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. Further detail.

Evidence

Intertidal species are exposed to extremes of high and low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period during the tidal cycle. In winter air temperatures are colder than the sea, conversely in summer air temperatures are much warmer than the sea. Species that occur in the intertidal are, therefore, generally adapted to tolerate a range of temperatures, with the width of the thermal niche positively correlated with the height of the shore that the animal usually occurs (Davenport & Davenport, 2005). The species present in this biotope are present in the mid and/or upper shore and therefore are likely to have broad temperature tolerances and to be tolerant of temperature fluctuations.

The amphipods that occur within this habitat are mobile and can avoid unfavourable conditions to some extent. The life cycle of Bathyporeia spp. vary between locations, which is related to temperature (Mettam, 1989). Preece (1971) tested the temperature tolerance of Bathyporeia pilosa in the laboratory. Individuals acclimated to 15°C for 24 hours were exposed to temperature increases (water temperature raised by 0.2°C/minute). As test temperatures were reached individuals were removed, placed in seawater at 4°C and allowed to recover for 24 hours at which point mortalities were tested. Amphipods were also allowed to bury into sediments and held at test temperatures for 24 hours of 32.5°C, 31.8°C and 29.5°C before being allowed to recover in fresh seawater at 15°C for a further 24 hours, before mortalities were assessed. The upper lethal temperature (the temperature at which 50% of individuals died) for adult males and gravid females of Bathyporeia pilosa was 39.4°C. These tests measures short-term exposure only and species had a lower tolerance for longer-term (24-hour exposure). No mortality occurred for Bathyporeia pilosa individuals held at 29.5°C and 30.8°C. However, 15% of individuals exposed to water temperatures of 31.8°C and 96% at 32.5°C died.

The tolerance of Bathyporeia pilosa to high temperatures may be because it is found in the intertidal, although its ability to withstand long-term changes in temperature is more difficult to discern. Whilst Bathyporeia pilosa can tolerate extremely high temperatures for short periods, this species has a limited distribution from the coast of Norway down to the French coast of the Bay of Biscay (Speybroeck et al., 2008), which suggests that it may be susceptible to ocean warming. As temperatures increase, non-lethal effects of heat stress may lead to changes in population dynamics and its biogeographic range moving northwards to compensate.

Corophium volutator occurs in the North-East Atlantic (NEA) from northern Norway to the Bay of Biscay, and in the North-West Atlantic in the Bay of Fundy and along the coast of Maine (www.obis.org). Whilst it is also thought to be present in the Mediterranean Sea, the Black Sea and the Azov Sea (Neal & Avant, 2006), it is expected that these records refer to Corophium orientale, which is a closely related species originally regarded as a subspecies of Corophium volutator (Crawford, 1937). Corophium arenarium has a distribution from Scotland to the French Coast of the Bay of Biscay. When Corophium volutator was exposed to choice experiments to assess temperature preferences between 0°C and 30°C, it was shown that this species preferred temperatures in the range of 15 - 20°C, although activity peaked at 25°C and it was able to withstand temperatures of 30°C (Meadows & Ruagh, 1981).

Temperature increases may lead to indirect effects on populations of the characterizing species, through changes in the distribution and prevalence of parasites and pathogens (see microbial pathogens pressure). Using a simulation model, Mouritsen et al. (2005) demonstrated that a 3.8°C increase in ambient temperature would probably result in a parasite‐induced collapse of the amphipod population in the Wadden Sea.

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. In the south of the UK, summer air temperatures generally reach a mean maximum of 22°C, and under the three scenarios (middle and high emission and extreme) this temperature is projected to increase to temperatures of 25, 26, and 27°C respectively.

Whilst all three species occur in the intertidal and are able to tolerate a wide range of temperatures, all three of these key characterizing species (Bathyporeia pilosa, Corophium volutator and Corophium arenarium) reach their southern limit at the Bay of Biscay, where summer sea temperatures reach up to 21°C (www.seatemperature.org) and air temperatures reach up to 24°C (www.climate-data.org). It is unknown whether its southern limit is controlled by sea or air temperatures, although as this biotope occurs on the mid and high shore, air temperatures may be more significant, and have been taken into primary consideration for this sensitivity assessment.

Under the middle and high emission scenario, this biotope (LS.LSa.MuSa.BatCare) is expected to experience mean maximum air temperatures of up to 25-26°C, which is greater than what this species experiences at its southerly limit and, therefore, some mortality and potential loss of these species from the south of the UK is possible. Therefore, under these scenarios, resistance has been assessed as ‘Medium’, and resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming under the middle and high emission scenarios. Under the extreme scenario, this biotope is predicted to experience mean maximum air temperatures of up to 27°C, and to experience temperatures exceeding those experienced at its southern limit across most of England. Therefore, under this scenario, resistance has been assessed as ‘Low’, whilst resilience has been assessed as ‘Very low’ due to the long term nature of ocean warming. Therefore, this biotope is assessed as having ‘High’ sensitivity to ocean warming under the extreme scenario.

Global warming (high)

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

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

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

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

Evidence

Intertidal species are exposed to extremes of high and low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period during the tidal cycle. In winter air temperatures are colder than the sea, conversely in summer air temperatures are much warmer than the sea. Species that occur in the intertidal are, therefore, generally adapted to tolerate a range of temperatures, with the width of the thermal niche positively correlated with the height of the shore that the animal usually occurs (Davenport & Davenport, 2005). The species present in this biotope are present in the mid and/or upper shore and therefore are likely to have broad temperature tolerances and to be tolerant of temperature fluctuations.

The amphipods that occur within this habitat are mobile and can avoid unfavourable conditions to some extent. The life cycle of Bathyporeia spp. vary between locations, which is related to temperature (Mettam, 1989). Preece (1971) tested the temperature tolerance of Bathyporeia pilosa in the laboratory. Individuals acclimated to 15°C for 24 hours were exposed to temperature increases (water temperature raised by 0.2°C/minute). As test temperatures were reached individuals were removed, placed in seawater at 4°C and allowed to recover for 24 hours at which point mortalities were tested. Amphipods were also allowed to bury into sediments and held at test temperatures for 24 hours of 32.5°C, 31.8°C and 29.5°C before being allowed to recover in fresh seawater at 15°C for a further 24 hours, before mortalities were assessed. The upper lethal temperature (the temperature at which 50% of individuals died) for adult males and gravid females of Bathyporeia pilosa was 39.4°C. These tests measures short-term exposure only and species had a lower tolerance for longer-term (24-hour exposure). No mortality occurred for Bathyporeia pilosa individuals held at 29.5°C and 30.8°C. However, 15% of individuals exposed to water temperatures of 31.8°C and 96% at 32.5°C died.

The tolerance of Bathyporeia pilosa to high temperatures may be because it is found in the intertidal, although its ability to withstand long-term changes in temperature is more difficult to discern. Whilst Bathyporeia pilosa can tolerate extremely high temperatures for short periods, this species has a limited distribution from the coast of Norway down to the French coast of the Bay of Biscay (Speybroeck et al., 2008), which suggests that it may be susceptible to ocean warming. As temperatures increase, non-lethal effects of heat stress may lead to changes in population dynamics and its biogeographic range moving northwards to compensate.

Corophium volutator occurs in the North-East Atlantic (NEA) from northern Norway to the Bay of Biscay, and in the North-West Atlantic in the Bay of Fundy and along the coast of Maine (www.obis.org). Whilst it is also thought to be present in the Mediterranean Sea, the Black Sea and the Azov Sea (Neal & Avant, 2006), it is expected that these records refer to Corophium orientale, which is a closely related species originally regarded as a subspecies of Corophium volutator (Crawford, 1937). Corophium arenarium has a distribution from Scotland to the French Coast of the Bay of Biscay. When Corophium volutator was exposed to choice experiments to assess temperature preferences between 0°C and 30°C, it was shown that this species preferred temperatures in the range of 15 - 20°C, although activity peaked at 25°C and it was able to withstand temperatures of 30°C (Meadows & Ruagh, 1981).

Temperature increases may lead to indirect effects on populations of the characterizing species, through changes in the distribution and prevalence of parasites and pathogens (see microbial pathogens pressure). Using a simulation model, Mouritsen et al. (2005) demonstrated that a 3.8°C increase in ambient temperature would probably result in a parasite‐induced collapse of the amphipod population in the Wadden Sea.

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. In the south of the UK, summer air temperatures generally reach a mean maximum of 22°C, and under the three scenarios (middle and high emission and extreme) this temperature is projected to increase to temperatures of 25, 26, and 27°C respectively.

Whilst all three species occur in the intertidal and are able to tolerate a wide range of temperatures, all three of these key characterizing species (Bathyporeia pilosa, Corophium volutator and Corophium arenarium) reach their southern limit at the Bay of Biscay, where summer sea temperatures reach up to 21°C (www.seatemperature.org) and air temperatures reach up to 24°C (www.climate-data.org). It is unknown whether its southern limit is controlled by sea or air temperatures, although as this biotope occurs on the mid and high shore, air temperatures may be more significant, and have been taken into primary consideration for this sensitivity assessment.

Under the middle and high emission scenario, this biotope (LS.LSa.MuSa.BatCare) is expected to experience mean maximum air temperatures of up to 25-26°C, which is greater than what this species experiences at its southerly limit and, therefore, some mortality and potential loss of these species from the south of the UK is possible. Therefore, under these scenarios, resistance has been assessed as ‘Medium’, and resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming under the middle and high emission scenarios. Under the extreme scenario, this biotope is predicted to experience mean maximum air temperatures of up to 27°C, and to experience temperatures exceeding those experienced at its southern limit across most of England. Therefore, under this scenario, resistance has been assessed as ‘Low’, whilst resilience has been assessed as ‘Very low’ due to the long term nature of ocean warming. Therefore, this biotope is assessed as having ‘High’ sensitivity to ocean warming under the extreme scenario.

Global warming (middle)

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

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

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

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

Evidence

Intertidal species are exposed to extremes of high and low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period during the tidal cycle. In winter air temperatures are colder than the sea, conversely in summer air temperatures are much warmer than the sea. Species that occur in the intertidal are, therefore, generally adapted to tolerate a range of temperatures, with the width of the thermal niche positively correlated with the height of the shore that the animal usually occurs (Davenport & Davenport, 2005). The species present in this biotope are present in the mid and/or upper shore and therefore are likely to have broad temperature tolerances and to be tolerant of temperature fluctuations.

The amphipods that occur within this habitat are mobile and can avoid unfavourable conditions to some extent. The life cycle of Bathyporeia spp. vary between locations, which is related to temperature (Mettam, 1989). Preece (1971) tested the temperature tolerance of Bathyporeia pilosa in the laboratory. Individuals acclimated to 15°C for 24 hours were exposed to temperature increases (water temperature raised by 0.2°C/minute). As test temperatures were reached individuals were removed, placed in seawater at 4°C and allowed to recover for 24 hours at which point mortalities were tested. Amphipods were also allowed to bury into sediments and held at test temperatures for 24 hours of 32.5°C, 31.8°C and 29.5°C before being allowed to recover in fresh seawater at 15°C for a further 24 hours, before mortalities were assessed. The upper lethal temperature (the temperature at which 50% of individuals died) for adult males and gravid females of Bathyporeia pilosa was 39.4°C. These tests measures short-term exposure only and species had a lower tolerance for longer-term (24-hour exposure). No mortality occurred for Bathyporeia pilosa individuals held at 29.5°C and 30.8°C. However, 15% of individuals exposed to water temperatures of 31.8°C and 96% at 32.5°C died.

The tolerance of Bathyporeia pilosa to high temperatures may be because it is found in the intertidal, although its ability to withstand long-term changes in temperature is more difficult to discern. Whilst Bathyporeia pilosa can tolerate extremely high temperatures for short periods, this species has a limited distribution from the coast of Norway down to the French coast of the Bay of Biscay (Speybroeck et al., 2008), which suggests that it may be susceptible to ocean warming. As temperatures increase, non-lethal effects of heat stress may lead to changes in population dynamics and its biogeographic range moving northwards to compensate.

Corophium volutator occurs in the North-East Atlantic (NEA) from northern Norway to the Bay of Biscay, and in the North-West Atlantic in the Bay of Fundy and along the coast of Maine (www.obis.org). Whilst it is also thought to be present in the Mediterranean Sea, the Black Sea and the Azov Sea (Neal & Avant, 2006), it is expected that these records refer to Corophium orientale, which is a closely related species originally regarded as a subspecies of Corophium volutator (Crawford, 1937). Corophium arenarium has a distribution from Scotland to the French Coast of the Bay of Biscay. When Corophium volutator was exposed to choice experiments to assess temperature preferences between 0°C and 30°C, it was shown that this species preferred temperatures in the range of 15 - 20°C, although activity peaked at 25°C and it was able to withstand temperatures of 30°C (Meadows & Ruagh, 1981).

Temperature increases may lead to indirect effects on populations of the characterizing species, through changes in the distribution and prevalence of parasites and pathogens (see microbial pathogens pressure). Using a simulation model, Mouritsen et al. (2005) demonstrated that a 3.8°C increase in ambient temperature would probably result in a parasite‐induced collapse of the amphipod population in the Wadden Sea.

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. In the south of the UK, summer air temperatures generally reach a mean maximum of 22°C, and under the three scenarios (middle and high emission and extreme) this temperature is projected to increase to temperatures of 25, 26, and 27°C respectively.

Whilst all three species occur in the intertidal and are able to tolerate a wide range of temperatures, all three of these key characterizing species (Bathyporeia pilosa, Corophium volutator and Corophium arenarium) reach their southern limit at the Bay of Biscay, where summer sea temperatures reach up to 21°C (www.seatemperature.org) and air temperatures reach up to 24°C (www.climate-data.org). It is unknown whether its southern limit is controlled by sea or air temperatures, although as this biotope occurs on the mid and high shore, air temperatures may be more significant, and have been taken into primary consideration for this sensitivity assessment.

Under the middle and high emission scenario, this biotope (LS.LSa.MuSa.BatCare) is expected to experience mean maximum air temperatures of up to 25-26°C, which is greater than what this species experiences at its southerly limit and, therefore, some mortality and potential loss of these species from the south of the UK is possible. Therefore, under these scenarios, resistance has been assessed as ‘Medium’, and resilience has been assessed as ‘Very low’ due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming under the middle and high emission scenarios. Under the extreme scenario, this biotope is predicted to experience mean maximum air temperatures of up to 27°C, and to experience temperatures exceeding those experienced at its southern limit across most of England. Therefore, under this scenario, resistance has been assessed as ‘Low’, whilst resilience has been assessed as ‘Very low’ due to the long term nature of ocean warming. Therefore, this biotope is assessed as having ‘High’ sensitivity to ocean warming under the extreme scenario.

Marine heatwaves (high)

High emission scenario benchmark: A marine heatwave occurring every two years, with a mean duration of 120 days, and a maximum intensity of 3.5°C. Further detail.

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). Both Bathyporeia pilosa and Corophium volutator are known to be able to withstand temperatures of 30°C (see Global Warming). Upper lethal temperatures (the temperature at which 50 % of individuals died) for adult males and gravid females of Bathyporeia pilosa were 39.4°C.

Many intertidal species can alter metabolic activity, burrow deeper in sediment or move further down the shore to tolerate or escape unfavourable conditions. Bathyporeia pilosa burrows to a depth of 10 cm (Nicolaisen & Kanneworff, 1969), whilst Corophium volutator and Corophium arenarium have burrows with mean depths of 4.35 cm and 5.27 cm respectively (Jensen & Kristensen, 1990), although sediment temperatures in the first few centimetres of sediment can exceed air temperatures (Harrison & Phizacklea, 1987). At low tide, air temperature becomes critically important to intertidal animals, and on sandy beaches, the substratum, from the surface to a depth of several centimetres, can experience large variations in temperature during a single cycle and throughout the year (Harrison & Phizacklea, 1987; Hayward, 1994). For instance, Khayrallah & Jones (1980b) reported the temperature range of sand at a depth of 1 cm during neap tides to be from -2°C in February 1973, to a maximum of 25°C in July 1977. Although adapted to temperature change, extremely warm temperatures that occur during a heatwave may result in a seasonal reduction in species richness and abundance. Negative impacts of heatwaves may not only be related to temperature stress but may be related to the resultant changes in other factors, especially reductions in oxygen resulting from enhanced microbial activity (Hayward, 1994; Eltringham, 1971).

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, this could lead to summer sea temperatures reaching up to 24°C in southern England, whilst mean maximum air temperatures could reach 27°C. These temperatures are higher than the characterizing species of this biotope experience at their southern limit in the Bay of Biscay (see Global Warming), although both these species have been shown to be able to tolerate short term temperatures of ≥ 30°C. Therefore, some mortality cannot be ruled out, although recovery is likely to be rapid. Resistance has been assessed as ‘Medium’ and recovery has been assessed as ‘High’, leading to an assessment of ‘Low’ for this biotope under the middle emission scenario.

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, this could lead to the heatwave lasting the entire summer with sea temperatures reaching up to 26.5°C, and mean maximum air temperatures exceeding 30°C across the south of the UK. Under this scenario, the characterizing species of this biotope may suffer significant mortality through acute physiological stress and, therefore, resistance has been assessed as ‘Low’. These species are known to have long reproductive seasons during their short lifespans, and as some individuals remain, recovery is likely to be rapid. Resilience has been assessed as ‘High’, leading to a sensitivity assessment of ‘Low’ under the 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. Further detail.

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). Both Bathyporeia pilosa and Corophium volutator are known to be able to withstand temperatures of 30°C (see Global Warming). Upper lethal temperatures (the temperature at which 50 % of individuals died) for adult males and gravid females of Bathyporeia pilosa were 39.4°C.

Many intertidal species can alter metabolic activity, burrow deeper in sediment or move further down the shore to tolerate or escape unfavourable conditions. Bathyporeia pilosa burrows to a depth of 10 cm (Nicolaisen & Kanneworff, 1969), whilst Corophium volutator and Corophium arenarium have burrows with mean depths of 4.35 cm and 5.27 cm respectively (Jensen & Kristensen, 1990), although sediment temperatures in the first few centimetres of sediment can exceed air temperatures (Harrison & Phizacklea, 1987). At low tide, air temperature becomes critically important to intertidal animals, and on sandy beaches, the substratum, from the surface to a depth of several centimetres, can experience large variations in temperature during a single cycle and throughout the year (Harrison & Phizacklea, 1987; Hayward, 1994). For instance, Khayrallah & Jones (1980b) reported the temperature range of sand at a depth of 1 cm during neap tides to be from -2°C in February 1973, to a maximum of 25°C in July 1977. Although adapted to temperature change, extremely warm temperatures that occur during a heatwave may result in a seasonal reduction in species richness and abundance. Negative impacts of heatwaves may not only be related to temperature stress but may be related to the resultant changes in other factors, especially reductions in oxygen resulting from enhanced microbial activity (Hayward, 1994; Eltringham, 1971).

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, this could lead to summer sea temperatures reaching up to 24°C in southern England, whilst mean maximum air temperatures could reach 27°C. These temperatures are higher than the characterizing species of this biotope experience at their southern limit in the Bay of Biscay (see Global Warming), although both these species have been shown to be able to tolerate short term temperatures of ≥ 30°C. Therefore, some mortality cannot be ruled out, although recovery is likely to be rapid. Resistance has been assessed as ‘Medium’ and recovery has been assessed as ‘High’, leading to an assessment of ‘Low’ for this biotope under the middle emission scenario.

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, this could lead to the heatwave lasting the entire summer with sea temperatures reaching up to 26.5°C, and mean maximum air temperatures exceeding 30°C across the south of the UK. Under this scenario, the characterizing species of this biotope may suffer significant mortality through acute physiological stress and, therefore, resistance has been assessed as ‘Low’. These species are known to have long reproductive seasons during their short lifespans, and as some individuals remain, recovery is likely to be rapid. Resilience has been assessed as ‘High’, leading to a sensitivity assessment of ‘Low’ under the 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. Further detail 

Evidence

Amphipods are generally thought to be less sensitive to ocean acidification than some other taxa and are found in greater numbers at naturally CO₂ enriched vents (Kroeker et al., 2011, Garrard et al., 2014, Vizzini et al., 2017). This increase in abundance is not directly related to CO₂ enrichment, but rather due to indirect effects such as reduced predation or increased food supply.  A laboratory study found that under CO₂ enrichment, the population size of the amphipod Gammarus locusta increased 20 fold and the proportion of gravid females doubled, suggesting that ocean acidification may confer an advantage to amphipods by relaxing environmental constraints on reproduction (Heldt et al., 2016). Further laboratory experiments show little effect of ocean acidification at levels expected for the high emission scenario at the end of this century (pH 7.8) (Hauton et al., 2009, Hale et al., 2011, Lim & Harley, 2018).

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005). When exposed experimentally to CO₂ enrichment that led to a decrease in pH of 0.2 units and 0.34 units, Corophium volutator showed no mortality, although at the higher pH decrease (0.34 units) there was a significant increase in DNA damage from 10 - 17 % (Roberts et al., 2013). There is no direct evidence of the impact of ocean acidification on amphipods from the genus Bathyporeia.

Sensitivity Assessment. Corophium volutator appears to be generally robust to levels of ocean acidification projected for the end of this century. Although increased DNA damage may lead to some negative effects, this is a short-lived, highly fecund species, and it is unlikely to lead to population-level effects. Direct evidence of the impact of ocean acidification on Bathyporeia pilosa is lacking. However, in general, amphipods appear to be tolerant. Therefore, it is likely that the characterizing species of this biotope will show a ‘High’ resistance to a decrease in pH. Hence, the biotope is assessed as ‘High’ resistance to ocean acidification, based on the evidence available under both the middle and high emission scenarios. Resilience is assessed as ‘High’ leading to an assessment of ‘Not sensitive’ at the benchmark level.

Ocean acidification (middle)

Middle emission scenario benchmark: a further decrease in pH of 0.15 (annual mean) and 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. Further detail.

Evidence

Amphipods are generally thought to be less sensitive to ocean acidification than some other taxa and are found in greater numbers at naturally CO₂ enriched vents (Kroeker et al., 2011, Garrard et al., 2014, Vizzini et al., 2017). This increase in abundance is not directly related to CO₂ enrichment, but rather due to indirect effects such as reduced predation or increased food supply.  A laboratory study found that under CO₂ enrichment, the population size of the amphipod Gammarus locusta increased 20 fold and the proportion of gravid females doubled, suggesting that ocean acidification may confer an advantage to amphipods by relaxing environmental constraints on reproduction (Heldt et al., 2016). Further laboratory experiments show little effect of ocean acidification at levels expected for the high emission scenario at the end of this century (pH 7.8) (Hauton et al., 2009, Hale et al., 2011, Lim & Harley, 2018).

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005). When exposed experimentally to CO₂ enrichment that led to a decrease in pH of 0.2 units and 0.34 units, Corophium volutator showed no mortality, although at the higher pH decrease (0.34 units) there was a significant increase in DNA damage from 10 - 17 % (Roberts et al., 2013). There is no direct evidence of the impact of ocean acidification on amphipods from the genus Bathyporeia.

Sensitivity Assessment. Corophium volutator appears to be generally robust to levels of ocean acidification projected for the end of this century. Although increased DNA damage may lead to some negative effects, this is a short-lived, highly fecund species, and it is unlikely to lead to population-level effects. Direct evidence of the impact of ocean acidification on Bathyporeia pilosa is lacking. However, in general, amphipods appear to be tolerant. Therefore, it is likely that the characterizing species of this biotope will show a ‘High’ resistance to a decrease in pH. Hence, the biotope is assessed as ‘High’ resistance to ocean acidification, based on the evidence available under both the middle and high emission scenarios. Resilience is assessed as ‘High’ leading to an assessment of ‘Not sensitive’ at the benchmark level.

Sea level rise (extreme)

Extreme scenario benchmark: a 107 cm rise in average UK by the end of this century (2018-2100). Further detail.

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 2001; Crooks, 2004; Fujii, 2012).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position within the coastal energy gradient (Cooks, 2004; Fujii, 2012).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (2001) suggested that a sea-level rise rate of 6 mm/yr.could result in landward movement of estuaries by 10 m/yr., long-shore migration of open coast landforms of 50 m/yr. and ebb-tidal deltas to extend laterally by 300 m/yr.  Of course, there is no guarantee that landward migration to compensate for sea-level rise will be possible.

The effects of sea-level rise and increased wave action increase 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).

The LS.LSa.MuSa.BatCare biotope occurs most commonly on medium to muddy sand on the upper and mid-shore of broad estuaries (JNCC, 2015). Therefore, an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. Beds may be able to expand their range and migrate landwards to compensate for sea-level rise, if not constrained by lack of suitable habitat or human modification of the environment (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Bathyporeia pilosa and Corophium arenarium in littoral muddy sand will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. The possibility to migrate inshore will be site-specific, hence, the assessment is based on a worst-case-scenario, i.e. that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), a third of this biotope may be lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small portion of this biotope may be lost. Under the medium and high emission scenarios, resistance has been assessed as ‘Medium’, as less than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘Medium’.  Under the high emission and extreme scenarios, resistance has been assessed as ‘Low’, as it is possible that more than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘High’, albeit with ‘Low’ confidence.

Sea level rise (high)

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

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 2001; Crooks, 2004; Fujii, 2012).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position within the coastal energy gradient (Cooks, 2004; Fujii, 2012).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (2001) suggested that a sea-level rise rate of 6 mm/yr.could result in landward movement of estuaries by 10 m/yr., long-shore migration of open coast landforms of 50 m/yr. and ebb-tidal deltas to extend laterally by 300 m/yr.  Of course, there is no guarantee that landward migration to compensate for sea-level rise will be possible.

The effects of sea-level rise and increased wave action increase 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).

The LS.LSa.MuSa.BatCare biotope occurs most commonly on medium to muddy sand on the upper and mid-shore of broad estuaries (JNCC, 2015). Therefore, an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. Beds may be able to expand their range and migrate landwards to compensate for sea-level rise, if not constrained by lack of suitable habitat or human modification of the environment (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Bathyporeia pilosa and Corophium arenarium in littoral muddy sand will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. The possibility to migrate inshore will be site-specific, hence, the assessment is based on a worst-case-scenario, i.e. that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), a third of this biotope may be lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small portion of this biotope may be lost. Under the medium and high emission scenarios, resistance has been assessed as ‘Medium’, as less than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘Medium’.  Under the high emission and extreme scenarios, resistance has been assessed as ‘Low’, as it is possible that more than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘High’, albeit with ‘Low’ confidence.

Sea level rise (middle)

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

Evidence

A rise in sea level increases the water depth at the shore and results in increased wave and tidal energy along the shore, due to the increase in fetch and reduction in wave attenuation (Pethick, 2001; Crooks, 2004; Fujii, 2012).  As a result, coast landforms (e.g. subtidal bedforms, intertidal flats, saltmarshes, shingle banks, sand dunes, cliffs and coastal lowlands) migrate along and parallel to the shore to maintain their position within the coastal energy gradient (Cooks, 2004; Fujii, 2012).  For example, mudflats migrate landwards to a lower energy position and may be replaced by sand flats that have themselves migrated landwards from exposed conditions (Crooks, 2004).  In effect, ‘coastal roll-over’ results as the shore moves landwards by the erosion of the landward, upper limit, of the shore and deposition at its lower limit (Crooks, 2004).  Pethick (2001) suggested that a sea-level rise rate of 6 mm/yr.could result in landward movement of estuaries by 10 m/yr., long-shore migration of open coast landforms of 50 m/yr. and ebb-tidal deltas to extend laterally by 300 m/yr.  Of course, there is no guarantee that landward migration to compensate for sea-level rise will be possible.

The effects of sea-level rise and increased wave action increase 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).

The LS.LSa.MuSa.BatCare biotope occurs most commonly on medium to muddy sand on the upper and mid-shore of broad estuaries (JNCC, 2015). Therefore, an increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope. Beds may be able to expand their range and migrate landwards to compensate for sea-level rise, if not constrained by lack of suitable habitat or human modification of the environment (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Bathyporeia pilosa and Corophium arenarium in littoral muddy sand will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. The possibility to migrate inshore will be site-specific, hence, the assessment is based on a worst-case-scenario, i.e. that landward migration is not possible.

Sensitivity assessment. The mean tidal range in the UK varies from 127 cm in the Shetland Islands to 972 cm at Avonmouth, in the Bristol Channel (Woodworth et al., 1991). This large difference in tidal amplitudes suggests that this biotope will be more affected in some parts of the UK than others. In Scotland and Ireland, where mean tidal range is generally less than 3 m (Woodworth et al., 1991), a third of this biotope may be lost under the extreme scenario, whereas in the Bristol Channel, where mean tidal range exceeds 9 m (Woodworth et al., 1991), only a small portion of this biotope may be lost. Under the medium and high emission scenarios, resistance has been assessed as ‘Medium’, as less than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘Medium’.  Under the high emission and extreme scenarios, resistance has been assessed as ‘Low’, as it is possible that more than 25% of this biotope could be lost. Resilience has been assessed as ‘Very low’, due to the long-term nature of sea-level rise.  Therefore, sensitivity is assessed as ‘High’, albeit with ‘Low’ confidence.

Temperature increase (local)

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

Evidence

Intertidal species are exposed to extremes of high and low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period of time during the tidal cycle. In winter air temperatures are colder than the sea, conversely in summer air temperatures are much warmer than the sea. Species that occur in the intertidal are therefore generally adapted to tolerate a range of temperatures, with the width of the thermal niche positively correlated with the height of the shore that the animal usually occurs at (Davenport & Davenport, 2005). The species present in this biotope are present in the mid and/or upper shore and therefore are likely to have broad temperature tolerances and to be tolerant of temperature fluctuations.

Many intertidal species can alter metabolic activity, burrow deeper in sediment or move to deeper water to tolerate or escape unfavourable conditions. At low tide, air temperature becomes critically important to intertidal animals, and on sandy beaches the substratum, from the surface to a depth of several centimetres, can experience large variations in temperature during a single cycle and throughout the year (Hayward, 1994). For instance, Khayrallah & Jones (1980b) reported the temperature range of sand at a depth of 1 cm during neap tides to be from -2°C in February 1973, to a maximum of 25°C in July 1977. Although adapted to temperature change, severe acute change may result in seasonal reduction in species richness and abundance, but the effects of an acute temperature increase are not necessarily direct and may be more related to the resultant changes in other factors, especially reductions in oxygen resulting from enhanced microbial activity (Hayward, 1994; Eltringham, 1971).

The amphipods that occur within this habitat are mobile and can avoid unfavourable conditions to some extent. Bathyporeia life cycles vary between locations and this is related to temperature (Mettam, 1989). Preece (1971) tested temperature tolerances of Bathyporeia pilosa in the laboratory. Individuals acclimated to 15°C for 24 hours were exposed to temperature increases (water temperature raised by 0.2°C/minute). As test temperature were reached individuals were removed, placed in seawater at 4°C and allowed to recover for 24 hours at which point mortalities were tested. Amphipods were also allowed to bury into sediments and held at test temperatures for 24 hours of 32.5°C, 31.8°C and 29.5°C before being allowed to recover in fresh seawater at 15°C for a further 24 hours, before mortalities were assessed. Upper lethal temperatures (the temperature at which 50% of individuals died for adult males and gravid females of Bathyporeia pilosa were 39.4°C. These tests measures short-term exposure only and species had lower tolerance for longer-term (24 hour exposure). No mortality occurred for Bathyporeia pilosa individuals held at 29.5°C and 30.8°C; however 15% of individuals exposed to water temperatures of 31.8°C and 96% at 32.5°C died. 

Corophium volutator is widely distributed in the north Atlantic, American and European coasts, from western Norway to the Mediterranean and the Black Sea and Azov Sea (Neal & Avant, 2006). The amphipod is subject to temperatures of 1°C in the winter to 17°C in the summer (Wilson & Parker, 1996) but can resist much higher temperatures (Meadows & Ruagh, 1981).

Temperature increases may lead to indirect effects on populations of the characterizing species, through changes in the distribution and prevalence of parasites and pathogens (see microbial pathogens pressure). Using a simulation model, Mouritsen et al. (2005) demonstrated that a 3.8 °C increase in ambient temperature would probably result in a parasite‐induced collapse of the amphipod population in the Wadden Sea. This indirect effect is assessed in the introduction of microbial pathogens pressure.

Sensitivity assessment. Typical surface water temperatures around the UK coast vary seasonally from 4-19°C (Huthnance, 2010). A chronic increase in temperature throughout the year of 2°C may fall within the normal temperature variation and an acute increase in water temperatures from 19 to 24°C for a month may be tolerated by the characterizing species supported by deeper burrowing and/or migration. For Bathyporeia spp. temperature increases above 30^oC appear to be critical based on Preece (1971). For intertidal biotopes air temperature may be more critical in the hottest parts of the year although nocturnal emergence of Bathyporeia spp. and isopods or burrowing life habit may also enhance survival. Biotope resistance is therefore assessed as ‘High’ and resilience as ‘High’ so that the biotope is assessed as ‘Not sensitive’. 

Temperature decrease (local)

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

Evidence

Intertidal species are exposed to extremes of high and low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period of time during the tidal cycle. In winter air temperatures are colder than the sea, conversely in summer air temperatures are much warmer than the sea. Species that occur in the intertidal are therefore generally adapted to tolerate a range of temperatures, with the width of the thermal niche positively correlated with the height of the shore that the animal usually occurs at (Davenport & Davenport, 2005).

Crisp (1964) reported that species of amphipod and isopods seemed to be unharmed by the severe winter of 1962-1963. This may be due to burial in sediments buffering temperature or seasonal migration to deeper waters to avoid freezing. In the winter migrations have also been observed for Bathyporeia spp. (Fish & Fish, 1978; Fish & Preece, 1970).

Preece (1971) tested temperature tolerances of Bathyporeia pilosa in the laboratory. Individuals acclimated to 15°C for 24 hours were placed in a freezer in wet sediment. As test temperatures were reached individuals were removed and allowed to recover for 24 hours at which point mortalities were tested. Amphipods were also allowed to bury into sediments and held at test temperatures of -1°C, -3°C and -5°C for 24 hours before being allowed to recover in fresh seawater at 15°C for a further 24 hours before mortalities were assessed. The lower lethal short-term tolerances of Bathyporeia pilosa were -13.6°C. Bathyporeia pilosa individuals could withstand temperatures as low as -1°C for 24 hours, at -3°C, 5% of Bathyporeia pilosa died but this rose to 82% at -5°C.

Corophium volutator is subject to temperatures of 1°C in the winter to 17°C in the summer (Wilson & Parker, 1996). The population may reduce activity and delay reproduction if the temperature drops below 7°C. Sudden pulses of very cold water can disrupt the circa-tidal rhythms of Corophium volutator by resetting the onset of swimming behaviour. For example, a 6 hour cold spell would lead to the population trying to swim at low tide and leave them vulnerable to increased predation. However, it took temperatures of 15-20°C below ambient temperature to induce this response (Holmström & Morgan, 1983b).

Drolet et al. (2013) sampled two intertidal mudflats in the upper Bay of Fundy, Canada, over two consecutive winters (2009–2011), where sediment temperature, 5 cm deep, reached −2°C. The authors concluded Corophium volutator survived winter through  physiological adaptations.  

Sensitivity assessment. Typical surface water temperatures around the UK coast vary seasonally from 4-19°C (Huthnance, 2010). A chronic decrease in temperature throughout the year of 2°C may fall within the normal temperature variation.  An acute decrease in water temperatures from 4°C to -1°C at the coldest part of the year may lead to freezing and lethal effects on exposed individuals but the characterizing species may survive by burrowing deeper into sediments and/or migration. For Bathyporeia pilosa seawater temperature decreases below -1°C appear to be critical based on Preece (1971). For intertidal biotopes air temperature may be more important than seawater temperatures in the coldest parts of the year and nocturnal emergence of Bathyporeia pilosa may reduce survival although these species are more likely to be quiescent during cold periods or to have migrated to the subtidal. Biotope resistance is therefore assessed as ‘High’ and resilience as ‘High’ so that the biotope is assessed as ‘Not sensitive’. 

Salinity increase (local)

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

Evidence

This biotope is found in full salinity (30-35 ppt) and variable salinity habitats (18-35 ppt) (JNCC, 2015). Biotope examples found in variable salinity are considered to tolerate a change to full salinity as this falls within the natural range. A change at the pressure benchmark is therefore assessed as a change to hypersaline conditions (>40 ppt) from full salinity. Little evidence was found to assess responses to hypersalinity. However, monitoring at a Spanish desalination facility where discharges close to the outfall reached a salinity of 53, found that amphipods were sensitive to the increased salinity and that species free-living in the sediment were most sensitive. The study area did not host any of the species characterizing this biotope but the results indicate a general sensitivity (De-la-Ossa-Carretero et al., 2016).

Bathyporeia pilosa appears to be confined to intertidal sand sediments. Corophium volutator occupies a greater variety of habitats and is found in fine sediments of mud flats, salt marsh pools and brackish ditches. It is therefore exposed in its natural range to salinity extremes, from freshwater flushing during periods of high rainfall and increases in salinity during periods of high evaporation. Corophium volutator is, therefore, an exceptionally euryhaline species able to tolerate 2-50 psu (McLusky, 1968) but growth is fastest at 15-20 psu (McLusky, 1967; McLusky, 1970 cited in Meadows & Ruagh, 1981). The interstitial salinity is more important for Corophium volutator than that of the overlying water. Sustained periods of increased salinity are required to alter that of the interstitial water and there is a lag between salinity changes and the response of Corophium volutator (McLusky, 1968). Sudden increases in salinity delay swimming activity (Harris & Morgan, 1984a). Corophium volutator will also migrate from areas of unfavourable salinity (McLusky, 1968). Corophium arenarium is more tolerant of Corophium volutator to short-term increases in salinity <45‰ (Mills & Fish, 1980).

Sensitivity assessment: The characterizing species of this biotope are euryhaline and likely to be resistant of an increase in salinity. However, a chronic increase at the pressure benchmark level is likely to result in the death of a portion of the individuals in the population. Resistance is therefore assessed as ‘Low’ and resilience is likely to be ‘High’ so the biotope is considered to have ‘Low’ sensitivity to an increase in salinity at the pressure benchmark level.

Salinity decrease (local)

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

Evidence

This biotope is found in full salinity (30-35 ppt) and variable salinity habitats (18-35 ppt) (JNCC, 2015). A change at the pressure benchmark refers to a change to low (<18 ppt salinity).  

Speybroeck et al. (2008) noted that Bathyporeia pilosa tends to occur subtidally in estuarine and brackish conditions. Bathyporeia pilosa is tolerant of low salinities and it is capable of reproducing at salinities as low as 2 (Khayrallah, 1977). Populations of Bathyporeia pilosa within the upper reaches of the Severn Estuary experience wide fluctuations in salinity ranging from 1-22 depending on the season and tidal cycle (Mettam, 1989). The physiological stress resulting from this environment, however, affects size and reproduction (Mettam, 1989).

Corophium volutator is an exceptionally euryhaline species able to tolerate 2-50 psu (McLusky, 1968) but growth is fastest at 15-20 psu (McLusky, 1970 cited in Meadows & Ruagh, 1981). Corophium volutator requires a salinity greater than 5 psu in order to moult, since osmoregulation is lost during moulting (McLusky, 1967). A salinity of at least 7.5 psu is required for reproduction (McLusky, 1968). Mills & Fish (1980), found that Corophium volutator is more tolerant of low salinity (2 to 10‰) than Corophium arenarium. Females undergoing a pre-copulatory moult failed to lay eggs below salinities of 3‰ (Corophium volutator) and 10‰ (Corophium arenarium), but in both species the lowest salinity at which all females moulted and laid eggs was 20‰. Changes in salinity are very unlikely to cause mortality but may alter population distribution and abundances within the biotope as the species is likely to move to more favourable conditions or to suffer some effects on reproduction with resulting decreases in abundance.

Sensitivity assessment: Both the characterizing Bathyporeia pilosa and Corophium spp. are likely to resist a decrease in salinity at the pressure benchmark level, although there may be some decreases in abundance and body size due to physiological stress, depending on the degree of decrease. Resistance is therefore assessed as ‘Medium’ and resilience as ‘High’ (following restoration of the usual salinity regime), and biotope sensitivity is assessed as ‘Low’.

Water flow (tidal current) changes (local)

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

Evidence

Water movement from tidal streams is a key factor physically structuring this biotope (JNCC, 2015). Sediment sorting and mobilization by tidal streams modifies the sediments present and creates a suitable habitat for Bathyporeia pilosa that is more typical of wave disturbed habitats. A change in water flow at the pressure benchmark may be significant as sand fractions are not cohesive (particularly where these occur higher on the shore and are coarser and drier) and are eroded at lower current velocities that muds and silts. Sand particles are most likely to be eroded at about 0.20 m/s (based on the Hjulström-Sundborg diagram, Sundborg, 1956).

Bathyporeia spp. are tolerant of disturbed and mobile sandy sediments and can re-position and burrow. They are therefore unlikely to be impacted by an increase at the pressure benchmark even if this results in more sediment disturbance and a change to coarser sands. Changes in sediment from sands to muddy sands following increased deposition or a change from muddy sand to clean sand may alter abundances and distribution of Corophium volutator and Corophium arenarium due to their differing sediment preferences (see physical change pressures). Decreases in flow rate are not considered to negatively impact Corophium volutator which inhabits muddy sediments in low energy environments such as saltmarsh pools.  

Sensitivity assessment. The sediments that characterize this biotope and sub-biotopes are mobile medium to fine sands and sometimes muddy sands. A change at the pressure benchmark (increase or decrease) may lead to some changes in sediment sorting. However, it is unlikely that a decrease at the pressure benchmark would lead to the development of a mud sediment as fine silts require lower velocities to deposit than erode. An increase at the pressure benchmark may lead to increased erosion of fine silts but the biotope is likely to persist and some sediment restoration is likely during the slower neap tides. Biotope resistance is therefore assessed as ‘High’ and resilience as ‘High’ (by default), so that the biotope is 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. Further detail

Evidence

This biotope is found on the lower, mid and upper shore (JNCC, 2015). On sandy shores (as in rocky shores) zonation of species occurs in relation to tidal height, in general, talitrid amphipods tend to occur higher on the shore where sediments dry out at low tide and macroalgae is deposited, while lower down the shore more species rich assemblages occur and suspension feeding bivalves are more abundant in the damper sediments (McLachlan et al., 1995). The zonation pattern is mediated by local hydrodynamics and the associated sediment characteristics. This biotope is typically found higher up the shore than sandflats with the cockle Cerastoderma edule in the large sandy estuaries of the west coast of England and Wales. In moderately exposed conditions, this biotope can occur on the mid shore below LS.LSa.St.Tal and/or LS.LSa.MoSa. In more sheltered conditions, this biotope may occur above LS.LMu.MEst.NhomLimStr (JNCC, 2015).

Changes in emergence are likely to result in some shifts in the shoreward or landward extent of the biotope group due to changes in the length of time the biotope is uncovered and covered. Decreased emergence could lead to biotopes occurring on the lower shore becoming sub-littoral or colonized by bivalves and more species rich assemblages. Similarly upper shore biotopes that are at the limit of the ebbing tide could revert to strandline biotopes LS.LSa.St.Tal or revert to the species-poor biotope LS.LSa.MoSa or the Scolelepis squamata dominated variant (LS.LSa.MoSa.AmSco.Sco) due to that species tolerance for well-drained habitats

Sensitivity assessment. Sandy beaches exhibit vertical zonation of biotopes due to species differing tolerances for emergence. Species within the biotope exhibit physiological and behavioural adaptations to changes and the biotope as a group is unlikely to be very sensitive to changes in emergence that do not alter the shore zone (from mid-eulittoral to supralittoral, for example). Changes in emergence may result in some changes in upper and lower extents of the biotope and the sub-biotopes and biotope resistance is therefore assessed as ‘Medium’, resilience is assessed as ‘High’ (following restoration of tidal regime) and biotope sensitivity is therefore ‘Low’. 

Wave exposure changes (local)

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

Evidence

This biotope occurrs in a range of estimated wave exposures, including; moderately exposed, sheltered, very sheltered and extremely sheltered (JNCC, 2015). Water movement is an important factor that physically structures this biotope through sediment sorting and disturbance, but tidal streams rather than wave action are the key element (JNCC, 2015).

Sensitivity assessment. As the biotope occurs across four wave exposure categories (JNCC, 2015) this is considered to indicate, by proxy, that a change in wave exposure at the pressure benchmark is less than the natural range of wave heights experienced, (particularly as wave height and hence the corresponding change will be small). Biotope resistance to this pressure is therefore assessed as ‘High’ and resilience as ‘High (by default) so that the biotope is considered to be ‘Not sensitive’ at the pressure benchmark.

Transition elements & organo-metal contamination

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

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Transitional metal or organometal' contaminants on selected amphipod species are summarized below. The full 'evidence reviews' 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 metal exposure was reported to result in ‘severe’ mortality in 11.4% of worst-case results examined, ‘significant’ in 73.4%, ‘some’ in 5.7%, ‘no’ mortality in 3.8% and only sublethal effects in 5.7% of the worst-case results examined.  The majority of studies used amphipods as a standard bioassay, based on standard 96-hour or 10-day assays, to derive LC50 endpoints. Hence, the results were dominated by ‘significant’ (i.e. 50%) mortality endpoints.

Metal toxicity was shown to vary with salinity, temperature, pH and season, as well as between species (Bigongiari et al., 2004; Kater et al., 2000; Prato et al., 2008; Re et al., 2009; Roberts et al., 2013). Roberts et al. (2013) concluded that exposure to pCO₂ (ocean acidification) increased the toxicity of the metal contaminated sediment, probably due to the effects of pH on their physiology rather than metal speciation. Shipp & Grant (2006) suggested that Corophium was more exposed to Cu in pore water than in sediment because of it lives in U-shaped burrows. Pérez-Landa et al. (2008) reported that Corophium multisetosum was more sensitive to ammonia and Cd in summer and less sensitive in the winter.

Strode & Balode (2013) was the only article to examine the effects of metal exposure on Bathyporeia directly. Sensitivity to metal exposure (Cd, Cu and Zn) varied significantly between species, while Cd was the most toxic metal tested. They reported that Bathyporeia pilosa exhibited moderate resistance compared to other tested amphipods, with the freshwater species Gammarus pulex being among the most sensitive. Strode et al. (2017) examined the effect of metal contaminated sediment on Bathyporeia pilosa and Corophium volutator. Bathyporeia pilosa exhibited survival rates ranging from 38% to 100%, while Corophium volutator had a survival range of 70 to 95%, indicating varied sensitivity between species, although the sediment was described as low toxicity. Warwick (2001) noted that Corophium volutator was absent from the heavy metal polluted areas of the Fal estuary, Cornwall.

Overall, ‘severe’ mortality was reported in 11.4% of the results, including Corophium and Ampelisca and ‘significant’ mortality in 73.4% of the results examined, including Bathyporeia. Therefore, the worst-case resistance to transitional metal exposure is assessed as ‘Low’ in Bathyporeia and ‘None’ in Corophium and Ampelisca. Resilience is probably ‘Medium’ for all species, and worst-case sensitivity to transitional metal exposure is assessed as ‘Medium’ in Bathyporeia, Corophium and Ampelisca.

Organometals. Stronkhorst et al. (1999) was the only study of the effects of TBT on Corophium volutator found. They reported 10-day LC50s at low (ng) levels of TBT in both pore water (329 ng/l TBT) and spiked sediment (2,185 ng/g TBT). They concluded that TBT was very toxic to Corophium volutator, especially in pore water.  No evidence on Ampelisca or Bathyporeia was found.  Therefore, the worst-case resistance to organotin is assessed as ‘Low’, resilience as ‘Medium’ and sensitivity as ‘Medium. The assessment is probably applicable to all three species of amphipod based on similarities in their physiology and endocrinology, but confidence in the assessment is ‘Low’ as evidence is limited to a single article.

Nanoparticulate metals. The toxicity of zinc oxide nanoparticulates (ZnO NPs) was studied in only four papers. In all cases, the articles compared the toxicity of ZnO NPs with inorganic zinc, as ZnO, ZnCl₂ or ZnSO₄ (Fabrega et al., 2012; Larner et al., 2012; Vimercati et al., 2020; Righi et al., 2023). In all cases, the authors concluded that zinc nanoparticulates were no more toxic than inorganic zinc and that the toxicity was due to Zn ions and that dissolution of the nanoparticulates contributed to their toxicity.  The articles reported a range of mortalities from ‘severe’ to ‘none’, depending on the concentration tested. For example, Vimercati et al. (2020) reported a 96-hour LC50 for ZnO NPs of 1.75 mg/l, and Righi et al. (2023) reported a 96-hour LC50 of 1.93 mg/l ZnO-NP in Corophium insidiosum. However, Larner et al. (2012) reported >80% survival in Corophium volutator exposed to ZnO-NPs (around 1,000 ng/g Zn). Nevertheless, Fabrega et al. (2012) noted that all forms of zinc tested significantly affected growth and could have long-term effects on populations of Corophium volutator.

Righi et al. (2023) also examined the toxicity of calcium phosphate nanoparticles (CaP-N). They concluded that CaP-N) were not toxic to Corophium insidiosum and Gammarus aequicauda as their 96-hour LC50s were >100 mg/l.  They also noted that Corophium insidiosum incorporated the CaP-N nanoparticles into their tubes during the 96-hour experiment. However, no evidence of the effects of other metal nanoparticulates was found.  Overall, the worst-case resistance to the effects of zinc nanoparticulates is assessed as ‘None’. Resilience is probably ‘Medium’. Hence, the worst-case sensitivity is assessed as ‘Medium’. Confidence in the assessment is ‘Low’ due to the limited number of studies found.

Overall sensitivity assessment of 'Transitional metals and organometals'. Bathyporeia sp. and Corophium spp are important characteristic species. The above evidence reported 'severe' mortality in Corophium sp. due to exposure to transitional metals and zinc-based nanoparticulates and 'significant' mortality due to organotin exposure, while Bathyporeia sp. experienced 'significant' mortality due to transitional metal exposure. Bathyporeia sp. was reported to be less sensitive than Corophium sp.; the overall worst-case resistance to transitional metals and organosmetals is assessed as 'None' based on the evidence for Corophium. Resilience is probably 'Medium', so sensitivity is assessed as 'Medium'. The toxicity of metals varies with salinity, temperature, pH, and season, as well as between species, and the metals themselves. Hence, the confidence in the assessment is 'Low'. 

Hydrocarbon & PAH contamination

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

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Hydrocarbon & PAH' contaminants on selected amphipod species are summarized below. The full 'evidence reviews' 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. 

Oil spills. Numerous studies reported the sensitivity of amphipods to oil spills (Suchanek, 1993). In particular, Ampelisca spp., Bathyporeia sp. and Corophium sp. were reported to be highly susceptible to oil pollution due to their sediment-dwelling habit and low dispersal (Cabioch et al., 1978; Dauvin, 1987, 1998; Gesteira & Dauvin, 2000; Poggiale & Dauvin, 2001). Amphipods were also reported to suffer declines after other oil spills. For example, significant declines in Corophium sp. after the Exxon Valdez oil spill (Wolfe, 1996) and Rose Bay oil spill (Roddie et al., 1994); and a ‘sharp decline’ in Ampelisca sp. after the Sea Empress spill (Nikitik & Robinson, 2003). Rostron (1998) also concluded that amphipods were sensitive to the effects of the Sea Empress spill. Smith (1968) reported that Corophium declined in heavily contaminated sites after the Torrey Canyon spill due to oil and dispersants. Widbom & Oviatt (1994) also reported significant declines in Ampelisca after the World Prodigy oil spill but no significant effects on Corophium. Kingston et al. (1997) reported that amphipods were completely absent from heavily impacted sites after the Braer oil spill, and all sites in the path of the oil had reduced numbers. Dauvin (1987) reported that Bathyporeia experienced severe declines after the Amoco Cadiz oil spills. However, Dauvin (1987) also reported that Bathyporeia recovered within three to five years due to its high mobility, while Corophium crassicorne had not returned after ca eight years, and Poggiale & Dauvin (2001) reported that recovery in Ampelisca sp. took up to 15 years.

Overall, 76% of the worst-case results reported ‘severe’ mortality, and 11.5% reported ‘significant’ mortality due to oil spills across the selected amphipods examined. The weight of evidence suggests that Amphipoda are highly sensitive to oil spills. However, recovery rates may vary between species but depend on the time taken for the contamination to disperse (or be removed) and recruitment from unaffected populations. Therefore, the resistance of amphipods to oil spills is assessed as ‘None’. The direct evidence of recovery after the Amoco Cadiz oil spills suggests that resilience should be assessed as ‘Medium’ in Bathyporeia but ‘Low’ for Ampelisca and Corophium.Therefore, sensitivity to oil spills is assessed as ‘Medium’ for Bathyporeia but ‘High’ for Corophium and Ampelisca.

Petrochemical hydrocarbons. The reported toxicity varied between studies, types, concentrations, and forms of oil fractions tested. However, 20% of the worst-case results reported ‘severe’ mortality, 35% reported ‘significant’ mortality, and 10% reported ‘some’ mortality, while ‘no’ mortality was reported in 15% of the results examined in the review. For example, the abundance of Corophium was significantly lower in the oiled treatments compared to controls, and experienced ‘severe’ mortality in sediment spiked with 60 mg/kg of Maya crude oil (Sanz-Lázaro & Marín, 2009). McClusky (1982) reported that Corophium volutator was absent within 500 m of the petrochemical effluent discharge in the intertidal mudflats of the Kinneil area, Forth Estuary, Scotland. McClusky & Martins (1998) reported that Corophium populations on the Kinneil mudflats increased in 1985 after the closure of the phenol plant. They noted that Corophium abundance was correlated significantly with a decrease in chemical effluents from the phenol or acrylonitrile plants and the resultant decrease in BOD (biological oxygen demand). Van Eenennaam et al. (2018) simulated the effects of the Marine Oil Snow Sedimentation and Flocculent Accumulation (MOSSFA) event that occurred after the Deepwater Horizon (DWH) oil spill. The MOSSFA transported ca 14% of the total released oil to the seabed. The oil treatment (10 g/m² for 16 days in laboratory mesocosms) severely reduced survival to 9% of controls (91% mortality) while the ‘snow and oil’ treatment reduced survival to 20% (80% mortality), probably because Corophium escaped the snow layer on the sediment, and possibly fed on the artificial snow.

Oil fractions also modified behaviour. Percy (1977) reported that Corophium clarencense actively avoided oil-spiked sediments in choice experiments. Kienle & Gerhardt (2008) noted that WAF affected locomotor activity and burial in Corophium volutator.Scarlett et al. (2007) reported that reproduction was significantly less in the DWAF and all oil treatments than in controls, Corexit and WAF treatments.

In his review, Suchanek (1993) identified amphipods, particularly ampeliscid species, as highly sensitive to oil contamination, often experiencing severe population declines and failing to recover for at least five years post-spill, likely due to the longevity of oil in sediments. Therefore, the worst-case resistance to petrochemical hydrocarbons is assessed as ‘None’ for amphipods as a group, and in particular, Corophium. No evidence on the resistance of Bathyporeia was found. Hence, resilience is assessed as ‘Medium’, and the worst-case sensitivity to petrochemical hydrocarbons in the selected amphipods is assessed as ‘Medium’.

Dispersants. Scarlett et al. (2005) examined the effects of the dispersants Corexit 9527 and Superdispersant-25 (SD-25) on a selection of invertebrates, including Corophium volutator. Mortality was 100% at 500 ppm of both dispersants after 48 hours, and at 375 ppm in Corexit 9527. The mortality from exposure to Corexit 9527 was significantly greater than SD-25, and recovery rates of survivors from SD-25 were higher than from Corexit 9527. Scarlett et al. (2005) noted that dispersant concentrations of ≥175 ppm reduced burrowing in the laboratory, which, in the wild, might allow Corophium to leave contaminated sediments.  Scarlett et al. (2007) reported some (25%) mortality in adults exposed to Corexit 9527 in acute tests, but no significant mortality in neonates in chronic tests.  Overall, the evidence suggests that Corophium is sensitive to the dispersants examined, although no information on Bathyporeia or Ampelisca was found. Therefore, the worst-case resistance of selected amphipods to dispersants (Corexit 9527 and Superdispersant-25) is assessed as ‘None’. Resistance is probably ‘Medium’, so the worst-case sensitivity to dispersants is assessed as ‘Medium’ but with ‘Low’ confidence due to the limited number of studies found.

Polyaromatic hydrocarbons (PAHs). PAH exposure was reported to result in ‘significant’ mortality in 87% of worst-case results examined, and only sublethal effects in 13% of the results examined. The majority of the results were derived from 10-day LC50s in sediment bioassays. For example, Boese et al. (1997) concluded that photoactivation significantly increased the toxicity of fluoranthene. In Corophium insidiosum, fluoranthene exposure alone resulted in a 96-hour LC50 of 85 µg/l, which decreased to 32 µg/l after UV exposure, while the EC50 for reburial declined from 54 µg/L to 20 µg/l, indicating an increase in toxicity after UV exposure. Similarly, Spehar et al. (1999) reported that fluoranthene was toxic under fluorescent light, with a 96-hour LC50 of 67 µg/l  in Ampelisca abdita, but under UV exposure, toxicity increased significantly, leading to complete mortality at concentrations that were too low to determine an LC50.

In sediment bioassays, Fisher et al. (2011) determined acute toxicity values (10-day LC50) for 12 PAHs in Corophium volutator, which ranged from 24 mg/kg for 4-methyldibenzothiophene to over 1,000 mg/kg for anthracene (see evidence summary spreadsheet). Fisher et al. (2011) concluded that PAH contamination in sediments can pose a significant risk to benthic organisms, with certain compounds demonstrating much higher toxicity than others. Swartz et al. (1990) exposed Corophium spinicorne and Rhepoxynius abronius to fluoranthene in sediment. Corophium spinicorne was less sensitive to fluoranthene exposure than Rhepoxynius abronius, with an interstitial water 10-day LC50 of 37.9 µg/l compared to 23.8 µg/l for Rhepoxynius abronius. Swartz et al. (1990) concluded that Corophium spinicorne was partly protected from the contaminated interstitial water because it lives in burrows connected to the overlying water. Ciarelli et al. (1999) noted that PAH exposure affected burrowing activity and bioturbation in Corophium volutator.  Overall, the worst-case resistance to PAH exposure in Corophium sp. and Ampelisca sp. is assessed as ‘Low’ based on the evidence above. Resilience is probably ‘Medium’, so sensitivity is assessed as ‘Medium’.

Petroleum hydrocarbons – others. Brown et al. (1999) reported that exposure to high concentrations of 5-nonylphenol resulted in mortality, while lower concentrations resulted in significant reduction in both survival and growth and interfered with reproduction, with potential population effects. Redmond & Scott (1987, cited from ECOTOX) reported 100% (‘severe’) mortality in Ampelisca after exposure to 150 mg/l of phenol.  In mating experiments, Krang (2007) reported that exposure to 0.5 and 5 µg/g naphthalene significantly reduced the ability of males to locate mature females, which could prevent mating and reproduction. The population dynamics of amphipods are marked by rapid population growth and significant predation. A reduction in mating success could significantly affect amphipod populations and the populations of other species dependent on them for food. Overall, the limited evidence above suggests that Corophium and Ampelisca have ‘Low’ and ‘None’ resistance to nonylphenol and phenol, respectively. Hence, resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’ but with ‘Low’ confidence.

Overall sensitivity to 'Hydrocarbons and PAHs'. The evidence of the effects of hydrocarbons on Bathyporeia is limited. However, the evidence above suggests that amphipods are highly sensitive to oils and petrochemical hydrocarbons, especially Ampelisca sp. Therefore, the worst-case resistance to 'Hydrocarbon & PAH' exposure is assessed as 'None', resilience as 'Medium' and sensitivity as 'Medium'.

Synthetic compound contamination

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

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Synthetic' contaminants on selected amphipod species are summarized below. The full 'evidence reviews' 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.

Pesticides/biocides. Pesticide exposure was reported to result in ‘severe’ mortality in 11.5% of worst-case results examined, ‘significant’ in 60.6%, ‘some’ in 0%, ‘no’ mortality in 0% and only sublethal effects in 27.8% of the worst-case results examined. ‘Severe’ (>75%) mortality was reported in five articles. Dumbauld et al. (2001) reported a 97% loss in abundance of Corophium in their large-scale experimental treatment of mudflats with carbamate pesticide, used to control mud and ghost shrimp, although not in all of their experiments. Collier & Pinn (1998) reported 100% mortality in Corophium volutator exposed to 80.0 mg/m² of ivermectin for 14 days (the highest concentration tested). Similarly, Davies et al. (1998) reported >92% mortality after exposure to sediment spiked with ≥0.7 mg/kg of ivermectin after 10 days, with a 10-day LC50 value of 0.18 mg/kg. Mayor et al. (2008) reported severe mortality in Corophium volutator due to exposure to several pesticides (cypermethrin, emamectin benzoate, and azamethipos). Thursby & Berry (1987; cited from ECOTOX) reported 100% mortality in Corophium after 96-hour exposure to 37.5 µg/l of the organophosphate pesticide hydroxydiazinon. ‘Significant’ mortality (25-75%) was reported by another 15 articles, mainly as LC50s based on 24-, 48-, 96-hour or 10-day bioassays. Overall, 72% of the worst-case results reported ‘severe’ to ‘significant’ mortality due to pesticide exposure. Hence, the worst-case resistance to pesticide exposure is assessed as ‘None’. Resilience is probably ‘Medium’, so sensitivity is assessed as ‘Medium’.

Pharmaceuticals. Mayor et al. (2008) reported 93% mortality in Corophium exposed to sediment spiked with 100 mg/kg (the highest concentration tested) of the antibiotic oxytetracycline in 10-day bioassay and a 10-day LC50 of 414 µg/kg of oxytetracycline. Gronlund et al. (2004) determined 10-day LC50s for the analgesic diclofenac and antidepressant citalopram in spiked sediment bioassays. They noted that Corophium volutator mortality increased drastically in a narrow range of concentrations. Diclofenac was more toxic than citalopram to Corophium volutator. However, Gronlund et al. (2024) suggested that citalopram is unlikely to directly pose a lethal threat to the Corophium volutator community based on citalopram concentrations in water and sediment/water partitioning observed in the experiments. Results from the mixture study found a synergistic effect, meaning the combined impact of the contaminants was greater than expected (Gronlund et al. 2024). Overall, the results indicate that the worst-case resistance to pharmaceutical exposure should be assessed as ‘None’. Hence, resilience is probably ‘Medium’ and sensitivity is assessed as ‘Medium’. However, the number of studies, species, and pharmaceuticals tested was limited, so the assessment is made with ‘Low’ confidence.

Polychlorinated biphenyls (PCBs). Reish (1993) reported a 96-hour LC50 of 0.009 mg/l PCB (Arochlor 1254) in water-only assays of Corophium insidiosum. Ho et al. (1997) reported that Aroclor 1242 was more toxic to Ampelisca abdita than Aroclor 1254. The 96-hour LC50 value for Aroclor 1242 was 10 ppb, and for Aroclor 1254 was 40 ppb, respectively. Ho et al. (1997) concluded that PCBs were likely to be the source of chronic and reproductive toxicity in the New Bedford Harbor (NBH) sediments studied. Overall, ‘significant’ mortality due to PCB exposure was reported in both studies and both PCBs examined. Therefore, the worst-case resistance to PCB exposure is assessed as ‘Low’, resilience as ‘Medium’ and sensitivity as ‘Medium’. However, the number of studies, species, and PCBs tested was limited, so the assessment is made with ‘Low’ confidence.

Phthalates. Tagatz et al. (1983) examined the effect of di-n-butyl phthalate (DBP) exposure on communities that colonized sediment in the laboratory. The number of Corophium acherusicum significantly decreased to eight individuals in the 0.34 mg/l DBP treatment. There were zero Corophium acherusicum individuals recorded in the 3.7 mg/l DBP treatment. Corophium acherusicum in laboratory colonized communities were significantly affected by as low a concentration as 0.34 mg/l DBP. Tagatz & Stanley (1987; cited from ECOTOX) reported a 96-hour LC50 of 450 µg/l DBP in Corophium acherusicum. Ho et al. (1997) examined the toxicity of bis(2-ethylhexyl) phthalate on Ampelisca abdita. They reported a 48-hour LC50 above 1,000 ppb of phthalate for Ampelisca abdita and suggested that the phthalate was not toxic to amphipods in the samples studied. Therefore, the worst-case resistance to phthalate exposure is assessed as ‘None’, resilience as ‘Medium’ and sensitivity as ‘Medium’. However, the number of studies, species, and phthalates tested was limited, so the assessment is made with ‘Low’ confidence.

Other synthetics. Hester et al. (1991) examined the effects of Arosurf MSF (monomolecular surface film) on a range of marine species, including Gammarus sp. No significant mortality was observed in any of the species tested after 96 hours at 47 ml/m² of Arosurf MSF. Picone et al. (2008) and Lera et al. (2008) examined the toxicity of sodium dodecyl sulphate (SDS) surfactants in Corophium orientale and reported 96-hour LC50s of 8.7 mg/l SDS and 3.14 mg/l SDS, respectively. Rico-Rico et al. (2009) examined the toxicity of the surfactant C12-2-LAS (2-(p-sulphophenyl)-dodecane) to Corophium volutator. They reported a 5-day LC50 of 0.73 mg/l but also noted that mortality was not significant at 50 mg/l but increased to 100% at 500 mg/l. Therefore, the worst-case resistance to surfactants may be considered ‘None’ (Rico-Rico et al.,2009) or ‘Low’ (Lera et al., 2008; Picone et al., 2008) depending on the surfactant. In both cases, resilience is probably ‘Medium’, so sensitivity is assessed as ‘Medium’. However, the number of studies, species, and chemicals tested was limited, so the assessment is made with ‘Low’ confidence.

Thursby & Berry (1987, cited from ECOTOX) examined the toxicity of Triethylene glycol diacetate on Ampelisca abdita and reported no mortality after 96 hours at 0.5 µg/l, but did not specify mortality at 156 µg/l. Harris & Morgan (1984b) reported that 1% ethanol was lethal to Corophium volutator after 24 hours. Fastelli & Renzi (2019) examined the toxicity of chemical-based sunscreen and nanoparticulate metal physical-based (TiO2-NPs; ZnO-NPs) sunscreen on Corophium orientalis. The 96-hour LC50 was 82 µl/l for chemical-based sunscreen and 87 µl/l for physical-based sunscreen at 4.0 ppt salinity. Fastelli & Renzi (2019) concluded that chemical-based and physical-based sunscreen types can exert different ecotoxicological responses on the tested species, suggesting that sensitivity toward nanoparticles in personal care products could be very different amongst different marine taxa. 

Overall sensitivity assessment for this pressure. No evidence of the effects of synthetic compounds on Bathyporeia sp. was found. However, Corophium sp. was reported to experience 'severe' or 'significant' mortality due to pesticide, pharmaceutical exposure and surfactants and 'significant' mortality due to PCB exposure. Therefore, the worst-case resistance of this biotope, based on the effect of pesticides/biocides on the important characteristic species Corophium sp., is assessed as ‘None’, resilience as ‘Medium’ and sensitivity is assessed as ‘Medium’. The confidence is assessed as ‘Low’ due to the limited evidence on some types of synthetic chemicals and the variation in toxicity between species and chemicals tested.

Radionuclide contamination

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

Evidence

Insufficient evidence was found to support an assessment of biotope sensitivity. Corophium volutator readily absorbs radionuclides such as americium and plutonium from water and contaminated sediments (Miramand et al., 1982). However, the effect of contamination on the individuals was not known, but accumulation through the food chain was assumed (Miramand et al., 1982).

Introduction of other substances

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

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Other' contaminants on selected amphipod species are summarized below. The full 'evidence reviews' 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.

Most studies of 'Other' substances in selected amphipods reported ‘significant’ mortality as LC50s (see evidence summary spreadsheet). Hydrogen peroxide was examined as a disinfectant in ballast water. Smit et al. (2008) concluded that Corophium volutator had a relatively low sensitivity to hydrogen peroxide compared to Artemia salina, Brachionus plicatilis, Dunaliella teriolecta, and Skeletonema costatum, which were the other species tested in the study. Caldwell (1975) concluded that Corophium salmonis is unable or barely able to tolerate 1.0 mg/l of hydrogen sulphide (as sodium sulphide) during continuous exposure for multiple days. Ferretti et al. (2004) examined the suitability of KCl as a reference toxicant in bioassays. Soto et al. (2000) exposed Ampelisca araucana juveniles to 38.5, 55, 78.5, 112, and 160 mg/l of potassium dichromate for 48 hours, resulting in a 48-hour LC50 value of 56.89 mg/l.  Overall, the worst-case resistance to 'other' substances could be assessed as 'Low' based on the reported LC50s, so resilience is assessed as 'Medium' and sensitivity as 'Medium'. However, the number of studies, species, and chemicals tested was limited, so the assessment is made with ‘Low’ confidence. No evidence for Bathyporeia sp. was found. 

De-oxygenation

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

Evidence

Brafield (1964) concluded that the most significant factor influencing the oxygenation is the drainage of the beach which, in turn, is determined by the slope and particle size. Oxygen depletion becomes a severe problem at all states of the tide on only the very finest grained beaches, and as a general rule, if the percentage of particles of less than 0.25 mm median diameter exceeds 10% of a sediment, then the oxygen concentration of its interstitial water will be less than 20% of the air saturation level, and will drop rapidly during low tide periods (Brafield, 1964). Fine sands tend to have lower oxygen levels because their lower permeability leads to the trapping of detritus which, together with the large surface area for microbial colonization, leads to higher oxygen uptake (Eagle, 1983).

Laboratory studies by Khayrallah (1977) on Bathyporeia pilosa, indicated that it has a relatively poor resistance to conditions of hypoxia in comparison to other interstitial animals. However, Mettam (1989) and Sandberg (1997) suggest that Bathyporeia pilosa can survive short-term hypoxia. In a series of experiments Gamenick et al. (1996) suggested that Corophium volutator is highly sensitive to hypoxia and suffers 50% mortality after just 4 hours in hypoxic conditions, or in 2 hours if there is rapid build-up of sulphide (Gamenick et al., 1996). These results are largely in concordance with other work by Gamble (1970) who found that survival reates were temperature dependent with individuals surviving longer at lower temperatures. The level of oxygen was not assessed by Gamenick et al. (1996) and the description of the experimental set-up suggests that anoxic test conditions were used rather than hypoxic. Gamble (1970) found that at 5°C most individuals were inactive after 30 minutes exposure to anaerobic seawater and that mortality occurred later, the inactivity may have allowed the species to survive longer (Gamble, 1070). At 10°C, Corophium volutator survived for 22 h while Corophium arenarium survived 25 h.

Sensitivity assessment.  Exposure to air and the mixing effect of tidal currents will limit the intensity and duration of exposure to deoxygenated waters. The species characterizing the biotope are also mobile and able to migrate vertically or shorewards to escape unsuitable conditions. Biotope resistance is therefore assessed as ‘High’ and resilience as ‘High’ (by default), so that the biotope is considered to be ‘Not sensitive’.

Nutrient enrichment

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

Evidence

In-situ primary production is limited to microphytobenthos within and on sediments and the high levels of sediment mobility due to resuspension in strong tidal streams may limit the level of primary production as abrasion would be likely to damage diatoms (Delgado et al., 1991).

Sensitivity assessment.  In general primary production is low and this biotope is species poor. Nutrient level is not a key factor structuring the biotope at the pressure benchmark. Biotope resistance is therefore assessed as ‘High’, resilience as ‘High’ (by default), and the biotope is considered to be ‘Not sensitive’. Changes in nutrient status may indirectly affect this biotope where these result in changes in diatom production and inputs of macroalagal debris.

Organic enrichment

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

Evidence

The biotope description for the biotope LS.LSa.MoSa.AmSco  (JNCC, 2015) notes that inputs of organic matter seeping from the driftline can result in high abundances of Bathyporeia pilosa. Organic matter is unlikely to accumulate as sediment mobility and re-suspension by tidal currents  will result in removal of fine organic matter deposits.  

An increase in organic enrichment that exceeded the pressure benchmark may impact the habitat and biological assemblage, particularly in more sheltered areas where deposits can accumulate. For instance, prior to the introduction of a sewage treatment scheme in the Firth of Forth (Scotland), the communities of several sandy beaches were considerably modified by gross sewage pollution (Read et al., 1983).

Sensitivity assessment.  At the pressure benchmark organic inputs are likely to represent a food subsidy for the characterizing species and are unlikely to significantly affect the structure of the biological assemblage or impact the physical habitat. Biotope sensitivity is therefore assessed as ‘High’ and resilience as ‘High’ (by default), and the biotope is therefore considered to be ‘Not sensitive’.

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site. Further detail

Evidence

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

Physical change (to another seabed type)

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

Evidence

The biotope is characterized by the sedimentary habitat (JNCC, 2015), a change to an artificial or rock substratum would alter the character of the biotope leading to reclassification and the loss of the sedimentary habitat and the characterizing amphipods and associated species such as the Baltic tellin, Macoma balthica.

Sensitivity assessment. Based on the loss of the biotope, resistance is assessed as ‘None’, recovery is assessed as ‘Very low’ (as the change at the pressure benchmark is permanent and sensitivity is assessed as ‘High’.

Physical change (to another sediment type)

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

Evidence

The pressure benchmark refers to the simplified Folk classification developed by Long (2006) and the UK Marine Habitat Classification Littoral and Sublittoral Sediment Matrices (Connor et al., 2004). The biotope is associated with medium to fine sands and muddy sand, a change at the pressure benchmark, therefore, refers to a change to muds and sandy muds or coarse sediments (gravelly sands and sandy gravels and gravels) and mixed sediments. The species present have specific sediment preferences related to burrowing or feeding. Experiments by Van Tomme et al. (2013) have shown that the optimal sedimentary habitats for Bathyporeia pilosa were fine sands, at a subtidal dredge disposal site the change to a finer muddy sediment led to a reduction in the abundance of Bathyporeia pilosa (Witt et al., 2004). Habitat partitioning has been observed between Corophium volutator and Corophium arenarium on the basis of sediments, although modified by disturbance from Arenicola marina and Cerastoderma edule (see removal of target species). Where the two Corophium species occur together Corophium arenarium is associated with areas of sandy sediment whilst Corophium volutator prefers muddier sediments (Watkin, 1941; Flach, 1993).

Sensitivity assessment. A change in sediment character would alter the character of the biotope which would be re-classified as another biotope type. A change in sediment type to a coarser, more mobile sediment such as coarse sands would favour Bathyporeia pilosa and other species present in mobile coarse sediments such as oligochaetes or the polychaetes Scolelepis squamata and the biotope would be reclassified within the LS.LSa.MoSa group. The presence of gravels could lead to the development of LS.LCS.Sh biotopes. Alternatively, where changes in hydrodynamic factors lead to the establishment of muddy sediment biotopes Corophium volutator already present in this biotope may be favoured while the habitat would be unsuitable for Corophium arenarium and Bathyporeia pilosa. The biotope present may revert to a more species rich, polychaete and bivalve dominated, shore depending on local conditions and larval transport.  Based on the loss of the biotope, resistance is assessed as ‘None’, recovery is assessed as ‘Very low’ (as the change at the pressure benchmark is permanent and sensitivity is assessed as ‘High’.

Habitat structure changes - removal of substratum (extraction)

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

Evidence

Removal of the substratum to 30 cm would result in the loss of the characterizing species Corophium volutator is characteristic of muddy substrata while Corophium arenarium prefers sand; both construct U-shaped burrows in the surface 10 cm of the sediment (Meadows, 1964). Bathyporeia pilosa burrows between 0-10 cm in sediments (Nicolaisen & Kanneworff, 1969)

Sensitivity assessment. Removal of the sediment to 30 cm would remove the key characterizing species and their habitat. Resistance is therefore assessed as ‘None’. Resilience is assessed as ‘High’ if adjacent populations are present to support recolonization through adult migration and reproduction, so that sensitivity is assessed as ‘Medium’. However, if populations were removed over a wide area recovery could be prolonged and is assessed as ‘Medium’, so that sensitivity is assessed as ‘Medium’. The less precautionary assessment is considered more representative but the caveats outlined in the resilience section should be considered and this assessment may underestimate sensitivity.

Abrasion / disturbance of the surface of the substratum or seabed

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

Evidence

The burrowing life habits of Corophium volutator are likely to provide some protection from abrasion at the surface only. However, any abrasion or physical disturbance is likely to reduce the density of Corophium spp. by emigration and increased mortality. For example, the sediment turnover caused by cockles and lugworms disturbs the burrows of Corophium volutator and caused a significant negative effect on Corophium volutator density as a result of increased rate of swimming making the amphipod more vulnerable to predation (Flach & De Bruin, 1993, 1994).  Corophium arenarium is also sensitive to sediment disturbance from bioturbating species (Flach, 1993).

A number of studies have assessed the effects of trampling on other intertidal amphipods and these assessments are used as a proxy. Comparisons between shores with low and high levels of trampling found that the amphipod Bathyporeia pelagica is sensitive to human trampling, other species including Pontocrates arenarius and the isopod Eurydice affinis also decreased in response to trampling but Bathyporeia pelagica appeared to be the most sensitive  (Reyes-Martínez et al., 2015). Changes in abundance of talitrid amphipods on urban beaches subject to high levels of recreational use was also observed by Bessa et al. (2014), this study compared abundances between samples taken ten years apart and thus the trends observed were not directly attributable to trampling vs beach cleaning or other pressures although they illustrate a general trend in density patterns as recreational use increases. Ugolini et al. (2008) carried out a controlled trampling experiment on Talitrus saltator. Plastic cylinders of 110 cm diameter (area 0.95 m²) were placed in the sand and all individuals trapped and counted, and 400 steps were made in a cylinder in 15 minutes after the amphipods had reburied. The trampling rate was based on observed number of beach users and therefore represents a realistic level of exposure. Alive individuals were counted at the end of the experiment and 24 hours after. Trampling significantly reduced abundance of the amphipods and after 24 hours the percentage of surviving amphipods dropped to almost zero, while survival rates of control (untrampled) amphipods were unaffected. Abrasion and compaction can therefore kill buried amphipods within sediments.

Sensitivity assessment. The trampling experiment (Ugolini et al., 2008) represents a high intensity of abrasion with multiple steps on the sediment within a short time period. The experiment does, however, demonstrate that amphipods are sensitive to abrasion and compaction of the sediment and these results are observed by comparisons between heavily and lightly used areas (Reyes-Martínez et al., 2015; Bessa et al., 2014). Corophium spp. are also sensitive to abrasion and sediment disturbance. Resistance to a single abrasion event is therefore assessed as ‘Low’ based on the characterizing species. Resilience is assessed as ‘High’, based on migration from adjacent populations and in-situ reproduction by surviving amphipods. Sensitivity is therefore assessed as ‘Low’. This assessment may underestimate sensitivity to high-levels of abrasion (repeated events within a short period). The trampling evidence and the evidence for penetration from mobile gears (see below) differ in the severity (resistance) of impact. This may be due to different levels of intensity (multiple trampling/abrasion events vs single penetration/towed gear impacts) or the nature of the pressure. Abrasion from trampling also involves a level of compaction that could collapse burrows and damage species through compression. Penetration may, however, break sediments open allowing mobile species to escape or species may be pushed forwards from towed gear by a pressure wave where this is deployed subtidally (Gilkinson et al., 1998). This suggests that exposure at different states of tide will mediate impacts. Both risk assessments are considered applicable to single events based on the evidence and the sensitivity assessment for both pressures is the same although resistance differs. 

Penetration or disturbance of the substratum subsurface

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

Evidence

In the Columbia river, no significant difference was found in Corophium volutator densities before and after dredging a channel and no difference between the dredged site and a control site (McCabe et al., 1998). Presumably, the dredging did cause mortality of Corophium volutator but recolonization was so rapid that no difference was found. The extraction of cockles by sediment raking and mechanical disturbance and digging for lugworms for bait is likely to cause significant mortality of Corophium volutator. Bait digging was found to reduce Corophium volutator densities by 39%, juveniles were most affected suffering a 55% reduction in dug areas (Shepherd & Boates, 1999).

Bergman and Santbrink (2000) found that direct mortality of gammarid amphipods, following a single passage of a beam trawl (in silty sediments where penetration is greater) was 28%. Similar results were reported from experiments s in shallow, wave disturbed areas, using a toothed, clam dredge. Bathyporeia spp. experienced a reduction of 25% abundance in samples immediately after intense clam dredging, abundance recovered after 1 day (Constantino et al., 2009). Experimental hydraulic dredging for razor clams resulted in  no statistically significant differences in Bathyporeia elegans abundances between treatments after 1 or 40 days (Hall et al., 1990), suggesting that recovery from effects was very rapid. Ferns et al. (2000) examined the effects of a tractor-towed cockle harvester on benthic invertebrates and predators in intertidal plots of muddy and clean sand. Harvesting resulted in the loss of a significant proportion of the most common invertebrates from both areas. In the muddy sand, the population of Bathyporeia pilosa remained significantly depleted for more than 50 days, whilst the population in clean sand recovered more quickly. These results agree with other experimental studies that clean sands tend to recover more quickly that other habitat types with higher proportions of fine sediment (Dernie et al., 2003).

Sensitivity assessment. Based on the evidence above it is considered that Bathyporeia spp. may be more resistant but Corophium species are likely to have ‘Low’ resistance to this pressure and resistance of the biotope is therefore assessed as ‘Low’. Recovery is assessed as ‘High’ and sensitivity is therefore categorised as ‘Low’.

Changes in suspended solids (water clarity)

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

Evidence

The characterizing species live within the sand and are unlikely to be directly affected by an increased concentration of suspended matter in the water column. Within the mobile sands habitat storm events or spring tides may re-suspend or transport large amounts of material and therefore species are considered to be adapted to varying levels of suspended solids.

Bathyporeia spp. feed on diatoms within the sand grains (Nicolaisen & Kanneworff, 1969), an increase in suspended solids that reduced light penetration could alter food supply. However, diatoms are able to photosynthesize while the tide is out and therefore a reduction in light during tidal inundation may not affect this food source, depending on the timing of the tidal cycle. Bathyporeia spp. may be regular swimmers within the surf plankton, where the concentration of suspended particles would be expected to be higher (Fincham, 1970a).

Sensitivity assessment. Increased organic solids in suspension may provide food to the key characterizing species. Increased inorganic suspended solids may increase abrasion but it is likely that the infaunal species would be unaffected. The biotope is considered to be ‘Not sensitive’ to a decrease in suspended solids that does not affect sediment transport and supply to the biotope. Biotope resistance is assessed as ‘Medium’ as some effects on feeding and diatom productivity may occur from increases in suspended solids, resilience is assessed as ‘High’, following a return to usual conditions and sensitivity is assessed as ‘Low’. This more precautionary assessment is presented in the table. Indirect effects such as deposition, erosion and associated sediment change that may result from changes in suspended solids in the long-term are assessed separately. 

Smothering and siltation rate changes (light)

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

Evidence

The pressure benchmark for light deposition refers to the addition of a relatively thin layer of deposits in a single event. Species adapted to coarse sediments may not be able to burrow through fine sediments, or experienced reduced burrowing ability. For example, Bijkerk (1988, results cited from Essink, 1999) found that the maximal overburden through which Bathyporeia could migrate was approximately 20 cm in mud and 40 cm in sand. No further information was available on the rates of survivorship or the time taken to reach the surface.

Corophium volutator was categorized in AMBI sedimentation Group III – ‘species insensitive to higher amounts of sedimentation, but don’t easily recover from strong fluctuations in sedimentation’ (Gittenberger & Van Loon, 2011). Experimental fences placed on mudflats caused sedimentation rates of 2-2.5 cm/month and reduced Corophium volutator densities from approximately 1700 m² to approximately 400 m². In areas without fences, Corophium volutator numbers increased from approximately 1700 per m² to 3500 per m² (Turk & Risk, 1981).

Sensitivity assessment.  As the biotope is associated with shores exposed to strong tidal streams (JNCC, 2015), some sediment removal will occur, mitigating the effect of deposition. Bathyporeia pilosa is likely to be able to burrow through a 5 cm layer of fine sediments. Burrowing ability of Corophium spp. was unclear. Biotope resistance was therefore assessed as a precaution as ‘Medium’ as a deposit of 5 cm could negatively affect Corophium spp. Resilience was assessed as ‘High’ (by default). Biotope sensitivity is therefore assessed as ‘Low’ to this pressure. Repeated deposits or deposits over a large area or in sheltered systems that were shifted by wave and tidal action may result in sediment change (see physical change pressure).

Smothering and siltation rate changes (heavy)

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

Evidence

Studies have found that beach ‘replenishment’ or ‘nourishment’ that involves the addition of sediments on beaches can have a number of impacts on the infauna (Peterson et al., 2000; Peterson et al., 2006). Impacts are more severe when the sediment added differs significantly in grain size or organic content (Nelson et al., 1989; Peterson et al., 2000). For example, Maurer et al. (1981) found that the amphipod Parahaustorius longimerus which occurs intertidally in clean, well-sorted sands and is an active, effective burrower was able to regain the surface after being buried by sand far more easily than when buried under silt/clay mixtures.

Little empirical information was found for the ability of characterizing species to reach the surface after burial. Bijkerk (1988, results cited from Essink, 1999) found that the maximal overburden through which Bathyporeia could migrate was approximately 20 cm in mud and 40 cm in sand. No further information was available on the rates of survivorship or the time taken to reach the surface and no information was available for other characterizing species. Corophium volutator was categorized in AMBI sedimentation Group III – ‘species insensitive to higher amounts of sedimentation, but don’t easily recover from strong fluctuations in sedimentation’ (Gittenberger & Van Loon, 2011). Experimental fences placed on mudflats caused sedimentation rates of 2-2.5 cm/month and reduced Corophium volutator densities from approximately 1700 m² to approximately 400 m². In areas without fences, Corophium volutator numbers increased from approximately 1700 per m² to 3500 per m² (Turk & Risk, 1981).

Repeated events are not considered at the pressure benchmark but it is noted that annual beach nourishment can alter beach sediments (see physical change pressure) and result in suppression of macroinvertebrate populations (Manning et al., 2014).

Sensitivity assessment. Sediment removal by wave action could mitigate the level of effect but overall smothering by fine sediments is likely to result in mortality of characterizing amphipods. Biotope resistance is therefore assessed as ‘Low’, resilience as High’, and biotope sensitivity is assessed as ‘Low’.

Litter

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

Evidence

Not assessed. Corophium volutator is widely used in ecotoxicological studies and know to uptake nanoplastics, but toxicity at the current environmental relevant concentrations has yet to be confirmed (Booth et al., 2015). However, Corophium volutator forms an important food source for several species of birds and mobile predators such as fish and crabs (Hughes, 1988; Jensen & Kristensen, 1990; Raffaelli et al., 1991; Flach & De Bruin, 1994; Brown et al., 1999), which is likely to result in transition of the particles up the marine food chain. Nevertheless, there was insufficient evidence on which to assess the sensitivity of this biotope to the introduction of litter.

Electromagnetic changes

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

Evidence

No evidence for the characterizing species was found to assess this pressure. For some amphipods there is evidence for geomagnetic orientation being inhibited or disrupted by the presence of electromagnetic fields or by changing magnetic fields. Arendse & Barendregt (1981) manipulated magnetic fields to alter orientation of the talitrid amphipod Orchestia cavimana. 

Deep-water amphipods Gondogenia arctica have been shown to be sensitive to even weak electromagnetic fields which cancel magnetic orientation (Tomanova & Vacha, 2016). Loss of orientation was observed at a radiofrequency electromagnetic field of 2 nT (0.002  µT) (Tomanova & Vacha, 2016).

Underwater noise changes

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

Evidence

Corophium spp. and Bathyporeia pilosa are probably sensitive to surface vibrations but little is known about the effects of noise on invertebrates. However, the species are unlikely to be affected by noise pollution and so the biotope is assessed as ‘Not sensitive’.

Introduction of light or shading

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

Evidence

As this feature is not characterized by the presence of primary producers it is not considered that shading would alter the character of the habitat. No specific evidence was found to assess sensitivity of the characterizing species to this pressure. Changes in light levels may, however, affect activity rhythms of the invertebrates. Bathyporeia spp. emerge at night and therefore an increase in light may inhibit activity, particularly at night when they emerge from the sediment and are most active (Jelassi et al., 2015; Ayari, 2015). Hartwick (1976), for example, found that artificial lighting interfered with learning or orientation cues by Talitrids. Corophium spp.  live in burrows and may not be affected by changes in light although males emerge and crawl over sediments (Meadows, 1964).

Orientation by light has been well studied for intertidal amphipods (particularly Talitrus saltator). Intertidal amphipods orientate themselves by a range of factors that include (but are not limited to) visual cues based on solar or astronomic cues such as the moon and the geomagnetic field (Scapini, 2014). Activity patterns are also linked to internal biological clocks that respond to diel, tidal, lunar and seasonal cycles, so that animals are active during the most suitable time of day or night (Scapini, 2014). The introduction of light or an increase in shading could therefore alter behavioural patterns and navigation. As responses may be species specific or vary according to local factors or individual needs such as feeding, mating, it is not possible to provide a simple assessment for this species. Some sensitivity is however likely if incident light levels were altered. This will depend, however, on the footprint and intensity of impact. Fanini et al. (2014) found no difference in abundance of Talitrus saltator between Greek beaches that frequently hosts small scale beach-parties with lights at night and those that were not used in this way. 

Changes in light and level of shade may indirectly affect the characterizing amphipods through changes in food supply via photosynthesis of diatoms within sediments. Benthic microalgae play a significant role in system productivity and trophic dynamics, as well as habitat characteristics such as sediment stability (Tait & Dipper, 1998). Shading could prevent photosynthesis leading to death or migration of sediment diatoms altering sediment cohesion and food supply to the characterizing amphipods.

Sensitivity assessment. Changes in light are not considered to directly affect the biotope, however, some changes in the behaviour of Bathyporeia spp. and food supply for the characterizing amphipods could result. Biotope resistance is assessed as ‘Medium’ and resilience is assessed as ’High’. Biotope sensitivity is therefore assessed as ‘Low’.

Barrier to species movement

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

Evidence

As the amphipods that characterize this biotope have benthic dispersal strategies (via brooding), water transport is not a key method of dispersal over wide distances, as it is for some marine invertebrates that produce pelagic larvae. Where populations were removed changes in water transport of adults may, however, be reduced by changes in local hydrodynamics preventing recolonization. Conversely, a barrier may enhance local connectivity by reducing the loss of adults from the system.

Sensitivity assessment. The biotope (based on the brooding life strategy of the characterizing species) is considered to have ‘High’ resistance to the presence of barriers that lead to a reduction in tidal excursion, resilience is assessed as ‘High’ (by default) and the biotope is considered to be ‘Not sensitive’.

Death or injury by collision

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

Evidence

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

Visual disturbance

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

Evidence

In substratum choice experiments, Corophium volutator preferred substrata independently of whether the dishes were in darkness or illuminated (Meadows, 1964). Bathyporeia pilosa emerges from the sediments at night and are unlikely to be disturbed although like many species they may flee from movements. Therefore, visual disturbance is probably 'Not relevant'.

Genetic modification & translocation of indigenous species

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

Evidence

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

Introduction or spread of invasive non-indigenous species

Benchmark. The introduction of one or more invasive non-indigenous species (INIS). Further detail

Evidence

The sand sediments characterizing this biotope are likely to be mobile due to strong tidal streams and contain little organic matter and the biotopes occurrence in variable salinity limits the establishment of marine and coastal invasive non-indigenous species as the habitat conditions are unsuitable for most species, as exemplified by the low species richness characterizing this biotope.

On the upper shore colonization of sandy mud flats by cord grass Spartina anglica may affect the distribution of amphipods. In the Dovey Estuary, the spread of this species has restricted the distribution of Corophium spp. to creeks and drainage channels (Fish & Mills, 1979).

The Pacific oyster, Magallana (syn. Crassostrea) gigas, is native to warm temperate regions from the northwest Pacific to Japan and northeast Asia, including Cape Mariya (Russia) to Hong Kong (China) (Carrasco & Baron, 2010; 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 northeastern Adriatic Sea, where it was not introduced for aquaculture (Ezgeta-Balic et al., 2019). It was also suggested that some Magallana gigas populations were established in southwest England from France possibly via fouling on ships (GBNNSS, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).

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

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

Magallana gigas requires hard substrata for successful settlement and establishment, including littoral rock, bedrock, chalk, bare boulders, cobbles and pebbles and shells (Kochmann, 2012; Kochmann et al., 2013; Mckinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). It also prefers mudflats with mixed sediment composed of shingle and sand, attaching to whatever hard substrata are available within otherwise unsuitable fine muddy sediment (Spencer et al., 1994; Mckinstry & Jensen, 2013; Tillin et al., 2020).

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

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

The oyster reefs, in the Wadden Sea and Brittany, on littoral muddy and sandy habitats formed predominantly at lower tidal levels from Mean Low Water levels to the shallow subtidal (Herbert et al., 2012, 2016). Pacific oyster spatfall was recorded in the estuarine intertidal zone on areas with hard substrata of stone and shell, particularly between the low water of spring tides and high water of neap tides, such as in the Menai Strait (Spencer et al., 1994).

At high densities the Pacific oyster reef smothers sediment, provides hard substrata in an otherwise sedimentary environment with additional niches for colonization by other species that require hard substratum (e.g. barnacles), and changes surface roughness and local hydrography (Troost, 2010; Herbert et al., 2012, 2016; Tillin et al., 2020). Lejart & Hily (2011) found the surface available for epibenthic species in the Bay of Brest, increased 4-fold when oysters were present on mud, for every 1 m² of colonized substrata the oyster reef added 3.87 m² of surface area on mud sediment. An increase in available settlement substrata, free of epibiota, could be the reason oyster reefs see an increase in macrofaunal abundance. This can change the community composition and habitat structure in reefs on soft mud sediments, creating new habitats for an increasing abundance of infaunal and epibenthic mobile species (Kochmann et al., 2008; Lejart & Hily, 2011; Zwerschke et al., 2018). Results have shown 38% of species present in the oyster reefs on mud were characteristic of rocky substratum habitats (Lejart & Hily, 2011).

In the Bay of Brest, Pacific oyster reefs had a higher diversity and species richness than surrounding mud habitats, including the mud underneath the reefs, where the population was dominated by carnivores rather than suspension the feeders found on the mudflats (Lejart & Hily, 2011; Herbert et al., 2012). In addition, in muddy habitats around the UK, Ireland and Northern France, macrofaunal diversity increased as Pacific oyster density increased but epifaunal diversity decreased as oyster densities increased (Zwerschke et al., 2018). It was suggested that the decrease in epifaunal diversity was due to a decrease in settlement space and a increase in habitat fragmentation because of dense oyster assemblages (Zwerschke et al., 2018).

Green & Crowe (2014) examined the effects of Magallana gigas density in experimental plots (0.25 m²) in Lough Swilly and Lough Foyle, Ireland. The number of species and species diversity increased with oyster cover on mudflats, depending on site and duration. The assemblage also changed due to the increased abundance of barnacles and bryozoans on the oyster shells and polychaetes within the sediment (Green & Crowe, 2014). Zwerschke et al. (2020) suggested that Pacific oyster beds could replace the ecosystem services provided by native oysters, in areas where native oysters had been lost. Morgan et al. (2021) suggested that the smothering of sediment habitats could prevent fish and bird species from feeding on infauna like worms, molluscs, and crustaceans. Also, the development of tidepools within mixed Pacific oyster and blue mussel reefs in soft sediment intertidal sites has been observed in the Wadden Sea, which can create new microhabitats within the reefs (Weniger et al., 2022).

Pacific oysters have been found to reduce the proportion of fine particles and increase the proportion of large particles in the mud under the reef (Lejart & Hily, 2011). The evidence suggests that Pacific oyster reefs change sediment characteristics, by affecting nutrient cycling and increasing the organic content of sediment, sand-to-silt ratio and levels of porewater ammonium (Kochmann et al., 2008; Padilla, 2010; Wagner et al., 2012 cited in Tillin et al., 2020; Green & Crowe, 2014; Herbert et al., 2012, 2016; Zwerschke et al., 2020; Hansen et al., 2023).  Zwerschke et al. (2020) found no significant differences in nutrient cycling rates of native oyster beds or Magallana gigas beds or their associated benthic communities, in experimental plots in Ireland. Persistent changes in the rates of nutrient cycling were driven by the density and presence of oysters (Zwerschke et al., 2020).

The deposition of faeces and pseudo-faeces by Magallana gigas can increase the toxic levels of sulphide in sediments and associated hypoxic sediment conditions, which can reduce photosynthesis and growth in eelgrass (Kelly & Volpe, 2007). Faecal deposition and hypoxia have also been suggested to explain a reduction in species diversity in the sediment underlying high density oyster reefs (Green & Crowe, 2013, 2014; Herbert et al., 2016). However, Lejart & Hily (2011) observed no organic or silt enrichment by Pacific oysters in mud beneath oyster reefs in the Bay of Brest, and no significant difference in the amount of organic matter found in the mud underneath oyster reefs and on bare mud not colonized by the oyster. The biodeposits excreted by the oyster may be washed away by powerful tides and currents seen in the Bay of Brest and the effects of organic enrichment at oyster reefs might be minimal due to wave action (Lejart & Hily, 2011).

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

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

High densities of Crepidula fornicata cause ecological impacts on sedimentary habitats. The species can form dense carpets that can smother the seabed in shallow bays, changing and modifying the habitat structure. 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 (Stiger-Pouvreau & Thouzeau, 2015, Blanchard, 2009). For example, in fine sands, the community is replaced by a reef of slipper limpets, that provide hard substrata for sessile suspension-feeders (e.g., sea squirts, tube worms and fixed shellfish), while mobile carnivorous microfauna occupy species between or within shells, resulting in a homogeneous Crepidula dominated habitat (Blanchard, 2009). Blanchard (2009) suggested the transition occurred and became irreversible at 50% cover of the limpet. De Montaudouin et al. (2018) suggested that homogenization occurred above a threshold of 20-50 Crepidula /m².

Impacts on the structure of benthic communities will depend on the type of habitat that Crepidula colonizes. De Montaudouin & Sauriau (1999) reported that in muddy sediment dominated by deposit-feeders, species richness, abundance and biomass increased in the presence of high densities of Crepidula (ca 562 to 4772 ind./m²), in the Bay of Marennes-Oléron, presumably because the Crepidula bed provided hard substrata in an otherwise sedimentary habitat. In medium sands, Crepidula density was moderate (330-1300 ind./m²) but there was no significant difference between communities in the presence of Crepidula. Intertidal coarse sediment was less suitable for Crepidula with only moderate or low abundances (11 ind./m²) and its presence did not affect the abundance or diversity of macrofauna. However, there was a higher abundance of suspension–feeders and mobile Crustacea in the absence of Crepidula (De Montaudouin & Sauriau, 1999). The presence of Crepidula as an ecosystem engineer has created a range of new niche habitats, reducing biodiversity as it modifies habitats (Fitzgerald, 2007). De Montaudouin et al. (1999) concluded that Crepidula did not influence macroinvertebrate diversity or density significantly under experimental conditions, on fine sands in Arcachon Bay, France. De Montaudouin et al. (2018) noted that the limpet reef increased the species diversity in the bed, but homogenised diversity compared to areas where the limpets were absent. In the Milford Haven Waterway (MHW), the highest densities of Crepidula were found in areas of sediment with hard substrata, e.g., mixed fine sediment with shell or gravel or both (grain sizes 16-256 mm) but, while Crepidula density increased as gravel cover increased in the subtidal, the reverse was found in the intertidal (Bohn et al., 2015). Bohn et al. (2015) suggested that high densities of Crepidula in high-energy environments were possible in the subtidal but not the intertidal, suggesting the availability of this substratum type is beneficial for its establishment. Hinz et al. (2011) reported a substantial increase in the occurrence of Crepidula off the Isle of Wight, between 1958 and 2006, at a depth of ca 60 m, on hard substrata (gravel, cobbles, and boulders), swept by strong tidal streams. Presumably, Crepidula is more tolerant of tidal flow than the oscillatory flow caused by wave action which may be less suitable (Tillin et al., 2020).

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

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

The density of Crepidula populations in northern Europe (Germany, Denmark, and Norway) was significantly lower (ca <100/ m2) than in southern waters. Thieltges et al. (2004) reported that the population of Crepidula was affected strongly by cold winters in the Wadden Sea. The winters of 2001 and 2003 resulted in ca 56-64% mortality of intertidal Crepidula and up to 97% on one mussel bed, compared to only 11-14% in southern areas without frost. Crepidula almost vanished from the Wadden Sea after the 1978/79 winter and took ten years to recover due to moderate winters which regularly affected the population. Similarly, 25% mortality was observed in Crepidula populations on the south coast of the UK after the extreme 1962/63 winter (Crisp, 1964, Bohn et al., 2012). Thieltges et al. (2003) suggested that global warming may allow Crepidula populations to become more abundant in northern Europe.

Sensitivity assessment. The above evidence suggests that this biotope is unsuitable for the colonization of Crepidula fornicata due to a lack of gravel, shells, or any other hard substrata used for larvae settlement (Tillin et al., 2020), especially in moderately exposed examples in which storms may mobilise the sediment (JNCC, 2022). Despite the extremely sheltered examples of the biotope that would otherwise be suitable for Crepidula, the mobility of the sediment is unsuitable and makes it unlikely for Crepidula to become established. There may be higher densities of Crepidula in the lower shore examples of the biotope but the densities may be lower in the mid-shore, and it may be absent from the upper shore due to the unsuitable extreme conditions in the intertidal zone preventing Crepidula post-settlement recruitment and mitigating colonization (Bohn et al. 2015). Therefore, resistance to colonization by Crepidula fornicata is assessed as 'High' and resilience as 'High' so the biotope is assessed as 'Not sensitive'. The confidence in the assessment is 'Low' because the sensitivity of this biotope to Crepidula is potentially site-specific, there is a risk of its introduction by artificial means, and there is a lack of direct evidence of Crepidula being reported to occur in the biotope. Further evidence is required.

The above evidence also suggests that this biotope is unsuitable for the colonization of Magallana gigas due to a lack of gravel, shell, or any other hard substrata required for successful settlement and establishment (Kochmann, 2012; Kochmann et al., 2013; Mckinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). Therefore, resistance to colonization by Magallana gigas is assessed as 'High' due to the lack of hard substrata in this biotope. Resilience is assessed as 'High', so the biotope is assessed as 'Not sensitive'. 

However, colonization of the upper shore by Spartina anglica may result in the loss of the characterizing amphipods, but this is likely to affect the upper shore extent of the biotope only. Therefore, resistance to colonization by Spartina anglica is assessed as 'Medium' and resilience as 'Very low' so the biotope is assessed as 'Medium'. The confidence in this assessment is 'Low' due to the lack of evidence. 

Introduction of microbial pathogens

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

Evidence

Amphipods may be infected by a number of parasites or pathogens that alter population numbers through changes in host condition, growth, behaviour and reproduction (Green Extabe & Ford, 2014). Infection by acanthocephalan larvae, for example, may alter behaviour and responses of gammarid amphipods (Bethel & Holmes, 1977). The amphipod Orchestia gammarellus is host to the parasitic protist Marteilia which has a feminizing effect on populations, with higher ratios of females and intersex males in infected, estuarine populations (Ginsburger-Vogel & Desportes, 1979).

Corophium volutator is parasitized by several species of trematodes in Europe and North American (McCurdy et al., 2000a; McCurdy et al., 2000b; Mouritsen & Jensen, 1997, cited in Shim et al., 2013). Mass mortalities of Corophium volutator have been associated with infestation by trematodes in the Wadden Sea (Jensen & Mouritsen, 1992). A dense field of Corophium volutator disappeared completely, and the density of the mud snail Hydrobia ulvae declined by 40% during spring 1990 as a result of an epizootic by trematodes. High spring temperature accelerated both the development rate and the release of infective larval stages of an infectious trematode from the snail Hydrobia ulvae (Poulin & Mouritsen, 2006). Studies conducted in the Baltic Sea suggested that increased parasitism by trematode species has a detrimental effect on local amphipods (Meissner & Bick, 1999; Mouritsen & Jensen, 1997 cited in Shim et al., 2013).  

Sensitivity assessment. Although no records of the biotope being affected by the introduction of microbial pathogens in the British Isles, mass mortality of characterizing species Corophium volutator has been observed (Jensen & Mouritsen, 1992). The biotope character is therefore considered to be at some risk and resistance is judged to be ‘Low’ as Bathyporeia spp. and other associated species such as Hydrobia ulvae may be affected increasing the overall impact of the biological assemblage. Resilience is assessed as ‘High’ (where the disease-causing organisms decline) and the biotope is assessed as having ‘Low’ sensitivity to the introduction of microbial pathogens. Sensitivity will be higher where pathogens persist, preventing recovery.

Removal of target species

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

Evidence

The characterizing species are not directly targeted and therefore this pressure is not considered relevant. Some polychaetes and bivalves that characterize intertidal sand flats and muddy sand habitats such as Arenicola marina and Cerastoderma edule may reduce abundance of Corophium volutator and Corophium arenarium (Flach, 1993). Removal of these species through recreational and commercial harvesting may therefore increase the suitability of some areas for the characterizing species. However, this effect is not assessed for this biotope. 

Removal of non-target species

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

Evidence

Direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures, while this pressure considers the ecological or biological effects of by-catch. The characterizing species in this biotope are highly likely to be damaged or directly removed by static or mobile gears that are targeting other species (see abrasion and penetration pressures).

The loss of the key characterizing species through unintentional removal would alter the character of the biotope and may lead to reclassification to a similar, species-poor biotope such as LS.LSa.MoSa.BarSa. The ecosystem services such as secondary production and food for higher trophic levels such as birds provided by the amphipods would be lost.  A dense field of Corophium volutator in the Wadden Sea disappeared completely during spring 1990 as a result of an epizootic by trematodes. The collapse of its population led to drastic changes in erosion patterns, sediment characteristics, and microtopography, as well as marked changes in the abundance of other macrofaunal species in the mudflat (Poulin & Mouritsen, 2006).

Sensitivity assessment. Biotope resistance to loss of the characterizing species is assessed as ‘Low’ as the burrowing lifestyle and mobility of characterizing species mean that a proportion of the population may escape incidental removal. Resilience is assessed as ‘High’ based on in-situ recovery and migration from adjacent populations and sensitivity is therefore assessed as ‘Low’.