported ‘significant’ mortality, 5.8% ‘some’ mortality, 1.9% reported no mortality and 1.9% reported sublethal effects.

Detergents and surfactants were the most toxic to larvae, which agrees with the finding of His et al. (2000).  The exposure of Ostrea edulis larvae to the detergents Kudos, Slix, Polyclens, Gamlen, Teepol, and Houghtosol halved the normal rate of larval development at concentrations ranging from 2.5-7.5 ppm (2.5-7.5 mg/l) (Smith, 1968).  Renzoni, (1973b; cited by His et al., 2000) also reported significant mortality in Ostrea edulis larvae exposed to tetrapropylene benzene sulphonate (with an LC50 of 2 mg/l). The remaining ‘Synthetic(other)’ chemicals were reported by only a few studies and varied in their sensitivity, although the effects of most types of chemicals would be assessed as ‘High’ sensitivity.

The resistance of the early life stages in oyster species to synthetic (others) is assessed as ‘None’.  However, the ‘worst-case’ resistance of adults and juveniles is probably ‘High’ (no mortality and/or sublethal effects).  Therefore, the resistance of oyster beds is assessed as ‘Low’ based on the assumption that loss of recruitment would lead to population decline.  Hence, resilience is assessed as ‘Low’ and sensitivity as ‘High’.

Polychlorinated biphenyls (PCBs). The effects of polychlorinated biphenyls (PCBs) were examined in only four studies (see above).  Exposure to either Aroclor 1016, 1254 or PCB 1254 did not result in mortality and/or only resulted in sublethal effects.  No studies on the effects of PCBs on oyster larvae were found.  Therefore, the resistance of oyster species to PCBs is assessed as ‘High’, resilience as ‘High’, and sensitivity as ‘Not sensitive’.

Flame retardants. Only two studies examined the effects of brominated flame retardants on oysters (Great Lakes Corporation, 1989; Xie et al., 2017b).  Sublethal effects (on shell deposition) were reported in immature oysters (Crassostrea sp.) but BDE-47 caused abnormal development of embryos and significant mortality in larvae.  Therefore, the resistance of early life stages to brominated flame retardants is probably ‘Low’ but immature oysters is ‘High’.  Hence, the resistance of oyster beds is assessed as ‘Low’ based on the assumption that loss of recruitment would lead to population decline.  Hence, resilience is assessed as ‘Low’ and sensitivity as ‘High’.  However, confidence in the assessment is ‘Low’ due to the disagreement in effect between the limited number of studies. 

Phthalates. The effects of phthalates were only examined in embryos and larvae.  All three of the studies reported significant mortality and/or abnormal development in embryos and larvae exposed to the phthalates studied.  Therefore, the resistance of early life stages to phthalates is probably ‘Low’.  Hence, the resistance of oyster beds is assessed as ‘Low’ based on the assumption that loss of recruitment would lead to population decline.  Hence, resilience is assessed as ‘Low’ and sensitivity as ‘High’.  However, confidence in the assessment is ‘Low’ due to the disagreement in effect between a limited number of studies.

Perfluoroalkyl substances (PFAS). Perfluoroalkyl substances (PFAS) were examined in two studies (Drottar & Krueger, 2000; OECD, 2002) neither of which specified the life stage of Crassostrea virginica examined.  Both studies reported sublethal effects at the concentrations tested.  The OECD (2002) suggested a 96-hour NOEC of 1.9 mg/l.  Therefore, resistance is assessed as ‘High’, resilience as ‘High’ and sensitivity as ‘Not sensitive’, albeit with ‘Low’ confidence due to the limited number of studies reviewed.

Overall sensitivity assessment. The effects of numerous (ca 200) pesticides/biocides, plus pharmaceuticals were examined in the literature reviewed, while PCBs, Flame retardants, Phthalates and PFAs were less studied. There was considerable variation between studies in experimental design as well as results. Nevertheless, the worst-case resistance of oyster species to 'synthetic compounds' is assessed as 'Low', especially due to the sensitivity of early life stages on the assumption that loss of recruitment would lead to population decline.  Therefore, the worst-case resistance of native oyster (Ostrea edulis) beds to 'synthetic compound' contamination is assessed as 'Low', resilience as 'Low' and sensitivity as 'High'. However, Ostrea spp. was the least studied species in the evidence review so confidence in the assessment is 'Low'. 

Radionuclide contamination

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

Evidence

The effects of exposure to tritium were reported by one article (Nelson, 1971; not accessed).  Nelson (1971) reported mortality in Crassostrea gigas larvae after exposure to 0.000001 Ci/l to 0.01 Ci/l Tritium (Ci = Curie – a non-SI unit of radioactive decay) but did not specify the larval stage or the level of mortality observed. Another paper that examined the effects of radioactive isotopes of chromium, strontium, zinc and yttrium on oyster larvae (Nelson, 1968) could also not be accessed.  Therefore, resistance is assessed as ‘Low’ as a precaution but with ‘Low’ confidence because the level of mortality was not specified. Hence, resilience is assessed as ‘Low’ and sensitivity as ‘High’.

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

The results of the Rapid Evidence Assessment on the effects of 'other substance' contaminants on oyster species (Crassostrea spp., Ostrea spp., Saccostrea spp. and Magallana gigas). are summarized below. The full 'Oyster species evidence review' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'other substance' contaminant examined, together with an overall pressure assessment. 

Inorganic chemicals. Most of the studies examined chemicals used in chlorination as a form of disinfectant. Bellance & Bailey (1977) reported a bioassay completed by the Virginia Institute of Marine Science that established an LC50 of 5 μg/L for Crassostrea virginica larvae exposed to chlorine for 48 hours in static testing.

Capuzzo (1979) investigated the effects of temperature on the toxicity of chlorine and chloramine to Crassostrea virginica in a flow-through system.  The oyster larvae were exposed to chlorine or chloramine for 30 minutes at either 20 or 25℃, after the 30-minute exposure the toxicant was removed from the solution and the temperature was reduced down to the acclimation temperature.  The mortality of the larvae was assessed 48 hours after exposure.  The temperature was shown to enhance the toxic effects of the chemicals.  The LC50 and LC100 values for exposure to chlorine at 20℃ were 120 and 1400 µg/l, respectively but at 25℃ the LC50 and LC100 values for chlorine were 80 and 860 µg/l, respectively.  The LC50 and LC100 values for exposure to chloramines at 20℃ were 10 and 480 µg/l, respectively but at 25℃ the LC50 and LC100 values for chloramine were <10 and 160 µg/l, respectively. Chien & Chou (1989) examined the effects of chlorine exposure on the development of Crassostrea gigas under various temperatures and salinities, at different stages of development.  Fertilized eggs at the first polar stage and four larval stages (blastula, trochophore, veliger, and D-larva) were exposed to combinations of five concentrations (0 to 2520 µg/l) of chlorine, at four temperatures (22, 25, 28℃) and three salinities (18, 26, 34 ppt) for one hour.  Chlorine exposure to all of the tested stages had lethal impacts on the larvae.  In general, the resistance to chlorine increased with salinity, with lower LC50s observed at lower salinities.  Larval sensitivity to chlorine generally increased with higher exposure temperatures. Crecelius (1979) examined the effects of ozonization of seawater on the production of bromate.  Crecelius (1979) reported that ozonization of seawater converted of all bromide to bromate within 60 mins.  Ozonization of sodium chloride solution did not result in significant oxidants while sodium bromide solution resulted in both bromide and bromate.  Nevertheless, they concluded that the levels of bromate produced by chlorination or ozonization of power plant cooling waters were not acutely toxic to Crassostrea gigas larvae, fish, shrimp, and clams by comparison with toxicity figures determined in their laboratory.  Crecelius (1979) reported that 1.0 mg/l bromate resulted in 90% mortality in Crassostrea virginica larvae and 30 mg/l bromate caused abnormal development in 50% of Crassostrea gigas larvae during the first 48-hour of larval development (48-hour EC50/LC50 of 30 mg/l bromate).  However, no experimental details were provided (Crecelius, 1979). 

Roberts et al. (1975) examined the toxicity of chlorine in estuarine water to a range of marine species, using both flow-through and static tank systems.  They examined oyster larvae and juveniles (Crassostrea virginica), Mercenaria mercenaria larvae, Acartia tonsa (copepod), Palaemonetes pugio and fish (Menidia menidia, Syngnathus fuscus, and Gobiosoma bosci). They noted that molluscan larvae and copepods were the most sensitive species with a 48-hour LC50 of less than 0.005 ppm.  Roberts & Gleeson (1978) exposed Crassostrea virginica larvae to bromine chloride (BrCl) during 48-hour exposure assays.  The concentration that caused 50% mortality (LC50) was calculated at 210 µg/l bromine chloride.  In addition, to the larvae tests, juvenile oysters were exposed to BrCl for 96 hours to assess the impacts on shell growth.  The 96-hour EC50s for the shell growth were 100 and 160 µg/l. Roosenburg et al. (1980) examined the effects of chlorine on two larval stages of Crassostrea virginica.  Straight-hinge veliger larvae were exposed to concentrations of 10, 50, 100 and 200 µg/l chlorine for 6, 12, 24 and 36 hours, and to 50, 100, 200 and 300 µg/l for 8, 24, 48, 72 and 96 hours.  Pediveliger larvae were exposed to 50, 100, 200 and 300 µg/l chlorine for 6, 24, 48, 72 and 96 hours.  Mortality increased with increasing concentration in both larval stages.  Straight hinge veliger larvae were more sensitive than pediveliger larvae with between 83-100% mortality at the highest tested concentration at 96 hours.  Pediveliger larvae had between 32.4-46.1% mortality under the same conditions.

Scott & Middaugh (1978) investigated the seasonal toxicity of chlorination on Crassostrea virginica.  Bioassays were conducted in the fall (45-day exposure), winter (75-day exposure) and spring (60-day exposure).  Adult oysters were collected, accumulated, and exposed to sodium hypochlorite at nominal concentrations of 5.6, 3.2, 1.8, and 1 mg/L.  During each of the seasonal bioassays survival, condition index, gonadal index, and faecal production were assessed.  Total (100%) mortality occurred in all of the treatments at the highest nominal concentration.  However, the winter assay had a delay in mortality with 100% mortality occurring on day 70, compared to day 22 for fall and day 32 for spring.  The condition index of the controls was higher than the exposed oysters.  Similarly, the gonadal index was significantly higher in the control oysters. Stewart et al. (1979) investigated the toxicity of by-products of oxidative biocides on oyster larvae.  Crassostrea virginica larvae were exposed to bromate, bromoform, and chloroform, at 0.05, 0.1, 1.0, and 10.0 mg/I for 48 hours.  Mortality was observed after the 48-hour exposure period, at all concentrations of the three substances larval mortality occurred.

All but one of the studies above examined the effect of chlorination, bromination, or ozonization of oyster larvae.  All the studies reported ‘severe’ or ‘significant’ larval mortality at the concentrations tested.  In addition, adult Crassostrea virginica experienced a reduction in condition due to exposure to chlorination and 100% mortality at 5.6 mg/l.  Therefore, resistance would be assessed as ‘None’, resilience as ‘Very low, and sensitivity as ‘High’, especially in larvae.

Fluoride.  Cardwell et al. (1979) (see above) examined the effects of a large number of different chemicals on the larvae of Crassostrea gigas.  Fluoride was reported to result in a 48-hour EC50 (abnormal development) of 58 mg/l and a 48-hour LC50 (larval mortality) of >100 mg/l.  Therefore, resistance to fluoride exposure would be assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’.  However, this assessment is based on a single study.

Phosphoric acid.  The effects of phosphoric acid on oyster larvae were reported by Daugherty (1951) and Kunigelis & Wilbur (1987).  Kunigelis & Wilbur (1987) could not be accessed and the level of mortality was not specified.  Daugherty (1951) reported 100% mortality in C. virginica after 28-hour exposure to 1,500 mg/l but the article could not be accessed for further detail.  .  Therefore, resistance would be assessed as ‘None’, resilience as ‘Very low, and sensitivity as ‘High’, especially in larvae.  However, this assessment is based on a single study that used a high concentration (1,500 mg/l).  It might represent the effects immediately after and in close proximity to a spill and confidence in the assessment is ‘Low’.

Potassium chloride (KCl) was included in two studies.  Da Cruz et al. (2007) reported that the exposure of Crassostrea rhizophorae embryos to potassium (as KCl) for 24 hours resulted in abnormal larval development and a 24-hour LC15 of 25.13 mg/l and a 24-hour LC50 35.56 mg/l.  Nell & Holliday (1986) examined the use of KCl and CuCl₂ to stimulate larval settlement in Saccostrea commercialis larvae, in static containers in the laboratory.  Settlement was stimulated by 8-12 mM KCl (160 - 210 µg/l) and no mortality was observed.  Therefore, low concentrations of KCl were used to induce larval settlement while high concentrations (mg/l) were reported to cause abnormal larval development.  Hence, resistance would be assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’.  However, the assessment is based on a single study using high concentrations of KCl so the confidence is assessed as ‘Low’.

Caldwell et al. (1975) investigated the effects of hydrogen sulphide on the survival and development of Crassostrea gigas.  The tests were run over a four-day period, the longer the oysters were exposed to hydrogen sulphide the lower the concentration was to cause 50% mortality.  The LC50 at 24, 48, and 96 hours were 3,300, 2,600, and 1,400 µg/l, respectively.  Therefore, resistance to sulphide exposure is assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’.

Okubo & Okubo, 1962 (cited by His, 2000) reported a 48-hour EC50 (abnormal development) of 32-100 µg/l in Crassostrea gigas embryos after exposure to sodium cyanide.  Therefore, the resistance of the early life stages of C. gigas to sodium cyanide would be assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’. 

Natural products. Daugherty (1951) reported that exposure to ‘starch’ at a concentration of 3 g/l resulted in 100% mortality in Crassostrea virginica.  Unfortunately, Daugherty (1951) could not be accessed and it is unclear how the starch was administered or how the effect was caused so no sensitivity assessment is made. Similarly, Cardwell, 1979b (cited by His et al., 2000) reported that exposure of the fertilized eggs of Crassostrea gigas to tannic acid resulted in abnormal development and a 48-hour EC50 of >10 mg/l.  Unfortunately, Cardwell, 1979b (cited by His et al., 2000) could not be accessed and it is unclear how the tannic acid was administered or how the effect was caused so no sensitivity assessment is made.

Explosives. Goodfellow et al. (1983, 1983b) examined the lethal and sublethal effects of picric acid and picramic acid on oysters as they were potential contaminants from industrial effluents and the manufacture of explosives.  Goodfellow et al. (1983) investigated the acute toxicity of picric acid and picramic acid on Crassostrea virginica.  The 144-hour LC5Os for picric and picramic acid were 254.9 and 69.8 mg/l, respectively.  No growth EC50s and shell deposition EC5Os showed that both contaminants caused adverse effects at much lower concentrations than indicated by the LC50s.  For example, the 144-hour shell deposition EC50s were 27.9 mg/l for picric acid and 5.6 mg/l for picramic acid. Goodfellow et al. (1983b) investigated the effects of picric acid and picramic acid on the growth of Crassostrea virginica.  Exposure to 0.45 and 0.05 mg/l (450 and 50 µg/l) picric acid and 0.24 and 0.02 mg/ (240 and 20 µg/l) picramic acid showed significant inhibition of shell deposition during the 42 days of exposure.  In addition, discolouration of the nacre layer of the shell and body mass was observed after exposure to both contaminants by the end of the 42-day trial.  Exposure to picric or picramic acids in the water column was reported to be lethal to C. virginica.  Therefore, resistance is assessed as ‘None’, resilience as ‘Very low, and sensitivity as ‘High’.

Lightsticks. De Araujo et al. (2015) determined the chemical composition and the toxicity of lightsticks that were recently activated, compared to lightsticks one year after activation and to lightsticks collected on beaches.  The effect of lightstick content on embryos of Crassostrea rhizophorae after 24 hours of exposure was assessed at various concentrations (0.32, 0.56, 1, 1.76, and 2.24% WSF).  The value of the WSF-effective concentration (24-hour EC50) that caused abnormal development of larvae was 0.35% for new light sticks but, after one year of activation, the toxicity of the light stick was even higher at 0.65%.  Therefore, resistance would be assessed as ‘Low’, resilience as ‘Low’ and sensitivity as ‘High’.  However, this assessment is based on a single study.

Rubber. Tallec et al. (2022) investigated the chemical toxicity of different types of new and used rubber products (tires, crumb rubber granulates, aquaculture rubber bands) on the early life stages of the Pacific oyster, Crassostrea gigas.  Leachates were obtained from the products at 0.1, 1, and 10 g/L. Sperm and embryos were exposed to leachates at 0.1, 1, and 10g/l for one hour before being assessed for viability.  The effect on fertilization was assessed by combining gametes with the different concentrations of each of the leachates for 1.5 hours before assessing the fertilization yields.  The impacts on the development of larvae were assessed by exposing embryos to leachates at 0.1, 1, and 10 g/l for 36 hours.  Abnormal D- larvae were classed as those with morphological malformations or those which had developmental arrest during embryogenesis.  The viability of oyster sperm was not significantly affected by exposure to leachates from new tires, used tires, new crumb rubber granulates, used crumb rubber granulates, and used oyster-farming rubber bands at any of the concentrations (0.1, 1 and 10g/l) tested when compared with the control treatments.  However, significant reductions in the percentage of live spermatozoa were observed at the highest tested concentrated leachate (10 g/l) from new oyster-farming rubber bands.  The viability of oyster oocytes was not significantly affected by exposure to any of the leachates at any of the tested concentrations.  The fertilization yield was not significantly affected by exposure to leachates from new tires, used tires, new crumb rubber granulates, used crumb rubber granulates, or used oyster-farming rubber bands when compared with the control treatment.  However, significant reductions in fertilization yield were observed at the highest tested concentration of leachate (10 g/l) from new oyster-farming rubber bands.  Embryo-larvae development was significantly reduced by 53% by exposure to new-tire leachate at 10 g/l.  Embryo-larval development was completely inhibited at the highest tested leachate concentration (10 g/L) of new crumb rubber granulates, used crumb rubber granulates, and used oyster-farming rubber bands.  Embryo-larval development was completely inhibited at 1g/l of new oyster-farming rubber bands leachate.  Therefore, the resistance of embryos and larvae to rubber leachates would be assessed as ‘None’, resilience as ‘Very low, and sensitivity as ‘High’. 

Overall sensitivity assessment for 'other substances'.  Most of the chemicals examined in the studies reviewed resulted in 'severe' or 'significant' mortality, especially in early life stages. Therefore, the worst-case resistance of native oyster (Ostrea edulis) beds to 'other substances' contamination is assessed as 'Low', resilience as 'Low' and sensitivity as 'High'. However, Ostrea spp. was not examined directly in the studies reviewed and the effects varied depending on the chemical examined so confidence in the assessment is 'Low'. 

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

Oysters are considered to be tolerant of periods of hypoxia due to their ability to survive out of water during transportation for long periods of time, and many weeks at low temperatures (Korringa, 1952; Yonge, 1960).  However, the sustained oxygen depletion typical of areas with high organic loading would probably have much more severe effects (Wilding & Hughes, 2010).  Although Ostrea edulis may be relatively tolerant of low oxygen concentrations other species within the community may be more intolerant (Tyler-Walters, 2008).

Lenihan (1999) reported that Crassostrea virginica could withstand hypoxic conditions (< 2mg O² /l ) for 7-10 days at 18 °C but last for several weeks at <5 °C.  However, Lenihan (1999) also suggested that many days (26) of hypoxia, contributed to the high rate of mortality observed at the base reefs at 6 m depth together with poor condition, parasitism and reduced food availability.  In addition, a prolonged period of hypoxia in the River Neuse (North Carolina) resulted in mass mortality of oysters (Lenihan, 1999).

Members of the characterizing species that occur in estuaries e.g. Ascidiella aspersa are probably tolerant of a degree of hypoxia and occasional anoxia. Similarly, most polychaetes are capable of a degree of anaerobic respiration (Diaz & Rosenberg, 1995).  However, periods of hypoxia and anoxia are likely to result in loss of some members of the infauna and epifauna within this biotope.

Sensitivity assessment.  Ostrea edulis is not affected by de-oxygenation at the level of the benchmark.  However, some of the associated species might be affected at the benchmark level.  For this reason the resistance and resilience are assessed as ‘High’, giving the biotope a ‘Not sensitive’ sensitivity.

Nutrient enrichment

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

Evidence

This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations. The nutrient enrichment of a marine environment leads to organisms no longer being limited by the availability of certain nutrients. The consequent changes in ecosystem functions can lead to the progression of eutrophic symptoms (Bricker et al., 2008), changes in species diversity and evenness (Johnston & Roberts, 2009) decreases in dissolved oxygen and uncharacteristic microalgae blooms (Bricker et al., 1999, 2008).

Moderate nutrient enrichment, especially in the form of organic particulates and dissolved organic material, is likely to increase food availability for all the suspension feeders within the biotope. However, long-term or high levels of nutrient or organic enrichment may result in eutrophication and have indirect adverse effects, such as increased turbidity, increased suspended sediment, increased risk of deoxygenation and the risk of algal blooms.

In the Mar Menor lagoon, Spain, nutrient enrichment resulting from agricultural runoff resulted in eutrophic waters, however, the Ostrea edulis population persists there. Oysters from this population were exposed to the same simulated eutrophic conditions within a laboratory to assess their potential for reducing eutrophication in the lagoon (Albentosa et al., 2023). Two levels of eutrophication were simulated using particulate organic matter (POM) comprised of phytoplankton cells and detritus, at concentrations that replicate those observed on two occasions in the Mar Menor in 2016: 2.50 and 4.83 mg POM/L. These were tested against an ‘oligotrophic’ environment of 0.82 mg POM/L. They found that the oysters could be grouped into ‘high’ and ‘low’ feeders despite no differences in size or tissue weight/composition. In high-feeding oysters, clearance rate was significantly lower (1.15 L/hour) in the highest concentration of POM compared to 1.73 L/hour at the intermediate POM level and 1.79 L/hour in the oligotrophic treatment. Bivalves are known to decrease their clearance rate in high POM concentrations as a mechanism to regulate ingestion. POM concentration had no effect on the clearance rate of low-feeding oysters, which remained significantly lower than that of the high-feeders (between 0.53 and 0.63 L/hour). Ingestion and absorption rate significantly increased as POM concentration increased for both low- and high-feeders, suggesting that maximum digestive capacity was not reached even at this level of POM. Therefore, increases in nutrient enrichment that may cause POM to reach 4.83 mg POM/L may not cause deleterious effects on individuals from this population of Ostrea edulis (Albentosa et al., 2023).

Ostrea edulis has been reported to suffer mortality due to toxic algal blooms, e.g. blooms of Gonyaulax sp. and Gymnodinium sp. (Shumway, 1990). Ostrea edulis has been shown to accumulate domoic acid (DA) in the presence of high abundances of Nitzschia bizertensi (known to form harmful algal blooms) in the south-west Mediterranean (Bouchouicha-Smida et al., 2015). Ostrea edulis accumulated more domoic acid in their tissue than Mytilus galloprovincialis in this study (Bouchouicha-Smida et al., 2015). Domoic acid is not necessarily harmful to Ostrea edulis, however, increases in occurrence of other toxic bloom-forming microalgae, likely facilitated by increased nutrient inputs, could prove harmful to Ostrea edulis due to their high accumulation rate of toxins.

The subsequent die-off of toxic and non-toxic algal blooms may also negatively impact benthic Ostrea edulis through the build-up of dead algal cells on the benthos. This may result in local de-oxygenation during decomposition, especially in sheltered areas with little water movement. Ostrea edulis may be relatively tolerant of low oxygen concentrations, however, other species within the biotope may be more intolerant.

Sensitivity assessment. A slight increase in nutrients may enhance food supply to Ostrea edulis and increase growth rates in the species. Excessive nutrient loading may lead to eutrophication. Ostrea edulis has been reported to survive the eutrophic waters of the Mar Menor lagoon and to have some tolerance to elevated levels of POM within the water, associated with nutrient enrichment. However, the formation of harmful algal blooms as a result could result in toxin accumulation in the oysters, and blooms of Gonyaulax sp. and Gymnodinium sp. have been reported to cause mortality in oysters (Shumway, 1990). Therefore, the above evidence suggests that oysters could survive eutrophic conditions, so resistance and resilience have been assessed as ‘High’, resulting in an assessment of ‘Not Sensitive’, albeit at ‘Low’ confidence as the assessment is based on a single study. However, if the nutrient enrichment resulted in an harmful algal bloom, oyster mortality could occur. Hence, the worst-case resistance is assessed as ‘Medium’, resilience as ‘Medium’ and sensitivity as ‘Medium’, but with ‘Low’ confidence.

Organic enrichment

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

Evidence

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

Nutrient enrichment of the water column is a potential impact arising from finfish aquaculture which can potentially lead to eutrophication and the alteration of the species composition of plankton with possible proliferation of potentially toxic or nuisance species (OSPAR, 2009b). However, the current consensus is that enrichment by salmon farm nutrients is generally too little, relative to natural levels, to have such an effect (SAMS and Napier University 2002, cited in Wilding & Hughes, 2010).

In the Mar Menor lagoon, Spain, nutrient enrichment resulting from agricultural runoff resulted in eutrophic waters, however, the Ostrea edulis population persists there. Oysters from this population were exposed to the same simulated eutrophic conditions within a laboratory to assess their potential for reducing eutrophication in the lagoon (Albentosa et al., 2023). Two levels of eutrophication were simulated using particulate organic matter (POM) comprised of phytoplankton cells and detritus, at concentrations that replicate those observed on two occasions in the Mar Menor in 2016: 2.50 and 4.83 mg POM/L. These were tested against an ‘oligotrophic’ environment of 0.82 mg POM/L. They found that the oysters could be grouped into ‘high’ and ‘low’ feeders despite no differences in size or tissue weight/composition. In high-feeding oysters, clearance rate was significantly lower (1.15 L/hour) in the highest concentration of POM compared to 1.73 L/hour at the intermediate POM level and 1.79 L/hour in the oligotrophic treatment. Bivalves are known to decrease their clearance rate in high POM concentrations as a mechanism to regulate ingestion. POM concentration had no effect on the clearance rate of low-feeding oysters, which remained significantly lower than that of the high-feeders (between 0.53 and 0.63 L/hour). Ingestion and absorption rate significantly increased as POM concentration increased for both low- and high-feeders, suggesting that maximum digestive capacity was not reached even at this level of POM. Therefore, increases in nutrient enrichment that may cause POM to reach 4.83 mg POM/L may not cause deleterious effects on individuals from this population of Ostrea edulis (Albentosa et al., 2023).

Johnston & Roberts (2009) undertook a review and meta-analysis of the effect of contaminants on species richness and evenness in the marine environment. Of the 49 papers reviewed relating to sewage as a contaminant, over 70% found that it had a negative impact on species diversity, < 5% found increased diversity, and the remaining papers finding no detectable effect. Due to the ‘remarkably consistent’ effect of marine pollutants on species diversity this finding relevant to this biotope (Johnston & Roberts, 2009). It was found that any single pollutant reduced species richness by 30 to 50% within any of the marine habitats considered (Johnston & Roberts, 2009). Throughout their investigation there were only a few examples where species richness was increased due to the anthropogenic introduction of a contaminant. These examples were almost entirely from the introduction of nutrients, either from aquaculture or sewage outfalls.

Sensitivity assessment. Little empirical evidence was found to support an assessment of this biotope at this benchmark. The lack of direct evidence for Ostrea edulis has resulted in this pressure being assessed as ‘Insufficient evidence’.

Physical loss (to land or freshwater habitat)

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

Evidence

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

Physical change (to another seabed type)

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

Evidence

Many Ostrea edulis reef restoration projects are now looking to use artificial substratum to promote larval recruitment. Studies have shown that Ostrea edulis larvae will settle on these artificial surfaces, providing they have ample rugosity, and those made with shells incorporated into the concrete had higher larval recruitment than those without (Potet et al., 2021; Kamermans et al., 2025). Often, granite is used as scour protection for offshore wind farms, and this has been suggested as a potential opportunity to incorporate reef restoration. Ter Hofstede et al. (2024) demonstrated that granite was a suitable settlement substratum for Ostrea edulis larvae, with higher average settlement per m² (206.8/m²) than mussel shells.

Despite the ability of Ostrea edulis larvae to settle on artificial hard substratum, this biotope occurs on sandy mud with some shells and occasionally gravel. Therefore, if there were a change from this substratum type, the physical conditions required for this biotope would no longer be present. Ultimately, a change to rock or artificial substrata would cause the biotope to be lost. 

Artificial hard substratum may also differ in character from natural hard substratum, so that replacement of natural surfaces with artificial may lead to changes in the biotope through changes in species composition, richness and diversity (Green et al., 2012; Firth et al., 2014) or the presence of non-native species (Bulleri & Airoldi, 2005).

Sensitivity assessment. The biotope has a resistance of ‘None’, a resilience of ‘Very low’, and consequently a sensitivity of ‘High’.

Physical change (to another sediment type)

Benchmark. Permanent change in one Folk class (based on UK SeaMap simplified classification) (Physical change in sediment type pressure definition). 

Evidence

Sensitivity assessment.  Ostrea edulis occur in a range of habitat types and hence are not considered sensitive to an increased sediment coarse faction.  Resistance and resilience are therefore assessed as ‘High’ resulting in this biotope being considered ‘Not sensitive’ at the pressure benchmark.

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

Ostrea edulis cements its lower valve permanently to solid pieces of substratum, such as pebbles, cobbles, boulders etc.  The removal of this layer of the substratum would lead to the loss of; the biogenic layer created by oysters and its biological community, the oyster cultch (which will remove an important chemical cue used by larvae when settling), and the substrate which provides a point of attachment for larvae. 

Sensitivity assessment.  The resistance to the removal of the substratum is ‘None’.  The resilience of the biotope to this pressure depends on what substratum lies 30 cm below the top layer.  If the substratum is the same as that which was removed, resilience is going to be ‘Very low’.  If the underlying substrate is not suitable for the recovery of this biotope i.e. bedrock, then the biotope will not be able to return at all.

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

Abrasion may cause damage to the shell of Ostrea edulis, particularly to the growing edge. Regeneration and repair abilities of the oyster are quite good. Power washing of cultivated oysters routinely causes chips to the edge of the shell increasing the risk of desiccation. This damage is soon repaired by the mantle. Oysters were often harvested by dredging in the past, during which their shells survived relatively intact. On mixed sediments, the dredge may remove the underlying sediment, and cobbles and shell material with effects similar to substratum loss above.

In a review of anthropogenic threats to restored Ostrea edulis broodstock areas, Woolmer et al. (2011) reported that, in general, fishing mortality arising from commercial fisheries (for oysters and other mobile gear fisheries) is a key pressure on native oyster populations and habitats. Impacts include stock removal, disturbance of spat (juvenile oysters) and habitat disturbances (to oyster banks and reefs). More specifically, Woolmer et al. (2011) stated that dredging over oyster beds removes both cultch material and target oysters. Over time, with sufficient effort, the net effect is a flattening of the bank and the creation of a flatter bed which is more susceptible to siltation and hypoxia in some water bodies (Woolmer et al., 2011).)). 

Despite this, when managed, disturbance to the bed by dredges or rakes may be beneficial. In oyster aquaculture, a process known as ‘harrowing’ can be employed, whereby machinery is used to clean the existing substratum or oyster bed to promote larval settlement.  Ostrea edulis larval settlement after harrowing has shown various levels of success. Cameron et al. (2023) demonstrated that, in the Blackwater estuary, Essex, there was a positive relationship between oyster recruitment and increased disturbance by harrowing between 2016 and 2018. They reported that in areas of 30 oysters per 100 m dredge that were harrowed for 40 minutes prior to spawning, 20 times more recruitment occurred (from five oysters to 100 oysters per 100 m dredge). They demonstrated however that this positive relationship ceased where oysters are found in densities of 60 oysters per 100 m dredge (or five oysters per m²), after which point, harrowing decreased recruitment success (Cameron et al., 2023). Therefore, in areas with high adult density of Ostrea edulis, harrowing may be counterproductive to facilitating recovery efforts, at least in muddy estuarine conditions such as the Blackwater estuary. At much lower oyster densities (0.037 oysters/ m²), harrowing may also have negative effects on settlement (Bromley et al., 2016), as observed in Lough Foyle, on the Ireland/Northern Ireland border.

Polychaetes and other segmented worms were reported to be badly affected by oyster dredging while any bivalves were displaced (Gubbay & Knapman, 1999).  In addition, the epifauna associated with horse mussel beds (Modiolus modiolus) was found to be particularly sensitive to abrasion due to scallop dredging (see A5.621; Service & Magorrian, 1997).  Therefore, Ostrea edulis and the other characterizing species are probably sensitive to physical disturbance at the benchmark level.

Sensitivity assessment. The characterizing species, Ostrea edulis, is somewhat resistant to some abrasion and is able to recover from some damage to shells e.g. chipping caused by pressure washers. However, damage caused to oyster beds and their habitats by commercial fishing is considered to be of importance to levels of mortality and health of oyster beds. Resistance has been assessed as ‘Low’, the resilience is assessed as ‘Low’. This gives the biotope a sensitivity of ‘High’.

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

In general, fishing activities that penetrate the substratum to a greater extent (e.g. beam trawls, scallop dredges and demersel trawls) will potentially damage these habitats to a greater degree than fishing activities using lighter gear (e.g. light demersel trawls and seines) (Hall et al., 2008).  One of the major reasons for the decline of the oyster population at Chesapeake Bay was mechanical destruction (Rothschild et al., 1994).

Sensitivity assessment.  The effect of sub-surface disturbance will be to displace, damage and remove individuals.  Shallow disturbance is considered to remove between 25-75% of the population so that resistance is assessed as ‘Low’.  Resilience is assessed as ‘Low’ and sensitivity is therefore 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

A decrease in turbidity and hence increased light penetration may result in increased phytoplankton production and hence increased food availability for suspension feeders, including Ostrea edulis.  Therefore, reduced turbidity may be beneficial.  However, increased fouling by red algae may result and compete with juveniles and settling spat for space.  In areas of high suspended sediment, a decrease may result in improved condition and recruitment due to a reduction in the clogging of filtration apparatus of suspension feeders and an increase in the relative proportion of organic particulates.  However, a decrease in suspended sediments in some areas may reduce food availability resulting in lower growth or reduced energy for reproduction (Tyler-Walters, 2008).

In a field experiment in Canada, the summer growth of Ostrea edulis, on coarse sandy substrata, was found to be enhanced at low levels of sediment resuspension and inhibited as sediment deposition increased (Grant et al., 1990, summarised in Ray et al., 2005).  In a review of the biological effects of dredging operations, Ray et al. (2005) stated that sediment chlorophyll in suspension at low levels may act as a food supplement, enhancing growth, but at higher concentrations may dilute planktonic food resources and suppress food ingestion (Jackson & Wilding, 2009, references therein).

Oysters respond to an increase in suspended sediment by increasing pseudofaeces production with occasional rapid closure of their valves to expel accumulated silt (Yonge, 1960) both of which exert an energetic cost. Korringa (1952) reported that an increase in suspended sediment decreased the filtration rate in oysters.  This study is supported by Grant et al. (1990) who found declining clearance rates in Ostrea edulis in response to an increase in suspended particulate matter.  Suspended sediment was also shown to reduce the growth rate of adult Ostrea edulis and to result in shell thickening (Moore, 1977).  Reduced growth probably results from increased shell deposition and an inability to feed efficiently.  Hutchinson & Hawkins (1992) reported that filtration was completely inhibited by 10 mg/l of particulate organic matter and significantly reduced by 5 mg/l.  Ostrea edulis larvae survived 7 days exposure to up to 4 g/l silt with little mortality.  However, their growth was impaired at 0.75 g/l or above (Moore, 1977). Yonge (1960) and Korringa (1952) considered Ostrea edulis to be intolerant of turbid (silt laden) environments.  Moore (1977) reported that variation in suspended sediment and silted substratum and resultant scour was an important factor restricting oyster spat fall, i.e. recruitment.  Therefore, an increase in suspended sediment may have longer term effects of the population by inhibiting recruitment, especially if the increase coincided with the peak settlement period in summer.  

The other suspension feeders characteristic of this biotope are probably tolerant of a degree of suspended sediment but an increase, especially of fine silt, would probably interfere with feeding mechanisms, resulting in reduced feeding and a loss of energy through mechanisms to shed or remove silt.  

Sensitivity assessment.  A short-term increase in sedimentation is unlikely to have an impact on this biotope and its characterizing species.  Ostrea edulis has a comping mechanism to remove increased levels of silt from within the mantle.  This behaviour is energetically expensive, and may cause a decrease in growth rate of the organism, but is unlikely to cause mortality.  For these reasons resistance and resilience are assessed as ‘High’ given a sensitivity score of ‘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

A decrease in turbidity and hence increased light penetration may result in increased phytoplankton production and hence increased food availability for suspension feeders, including Ostrea edulis.  Therefore, reduced turbidity may be beneficial.  However, increased fouling by red algae may result and compete with juveniles and settling spat for space.  In areas of high suspended sediment, a decrease may result in improved condition and recruitment due to a reduction in the clogging of filtration apparatus of suspension feeders and an increase in the relative proportion of organic particulates.  However, a decrease in suspended sediments in some areas may reduce food availability resulting in lower growth or reduced energy for reproduction (Tyler-Walters, 2008).

In a field experiment in Canada, the summer growth of Ostrea edulis, on coarse sandy substrata, was found to be enhanced at low levels of sediment resuspension and inhibited as sediment deposition increased (Grant et al., 1990, summarised in Ray et al., 2005).  In a review of the biological effects of dredging operations, Ray et al. (2005) stated that sediment chlorophyll in suspension at low levels may act as a food supplement, enhancing growth, but at higher concentrations may dilute planktonic food resources and suppress food ingestion (Jackson & Wilding, 2009, references therein).

Oysters respond to an increase in suspended sediment by increasing pseudofaeces production with occasional rapid closure of their valves to expel accumulated silt (Yonge, 1960) both of which exert an energetic cost. Korringa (1952) reported that an increase in suspended sediment decreased the filtration rate in oysters.  This study is supported by Grant et al. (1990) who found declining clearance rates in Ostrea edulis in response to an increase in suspended particulate matter.  Suspended sediment was also shown to reduce the growth rate of adult Ostrea edulis and to result in shell thickening (Moore, 1977).  Reduced growth probably results from increased shell deposition and an inability to feed efficiently.  Hutchinson & Hawkins (1992) reported that filtration was completely inhibited by 10 mg/l of particulate organic matter and significantly reduced by 5 mg/l.  Ostrea edulis larvae survived 7 days exposure to up to 4 g/l silt with little mortality.  However, their growth was impaired at 0.75 g/l or above (Moore, 1977). Yonge (1960) and Korringa (1952) considered Ostrea edulis to be intolerant of turbid (silt laden) environments.  Moore (1977) reported that variation in suspended sediment and silted substratum and resultant scour was an important factor restricting oyster spat fall, i.e. recruitment.  Therefore, an increase in suspended sediment may have longer term effects of the population by inhibiting recruitment, especially if the increase coincided with the peak settlement period in summer.  

The other suspension feeders characteristic of this biotope are probably tolerant of a degree of suspended sediment but an increase, especially of fine silt, would probably interfere with feeding mechanisms, resulting in reduced feeding and a loss of energy through mechanisms to shed or remove silt.  

Sensitivity assessment.  A short-term increase in sedimentation is unlikely to have an impact on this biotope and its characterizing species.  Ostrea edulis has a comping mechanism to remove increased levels of silt from within the mantle.  This behaviour is energetically expensive, and may cause a decrease in growth rate of the organism, but is unlikely to cause mortality.  For these reasons resistance and resilience are assessed as ‘High’ given a sensitivity score of ‘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

No direct evidence was found to assess this pressure at the benchmark.  A deposit at the pressure benchmark would cover all species with a thick layer of fine materials.  Species associated with this biotope would not be able to escape and would likely suffer mortality (see evidence for light siltation).  Ostrea edulis would be unable to survive this pressure at the benchmark.  As filter feeders that are permanently attached to the substratum they would be unable to borrow up to the surface to enable basic life functions to occur.  The low tidal streams within this biotope, in addition to the extremely sheltered to sheltered wave exposure mean that there would be low levels of sediment resuspension.  This could possibly exacerbate the negative impacts of this pressure.  The same assessment has been used for this pressure as in the light pressure benchmark.  Resistance to the pressure is ‘None’, resilience is ‘Very low’ and sensitivity is given as ‘High’.

Litter

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

Evidence

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. 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., 2020a), depending on the study species and duration and intensity of exposure. There have been no studies investigating the effect of EMFs at the population or community level for benthic organisms. 

Sensitivity assessment. No studies have examined the effect of EMFs on Ostrea edulis, therefore, there is ‘Insufficient evidence’ on which to base an assessment of the likely sensitivity this biotope to EMFs.

Underwater noise changes

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

Evidence

Species characterizing this habitat do not have hearing perception but vibrations may cause an impact, however no studies exist to support an assessment.

Introduction of light or shading

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

Evidence

Studies have shown that valve gaping behaviour of Ostrea edulis appears to be linked to light intensity. The duration and extent to which Ostrea edulis open their valves determine the rate at which they breathe and feed, and in closing their gape, protect themselves from predators and poor-quality water. Le Moal et al. (2025) demonstrated that under normal light-dark cycles (12 hours each), Ostrea edulis showed a strong 24-hour circadian rhythm for valve opening amplitude and duration. This rhythm persisted even when the light cycle was shifted by four hours, and when under constant darkness, demonstrating an internal circadian rhythm. However, under constant dim light and when exposed to 12 hours of artificial light at night (ALAN) rather than darkness, this rhythm was disrupted and/or weakened (Le Moal et al., 2025). This suggested that constant light and ALAN alter the natural circadian valve gaping behaviour of Ostrea edulis. Fabra et al. (2025) investigated the clearance rates Ostrea edulis over 24 hours in March (8°C), April (10°C) and May (12°C). Oysters had higher filtration rates during hours of darkness (up to 4.5 litres per hour), further suggesting that the feeding behaviour of this species may be diurnal and influenced by light, at least in laboratory conditions (Fabra et al., 2025).

The disruption of this circadian rhythm may be exacerbated when coupled with other stressors such as low salinity and changes in temperature. Bamber (2023) examined valve gape behaviour in response to full (1000 lx) and reduced (> 2 lx) light intensity and different salinities and temperatures. Oysters were shown to exhibit maximum gaping behaviour during hours of reduced light intensity, apart from when salinity was lowered to 18.6 as oysters had closed their valves. Once salinity had increased above 24, oysters returned to gaping significantly more in the low light than the full light intensity. While increasing temperatures from 15.8°C to 20.1°C had no effect on gaping behaviour, decreasing between these temperatures significantly altered how long oysters kept their valves open in low light conditions. Oysters returned to their expected valve gaping behaviour once temperature had stabilised at 15.8°C after four days (Bamber, 2023).

There is also evidence that increased periods of light may impact reproduction in Ostrea edulis. Cowen et al. (2025) demonstrated that oysters kept under constant light transitioned from female to male earlier than those kept in a simulated natural photoperiod. Extended photoperiods also resulted in higher gonadal development and higher larval production compared to photoperiods shorter than the natural cycle, which were seen to delay spawning by up to four weeks (Maneiro et al., 2017).

The prevention of light reaching the seabed may affect Ostrea edulis indirectly through changes in phytoplankton abundance and primary production. Red algae found in the biotope Ostrea edulis beds on shallow sublittoral muddy mixed sediments will be affected by a reduction in primary production. Red algae are probably shade tolerant but may be lost from deeper examples of this biotope (Tyler-Walters, 2008).

Sensitivity assessment. Changes to light exposure for Ostrea edulis may impact their feeding and respiration rate and alter the chronology of their reproductive processes. However, the long-term effect of these changes on the survival of this species is unknown. Therefore, there is currently ‘Insufficient evidence’ from which to assess the sensitivity of this biotope.

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 – this pressure is considered applicable to mobile species, e.g. fish and marine mammals rather than seabed habitats. Physical and hydrographic barriers may limit propagule dispersal.  But propagule dispersal is not considered under the pressure definition and benchmark.

Death or injury by collision

Benchmark. Injury or mortality from collisions of biota with both static or moving structures due to 0.1% of tidal volume on an average tide, passing through an artificial structure (Death for collision pressure definition).

Evidence

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

Visual disturbance

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

Evidence

Not relevant. 

Genetic modification & translocation of indigenous species

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

Evidence

Organisms are frequently transplanted from one location to another in marine aquaculture and these transplanted species may pose potentially serious impacts to native populations through interbreeding and thus alteration of the gene pool. The Pacific oyster (Magallana gigas) has been intentionally imported from Japan into Ireland because it is larger and faster growing than the native oyster (Ostrea edulis).  Pacific oysters cannot hybridize with the native oyster but indirect effects may occur through alterations in gene frequencies as a result of ecological interactions with the Pacific oyster (Heffernan, 1999).

Sensitivity assessment.  Very little information is available on the effect of this pressure on the characterizing species Ostrea edulis.  Ostrea edulis may be translocated, resistance to genetic impacts is assessed as ‘None’ and recovery as ‘Low’ due to the potential for permanent effects.  Sensitivity is therefore categorised as ‘High’.

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

Numerous diseases and parasites have been identified in oysters, partly due to their commercial importance and partly because of incidences of disease related mass mortalities in oyster beds. Diseases in oysters and other commercial bivalve species may be caused by bacteria (especially in larvae), protists, fungi, coccidians, gregarines, trematodes, while annelids and copepods may be parasite. The reader should refer to reviews by Lauckner (1983) and Bower & McGladdery (1996) for further detail. The following species have caused mortalities in Ostrea edulis populations in the UK. 

The parasitic worm, Polydora ciliata, burrows into Ostrea edulis shell, potentially weakening the shell and increasing the oyster’s vulnerability to predation and physical damage. However, Polydora ciliata infestation did not appear to reduce the condition or strength of shells of Ostrea edulis collected from Plymouth Sound (Lemasson & Knights, 2019). This study also showed that Polydora ciliata had a significant preference for infesting Ostrea edulis (99% infested) compared to the Pacific oyster, Magallana gigas (25% infested) during a laboratory study (Lemasson & Knights, 2019). This could have implications for Ostrea edulis competition with Magallana gigas in areas where they both inhabit, potentially hampering recovery of the native species. Another parasitic worm, Polydora hoplura, can also damage Ostrea edulis shells by causing shell blisters. Boring sponges of the genus Cliona may also bore the shell of oysters causing shell weakening, especially in older specimens. 

The flagellate protozoan Heximata sp. resulted in mass mortalities of natural and cultivated beds of oysters in Europe between 1920 and 1921, from which many populations did not recover (Yonge, 1960). Another protozoan parasite Marteilia refingens, present in France has not yet affected stocks in the British Isles (Kerr et al., 2018), and the copepod parasite, Mytilicola intestinalis, of mussels, has also been found to infect Ostrea edulis potentially causing considerable loss of condition, although in most infections there is no evidence of pathology. Perkinsus spp. is known to cause mortalities in bivalves globally and has been identified infecting Ostrea edulis populations with the Mediterranean (Ramilo et al., 2015). However, there have been no reports of this pathogen parasitizing Ostrea edulis around the UK

The parasitic protozoan Bonamia ostreae caused mass mortalities in France, the Netherlands, Spain, Iceland and England after its accidental introduction in the 1980's resulting in a further reduction in oyster production (Edwards, 1997). It is a serious threat to Ostrea edulis in the UK (Laing et al., 2005, cited in Woomer et al. 2011) having caused mortality of Ostrea edulis throughout northern Europe, with disease events reducing populations by 80% or higher (Heffernan, 1999). Disease transmission can occur from oyster to oyster. However, Bonamia ostreae is also found in other marine invertebrates, including zooplankton (indicating the possibility of interspecies transmission; Lynch et al., 2007 cited in Woolmer et al., 2011). Ostrea edulis larvae may also be vectors for disease between populations; Arzul et al., 2011 cited in Woolmer et al., 2011). Horizontal transmission of Bonamia ostreae was observed within Ostrea edulis larvae, as brooding larvae were shown to test positive for Bonamia ostreae DNA when parents were not, suggesting this pathogen can be taken up from the surrounding water column, not necessarily passed on from parent to offspring (Flannery et al., 2016). There is evidence that Ostrea edulis has some resistance to Bonamia. Populations of Ostrea edulis can be resistant (avoids infection), tolerant (remains healthy despite infection), and/or resilient (can recover from impacts of disease over time) to this pathogen (as reviewed in Holbrook et al., 2021). Populations that have been exposed to it for a long time (> 20 years) can survive infection to this pathogen better than those that have not. This may be due to changes in gene expression that induce the death of infected cells (as reviewed in Sas et al., 2020). Egerton et al. (2020) suggested that oysters in Loch Ryan, Scotland, have adapted to become S-strategists (also known as K-strategists) rather than R-strategists, such that they are characterized by slower growth with lower reproductive output, and invest more energy into durability and immunity. As such, the population here shows some resistance to infection by Bonamia ostreae, with only 5% of the population testing positive for infection and only at low levels (Egerton et al., 2020). Therefore, this species may be able to undertake local adaptations that make them more resilient to this pathogen.

The transportation of Pacific oysters from Japan to the west coast of North America is thought to have resulted in the introduction of the bacterium Nocardia crassostreae leading to nocardiosis (bacterial infection that can invade every tissue) in Pacific oysters (Magallana gigas) and Ostrea edulis (Forrest et al., 2009; taken from Tillin et al., 2013). Ostreid herpesvirus 1 is known to have caused bivalve mortality and is also known to cause mortality in Magallana gigas, which can often be found coexisting with Ostrea edulis. Ostrea edulis was shown to be susceptible to this virus in a laboratory study that directly infected juvenile native oyster muscle, with mortality occurring after only four days, and cumulative mortality reaching 25% after 10 days (end of experiment) (Sanmartín et al., 2016). No reports of this virus causing mortality in wild oyster populations around the UK were found.

Sensitivity assessment. Although the impact of individual species of microbial pathogen on Ostrea edulis varies, pathogens known to affect this species in the UK can cause significant mortality. Bonamia ostrea is known to cause more than 80% mortality of oyster beds within the UK. Therefore, both the resistance and resilience have been assessed as ‘Low’ and sensitivity as ‘High’.

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

Ostrea edulis is long lived, has notably unreliable reproduction, and low levels of recruitment, which makes it vulnerable to overfishing (Orton, 1927; Spärck, 1951; Laing et al., 1951; taken from Gravestock et al., 2014). British native oyster beds were exploited in Roman times. The introduction of oyster dredging in the mid-19th century developed the oyster beds into one of Britain's largest fisheries, employing about 120,000 men around the coast in the 1880's. However, by the late 19^th century, stocks were beginning to be depleted so that by the 1950s the native oyster beds were regarded as scarce (Korringa, 1952; Yonge, 1960; Edwards, 1997). This biotope is still regarded as scarce today. Overfishing, combined with reductions in water quality, cold winters (hence poor spat fall), flooding, the introduction of non-native competitors and pests, outbreaks of disease and severe winters were blamed for the decline (Korringa, 1952; Yonge, 1960; Edwards, 1997). As a result, although 700 million oysters were consumed in London alone in 1864, the catch fell from 40 million in 1920 to 3 million in the 1960s; from which the catch has not recovered (Edwards, 1997). Most populations are now artificially laid for culture (Fowler, 1999; Edwards, 1997).

The Ostrea edulis fishery in The Solent was once considered to be the largest self-sustaining fishery in Europe (Gravestock et al., 2014). However, since the turn of the 20^th century, the population has collapsed significantly three times. The first collapse occurred between 1919 and 1921 due to a disease epidemic caused by the flagellate protozoan Hexamita (Tubbs, 1999). The second collapse was caused by the 1962/1963 winter, during which temperatures were significantly below average (Kamphausen, 2012). The final collapse was in 2006 when poor recruitment led to a sharp drop in the population (Gravestock et al., 2014). Although a number of potential causes of recruitment failure have been suggested (see Gravestock et al., 2014), overfishing likely exacerbated the effect of poor recruitment.

Populations of Ostrea edulis in Strangford Lough, Northern Ireland, are the only known examples to demonstrate natural, unassisted recovery from harvesting (Smyth et al., 2009). In the northern basin, there was an increase in littoral oysters from > 100,000 in 1998 to 1.2 million in 2003. However, this number then declined to 650,000 in 2005, mainly through the loss of medium and large oysters, suggesting that this decrease in abundance was due to unregulated harvesting. Oyster abundance increased again to over 1 million by 2007 due to successful larval recruitment, as evidenced by the replacement of larger oysters by smaller ones. Larvae were likely provided by subtidal and/or less accessible populations that were not able to be harvested (Smyth et al., 2009). Strangford Lough has been afforded multiple protections including Marine Conservation Zone and Marine Protected Area. In 2008, after a period of heavy trawling for Modiolus modiolus, the northern area of the Lough was designated a closed fishing zone, whereby increases in government enforcement officers were enacted, particularly on islands only accessible by boat. Smyth et al. (2023) reported that in these closed zones with increased patrols, abundance of Ostrea edulis increased from around 1000 oysters in 2004 to > 88,000 in 2021. In contrast, they showed that at easily accessible and unpatrolled intertidal sites, abundance of Ostrea edulis severely declined from 964,000 oysters in 2004 to 58,000 oysters in 2021 (94% decline). Oysters at the intertidal sites were also found in lower density (3.36 oysters/m² in 2004 to 0.44 oysters/m² in 2021) and more fragmented, which could reduce fertilization success (Guy et al., 2019) and hamper recovery. In addition, the intertidal sites in 2021 were dominated by small oysters, approximately one to three years old, with low larval output, whereas at the island sites, oysters were mainly four to seven years old (Smyth et al., 2023) and more highly fecund compared to those at the intertidal sites (Smyth, 2022).

There is evidence, however, that harvesting of Ostrea edulis can be done sustainably. In the Skagerrak region of Sweden, the population of Ostrea edulis has been estimated at around 36 million, with annual landings of this species representing only 0.24% of the population. Exploitation of this species appears to have been made sustainable by the implementation of bans on dredges and trawls, only allowing hand-diving, and private ownership of the habitat (requiring a licence for collection) (Thorngren et al., 2019). In Loch Ryan, Scotland, the oyster fishery is harvested in a rotation, whereby one area is fished and then left to recover for six years before re-harvesting. Kennon et al. (2023) looked at the recovery of harvested areas after one, two and six years. Oyster shell density and percentage cover, along with macrofaunal biodiversity, significantly differed between time after harvesting. Shell density and percentage cover and biodiversity were highest six years after harvesting, with shell density averaging 8.5, 21.4 and 71.8 oyster/m² one-, two- and six-years post harvesting, respectively. The authors modelled that full recovery of associated macrofaunal biodiversity would take 10 years (Kennon et al., 2023), however, this does not necessarily imply an equivalent recovery in shell density as biodiversity rose even when shell density plateaued (Kennon et al., 2023).

Sensitivity assessment. The current rarity of oyster beds in the UK is due to the pressure the populations were put under due to commercial fishing. Stock from beds can remain sustainable under commercial fishing pressure as long as there is effective management, non-destructive fishing gear, suitable recovery time between harvests, and good larval recruitment. However, if these populations have a period of bad recruitment or are affected by another negative pressure, then fishing can compound this effect. Ostrea edulis have no ability to remove themselves from fishing pressure as they are permanently attached to the substrate once they have settled from larvae. For this reason, resilience of this biotope is given as ‘None’. A number of native oyster beds in the UK have been destroyed by fishing and have had to undergo human intervention to return the oyster population. In some areas oysters have not returned. Resilience is assessed as ‘Very low’, resulting in a ‘High’ sensitivity score.

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

Direct, physical impacts from harvesting are assessed through the abrasion and penetration of the seabed pressures. Ostrea edulis is the dominant species within this biotope so they could easily be incidentally removed from this biotope as by-catch when other species are being targeted. The loss of these species and other associated species would decrease species richness and negatively impact on the ecosystem function.

Fishing pressure (from bottom-towed gears) was identified as the most influential variable inhibiting potential habitat suitability of oysters within the German Bight area of the North Sea (Pogoda et al., 2023), and protection from destructive fishing practices was considered essential by experts to make habitats suitable for restoration purposes around the UK and the within the North Sea (Hughes et al. 2023).

Ellrich et al. (2025) examined the predator-prey interactions of brown crab (Cancer pagurus) and the European lobster (Homarus gammarus) on Ostrea edulis that co-occur in the German Bight, North Sea. They observed that brown crabs preferentially consumed smaller oysters, and that those with a shell length of > 7 cm were generally safe from predation, whereas lobster preferably consumed medium-sized oysters (< 12 cm). Furthermore, they determined that oysters were less likely to be preyed upon by these species when Mytilus spp. are present as they are an alternative and more beneficial food source. They also showed that when Magallana gigas oysters existed as ‘clumps’ (four to six oysters), they were less likely to be eaten given the more difficult nature of extraction, which may be the same for Ostrea edulis (no clumps were available for this study). Finally, this study showed that the presence of larger predators, such as the lobsters, deterred feeding from brown crabs, likely through waterborne cues Ellrich et al. (2025). The removal of other commercial species such as Mytilus spp. may therefore increase the likelihood of Ostrea edulis being preyed upon.

Sensitivity assessment. Removal of a large percentage of Ostrea edulis as by-catch from other targeted fisheries would alter the character of the biotope, and fishing pressure from bottom-towed gears may prevent establishment of this species in otherwise suitable habitats. However, the removal of commercial species that predate upon Ostrea edulis may facilitate their survival, depending on the method through which they are collected. The resistance to removal is ‘Low’ due to the easy accessibility of the biotopes and the inability of these species to evade collection. Resilience is ‘Medium’, with recovery only being able to begin when the harvesting pressure is removed altogether. Therefore, gives an overall sensitivity score of ‘Medium’ is recorded.

The American slipper limpet, Crepidula fornicata

Evidence

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and the Essex coast in 1887-1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland, including native and cultured oyster beds (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 1999, 2018; Hinz et al., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). It occurs in large numbers in most of the oyster-producing areas of England and Wales (Blanchard, 1997; Thieltges, 2005; Powell-Jennings & Calloway, 2018). 

High densities of Crepidula fornicata cause ecological impacts on sedimentary habitats. The species can smother the seabed in shallow bays, changing and modifying the habitat structure (Blanchard, 1997, 2009; Helmer et al., 2019; Tillin et al., 2020). At high densities, the species physically smothers the sediment, and the resultant build-up of silt, pseudofaeces, and faeces is deposited and trapped within the bed (Tillin et al. 2020, Fitzgerald, 2007, Blanchard, 2009, Stiger-Pouvreau & Thouzeau, 2015). The biodeposition rates of Crepidula are extremely high and once deposited, form an anoxic mud, making the environment suitable for other species, including most infauna (Blanchard, 2009; Stiger-Pouvreau & Thouzea, 2015). The resultant modification of the substratum can render it unsuitable for native oysters (Blanchard, 1997, 2009).

For example, the Bay of Mont Saint-Michel supports shellfisheries for the Pacific oyster (Magallana gigas), flat oyster (Ostrea edulis) and blue mussels (Mytilus edulis). The flat oyster fishery was impacted by Crepidula which spread rapidly into the oyster beds and became attached to the oyster shells (Blanchard, 2009). After forty years, Crepidula occupied most of the subtidal area of the Bay, its spread facilitated by oyster farming and shellfish dredging (Blanchard, 2009). However, it occurred at lower densities and co-existed with other shellfish in Arcachon Bay (De Montaudouin et al. 1999, 2018) and Bourgnerf Bay (Decottignies, 2006, 2007 cited in Blanchard, 2009). 

Similarly, Helmer et al. (2019) reported that Ostrea edulis numbers in the Chichester Harbour had decreased by 96% in a 19-year period since 1998, while Crepidula numbers had increased by 441% in the same period. Low densities of Ostrea edulis and high densities of Crepidula were also reported in Portsmouth and Langstone Harbours, which are within one of the few remaining oyster fisheries in the UK. Helmer et al. (2019) suggested that the lack of recovery of Ostrea edulis populations was probably due to a lack of habitat heterogeneity and suitable settlement substratum, together with ongoing fishing activity and disease. Helmer et al. (2019) suggested that the native oyster population in the Solent was "on the brink of ecological collapse" without active management to mitigate the dominance of Crepidula. Preston et al. (2020) reported that Crepidula larvae probably competed with Ostrea larvae for food in the plankton, with resultant negative effects on recruitment, and that the Solent is probably substratum-limited for Ostrea due to larval preference for native oyster shell and mixed sediment rather than the muddy sediments created by dense populations of Crepidula. Preston et al. (2020) also noted that the biogenic habitat created by Crepidula was less diverse and species-rich than that provided by native oysters. 

Crepidula fornicata has been implicated in the decline of Ostrea edulis beds across the North Sea. However, Hayer et al. (2019) concluded that the decline of native oyster beds was almost complete before Crepidula began to invade the North Sea. The decline in native oyster beds by the 1940s was most probably due to overfishing combined with reductions in water quality, cold winters (hence poor spat fall), flooding, the introduction of non-native competitors and pests, and outbreaks of disease (Korringa, 1952; Yonge, 1960; Edwards, 1997). The population dynamics of oyster populations are dependent on positive feedback between adult abundance and recruitment via the provision of reef habitat for the settlement of larvae (e.g. adult shell), and the growth of the height of the reef about the sediment and the supply of food (facilitated by current flow) (Bayne, 2017). Nevertheless, the presence of high densities of Crepidula, modification of the substratum, competition for food as larvae, and exclusion from suitable settlement substratum, probably prevents the recovery of Ostrea edulis beds (Blanchard, 2009; Helmer et al., 2018; Preston et al., 2020). 

Sensitivity assessment. Crepidula is reported to damage oyster culture and is thought to prevent oyster bed recovery and compete for habitats. Where abundant, Crepidula fornicata is likely to change the entire biotope to produce a Crepidula fornicata dominated biotope such as SS.SMx.SMxVS.CreMed or SS.SMx.IMx.CreAsAn (JNCC, 2015, 2022). Therefore, resistance is assessed as ‘Low’. Resilience is assessed as 'Very low' because the successful removal of an INIS is extremely rare. Hence, sensitivity is assessed as 'High'. Due to the constant risk of new invasive species, the literature on this pressure should be revisited

The carpet sea squirt, Didemnum vexillum

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009). Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024). In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; 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 11 to 19 times after 30 days. Valentine et al. (2007a) concluded fragments could successfully grow, survive, and help to spread Didemnum vexillum.

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

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

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

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

Didemnum vexillum requires suitable hard substrata for successful settlement and the establishment of colonies. It can grow quickly and establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems because of its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock gravel, pebble, cobble, or boulders or artificial substrata such as a variety of maritime structures such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ship hulls and at aquaculture facilities (Valentine et al., 2007 a&b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012; Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent, and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020).

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

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

However, some studies on Georges Bank, USA and Sandwich, Massachusetts observed colonies were able to survive partial covering by sand (Bullard et al., 2007; Valentine et al., 2007a). Gittenberger et al. (2015) reported that Didemnum vexillum was able to overgrow sandy bottom (cited Gittenberger, 2007). In northern Kent, Didemnum vexillum has been recorded covering London clay boulders on Whitstable Flats, West Beach, north Kent, covering tabulate sandstone boulders (0.5 to 2 m across) on the mid-shore and colonizing sediment mounds on the low shore characterized by larger areas of sand, mud and low-lying sediment at Reculver and Bishopstone, north Kent (Hitchin, 2012). It was also recorded from 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 on 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.

In contrast, Didemnum vexillum’s preference for sheltered conditions, established colonies observed in Georges Bank and Long Island Sound were exposed to moderately strong tidal currents (1 to 2 knots; ca 0.5 to 1 m/s recorded at both sites) that may mobilise sediment (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020). However, Valentine et al. (2007b) describe the substratum as immobile, presumably consolidated, gravel, cobbles, and pebbles. Kleeman (2009), stated that the presence of a consistent mild wave action or ‘swash zone’ appears to favour Didemnum sp. establishment in the intertidal. Although some evidence suggests that waves and currents can facilitate the fragmentation and spread of Didemnum vexillum (Mckenzie et al., 2017), the tidal current velocities at some sites where Didemnum vexillum has been reported (for example, New England, where current velocities reach up to around 3 m/s) is lower than the current velocity required for the dislodgement of Didemnum vexillum fragments (around 7.6 m/s) (Reinhardt et al., 2012). This suggests that not all tidal currents are likely to dislodge Didemnum vexillum fragments. When on boat hulls the species can experience higher current velocities which is enough to cause dislodgement (Reinhardt et al., 2012).  

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

Colonies of Didemnum vexillum dominated sublittoral Pacific oyster reefs off Terschelling and Texel, in the Wadden Sea (Gittenberger et al., 2015) and fouled Pacific oysters on west Canadian coastline (Valentine et al., 2007a cited by Tillin et al., 2020). Gittenberger et al. (2015) reported that Didemnum vexillum did not seem to kill the oysters, but invasive colonies probably inhibited the settlement of mussels and other sessile organisms on the oysters. However, Gittenberger (2007) listed Magallana (syn. Crassostrea) gigas and Ostrea edulis among organisms that "died on contact" with Didemnum sp.

Sensitivity assessment. Colonies of Didemnum vexillum have been reported to grow over (smother) Pacific oyster beds in the Wadden Sea and Canada (Valentine et al., 2007a; Gittenberger et al., 2015). Evidence of mortality is limited. Gittenberger (2007) suggested oysters "die on contact" with Didemnum, while Gittenberger et al. (2015) noted that Pacific oysters were not killed, although smothering will result in lower growth rates and productivity. The population dynamics of oyster populations are dependent on positive feedback between adult abundance and recruitment via the provision of reef habitat for the settlement of larvae (e.g. adult shell), and the growth of the height of the reef above the sediment and the supply of food (facilitated by current flow) (Bayne, 2017). Therefore, smothering by mats of Didemnum may adversely affect recruitment and contribute to the long-term decline in the oyster beds. Therefore, resistance is assessed as 'Medium' (some; <25% mortality), albeit with 'Low' confidence due to the lack of direct evidence. Resilience is assessed as 'Very low' as Didemnum would need to be physically removed to enable recovery. Hence, sensitivity to colonization by Didemnum sp. is assessed as 'Medium'.   

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

Since introduction, the species has invaded and established self-sustaining natural populations throughout Europe from the North Sea, Wadden Sea and Scandinavian coastlines to the Atlantic coastlines of Spain and Portugal, as well as the Mediterranean and Adriatic Sea (Wrange et al., 2010; GBNNSS, 2011, 2012; Ezgeta-Balic et al., 2019; Spagnolo et al., 2019; Bergstrom et al., 2021; Hansen et al., 2023). In the UK, the species predominantly occurs around the southern and western coastlines (OBIS, 2024; NBN, 2024). Shipping activity has also been associated with the introduction of Magallana gigas in the north-eastern Adriatic Sea, where it was not introduced for aquaculture (Ezgeta-Balic et al., 2019) and possibly in south-west England from France possibly via fouling on ships (GBNNSS, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).

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 where the Pacific oyster occurs sporadically, no immediate ecosystem risk is observed.

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). In Lim Bay, Adriatic Sea, Magallana gigas is only found in the intertidal and on the sublittoral edge (at a depth of 1 m) and not at 3 m or 6 m depth (Stagličić et al., 2020; Tillin et al., 2020). It coexists here with Ostrea edulis which is abundant in the subtidal (Stagličić et al., 2020). Bergstrom et al. (2021) found that depth was one of the most important predictors of the occurrence of Magallana gigas in the Skagerrak and suggested the optimal depth of the species was 0.5 m in the shallow subtidal, although it occurred down to 5 m. On littoral rock in Brittany, the Pacific oyster colonizes all intertidal levels from Mean High Water to Mean Low Water on sheltered (low energy), moderately exposed (moderate energy) and high energy rock shores (Herbert et al., 2012). 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 (Smaal et al., 2009). Magallana gigas has not been found at extreme low water or subtidally beneath rocky habitats, as it has been in soft sediment areas (Herbert et al., 2012).

It has been suggested that recruitment is enhanced and abundances are higher in wave-sheltered conditions (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. Similarly, Bergstrom et al. (2021) noted that the occurrence of high densities of both Ostrea edulis and Magallana gigas decreased with increasing wave exposure.

Magallana gigas has been recorded co-occurring with Ostrea edulis at five intertidal sites around the coast of Ireland (Zwerschke et al., 2018b). Magallana gigas occurred at consistently higher densities than Ostrea edulis at all five sites, with the native oyster always present at comparatively low densities (Zwerschke et al., 2018b). At one site, Magallana gigas abundance was approximately 10 times higher than Ostrea edulis (Zwerschke et al., 2018b). Where both species were observed coexisting around Ireland, higher abundances of Magallana gigas tended to occur at intertidal sites, whereas Ostrea edulis dominated at shallow subtidal sites (Tully & Clarke, 2012). Competitive interactions between Ostrea edulis and Magallana gigas in areas where they coexist is not fully understood. Both species occupy the same niche and overlap in their phytoplankton consumption was observed by Ezgeta-Balic et al. (2020b), demonstrating competition for their primary food source. There was, however, variation in zooplankton consumption between the species throughout the year, with Ostrea edulis consuming higher concentrations in spring and autumn compared to Magallana gigas. It was confirmed through DNA analysis that Magallana gigas consumes Ostrea edulis larvae during spawning season (Ezgeta-Balic et al., 2020b). Therefore, the presence of the invasive species could hamper the recovery of the native species due to competition for food and larval predation. However, this study did not confirm whether Ostrea edulis also consumes Magallana gigas larvae.

Magallana gigas may have a competitive advantage as environmental conditions change. Magallana gigas was shown to tolerate fluctuating temperature and salinity conditions and starvation better than Ostrea edulis (Stechele et al., 2022). Both species were shown to survive in 30°C for one week (Kamermans & Saurel, 2022). However, Magallana gigas maintained a consistently higher body condition index (biomass relative to its shell) than Ostrea edulis (Kamermans & Saurel, 2022). In addition, Magallana gigas were able to survive better than Ostrea edulis in reduced pH, and their physiological functioning only declined once pH had reached a certain threshold, whereas the functioning of Ostrea edulis decreased linearly with pH. In addition, Ostrea edulis were shown to have overall lower clearance rates than Magallana gigas and respired more in the presence of the invasive species (Green et al., 2017). These results suggest that Magallana gigas may have a better tolerance for decreased pH, changing temperature and salinity, and reduced food availability than Ostrea edulis, which may give them a competitive advantage where they are found co-occurring.

However, interactions between these two species may not always be competitive. Zwerschke et al. (2018c) investigated the interactions between these species in Strangford Lough in the intertidal (high abiotic stress) and subtidal (12m, low abiotic stress), on horizontal and vertical substrata. They demonstrated that the growth rate of Ostrea edulis was not affected by the presence of Magallana gigas at the intertidal site. At the subtidal site however, their growth rate was reduced on horizontal rock in the presence of Magallana gigas. Irrespective of site, the biomass of Magallana gigas was significantly lower when in the presence of Ostrea edulis on horizontal substrata, suggesting higher competition for food, whereas biomass of both species was higher on vertical substrata. It appears, therefore, that interactions between Ostrea edulis and Magallana gigas depend on the level of abiotic stress and the orientation of substrata, whereby, on horizontal substrata, competition may be intensified, but on vertical substrata, interactions may be facilitative such that both species are able to coexist (Zwerschke et al., 2018c). In Lim Bay, Adriatic Sea, where Magallana gigas coexists with Ostrea edulis, Stagličić et al. (2020) suggested that Magallana occupied the available niche in the intertidal rather than was competitively excluded.

Another facilitative interaction between these two species was observed in the Dutch North Sea. through the return of a shellfish reef consisting of the native oyster, where it was thought this species was ecologically extinct given its absence for nearly a century (Christianen et al., 2018). The reef was formed of Ostrea edulis, but mostly Magallana gigas and Mytilus edulis, and 81% of Ostrea edulis were found attached to the shells of Magallana gigas. Therefore, the presence of the Pacific oyster here appeared to have facilitated the re-establishment of native oysters by providing suitable attachment substratum.

Ostrea edulis are known to perform a variety of ecological functions within their ecosystem, and it is unknown how these functions might change should Magallana gigas outcompete the native species. In Lough Foyle, Ireland, and Strangford Lough, Northern Ireland, in-situ studies revealed no significant differences in the nutrient cycling and habitat provisioning of these species, suggesting that, should Magallana gigas outcompete Ostrea edulis, it may not result in the loss of these functions in these areas (Zwerschke et al., 2016, 2018, 2020). Nutrient cycling and habitat provision were associated with the density of oysters rather than the species. There may, however, be changes to the associated epifaunal community should Magallana gigas replace Ostrea edulis. Guy et al. (2018) observed from Strangford Lough populations that Magallana gigas had significantly fewer epibionts, and lower species richness within its epifaunal community compared to those found on Ostrea edulis. This may be due to Ostrea edulis shells having higher rugosity allowing for more microhabitats. Therefore, should Magallana gigas outcompete Ostrea edulis on reefs, its dominance may also alter the associated epifaunal biodiversity, with potential trophic implications.

Sensitivity assessment. Most of the evidence suggests that Magallana gigas is limited to 10 m but they can form beds down to 42 m where conditions allow (Smaal et al., 2009; Herbert et al., 2012, 2016; Tillin et al., 2020). In addition, Magallana gigas is known to prefer wave-sheltered conditions. Therefore, the depth and sheltered to extremely sheltered conditions in which this biotope is found make it suitable for Magallana gigas colonization. There has already been evidence of Magallana gigas co-existing with Ostrea edulis beds around Europe and the UK and Ireland, however, the extent to which this negatively impacts the survivability of Ostrea edulis is unknown. The presence of Magallana gigas may facilitate the recovery of Ostrea edulis by providing suitable surface for larval settlement, however, in areas of low abiotic stress and particular substratum, their coexistence may result in competition for food and space. Ostrea edulis beds below 10 m are likely to be ‘Not sensitive’ to Magallana gigas, however, in shallow water where they overlap, they may form mixed beds. A mixed bed of native and Pacific oysters may represent a change in the biotope requiring reclassification depending on the relative proportions of the two species or their vertical distribution with depth, which will probably vary between sites. It may also decrease the biodiversity of the biotope. Therefore, a resistance of ‘Low’ is suggested as a precaution to represent changes to the biotope. Resilience is 'Very low' as the Magallana gigas population would need to be removed for recovery to occur. Therefore, sensitivity is assessed as ‘’High’ albeit with ‘Low’ confidence pending further evidence.

Wireweed, Sargassum muticum

Evidence

Wireweed, Sargassum muticum  is known to grow in the shallow subtidal around the UK, usually in areas sheltered from wave action. Its distribution is limited by the availability of hard substratum (e.g. stones > 10 cm) and light (Staeher et al., 2000; Strong & Dring 2011; Engelen et al., 2015). It is most abundant between 1 and 3 m below mean water, is a poor competitor under low light and only develops dense canopies in shallow areas (Engelen et al., 2015). This biotope exists in suitability sheltered environment butit consists of muddy mixed sediment which, other than the oysters themselves, lacks the hard attachment substrata required by Sargassum. In addition, this biotope is often found at depths that would be unfavourable to Sargassum (5 to 20 m), making it unlikely that Sargassum muticum would colonize this biotope. Resistance and resilience are therefore 'High' and the biotope 'Not sensitive' to Sargassum muticum. 

Wakame, Undaria pinnatifida

Evidence

Wakame, Undaria pinnatifida is known to grow in the shallow subtidal around the UK but is usually found in areas sheltered from wave action, with a depth range of -1 to 4 m. This biotope exists in suitability sheltered environments; however, it consists of muddy mixed sediment which, other than the oysters themselves, lacks the hard attachment substrata required by Undaria. In addition, this biotope is often found at depths that would be unfavourable to Undaria (5 to 20 m). Resistance and resilience are therefore 'High' and the biotope 'Not sensitive' to Undaria pinnatifida. 

Other INIS

Evidence

The American oyster drill, Urosalpinx cinerea. Urosalpinx cinerea are known to exist from the mid-intertidal down to 36 m, from sheltered to wave exposed habitats (Tillin et al., 2020). It was first recorded in 1927 and occurs in the south-east and south-west of the UK. Urosalpinx cinerea is a major predator of oyster spat and was considered to be a major pest on native and cultured oyster beds (Korringa, 1952; Yonge, 1960) and contributed to the decline in oyster populations in the first half of the 20th century. For example, in the Oosterschelde, Korringa (1952) reported 90% mortality in oyster spat by their first winter, with up to 75% being taken by Urosalpinx cinerea, while Hancock (1955) noted that 73% of spat settling in the summer of 1953 died by December, 55 to 58% being taken by Urosalpinx cinerea. The presence of this species has been identified as a factor preventing recovery of an oyster reef in Washington, USA (Buhle & Ruesink, 2009, cited in Tillin et al., 2020). Predation on adult Ostrea edulis is considered to be moderate due to their size, however high densities of Urosalpinx cinerea may lead to reduced recruitment (Tillin et al., 2020). Given its known association with oysters and reported reduction in oyster spat, the resistance of this biotope is ‘Low’. The resilience is considered ‘Very Low’ as Urosalpinx cinerea would have to be physically removed, making the sensitivity of this biotope ‘High’ to Urosalpinx cinerea.

Compass sea squirt, Asterocarpa humilis. This species has been observed colonizing intertidal and shallow subtidal habitats down to 26 m, can tolerate fully marine to low estuarine salinities and appears to prefer sheltered sites such as harbours and marinas (Tillin et al., 2020). It has been recorded attached to bivalves (Page et al., 2016) and could therefore pose competition for food and space resources Ostrea edulis beds. Given the depth ranges of this biotope (0 to 20 m) and the availability of suitable attachment substrata (i.e., the oysters themselves), Ostrea edulis beds in this biotope are potentially suitable habitat for Asterocarpa humilis (Tillin et al., 2020). Little is known about the potential impacts of Asterocapa humilis colonizing biogenic reefs, and no reports of it interacting with Ostrea edulis have been found.

Chinese mitten crab, Eriocheir sinensis. Adult Eriocheir sinensis are euryhaline, existing in both fresh and full salinity waters. The depth range of this species is not fully known, but it is usually found in low energy estuarine and freshwater habitats, or shallow bays, often with underwater vegetation (Tillin et al., 2020). This species is known to predate on mussels, and, given its generalist feeding strategy, would also likely consume oysters, though this has not yet been documented for Ostrea edulis. This biotope has been suggested as potentially suitable habitat for Eriocheir sinensis given the depth and sheltered nature in which it is found (Tillin et al., 2020) but no adverse effects on oysters have yet been reported.

Red ripple bryozoan, Watersipora subatra. This species colonizes a variety of substrata and is often found in the intertidal and shallow subtidal, though has been recorded deeper than 10 m, and in salinities between 18 and 49 psu as well as a wide range of wave exposures from sheltered to exposed (Tillin et al., 2020). It is known to associate with mussels and shellfish. Therefore, given that they can be found at similar depths, and the availability of suitable attachment substrata, Ostrea edulis beds are considered suitable habitat for Watersipora subatra (Tillin et al., 2020). However, there have been no records of Watersipora subatra colonizing this biotope around the UK and Ireland.

Bonnemaison’s hook weed, Bonnemaisonia hamifera. This species can be found in salinities of 14.26 to 37.55 psu and is usually found growing as epifauna on macroalgae in the lower littoral down to 20 m (Tillin et al., 2020), It prefers very sheltered conditions, such as those in which tis biotope is found. Tillin et al. (2020) reported that this biotope could be potentially suitable habitat for Bonnemaisonia hamifera given the availability of suitable attachment substrata. No adverse effects on oysters have yet been reported.

Japanese skeleton shrimp, Caprella mutica. This species has been found on a range of different substrata between 18 to 35 ppt salinity, intertidally and subtidally up to 20 m deep and in sheltered areas to those with high tidal flow regimes (Tillin et al.,2020). While this species does not usually associate directly with hard surfaces such as bivalves, it favours filamentous structures like hydroids and turf algae that it can hold onto, which are often found growing as epifauna on oysters. Therefore, this biotope has been considered as potentially suitable habitat for Caprella mutica due to its associated epifaunal community and suitable depth, salinity, and hydrodynamic regimes (Tillin et al., 2020). However, there have been no reports of Caprella mutica on Ostrea edulis beds.

A red seaweed, Agarophyton vermiculophyllum. This species is known to tolerate a wide salinity range of 5 to 60 psu and inhabit shallow habitats in sheltered areas (Tillin et al., 2020). Populations of Agarophyton vermiculophyllum are common in shellfish aquaculture sites around the British Isles as it is known to associate with bivalves and biogenic reefs, including attaching itself to mussel byssal threads and oysters (Tillin et al., 2020). Once established, it may prevent settlement of oyster spat by overgrowing the reef, thus reducing available substrata. The extent to which Agarophyton vermiculophyllum colonization negatively impacts adult populations of Ostrea edulis is not known, however, their existence may only be ephemeral (Tillin et al., 2020). This biotope is considered as suitable habitat for this species given its association with biogenic reefs,

Orange striped anemone, Diadumene lineata. This species tends to be found in brackish waters, particularly in bays, estuaries, and marinas where its only requirement is hard substrata on which to attach. As such, is often associated with mussels and oysters (Tillin et al., 2020). It can tolerate a large salinity range, from 0.5 to 35 ppt, and is found in shallow waters to depths of a few hundred meters (Cohen, 2011), preferring sheltered areas with low wave exposure (Fofonoff et al., 2003). This species has not been shown to cause negative impacts on the habitats that it colonizes (Fofonoff et al., 2003). Ostrea edulis beds are found within suitable depths and salinities, and the association of Diadumene lineata with oysters makes this biotope suitable habitat (Tillin et al., 2020).

American jack knife clam, Ensis leei. As a burrowing species, Ensis leei requires sedimentary habitats and can persist in fine to coarse grained substratum. It appears to prefer estuarine conditions but survives in salinities of 7 to 32 ppt. This species tends to be found from the lower shore to fully subtidal and has been reported as deep at 100 m in Canada, preferring moderately high bed shear stress (Tillin et al., 2020). This biotope may confer potentially suitable habitat for Ensis leei based on suitable depths, salinity, and substratum (Tillin et al., 2020). Ensis leei may compete for food with Ostrea edulis, however, this impact was considered ‘minor’ (Tillin et al., 2020).

Asian rapa whelk, Rapana venosa. This species colonizes subtidal habitats up to 90 m deep and in salinities of 16 to 35 ppt (Tillin et al., 2020). It can live in a variety of substrata including mixed sediment. They are known to prey on bivalves and have been seen to significantly reduce bivalve fisheries through predation (Sewell & Sweet, 2011b). In addition, high densities of this species have resulted in the loss of native bivalves, including Ostrea edulis, in the Black Sea (Zolotarev, 1996). Its preferred wave exposure and tidal currents are not known. Ostrea edulis beds in the UK have been suggested as potentially suitable habitat for Rapana venosa given their favourable depths and occurrence on sublittoral mixed sediment (Tillin et al., 2020). The potential impact of colonization is not known However, given its ability to predate and significantly reduce bivalve abundance, colonization at high enough densities by Rapana venosa may negatively impact Ostrea edulis beds. However, no adverse effects on oysters have yet been reported in the UK.

Asian/Japanese oyster drill, Ocinebrellus inornatus (formerly Ocenebra inornata). This species is known to exist on oyster beds, favouring natural oyster beds over those formed in aquaculture. It prefers sand, mud and gravelly substratum, and has a low tolerance to reduced salinity, although can survive at 23 psu for some months. It exists on the mid-shore, with a maximum depth likely around 5 to 6 m (Tillin et al., 2020). Its wave exposure tolerances are unknown. This biotope, therefore, provides suitable habitat for Ocinebrellus inornatus (Tillin et al., 2020). Ocinebrellus inornatus predates on Ostrea edulis spat, therefore, high densities of this invader could lead to reduced recruitment. However, predation on adults is considered to be moderate due to their size and shell thickness (Tillin et al., 2020).

Sensitivity assessment. Several INIS could potentially impact oyster beds, however, for all mentioned species (other than Urosalpinx cinerea), there is ‘Insufficient evidence’ on which to determine the sensitivity of Ostrea edulis. However, Urosalpinx cinerea has be reported to adversely affect native oyster beds. Hence, resistance is considered ‘Low’. The resilience is considered ‘Very Low’ as Urosalpinx cinerea would have to be physically removed, and sensitivity to invasion by Urosalpinx cinerea is assessed as  ‘High’.