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

Zostera marina is widely distributed on sandy shores and estuaries throughout the northern hemisphere and is known to survive at temperatures of between 0-30°C (Zimmerman et al., 2017), although optimum conditions for growth are generally restricted to between 13-24°C (Lee et al., 2007). In the UK, sea surface temperatures are currently between 6-19°C (Huthnance, 2010), and Zostera marina is in the middle of its range (Potouroglou et al., 2014). Zostera marina beds are characterized by high periods of growth in springtime followed by late-summer die-offs (Zimmerman et al., 1989). Populations are maintained primarily by vegetative growth of the surviving plants (Potouroglou et al., 2014). Zostera marina die-offs are triggered by warm summer temperatures, and in Chesapeake Bay shoots were found to stop growing at 20°C and start to die off at 25°C (Reusch et al., 2005). An increase in water temperature in the Chesapeake Bay since the 1950s has led to a shift in historical growth patterns, with Zostera marina biomass peaking earlier and summer declines beginning earlier (Shields et al., 2018).  Biomass varies between years, with beds declining some years and recovering other years.  Recovery is associated with years of greater spring water clarity and water temperatures that rarely exceed 28°C (Shields et al., 2018). Whilst rhizomes generally survive these die-off events and recovery occurs, an extreme increase in water temperature has led to high levels of die-off and a lack of recovery for some populations. In the USA, increases in temperature have led to Zostera marina competing with the tasselweed Ruppia maritima  (Johnson et al., 2003, Moore et al., 2014), which has a greater tolerance to an increase in temperatures (Evans et al., 1986). In some estuaries and lagoons in the UK, Zostera marina can form mixed stands with Ruppia maritima (Davison & Hughes, 1998), and as temperatures increase, there is potential for Ruppia maritima to outcompete Zostera marina due to its lack of sensitivity to high temperatures.

Flowering is important for the maintenance of genetic diversity which may enhance the ability of seagrasses to cope with increasing temperatures (Björk et al., 2008, Ehlers et al., 2008). Flowering has been shown to increase with increases in temperature (Potouroglou et al., 2014) and ocean acidification (Palacios & Zimmerman, 2007), which may benefit this species in the future. Optimum seedling development for North Sea populations of Zostera marina occurs at 10°C (Hootsmans et al., 1987). However, in populations at the southern extent of their global distribution (e.g. Sea of Cortez, Mexico), germination occurs at 18-20°C (McMillan, 1983), which suggests that this species can tolerate and adapt to a wide range of temperatures. A temperature of 28°C is thought to be a critical threshold for this species, as seedlings are unable to survive at higher temperatures (Abe et al., 2008) and, if summer temperatures exceed 28°C for long periods of time, the species may not recover from seasonal die-offs (Shields et al., 2018).

Intertidal populations of Zostera marina var. angustifolia live in less stable conditions than subtidal populations, with greater daily temperature fluctuations. They have more slender shoots and are known to be more salinity tolerant than Zostera marina (Jackson et al., 2013).  While the south of the UK has a mean summer daily high temperature of 21°C, temperatures can often reach ≥30°C (Met_Office, 2016). Temperature loggers on the west coast of Scotland recorded intertidal temperatures on the high shore exceeding 40°C in 7 of the 11 years it was recorded (Burrows, 2017), and yet intertidal populations of Zostera marina var. angustifolia are common in Scotland (Lyndon et al., 2016), suggesting that these populations are able to cope with exposure to high air temperatures. There is currently no evidence on threshold air temperatures for exposed plants, although it is likely that rising air temperatures may have a similar effect to rising seawater temperatures.

Sensitivity assessment. UK populations of Zostera marina are in the middle latitude of their global distribution with populations elsewhere, such as those in Chesapeake Bay and Venice lagoon able to withstand temperatures up to 30°C (Zharova et al., 2001, Shields et al., 2018), although inhibition of growth occurs at 25°C (Zharova et al., 2001). While genetic differences may accompany this higher thermal tolerance, evolutionary change can occur within a few generations in plants (Rice & Emery, 2003). Therefore, with the pace of ocean warming over the next 50-80 years, UK Zostera marina populations may have the opportunity to adapt to withstand temperatures similar to those observed in Chesapeake Bay. With sea surface temperature around the UK of between 6-19°C (Huthnance, 2010), populations of Zostera marina may be able to adapt to cope with a gradual rise in ocean temperatures of both 3°C (middle emission scenario) and 4°C (high emission scenario) by the end of this century, leading to maximum summer high temperatures in the south of the UK of 22-23°C.  However, some mortality from the increased temperature cannot be ruled out, particularly in the south, therefore resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very Low’, as loss is likely to be a long-term decline, due to the long-term nature of ocean warming. For the extreme scenario, whereby sea temperatures rise by 5°C to potential southern summer temperatures of 24°C by the end of this century and air temperatures rise by 6°C to potential average summer temperatures of 26°C, some mortality is expected, as some genotypes may fail to adapt to increasing temperatures, but as the water temperature is still below the critical threshold for this species, and both seawater and air temperatures are still lower than those found in the Venice lagoon and Chesapeake Bay, where extensive Zostera marina beds occur, resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very low’. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming in both emission scenarios and 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

Zostera marina is widely distributed on sandy shores and estuaries throughout the northern hemisphere and is known to survive at temperatures of between 0-30°C (Zimmerman et al., 2017), although optimum conditions for growth are generally restricted to between 13-24°C (Lee et al., 2007). In the UK, sea surface temperatures are currently between 6-19°C (Huthnance, 2010), and Zostera marina is in the middle of its range (Potouroglou et al., 2014). Zostera marina beds are characterized by high periods of growth in springtime followed by late-summer die-offs (Zimmerman et al., 1989). Populations are maintained primarily by vegetative growth of the surviving plants (Potouroglou et al., 2014). Zostera marina die-offs are triggered by warm summer temperatures, and in Chesapeake Bay shoots were found to stop growing at 20°C and start to die off at 25°C (Reusch et al., 2005). An increase in water temperature in the Chesapeake Bay since the 1950s has led to a shift in historical growth patterns, with Zostera marina biomass peaking earlier and summer declines beginning earlier (Shields et al., 2018).  Biomass varies between years, with beds declining some years and recovering other years.  Recovery is associated with years of greater spring water clarity and water temperatures that rarely exceed 28°C (Shields et al., 2018). Whilst rhizomes generally survive these die-off events and recovery occurs, an extreme increase in water temperature has led to high levels of die-off and a lack of recovery for some populations. In the USA, increases in temperature have led to Zostera marina competing with the tasselweed Ruppia maritima  (Johnson et al., 2003, Moore et al., 2014), which has a greater tolerance to an increase in temperatures (Evans et al., 1986). In some estuaries and lagoons in the UK, Zostera marina can form mixed stands with Ruppia maritima (Davison & Hughes, 1998), and as temperatures increase, there is potential for Ruppia maritima to outcompete Zostera marina due to its lack of sensitivity to high temperatures.

Flowering is important for the maintenance of genetic diversity which may enhance the ability of seagrasses to cope with increasing temperatures (Björk et al., 2008, Ehlers et al., 2008). Flowering has been shown to increase with increases in temperature (Potouroglou et al., 2014) and ocean acidification (Palacios & Zimmerman, 2007), which may benefit this species in the future. Optimum seedling development for North Sea populations of Zostera marina occurs at 10°C (Hootsmans et al., 1987). However, in populations at the southern extent of their global distribution (e.g. Sea of Cortez, Mexico), germination occurs at 18-20°C (McMillan, 1983), which suggests that this species can tolerate and adapt to a wide range of temperatures. A temperature of 28°C is thought to be a critical threshold for this species, as seedlings are unable to survive at higher temperatures (Abe et al., 2008) and, if summer temperatures exceed 28°C for long periods of time, the species may not recover from seasonal die-offs (Shields et al., 2018).

Intertidal populations of Zostera marina var. angustifolia live in less stable conditions than subtidal populations, with greater daily temperature fluctuations. They have more slender shoots and are known to be more salinity tolerant than Zostera marina (Jackson et al., 2013).  While the south of the UK has a mean summer daily high temperature of 21°C, temperatures can often reach ≥30°C (Met_Office, 2016). Temperature loggers on the west coast of Scotland recorded intertidal temperatures on the high shore exceeding 40°C in 7 of the 11 years it was recorded (Burrows, 2017), and yet intertidal populations of Zostera marina var. angustifolia are common in Scotland (Lyndon et al., 2016), suggesting that these populations are able to cope with exposure to high air temperatures. There is currently no evidence on threshold air temperatures for exposed plants, although it is likely that rising air temperatures may have a similar effect to rising seawater temperatures.

Sensitivity assessment. UK populations of Zostera marina are in the middle latitude of their global distribution with populations elsewhere, such as those in Chesapeake Bay and Venice lagoon able to withstand temperatures up to 30°C (Zharova et al., 2001, Shields et al., 2018), although inhibition of growth occurs at 25°C (Zharova et al., 2001). While genetic differences may accompany this higher thermal tolerance, evolutionary change can occur within a few generations in plants (Rice & Emery, 2003). Therefore, with the pace of ocean warming over the next 50-80 years, UK Zostera marina populations may have the opportunity to adapt to withstand temperatures similar to those observed in Chesapeake Bay. With sea surface temperature around the UK of between 6-19°C (Huthnance, 2010), populations of Zostera marina may be able to adapt to cope with a gradual rise in ocean temperatures of both 3°C (middle emission scenario) and 4°C (high emission scenario) by the end of this century, leading to maximum summer high temperatures in the south of the UK of 22-23°C.  However, some mortality from the increased temperature cannot be ruled out, particularly in the south, therefore resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very Low’, as loss is likely to be a long-term decline, due to the long-term nature of ocean warming. For the extreme scenario, whereby sea temperatures rise by 5°C to potential southern summer temperatures of 24°C by the end of this century and air temperatures rise by 6°C to potential average summer temperatures of 26°C, some mortality is expected, as some genotypes may fail to adapt to increasing temperatures, but as the water temperature is still below the critical threshold for this species, and both seawater and air temperatures are still lower than those found in the Venice lagoon and Chesapeake Bay, where extensive Zostera marina beds occur, resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very low’. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming in both emission scenarios and 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

Zostera marina is widely distributed on sandy shores and estuaries throughout the northern hemisphere and is known to survive at temperatures of between 0-30°C (Zimmerman et al., 2017), although optimum conditions for growth are generally restricted to between 13-24°C (Lee et al., 2007). In the UK, sea surface temperatures are currently between 6-19°C (Huthnance, 2010), and Zostera marina is in the middle of its range (Potouroglou et al., 2014). Zostera marina beds are characterized by high periods of growth in springtime followed by late-summer die-offs (Zimmerman et al., 1989). Populations are maintained primarily by vegetative growth of the surviving plants (Potouroglou et al., 2014). Zostera marina die-offs are triggered by warm summer temperatures, and in Chesapeake Bay shoots were found to stop growing at 20°C and start to die off at 25°C (Reusch et al., 2005). An increase in water temperature in the Chesapeake Bay since the 1950s has led to a shift in historical growth patterns, with Zostera marina biomass peaking earlier and summer declines beginning earlier (Shields et al., 2018).  Biomass varies between years, with beds declining some years and recovering other years.  Recovery is associated with years of greater spring water clarity and water temperatures that rarely exceed 28°C (Shields et al., 2018). Whilst rhizomes generally survive these die-off events and recovery occurs, an extreme increase in water temperature has led to high levels of die-off and a lack of recovery for some populations. In the USA, increases in temperature have led to Zostera marina competing with the tasselweed Ruppia maritima  (Johnson et al., 2003, Moore et al., 2014), which has a greater tolerance to an increase in temperatures (Evans et al., 1986). In some estuaries and lagoons in the UK, Zostera marina can form mixed stands with Ruppia maritima (Davison & Hughes, 1998), and as temperatures increase, there is potential for Ruppia maritima to outcompete Zostera marina due to its lack of sensitivity to high temperatures.

Flowering is important for the maintenance of genetic diversity which may enhance the ability of seagrasses to cope with increasing temperatures (Björk et al., 2008, Ehlers et al., 2008). Flowering has been shown to increase with increases in temperature (Potouroglou et al., 2014) and ocean acidification (Palacios & Zimmerman, 2007), which may benefit this species in the future. Optimum seedling development for North Sea populations of Zostera marina occurs at 10°C (Hootsmans et al., 1987). However, in populations at the southern extent of their global distribution (e.g. Sea of Cortez, Mexico), germination occurs at 18-20°C (McMillan, 1983), which suggests that this species can tolerate and adapt to a wide range of temperatures. A temperature of 28°C is thought to be a critical threshold for this species, as seedlings are unable to survive at higher temperatures (Abe et al., 2008) and, if summer temperatures exceed 28°C for long periods of time, the species may not recover from seasonal die-offs (Shields et al., 2018).

Intertidal populations of Zostera marina var. angustifolia live in less stable conditions than subtidal populations, with greater daily temperature fluctuations. They have more slender shoots and are known to be more salinity tolerant than Zostera marina (Jackson et al., 2013).  While the south of the UK has a mean summer daily high temperature of 21°C, temperatures can often reach ≥30°C (Met_Office, 2016). Temperature loggers on the west coast of Scotland recorded intertidal temperatures on the high shore exceeding 40°C in 7 of the 11 years it was recorded (Burrows, 2017), and yet intertidal populations of Zostera marina var. angustifolia are common in Scotland (Lyndon et al., 2016), suggesting that these populations are able to cope with exposure to high air temperatures. There is currently no evidence on threshold air temperatures for exposed plants, although it is likely that rising air temperatures may have a similar effect to rising seawater temperatures.

Sensitivity assessment. UK populations of Zostera marina are in the middle latitude of their global distribution with populations elsewhere, such as those in Chesapeake Bay and Venice lagoon able to withstand temperatures up to 30°C (Zharova et al., 2001, Shields et al., 2018), although inhibition of growth occurs at 25°C (Zharova et al., 2001). While genetic differences may accompany this higher thermal tolerance, evolutionary change can occur within a few generations in plants (Rice & Emery, 2003). Therefore, with the pace of ocean warming over the next 50-80 years, UK Zostera marina populations may have the opportunity to adapt to withstand temperatures similar to those observed in Chesapeake Bay. With sea surface temperature around the UK of between 6-19°C (Huthnance, 2010), populations of Zostera marina may be able to adapt to cope with a gradual rise in ocean temperatures of both 3°C (middle emission scenario) and 4°C (high emission scenario) by the end of this century, leading to maximum summer high temperatures in the south of the UK of 22-23°C.  However, some mortality from the increased temperature cannot be ruled out, particularly in the south, therefore resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very Low’, as loss is likely to be a long-term decline, due to the long-term nature of ocean warming. For the extreme scenario, whereby sea temperatures rise by 5°C to potential southern summer temperatures of 24°C by the end of this century and air temperatures rise by 6°C to potential average summer temperatures of 26°C, some mortality is expected, as some genotypes may fail to adapt to increasing temperatures, but as the water temperature is still below the critical threshold for this species, and both seawater and air temperatures are still lower than those found in the Venice lagoon and Chesapeake Bay, where extensive Zostera marina beds occur, resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very low’. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming in both emission scenarios and 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).  UK populations of Zostera marina may be able to withstand a gradual increase in temperatures over the next century due to their ability to adapt and the fact that UK populations occur in the middle of their biogeographical range (see ‘Global warming’ above). Their ability to tolerate marine heatwaves may be more problematic and these extreme temperature events have been reported to cause widespread mortality to seagrasses (e.g. Marba & Duarte, 2010, Fraser et al., 2014, Arias-Ortiz et al., 2018). Following the high shoot mortality that occurred during the 2003 Mediterranean marine heatwave, the seagrass Posidonia oceanica exhibited exceptionally high recruitment, which enabled it to recover shoot density by the following year (Marba & Duarte, 2010), although it is not known whether Zostera marina would exhibit the same response.

In experimental conditions, a northern (Denmark) population of Zostera marina was subjected to a simulated three week heatwave, with an increase in temperature from 18°C to 27°C, that resulted in mortality of two thirds of shoots and a decrease in the biomass of both shoots and rhizomes (Höffle et al., 2011). Similarly, in experimental conditions, when specimens from northern (Denmark) and southern (Italy) populations were subjected to a simulated three-week heatwave (temperatures increased from 19°C to 27°C), Zostera marina experienced thermal stress and a decrease in photochemical activity. Whereas photochemical activity recovered in the southern populations, the northern populations continued to exhibit negative effects (Winters et al., 2011). Recoverability in southern populations was mirrored in a study of genetic resilience of Zostera marina to a simulated heatwave (temperatures increased from 19°C to 26°C), where both a northern (Denmark) and southern (Italy) population exhibited signs of acute heat stress when temperatures were rapidly increased to 26°C. The return to control gene expression (rate of recovery) was immediate for the southern population, but no recovery occurred in the northern population genes involved in protein degradation, which indicated that metabolic compensation to high sea-surface temperature had failed (Franssen et al., 2011).

These results suggest that Mediterranean populations have adapted to acute heat stress, whereas northern populations may be more susceptible. Genetic diversity is thought to be the most important factor in enhancing both resistance (Ehlers et al., 2008) and recoverability (Reusch et al., 2005) of Zostera marina beds after extreme heat events.   Cunha & Santos (2009)  evaluated temporal persistence of seagrass coverage in Ria Formosa, Portugal, suggesting that large and non-fragmented seagrass meadows had higher persistence values than small, fragmented meadows and, hence, that smaller patches were more vulnerable to disturbance. In the UK,  recovery of Zostera marina from the wasting disease in the 1930s is limited, and beds are still scarce, and often small (Davison & Hughes, 1998), which may reduce their ability to withstand heatwaves. Furthermore, in disturbed meadows, a decrease in sexual reproduction is observed, with beds maintained through vegetative spread, which will lead to decreased genetic diversity and therefore resistance (Potouroglou et al., 2014).

Increasing levels of carbon dioxide are causing the pH of surface waters to decrease (see Ocean Acidification). Ocean acidification has been shown to counteract the negative impacts of increasing temperatures on Zostera marina survival and growth and to enhance sexual reproduction, and the co-occurrence of these climate change pressures may be beneficial to Zostera marina (Zimmerman et al., 2017).

Sensitivity assessment. The ability of UK populations to withstand future marine heatwaves will depend on their ability to adapt to rising temperatures (see Global Warming) and acute heat stress, as shown by Mediterranean populations. In the Mediterranean, Zostera marina populations appear to be able to adapt to higher temperatures, and there is some evidence that they may be able to withstand marine heatwaves (Franssen et al., 2011), although UK populations may be more sensitive. Their ability to withstand marine heatwaves, whereby both seawater and air temperatures (for intertidal populations) are increased, will depend on the duration and severity of the heatwave, and the ability and time available for subsequent recovery before the next heatwave hits. Under the middle emission scenario, if heatwaves were occurring at a frequency of every three years by the end of this century, with heatwaves reaching a maximum intensity of 2°C for a period of 80 days, this could lead to temperatures reaching up to 24°C in summer months and is likely to lead to some seagrass mortality, although recovery should occur before the next heatwave. Under the high emission scenario, if heatwaves occur at a frequency of every two years by the end of this century, reaching a maximum intensity of 3.5°C for a period of 120 days, this could lead to the heatwave lasting the entire summer with seawater temperatures reaching up to 26.5°C, and air temperatures exceeding 30°C. In heat adapted populations of Zostera marina in Chesapeake Bay, die-offs start to occur at water temperatures of 25°C (Reusch et al., 2005). Even if Zostera marina is able to adapt to gradual ocean warming, an increasing length of stressful high summer temperatures such as this is likely to trigger an earlier die-off. This pattern has been seen in Chesapeake Bay, with higher summer temperatures leading to a shift in historical growth patterns, with summer declines beginning earlier (Shields et al., 2018). Under the middle emission scenario, there may be some mortality so resistance has been assessed as ‘Medium, and resilience is assessed as ‘Medium’, as recovery is expected within 2-3 years. Under the high emission scenario it is likely that marine heatwaves will lead to an increase in seagrass decline through earlier summer die-offs, therefore resistance has been assessed as ‘Low’, and resilience is assessed as ‘Low’, as heatwaves will be more common, occurring biannually, and recovery may take longer if a further heatwave occurs before full recovery has been achieved.  Therefore, this biotope is assessed as ‘Medium’ sensitivity to marine heatwaves under the middle emission scenario, and ‘High’ sensitivity to marine heatwaves 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).  UK populations of Zostera marina may be able to withstand a gradual increase in temperatures over the next century due to their ability to adapt and the fact that UK populations occur in the middle of their biogeographical range (see ‘Global warming’ above). Their ability to tolerate marine heatwaves may be more problematic and these extreme temperature events have been reported to cause widespread mortality to seagrasses (e.g. Marba & Duarte, 2010, Fraser et al., 2014, Arias-Ortiz et al., 2018). Following the high shoot mortality that occurred during the 2003 Mediterranean marine heatwave, the seagrass Posidonia oceanica exhibited exceptionally high recruitment, which enabled it to recover shoot density by the following year (Marba & Duarte, 2010), although it is not known whether Zostera marina would exhibit the same response.

In experimental conditions, a northern (Denmark) population of Zostera marina was subjected to a simulated three week heatwave, with an increase in temperature from 18°C to 27°C, that resulted in mortality of two thirds of shoots and a decrease in the biomass of both shoots and rhizomes (Höffle et al., 2011). Similarly, in experimental conditions, when specimens from northern (Denmark) and southern (Italy) populations were subjected to a simulated three-week heatwave (temperatures increased from 19°C to 27°C), Zostera marina experienced thermal stress and a decrease in photochemical activity. Whereas photochemical activity recovered in the southern populations, the northern populations continued to exhibit negative effects (Winters et al., 2011). Recoverability in southern populations was mirrored in a study of genetic resilience of Zostera marina to a simulated heatwave (temperatures increased from 19°C to 26°C), where both a northern (Denmark) and southern (Italy) population exhibited signs of acute heat stress when temperatures were rapidly increased to 26°C. The return to control gene expression (rate of recovery) was immediate for the southern population, but no recovery occurred in the northern population genes involved in protein degradation, which indicated that metabolic compensation to high sea-surface temperature had failed (Franssen et al., 2011).

These results suggest that Mediterranean populations have adapted to acute heat stress, whereas northern populations may be more susceptible. Genetic diversity is thought to be the most important factor in enhancing both resistance (Ehlers et al., 2008) and recoverability (Reusch et al., 2005) of Zostera marina beds after extreme heat events.   Cunha & Santos (2009)  evaluated temporal persistence of seagrass coverage in Ria Formosa, Portugal, suggesting that large and non-fragmented seagrass meadows had higher persistence values than small, fragmented meadows and, hence, that smaller patches were more vulnerable to disturbance. In the UK,  recovery of Zostera marina from the wasting disease in the 1930s is limited, and beds are still scarce, and often small (Davison & Hughes, 1998), which may reduce their ability to withstand heatwaves. Furthermore, in disturbed meadows, a decrease in sexual reproduction is observed, with beds maintained through vegetative spread, which will lead to decreased genetic diversity and therefore resistance (Potouroglou et al., 2014).

Increasing levels of carbon dioxide are causing the pH of surface waters to decrease (see Ocean Acidification). Ocean acidification has been shown to counteract the negative impacts of increasing temperatures on Zostera marina survival and growth and to enhance sexual reproduction, and the co-occurrence of these climate change pressures may be beneficial to Zostera marina (Zimmerman et al., 2017).

Sensitivity assessment. The ability of UK populations to withstand future marine heatwaves will depend on their ability to adapt to rising temperatures (see Global Warming) and acute heat stress, as shown by Mediterranean populations. In the Mediterranean, Zostera marina populations appear to be able to adapt to higher temperatures, and there is some evidence that they may be able to withstand marine heatwaves (Franssen et al., 2011), although UK populations may be more sensitive. Their ability to withstand marine heatwaves, whereby both seawater and air temperatures (for intertidal populations) are increased, will depend on the duration and severity of the heatwave, and the ability and time available for subsequent recovery before the next heatwave hits. Under the middle emission scenario, if heatwaves were occurring at a frequency of every three years by the end of this century, with heatwaves reaching a maximum intensity of 2°C for a period of 80 days, this could lead to temperatures reaching up to 24°C in summer months and is likely to lead to some seagrass mortality, although recovery should occur before the next heatwave. Under the high emission scenario, if heatwaves occur at a frequency of every two years by the end of this century, reaching a maximum intensity of 3.5°C for a period of 120 days, this could lead to the heatwave lasting the entire summer with seawater temperatures reaching up to 26.5°C, and air temperatures exceeding 30°C. In heat adapted populations of Zostera marina in Chesapeake Bay, die-offs start to occur at water temperatures of 25°C (Reusch et al., 2005). Even if Zostera marina is able to adapt to gradual ocean warming, an increasing length of stressful high summer temperatures such as this is likely to trigger an earlier die-off. This pattern has been seen in Chesapeake Bay, with higher summer temperatures leading to a shift in historical growth patterns, with summer declines beginning earlier (Shields et al., 2018). Under the middle emission scenario, there may be some mortality so resistance has been assessed as ‘Medium, and resilience is assessed as ‘Medium’, as recovery is expected within 2-3 years. Under the high emission scenario it is likely that marine heatwaves will lead to an increase in seagrass decline through earlier summer die-offs, therefore resistance has been assessed as ‘Low’, and resilience is assessed as ‘Low’, as heatwaves will be more common, occurring biannually, and recovery may take longer if a further heatwave occurs before full recovery has been achieved.  Therefore, this biotope is assessed as ‘Medium’ sensitivity to marine heatwaves under the middle emission scenario, and ‘High’ sensitivity to marine heatwaves 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

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). Many seagrasses are reported to respond positively to ocean acidification, as most seagrasses (>85% of species) exhibit C3 carbon fixation photophysiology and preferentially utilize aqueous CO₂ over bicarbonate (HCO₃^-) (Koch et al., 2013), although they are able to utilize both through the presence of carbon concentrating mechanisms (CCMs) and use of carbonic anhydrase enzymes to dehydrate HCO₃^- to aqueous CO₂ (Beer & Rehnberg, 1997). This may be particularly important for deeper stands of seagrass at low light levels where metabolic energy availability is low, as the use of carbonic anhydrase or a carbon concentrating mechanism (CCM) is energy dependent, increasing their reliance on CO₂ diffusion and making them respond positively to CO₂ enrichment (Koch et al., 2013). This said, at extremely low levels of light, Zostera marina is light limited, not carbon limited, and an increase in CO₂ will not be beneficial (Palacios & Zimmerman, 2007).

Under light-replete conditions, results of increasing CO₂ have generally been positive, and show an increase in photosynthesis, growth and sugar levels in response to increasing CO₂ (Zimmerman et al., 1997, Palacios & Zimmerman, 2007, Zimmerman et al., 2017), although beneficial effects are not always apparent (Miller et al., 2017). A long-term (1 year) experiment of the impact of CO₂ enrichment on light-replete Zostera marina plants found that enhanced CO₂ boosted carbon accumulation, which led to an increase in rhizome growth, vegetative proliferation and sexual reproduction (Palacios & Zimmerman, 2007). A rise in rates of sexual reproduction could increase genetic diversity of Zostera beds, and hence their ability to adapt and withstand other stressors such as temperature rise (Rice & Emery, 2003). Further experimentation showed that plants exposed to elevated CO₂ levels were able to withstand greater heat stress than plants at ambient CO₂, although it was uncertain whether this was a direct effect, or a side effect of increased carbon reserves that allowed the plant to support metabolic repair without negatively impacting growth (Zimmerman et al., 2017).

Sensitivity Assessment. An increase in CO₂ and the subsequent decrease in pH as the oceans acidify is likely to have a net beneficial impact on Zostera marina beds globally, except in light-limited, deeper or more turbid waters. No mortality is expected at both the middle emissions scenario benchmark of pH 8.0 and the high emissions scenario benchmark of pH7.8. Therefore, resistance is assessed as ‘High’. No recovery is required, and resilience is assessed as ‘High’ so that the biotope is considered ‘Not sensitive’ to ocean acidification 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

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). Many seagrasses are reported to respond positively to ocean acidification, as most seagrasses (>85% of species) exhibit C3 carbon fixation photophysiology and preferentially utilize aqueous CO₂ over bicarbonate (HCO₃^-) (Koch et al., 2013), although they are able to utilize both through the presence of carbon concentrating mechanisms (CCMs) and use of carbonic anhydrase enzymes to dehydrate HCO₃^- to aqueous CO₂ (Beer & Rehnberg, 1997). This may be particularly important for deeper stands of seagrass at low light levels where metabolic energy availability is low, as the use of carbonic anhydrase or a carbon concentrating mechanism (CCM) is energy dependent, increasing their reliance on CO₂ diffusion and making them respond positively to CO₂ enrichment (Koch et al., 2013). This said, at extremely low levels of light, Zostera marina is light limited, not carbon limited, and an increase in CO₂ will not be beneficial (Palacios & Zimmerman, 2007).

Under light-replete conditions, results of increasing CO₂ have generally been positive, and show an increase in photosynthesis, growth and sugar levels in response to increasing CO₂ (Zimmerman et al., 1997, Palacios & Zimmerman, 2007, Zimmerman et al., 2017), although beneficial effects are not always apparent (Miller et al., 2017). A long-term (1 year) experiment of the impact of CO₂ enrichment on light-replete Zostera marina plants found that enhanced CO₂ boosted carbon accumulation, which led to an increase in rhizome growth, vegetative proliferation and sexual reproduction (Palacios & Zimmerman, 2007). A rise in rates of sexual reproduction could increase genetic diversity of Zostera beds, and hence their ability to adapt and withstand other stressors such as temperature rise (Rice & Emery, 2003). Further experimentation showed that plants exposed to elevated CO₂ levels were able to withstand greater heat stress than plants at ambient CO₂, although it was uncertain whether this was a direct effect, or a side effect of increased carbon reserves that allowed the plant to support metabolic repair without negatively impacting growth (Zimmerman et al., 2017).

Sensitivity Assessment. An increase in CO₂ and the subsequent decrease in pH as the oceans acidify is likely to have a net beneficial impact on Zostera marina beds globally, except in light-limited, deeper or more turbid waters. No mortality is expected at both the middle emissions scenario benchmark of pH 8.0 and the high emissions scenario benchmark of pH7.8. Therefore, resistance is assessed as ‘High’. No recovery is required, and resilience is assessed as ‘High’ so that the biotope is considered ‘Not sensitive’ to ocean acidification 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

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). Dense stands of Zostera marina can be found on the lower shore down to about 5 m, whilst Zostera marina var. angustifolia is found in the intertidal. Depth limitation is due to light availability, with light penetration decreasing with depth and/or turbidity (Nielsen et al., 2002). Shoot density is highest in shallow waters and declines thereafter, down to a depth limited by surface irradiance. Estimates from across the biological range of Zostera marina suggest it requires between 12 – 37% surface irradiance  to survive in the long-term with a mean of 18% (Erftemeijer & Robin Lewis, 2006), with photo-acclimation to local light regimes appearing to be the main cause of the high level of variation (Lee et al., 2007). Plant biomass gradually increases up to the middle depth range due to an increase in shoot weight, leading to maximum percentage coverage of seagrass at middle depth ranges (Krause-Jensen et al., 2000).

Sensitivity assessment. An increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of current Zostera marina beds. Beds may be able to expand their range and migrate landwards to compensate for sea-level rise, if not constrained by steep topography, lack of suitable sediment, or human-modified shorelines (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Zostera marina beds will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. As intertidal beds become sublittoral, Zostera marina var. angustifolia may be replaced by Zostera marina, which has wider shoots, and may lead to an increase in biomass, partially offsetting the loss in depth distribution. For the middle and high emission scenario (50 and 70 cm rise) the resistance has been assessed as ‘Medium’ while resilience is assessed as ‘Very low’.  Therefore, sensitivity is assessed as ‘Medium’ sensitivity to sea-level rise predicted for the end of this century in these scenarios. Under the extreme sea-level rise scenario of 107 cm, there is potential that more than 25% of the bed could be lost, dependent on biotope slope.  Therefore, resistance has been assessed as ‘Low’, and resilience as ‘Very low’, albeit with ‘Low’ confidence.  Therefore, sensitivity is assessed as ‘High’ under the extreme sea-level rise scenario predicted for the end of this century. 

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

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). Dense stands of Zostera marina can be found on the lower shore down to about 5 m, whilst Zostera marina var. angustifolia is found in the intertidal. Depth limitation is due to light availability, with light penetration decreasing with depth and/or turbidity (Nielsen et al., 2002). Shoot density is highest in shallow waters and declines thereafter, down to a depth limited by surface irradiance. Estimates from across the biological range of Zostera marina suggest it requires between 12 – 37% surface irradiance  to survive in the long-term with a mean of 18% (Erftemeijer & Robin Lewis, 2006), with photo-acclimation to local light regimes appearing to be the main cause of the high level of variation (Lee et al., 2007). Plant biomass gradually increases up to the middle depth range due to an increase in shoot weight, leading to maximum percentage coverage of seagrass at middle depth ranges (Krause-Jensen et al., 2000).

Sensitivity assessment. An increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of current Zostera marina beds. Beds may be able to expand their range and migrate landwards to compensate for sea-level rise, if not constrained by steep topography, lack of suitable sediment, or human-modified shorelines (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Zostera marina beds will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. As intertidal beds become sublittoral, Zostera marina var. angustifolia may be replaced by Zostera marina, which has wider shoots, and may lead to an increase in biomass, partially offsetting the loss in depth distribution. For the middle and high emission scenario (50 and 70 cm rise) the resistance has been assessed as ‘Medium’ while resilience is assessed as ‘Very low’.  Therefore, sensitivity is assessed as ‘Medium’ sensitivity to sea-level rise predicted for the end of this century in these scenarios. Under the extreme sea-level rise scenario of 107 cm, there is potential that more than 25% of the bed could be lost, dependent on biotope slope.  Therefore, resistance has been assessed as ‘Low’, and resilience as ‘Very low’, albeit with ‘Low’ confidence.  Therefore, sensitivity is assessed as ‘High’ under the extreme sea-level rise scenario predicted for the end of this century. 

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

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). Dense stands of Zostera marina can be found on the lower shore down to about 5 m, whilst Zostera marina var. angustifolia is found in the intertidal. Depth limitation is due to light availability, with light penetration decreasing with depth and/or turbidity (Nielsen et al., 2002). Shoot density is highest in shallow waters and declines thereafter, down to a depth limited by surface irradiance. Estimates from across the biological range of Zostera marina suggest it requires between 12 – 37% surface irradiance  to survive in the long-term with a mean of 18% (Erftemeijer & Robin Lewis, 2006), with photo-acclimation to local light regimes appearing to be the main cause of the high level of variation (Lee et al., 2007). Plant biomass gradually increases up to the middle depth range due to an increase in shoot weight, leading to maximum percentage coverage of seagrass at middle depth ranges (Krause-Jensen et al., 2000).

Sensitivity assessment. An increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of current Zostera marina beds. Beds may be able to expand their range and migrate landwards to compensate for sea-level rise, if not constrained by steep topography, lack of suitable sediment, or human-modified shorelines (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of Zostera marina beds will shrink in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery. As intertidal beds become sublittoral, Zostera marina var. angustifolia may be replaced by Zostera marina, which has wider shoots, and may lead to an increase in biomass, partially offsetting the loss in depth distribution. For the middle and high emission scenario (50 and 70 cm rise) the resistance has been assessed as ‘Medium’ while resilience is assessed as ‘Very low’.  Therefore, sensitivity is assessed as ‘Medium’ sensitivity to sea-level rise predicted for the end of this century in these scenarios. Under the extreme sea-level rise scenario of 107 cm, there is potential that more than 25% of the bed could be lost, dependent on biotope slope.  Therefore, resistance has been assessed as ‘Low’, and resilience as ‘Very low’, albeit with ‘Low’ confidence.  Therefore, sensitivity is assessed as ‘High’ under the extreme sea-level rise scenario predicted for the end of this century. 

Temperature increase (local)

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

Evidence

Temperature is considered the overall parameter controlling the geographical distribution of seagrasses.  All enzymatic processes, related to plant metabolism are temperature dependent and specific life cycle events, such as flowering and germination, are also often related to temperature (Phillips et al., 1983). For seagrasses, temperature affects biological processes by increasing reaction rates of biological pathways.  Photosynthesis and respiration increase with higher temperature until a point where enzymes associated with these processes are inhibited.  Beyond a certain threshold, high temperatures will result in respiration being greater than photosynthesis resulting in a negative energy balance.  Increased temperatures do also encourage the growth of epiphytes increasing the burden upon seagrass beds and making them more susceptible to disease (Rasmussen, 1977). Zostera marina can tolerate temperatures between -1 to 25°C with optimum conditions for growth being around 10 to 15°C, and 10°C for seedling development (Hootsmans et al., 1987). Nejrup & Pedersen (2007) found that temperatures between 25 and 30°C lowered photosynthetic rates by 50% as well as growth (production of new leaves by 50% and leaf elongation rate by 75%).  High temperatures also resulted in a 12-fold increase in mortality of Zostera marina plants. Moore et al. (2014) found that short-term exposures to a rapid increase of 4–5°C above normal temperature (25°C) during summer months resulted in widespread diebacks of Zostera marina. Recovery was observed to be minimal as the seagrass was replaced by Ruppia maritima. Similarly, Salo & Peterson (2014) found that exposure to high temperature for five weeks led to enhanced mortality, reduced formation of new leaves and a lower number of standing leaves per shoot. Orth & Moore (1983) reported that the majority (68%) of Zostera marina seeds germinated in the winter months between 0-10°C, and that germination was most rapid between 5-10°C but virtually no germination was observed when temperatures were above 20°C, in Chesapeake Bay, USA. 

Other species associated with seagrass habitats are also affected by changes in temperature. For instance, the gastropod Lacuna vincta, an important grazer found in seagrass beds, is near its southern range limit in the British Isles.  Long-term increases in temperature due to human activity may limit the survival of the snail and restrict subsequent distribution whilst a short-term acute temperature increase may cause death.  The loss of grazers could have detrimental effects on seagrass beds as the leaves provide a substratum for the growth of many species of epiphytic algae.  These epiphytes may smother the Zostera plants unless kept in check by the grazing activities of gastropods and other invertebrates.  Healthy populations of epiphyte grazers are therefore essential to the maintenance of seagrass beds.

Sensitivity assessment. High temperatures during hot summer months have caused massive die-off events among seagrasses worldwide (Moore & Jarvis, 2008; Reusch et al., 2005). A 5°C change in temperature over one month or a 2°C change over the period of a year is thus likely to result in some Zostera marina mortality. In addition, a longer-term or persistent increase in temperature may reduce germination rates and hence reduce recruitment and resilience (Jackson, pers comm., 2019). Therefore, resistance is assessed as ‘Medium’.  Recovery will be fairly rapid once conditions return to normal resulting in a ‘Medium’ resilience score. If however, temperatures remain elevated for a prolonged period of time, Zostera marina can be out-competed and subsequently excluded from the habitat by other species such as Ruppia maritima. Overall, the biotope is assessed as ‘Medium’ sensitivity to an increase in temperature at the pressure benchmark.

Temperature decrease (local)

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

Evidence

Temperature is considered the overall parameter controlling the geographical distribution of seagrasses.  All enzymatic processes related to plant metabolism are temperature dependent and specific life cycle events, such as flowering and germination, are also often related to temperature (Phillips et al., 1983). For seagrasses, temperature affects biological processes by increasing reaction rates of biological pathways.  Photosynthesis and respiration increase with higher temperature until a point where enzymes associated with these processes are inhibited.  Beyond a certain threshold, high temperatures will result in respiration being greater than photosynthesis resulting in a negative energy balance.  Increased temperatures do also encourage the growth of epiphytes increasing the burden upon seagrass beds and making them more susceptible to disease (Rasmussen, 1977). Zostera marina can tolerate temperatures between -1 to 25°C with optimum conditions for growth being around 10 to 15°C, and 10°C for seedling development (Hootsmans et al., 1987). Nejrup & Pedersen (2007) found that low water temperatures (5°C) slowed down the photosynthetic rate by 75%; growth was also affected, with the production of new leaves reduced by 30% and leaf elongation rate reduced by 80% compared to the control, however, mortality was not affected. 

Other species associated with seagrass habitats are also affected by changes in temperature. For instance, the gastropod Lacuna vincta, an important grazer found in seagrass beds, is near its southern range limit in the British Isles.  Long-term change in temperature due to human activity may limit the survival of the snail and restrict subsequent distribution whilst a short-term acute temperature increase may cause death, although it may be replaced by other grazers. Healthy populations of epiphyte grazers are therefore essential to the maintenance of seagrass beds.

Sensitivity assessment. Overall, a decrease in temperature is likely to reduce growth rates but not to cause mortality directly. Frost damage could occur to plants exposed at extreme low tides in the winter months but as the seagrass dies back in winter this is unlikely to be significant.  Therefore, a 5°C decrease in temperature over one month or a 2°C decrease over the period of a year is thus unlikely to result in some Zostera marina mortality. Resistance is therefore considered ‘High’. Recovery will be rapid once conditions return to normal resulting in a ‘High’ resilience score. Hence, the biotope is considered be ‘Not sensitive’ to a decrease in temperature at the pressure benchmark.

Salinity increase (local)

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

Evidence

In general, seagrass species have a wide salinity tolerance. Nejrup & Pedersen (2008) reported optimum salinities between 10 and 25 ppt. Hypersaline conditions can affect seagrass performance as changes in salinity may increase the energy requirements due to demanding osmotic adjustments (Touchette, 2007). Den Hartog (1997) stated that Zostera noltei has a greater tolerance to extremes salinities compared to Zostera marina due to its intertidal habitat. Vermaat et al. (2000) investigating salinity tolerance in Zostera noltei found considerable mortalities of plants at a salinity of 35 ppt. These findings suggest that both Zostera species are ill-equipped to withstand high saline conditions. A review by d’Avack et al. (2014) reported that phenotypic plasticity can play an important role in the ability of seagrasses to withstand external pressures such as changes in salinity.  Changes in physiological and morphological characteristics of seagrass plants will enable species to cope with varying degrees of stress for an extended period of time (Maxwell et al., 2014).

Sensitivity assessment. Even though Zostera plants display a wide tolerance to a range of salinities, an increase from 35 to 38 units for the period of one year will cause some mortality in Zostera marina. The subtidal habitat makes the species more vulnerable to salinity extremes compared to the intertidal Zostera noltei resulting in a ‘Low’ resistance score. Zostera marina will thus be adversely affected by activities such as brine discharges from seawater desalination plant. Recovery, enabled by recolonization from surrounding communities, will be fairly rapid once conditions return to normal resulting in a ‘Medium’ resilience score. The biotope is therefore considered to have a ‘Medium’ sensitivity to this pressure at the pressure benchmark.

Salinity decrease (local)

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

Evidence

In general, seagrass species have a wide salinity tolerance.  Nejrup & Pedersen (2008) reported optimum salinities between 10 and 25 ppt, while den Hartog (1970) reported tolerance to salinities as low as 5 ppt. Hyposaline conditions (reduced salinity) can, however, affect seagrass performance as changes in salinity may increase the energy requirements due to demanding osmotic adjustments (Touchette, 2007).

A study by Salo et al. (2014) found that hyposaline conditions can seriously impair plant performance and survival rates. The study determined that the severity of impact will be population specific as seagrass populations from different areas may substantially differ in their salinity tolerance range with population naturally occurring in low saline areas having greater resistance to this pressure.

Salo & Petersen (2014) experimentally tested the effects of different combinations of salinity and temperature on the physiological performance of Zostera marina. The study found that the combination of high temperature and low salinity resulted in high mortality rates, highlighting negative synergistic effects when seagrasses are exposed to multiple pressures.

A review by d’Avack et al. (2014) determined that phenotypic plasticity can play an important role in the ability of seagrasses to withstand external pressures such as changes in salinity.  Changes in physiological and morphological characteristics of seagrass plants will enable species to cope with varying degrees of stress for an extended period of time (Maxwell et al., 2014).

Sensitivity assessment. Zostera marina has a wide salinity tolerance. Reduced salinity will, however, impact performance causing some mortality. Resistance is therefore considered ‘Medium’. Effects can be exasperated when the seagrass is exposed to multiple stressors at the same time, highlighting the importance to consider negative synergistic effects when conduction assessments. Recovery is considered fairly rapid once conditions return to normal resulting in a ‘Medium’ resilience score. The biotope is therefore considered to have a ‘Medium’ sensitivity to this pressure at the pressure benchmark.

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

A complex interaction exists between seagrass beds and water flow.  Water flow determines the upper distribution of plants on the shore whilst plants mediate the velocity of the flow by extracting momentum from the moving water.  Reducing the flow increases water transparency (see ‘changes in suspended sediments’ pressure) and causes the deposition and retention of fine sediments. Increased flow rates, on the other hand, are likely to erode sediments, expose rhizomes and lead to loss of plants.

The highest current velocity a seagrass can withstand is determined by a threshold beyond which sediment re-suspension and erosion rates are greater than the seagrasses ability to bind sediment and attenuate currents.  In very strong currents, leaves might lie flat on the sea bed reducing erosion under the leaves but not on the unvegetated edges which begin to erode.  High velocity currents can thus change the configuration of patches within a meadow, creating striations and mounding in the seagrass beds.  Such turreted profiles destabilise the bed and increase the risk of 'blow outs' (Jackson et al., 2013).  Populations found in stronger currents are usually smaller, patchy and more vulnerable to storm damage.

A review by Koch (2001) determined that the range of current velocities tolerated by seagrass lies approximately between a minimum of 5 cm/s and a maximum of 180 cm/s. Fonseca et al. (1983) found a lower maximum for Zostera marina and estimated the highest current velocity at approximately 120–150 cm/s.

Human activities in coastal waters which alter hydrology have been implicated in the disappearance of seagrass beds.  For instance, van der Heide et al. (2007) noted that the construction of a dam in the Wadden Sea influencing the hydrological regime inhibited the recovery of Zostera plants after their initial decline following the wasting disease in the 1930s.  Aquaculture installations can also change water flow and have shown to directly impact seagrass habitats. Everett et al. (1995) experimentally altered water flow to investigate the effects of the commercial culture of the oyster Magallana gigas on Zostera marina, using both stake and rack methods.  The study found that both culture methods caused a sharp decline in Zostera marina plants with cover being less than 25% compared to control plots after one year of culture due to changes in local hydrological regime. Both culture methods produced strong, although dissimilar, changes in local hydrological conditions, which had clear effects on sediment characteristics.  In general, stakes resulted in local sediment deposition while racks produced local erosion, both leading to the reduction and eventual death of nearby seagrass beds.

Sensitivity assessment. Any changes in hydrology will have a considerable impact on the integrity of seagrass habitat.  A change in water flow at the level of the benchmark of 10 to 20 cm/s for more than 1 year would cause some mortality in seagrasses resulting in a ‘Medium’ resistance score. Recovery will depend on the species capacities to adapt to changes in water flow regime but is considered to be fairly rapid. Resilience is thus assessed as ‘Medium’. The biotope scores a ‘Medium’ sensitivity to changes in water flow at the pressure benchmark.

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

Seagrasses are generally not tolerant to exposure to aerial conditions, suggesting that the shallowest distribution should be at a depth below mean low water (MLW) (Koch, 2001). Zostera noltei grows predominantly in the intertidal zone and demonstrate higher resistance to desiccation than Zostera marina which occurs more frequently in the subtidal.  To understand the differences in desiccation tolerance between the two Zostera species, Leuschner et al. (1998) investigated the photosynthetic activity of emerged plants. The study found that after 5 hours of exposure to air during low tide, leaves of Zostera noltei had lost up to 50% of their water content.  Decreasing leaf water content resulted in a reversible reduction in light-saturated net photosynthesis rate of the plant.  The experiment further showed that photosynthesis was more sensitive to desiccation in Zostera marina plants than in Zostera noltei under a given leaf water content.  The experiment confirmed that Zostera marina is most susceptible to local changes in emergence regimes by being less tolerant to desiccation pressure.

Tolerances vary not only between species but also within species.  For instance, annual and perennial forms of Zostera marina were observed to tolerate desiccation to different extents. Van Katwijk & Hermus (2000) noted that in intertidal areas of the Wadden Sea, annual Zostera marina plants tended to lie flat on the moist sediment when exposed at low tide.  Perennial plants, on the other hand, had stiffer stems inhibiting contact with the sediment.  These upright sheaths desiccate more rapidly when exposed.  Morphology is, therefore, a factor partly determining tolerance to desiccation.  The same phenomenon was observed by Boese et al. (2003) on Zostera marina in Aquinas Bay, USA.

The overall low tolerance of seagrass species to aerial exposure means that an increase in tidal amplitudes could force seagrass to grow deeper where there was less chance of exposure to the air.  As the depth limit of seagrasses is set by light penetration, this change is likely to reduce the extent of suitable habitat.  Changes in seagrass distribution along a depth gradient will have an impact further down the food chain.

Sensitivity assessment. Sensitivity to changes in emergence regimes varies between species and habitats. Species growing in intertidal habitats have greater tolerance to exposure to air than species inhabiting subtidal beds. The resistance of Zostera marina to this pressure is therefore assessed as ‘Low’. Recovery will be enabled by recolonization from surrounding communities located further down the shore and via the remaining seed bank.  Recovery is therefore considered to be fairly rapid resulting in a ‘Medium’ resilience score. The biotope is therefore considered to have a ‘Medium’ sensitivity to this pressure at the pressure benchmark.

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

An absolute wave exposure limit and maximum wave height for Zostera has not been established (Short et al., 2002) but an increase in wave action can harm the plants in several ways.  Seagrasses are not robust.  Strong waves can cause mechanical damage to leaves and rhizomes.  By losing above ground biomass due to increased wave action, the productivity of seagrass plants is limited.  Small and patchy populations, as well as seedlings, will be particularly vulnerable to wave exposure as they lack extensive rhizome systems to effectively anchor the plant to the seabed.

Wave action also continuously mobilises sediments in coastal areas causing sediment re-suspension which in turn leads to a reduction in water transparency (Koch, 2001) (see ‘changes in suspended sediments’ pressure).  Photosynthesis can be further limited by breaking waves inhibiting light penetration to the seafloor. Wave exposure can also influence the sediment grain size, with areas of high wave exposure having coarser sediments with lower nutrient concentrations.  Coarser sediments reduce the vegetative spreading of seagrasses and inhibit seedling colonisation (Gray & Elliott, 2009).  Changes in sediment type can, therefore, have wider implications for the sensitivity of the beds on a long-term scale.

Sensitivity assessment. No evidence was available to determine the impact of this pressure at the benchmark level.  However, exposure models from Studland Bay and Salcombe, where seagrass beds are limited to low wave exposure, show that even a change of 3% is likely to influence the upper shore limits as well as beds living at the limits of their wave exposure tolerance (Rhodes et al., 2006; Jackson et al., 2013). Change in wave exposure will impact the upper limit of seagrass and thus influence its wider distribution. At the benchmark level, an increase in wave exposure is likely to remove surface vegetation and the majority of the root system causing some mortality. Resistance is thus assessed as ‘Medium’. Recovery will depend on the presence of adjacent seagrass beds and is considered to be fairly rapid scoring a ‘Medium’ resilience.  The biotope, therefore, scores a ‘Medium’ sensitivity to changes in wave exposure 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 elements & organometal' contaminants on seagrasses are summarized below. The full 'seagrass evidence review' should be consulted for details of the studies examined and their results. 

Seagrasses were reported to be relatively tolerant of heavy metals contamination, accumulate metals in their tissues, act as useful bioindicators of heavy metals in the environment, and trap heavy metals in seagrass bed sediments (Lyngby & Brix, 1984; Ward, 1987; Williams et al., 1994; Davison & Hughes, 1998; Prange & Dennison, 2000; Govers et al., 2014).  The tissue accumulation varied between the heavy metals, season, and species of seagrass tested.  The number of articles that report mortalities due to metal, organometals, and nanoparticulate metals is summarized in the 'seagrass evidence review'; (see Figure 1.1 and Table 1.3).

Halophila serratus was the only seagrass species reported to exhibit mortality due to exposure to copper under laboratory conditions (6 days at 1 mg/l Cu) (Prange & Dennison, 2000).  The remaining articles reported ‘toxicity’ in terms of sublethal effects, primarily on photosynthetic efficiency (e.g. effective and maximum quantum yield, fluorescence, or photosystem II (PSII) function, photosynthetic pigment ratios, and growth (e.g. leaf extension).  Ralph & Burchett (1998b) suggested that the relative toxicity was Cu >Zn >Cd > Pb based on weight or Zn >Cu >Cd >Pb based on molarity.  Nevertheless, Cu was more toxic than Zn based on the lethal response at lower molarity.  They also suggested that Cu and Zn were the most toxic as they were essential trace metals in plant metabolism and hence actively taken up, while Cd and Pb were less toxic as they were excluded.  Toxicity increased with exposure time and concentration but most papers noted that the concentrations studied were higher than those reported in the environment (e.g. Lyngby & Brix, 1984; Ward, 1987). 

There was also some evidence that prior exposure to heavy metals affected the toxic response, for example, Macinnis-Ng & Ralph (2004b) noted that seagrasses (Zostera capricorni) from their pristine site were more sensitive than those from contaminated sites.  Few articles examined the effect on seagrass beds and their associated community.  The reduction in photosynthetic efficiency and growth demonstrated in the evidence would be expected the cause stress on seagrasses and had the potential to cause loss at the population level this was not demonstrated in the evidence.  For example, Marin-Guirao et al. (2005) compared the metal contaminated Cymodocea nodosa seagrass beds with uncontaminated reference areas in Mar Menor lagoon, Spain and found but few differences in seagrass metrics between sites.  However, there were differences in the macroinvertebrate community.

Mauro et al. (2013) examined the condition of a Posidonia oceanica bed in a lagoon exposed to human impacts for ca 40 years and found that the bed did not show any sign of regression, and may have been extending seaward, even though the sediment was contaminated with PAHs and metals.  Wang et al. (2019) concluded that both the natural and restored Zostera marina beds had similar growth characteristics but that differences in chemical parameters (metals, petroleum, and nutrients) may affect long-term growth and restoration.  And Ward (1984) concluded that the acute toxicity of metals played a minor role in structuring the seagrass faunal community.

Similarly, Ward (1987) reported that seagrass (Posidonia australis) beds exhibited the lowest density, standing crop and leaf growth at a site contaminated by smelter effluent in Spence Gulf, South Australia when compared with sites further away from the effluent discharge.  But the differences were not always significant.  Posidonia australis was not sensitive to heavy metals as it maintained its distribution in highly contaminated areas.  Lafratta et al. (2019) also reported Posidonia beds surviving downstream of smelter effluent in Spence Gulf, South Australia and accumulating heavy metals in the sediment over a 15-year period.

Sensitivity assessment. Therefore, the weight of evidence presented suggests that seagrasses are probably resistant and, hence, ‘Not sensitive’ to heavy metal contamination, especially those concentrations reported in the environment.  Halophila spinulosa is an exception when exposed to high concentrations (1 mg/l for 6 days) of copper.  Technically, the response of Halophila spinulosa could be interpreted as the ‘worst-case’ scenario.  But the overall weight of evidence suggests it was an exception, and it is unwise to extrapolate this to the entire dataset based on one observation in a single study.  Nevertheless, studies of Zostera spp. dominated the evidence review (50% of records) so that the sensitivity assessment is probably representative of Zostera spp.  All the papers examined were of High quality, and ‘High or Medium’ applicability and all (except one) did not report mortality.  Therefore, confidence is assessed as ‘Medium’.

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 'Hydrocarbons and PAH' contaminants on seagrasses are summarized below. The full 'seagrass evidence review' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Hydrocarbon' contaminant examined, together with an overall assessment for the pressure. 

Oil spills.  The effects of the oil spills on seagrass meadows were inconsistent and variation was reported between seagrass species and oil types.  Studies have shown some seagrass meadows to be tolerant to oil spill exposure and others have resulted in severe mortality.

Zostera marina is tolerant to oiling (in the absence of dispersants or other cleaning treatments).  All reported effects on Zostera marina after exposure to spilt crude oil and fuel oil were sublethal.  Only sublethal, short-term damage was reported in the form of a decline in abundance in shoots, blades, and flowering shoots in the Exxon Valdez oil spill and blackened/burnt leaves in the Amoco Cadiz oil spill.

Other species are less tolerant. ‘Severe’ mortality was reported in 20% of the results of oil spills and is recorded in the species Phyllospadix torreyi, Posidonia oceanica, Thalassia testudinum and in unspecified Seagrass (var.) located in the Gulf of Mexico, after exposure to spilt crude oil.  ‘Some’ mortality was also seen in Thalassia hemprichii after the fuel oil Taklong Island National marine reserve oil spill.  In addition, the Deepwater Horizon oil spill report also recorded large-scale seagrass mortality/population loss but did not quantify the scale of losses.  Sublethal effects were reported in 65% of the results on oil spill damage to seagrass.  These ranged from reduced growth rates, bleaching, decreased density of shoots, reduced flowering success (Den Hartog & Jacobs 1980; Jacobs 1980; Dean et al. 1998; Keesing et al., 2018), blackening leaves, leaf loss (Den Hartog & Jacobs 1980; Jacobs 1980; Keesing et al., 2018) and reduced growth rate (Kenworthy et al., 1993).

Due to the low solubility of oil, subtidal seagrass species, such as Zostera marina, are exposed only to the water accommodating fraction (WAF) of oil or dispersed oil droplets meaning they are less susceptible to damage than intertidal seagrass beds that experience physical contact with oil leading to greater amounts of damage and mortality (Lopez, 1978; Zieman et al. 1984; Zieman & Zieman, 1989; Keesing et al. 2018; Fonseca et al. 2017). Other factors influencing the effect of oil on seagrass include seagrass species, oil type, intensity, duration, and circumstance of the exposure (Keesing et al., 2018).

Seagrass situated near an oil refinery in Milford Haven showed no chronic sensitivity or long-term effects to the exposure to the oil effluent.  However, this may have been due to little penetration of the effluent (Hiscock, 1987, cited in; Holt et al., 1995).  In addition, oil spills can cause indirect effects and mortalities to seagrass communities.  Heavy oiling can lead to an increase in algal growth resulting in heavy fouling that persists for several months after an oil spill has occurred due to the mortality of grazers (Jackson et al. 1989). Jacobs (1980) noted a larger algal bloom than in previous years after the Amoco Cadiz spill in Roscoff, probably as a result of increased nutrients (from dead organisms and breakdown of oil) and the reduction of algal grazers. However, herbivores recolonized and the situation returned to 'normal' within a few months.

Overall, based on the ‘worst case' scenario for oil spills, the resistance is assessed as ‘None’ for seagrasses as a group. Resilience is probably ‘Low’ so sensitivity to petroleum-based oil spills is assessed as ‘High’.  But the above evidence also suggests that Zostera spp. (and by inference Zostera dominated habitats), are ‘Not sensitive’ to oil spills (in the absence of dispersants or other cleaning treatments). The confidence in the assessment is probably ‘High’ because all of the reported effects on Zostera marina after exposure to spilt crude oil and fuel oil were sublethal.  However, the impact on the community living in the seagrass is often greater than the impact on the seagrass itself (Holt et al., 1995; Jacobs, 1988).

Petroleum hydrocarbons (oils).  The reported results to the exposure of petroleum oils on seagrass suggested that 6.4% of cases resulted in ‘Severe’ (>75%) mortality while another 6.25% of the articles reported ‘significant’ (25-75%) mortality and 18.75% of articles reported ‘some’ (<25%) mortality depending on the species of seagrass, type of oil and its concentration.

The majority of the reported effects of oil on seagrass were generally sublethal (64.5%).  These include reduced photosynthetic efficiency, loss of leaf pigmentation, reduced growth rate and leaf loss.  Exposure to oil was reported to cause ‘severe’ mortality in only 6.4% of the results.  The result of exposure differed depending on the type of oil used.  Louisiana crude caused ‘severe’ mortality in all reports of exposure of the seagrasses Syringodium filiforme and Halodule wrightii. Murban crude was less toxic to seagrass than Louisiana crude, causing only ‘some’ damage to these species.  Hence, oils from various sources have different levels of toxicity on seagrass and, therefore, may explain some of the different results.  Fuel oil was reported to only cause sublethal effects on seagrass (Costa, 1982; Wilson & Ralph, 2012).  However, both Zieman & Zieman (1989) and Keesing et al. (2018) noted that refined oils, diesel and bunker fuels were more toxic than crude oil.  The exposure of seagrass to the simulated coal dust spill resulted in only sublethal effects.

The differences seen between species were greater than that seen between oil types.  ‘Severe’ and ‘significant’ mortality were reported more often in the tropical species Syringodium filiforme and Halodule wrightii and Thalassia testudinum than Zostera marina and Zostera capricorni where exposure only led to sublethal effects.  There was ‘Some’ mortality reported when Zostera spp. were exposed to crude oil in a field experiment (Howard et al., 1989). However, these Zostera spp. were most likely the intertidal species Zostera noltei or the shallow extent of Zostera marina (as syn. Zostera angustifolia), which were more likely to have been into direct contact with the oil, and to experience more damage than subtidal species (Howard et al., 1989).

Technically, the worst-case sensitivity of seagrass to 'oils', as a group, would be assessed as ‘High’ (see the 'seagrass evidence review'; Table 1.2) based on the response of tropical species.  Native Zostera spp. are probably less sensitive and a sensitivity of ‘Medium’ is suggested in the intertidal based on the evidence presented by Howard et al. (1989), while subtidal species (and beds) are probably ‘Not sensitive’.  Confidence in the assessment is ‘Low’ due to the variation in effect shown in the evidence.

Dispersants. Across six dispersant treatments recorded, only two dispersants (BP 1100 WD and Corexit 9527) were reported to cause lethal effects in seagrasses.  Corexit 9527 was the most lethal dispersant.  Two records of ‘severe’ mortality in Syringodium filiforme and Halodule wrightii were recorded and two records of ‘significant’ mortality in Thalassia testudinum.  There was one report of ‘significant’ mortality in Zostera spp. after exposure to BP 1100 WD.  All other responses were sublethal.  Therefore, sensitivity to dispersants is assessed as ‘Medium’ for Zostera spp. and ‘High’ for seagrasses as a group.  However, confidence is assessed as ‘Low’ because of the variation in response between species, and the limited number of dispersants examined in the evidence review.

Dispersed oils. Overall, the reported results on the exposure to dispersed oils (oil and dispersant mixtures) suggest that 29.8% of cases could result in ‘Severe’ (>75%) mortality while another 33.3% of the articles reported ‘Significant’ (25-75%) mortality and 24.6% of articles reported ‘Some’ (<25%) mortality depending on the species of seagrass, type of oil, dispersant and the concentration of both.  Dispersed oil was reported to have a variety of effects on seagrass from ‘no observed’ mortality to 100% mortality.  Dispersed oil was more toxic than both oil and dispersant treatments alone with 89% of dispersed oil exposure resulting in a lethal effect on the seagrasses.  Different dispersant oil mixtures had various levels of toxicity.  The most toxic recorded dispersant mixed with crude oils was ConcoK(K), which had the highest number of results of ‘severe’ and ‘significant’ mortality (Thorhaug & Marcus, 1987b).

Dispersants can break down the waxy epidermal coating on the leaves allowing the toxic components to access the cellular membrane.  This allows for greater absorption of aliphatic oil fractions which increases the toxic damage and leads to a decreased tolerance to other stress factors (Zieman et al., 1984; Howard et al., 1989; Ralph & Burchett, 1998b; Wilson & Ralph, 2012).  In addition, Wilson & Ralph (2012) noted that the addition of dispersants increases the total petroleum hydrocarbon (TPH) concentration in the water column from 12 mg/l to 101 mg/l in crude oil and 3 mg/l to 522 mg/l in fuel oil.  These were considered realistic to those reported in oil spills with the higher concentrations being ‘worse-case’ scenarios (Wilson & Ralph, 2012).  However, they resulted in no recorded mortality in Zostera capricorni.  No mortality was also recorded in Zostera marina and Halophila ovalis after exposure to dispersed oils, which only experienced sublethal effects.  Sublethal effects were mostly short-term negative impacts on the photosynthetic efficiency and decreased pigmentation of leaves after exposure.  However, some species of seagrass were less tolerant of exposure to dispersed oil.  The tropical species of seagrasses showed a low resistance to dispersed oil exposure with ‘severe’ mortality reported in 2.6% of the results of exposure in Thalassia testudinum, 14.9% in Syringodium filiforme and 14% in Halodule wrightii (Thorhaug et al. 1986; Thorhaug & Marcus, 1987; Thorhaug & Marcus, 1987b).

However, Howard (1986) reported that treatment of Zostera spp. (probably Zostera noltei or lower shore intertidal Zostera marina) with premixed oil and dispersant treatment showed a significant decrease in cover within the first week that resulted in a decrease in cover from 55% to 15% after 18 months (Howard et al., 1989).

The worst-case sensitivity of seagrass, as a group, would be assessed as ‘High’ based on the response of tropical species.  Native Zostera spp. are probably less sensitive depending on the exposure. Intertidal Zostera noltei and lower shore intertidal Zostera marina beds may exhibit a ‘Medium ‘ sensitivity to dispersed oils based on the evidence presented by Howard et al. (1989), while subtidal species (and beds) are probably ‘Not sensitive’.  Confidence in the assessment is ‘Low’ due to the variation in effects shown in the evidence.

Polyaromatic hydrocarbons (PAHs).  The evidence on the effects of PAH contaminants on seagrass was limited, with only two relevant papers (Faganeli et al., 1997; Mauro et al., 2013).  In these papers, environmental exposure to PAH was recorded but no mortality or sublethal effects were reported.  Therefore, the resistance is assessed as ‘High’ and resilience as ‘High’, so that the sensitivity of seagrasses to PAH exposure is assessed as ‘Not sensitive’.

Sensitivity to 'Hydrocarbons and PAH' contamination.  Overall, seagrasses are probably highly sensitive to exposure to hydrocarbons as via oil spills, water accommodated fractions of oils and, in particular, oil and dispersant mixtures. However, the evidence on the effects of PAHs is limited. The native Zostera species were amongst the least sensitive species reviewed. Zostera marina may be partially protected from direct contact with oil due to its subtidal habitat. However, the 'worst-case' evidence suggests that intertidal Zostera noltei and lower shore intertidal Zostera marina beds may exhibit 'Medium' sensitivity to water accommodated oils and ‘Medium' sensitivity to dispersed oils based on the evidence presented by Howard et al. (1989), while subtidal beds are probably ‘Not sensitive’.  Confidence in the assessment is ‘Low’ due to the variation in effects shown in the evidence (see 'seagrass evidence review; Table 1.2). 

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 compound' contaminants on seagrasses are summarized below. The full 'seagrass evidence review' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Synthetic' contaminant examined, together with an overall assessment for the pressure. 

The effects of herbicides were examined in 92% of the results in the evidence review of pesticides and the antifoulant (pesticide) Irgarol was examined in the remaining 8% of results.  The number of articles that report mortalities due to synthetic contaminants is summarized in the 'seagrass evidence review' (Figure 1.7 and in Table 1.4).  Herbicides are released into the water column via spraying and via runoff from agriculture or land management.  In a couple of studies (Patten, 2003, Major et al., 2004) the articles examined the effect of herbicides used to control Spartina in the past.  Both studies concluded that the effect of the herbicide was limited and the potential effect of Spartina on seagrass beds was worse.

It is not surprising that most papers examined the effects of herbicides on photosynthesis and, hence, growth in seagrasses, as many herbicides specifically target the PSII (photosystem II) of plants.  The effects varied with concentration, duration of exposure, type of herbicide, seagrass species and mode of application.  Nevertheless, 76% of the reported effects were sublethal, ‘some’ mortality was only reported in a single article and ‘severe’ (>75%) mortality in seven articles (18% of reported effects).  Therefore, the resistance to herbicides is probably ‘None’ based on the examples of ‘severe’ mortality reported in the evidence review.  Hence, an overall sensitivity of ‘High’ is suggested for herbicides and pesticides in general for seagrasses.  In addition, 72% of the reported effects of herbicides examined Zostera spp. and all the ‘severe’ mortality results were from studies of Zostera spp.  Therefore, the assessment is probably made with ‘High’ confidence.

This sensitivity assessment agrees with Bester (2000) who reported high concentrations of pesticides in areas of the German Bight where seagrass beds had been destroyed, with the caveat that further experimental evidence was required, and that other contaminants might have been involved.  However, several authors suggested that the sublethal effects on photosynthesis and growth would probably render the seagrass vulnerable to other adverse effects. 

The remaining evidence on the effect of pharmaceuticals, and other synthetics was each limited to a single article in the review.  Zostera marina was reported to be not affected by exposure to methanol but only as a control in a study on the effects of herbicides (Hershner et al., 1982).  The pharmaceutical study did not report any effect of the artificial auxin hormone on Zostera marina.  However, no evidence of the effect of human pharmaceuticals or maricultural or agricultural chemotherapeutics was found.  Therefore, Zostera marina is probably ‘Not sensitive’ to the pharmaceuticals, and other synthetic contaminants reviewed but with ‘Low’ confidence due to the limited evidence recovered.

Sensitivity assessment. Overall, resistance to the effect of ‘Synthetic compound’ contaminants on Zostera spp. is assessed as ‘None’ so that Zostera spp. beds (Zmar and Znol) are assessed as ‘High’ sensitivity, although the weight of evidence is based on the effect of pesticides and, in particular, herbicides.  The evidence on other types of synthetic contaminants is limited so overall confidence is assessed as ‘Medium’.

Radionuclide contamination

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

Evidence

No evidence found. 

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

Portillo et al. (2014) examined the effect of a disinfectant (SMBS) in the effluent for a desalination plant on Cymodocea nodosa seagrass bed.  They also concluded that exposure to SMBS affected significantly the survival and vitality of seagrass seedlings, probably as SMBS reduces the pH and dissolved oxygen concentration of the water column, and that its effect was greater under hypersaline conditions. But it was the hypersaline conditions (39 psu) that excluded the seagrass from the vicinity of the discharge.  However, no evidence of similar effects on Zostera spp. 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

The effects of oxygen concentration on the growth and survivability of Zostera marina are not reported in the literature. Zostera sp. leaves contain air spaces (lacunae). Oxygen is transported to the roots where it permeates into the sediment, resulting in an oxygenated microzone, enhancing the uptake of nitrogen. The presence of air spaces suggests that seagrass may be tolerant of low oxygen levels in the short-term, however, prolonged deoxygenation, especially if combined with low light penetration and hence reduced photosynthesis will have an adverse effect. 

Epifaunal gastropods may be tolerant of hypoxic conditions, especially Littorina littorea and Hydrobia ulvae. Infaunal species are likely to be exposed to hypoxic conditions, especially at low tide when they can no longer irrigate their burrows e.g. Arenicola marina can survive for 9 days without oxygen (Hayward, 1994). Conversely, possibly since it occupies the top few centimetres of sediment, Cerastoderma edule may be adversely affected by anoxia and would probably be killed by exposure to 2 mg/l oxygen for a week (benchmark). Loss of grazers would result in unchecked growth of epiphytes and other algae which may smother Zostera marina.

Sensitivity assessment. De-oxygenation is not likely to adversely affect seagrass beds in areas of adequate light.  The loss of grazing gastropods could result in smothering and potential reduction in the extent of the seagrass. At the level of the benchmark, both resistance and resilience are assessed as 'High' (no impact to recover from). Overall, the biotope is therefore assessed as 'Not Sensitive' to de-oxygenation at the pressure benchmark. 

Nutrient enrichment

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

Evidence

During the past several decades, important losses in seagrass meadows have been documented worldwide related to an increase in nutrient load.  Seagrasses are typically found in low energy habitats such as estuaries, coastal embayments and lagoons with reduced tidal flushing where nutrient loads are both concentrated and frequent.  A typical response to nutrient enrichment is a decline in seagrass populations in favour of macroalgae or phytoplankton (Baden et al., 2003).  Nutrient enrichment, especially of nitrogen and phosphorus, can lead to eutrophication.

The mechanisms responsible for seagrass decline under eutrophication are complex and involve direct and indirect effects relating to changes in water quality, smothering by macroalgal blooms (Den Hartog & Phillips, 2000), and competition for light and nutrients with epiphytic microalgae and with phytoplankton (Nienhuis, 1996).  In the Mondego estuary (Portugal), eutrophication triggered serious biological changes, which led to an overall increase in primary production and to a progressive replacement of seagrass Zostera noltei beds by coarser sediments and opportunistic macroalgae (Cardoso et al., 2004). Nutrients stimulate phytoplankton blooms that compete for nutrients but more importantly increase the turbidity and absorb light, reducing seagrass productivity (discussed in ‘changes in suspended solids’).  In general terms, algae are able to out-compete seagrasses for water column nutrients since they have a higher affinity for nitrogen (Touchette & Burkholder, 2000). Short and Burdick (1996) found that excessive nitrogen loading stimulated the proliferation of algal competitors that caused shading and thereby stressed Zostera plants. Many seagrasses have a positive response to nitrogen and/or phosphorous enrichment (Peralta et al., 2003), but excessive loads can inhibit seagrass growth and survival, not only indirectly through light reduction resulting from increased algal growth but also directly in terms of the physiology of the seagrass.  Direct physiological responses include ammonium toxicity and water column nitrate inhibition through internal carbon limitation (Touchette & Burkholder, 2000). Indirect effects of nutrient enrichment can accelerate decreases in seagrass beds such as sediment re-suspension from seagrass loss (see pressure on ‘changes in suspended solids’). Jones & Unsworth (2015) concluded that seagrass habitats in the British Isles were nutrient-enriched, with nitrogen levels 75% higher than the global average for Zostera marina, yet phosphate-limited, and concluded that many beds in the vicinity of human populations were in a poor state.

Sensitivity assessment. The loss of seagrass beds worldwide has been attributed to nutrient enrichment, due in part to the likeliness of smothering by epiphytes, and the effects of reduced light penetration caused by eutrophication.  For instance, a study by Greening & Janicki (2006) found that in Florida, the USA, recovery of seagrass beds was incomplete 20 years after nutrient enrichment caused an eutrophication event.  Seagrass beds are regarded as highly intolerant or 'Low' resistance to this pressure. However, the benchmark of this pressure (compliance with WFD ‘good’ status) allows for a 30% loss of intertidal seagrass beds under the WFD criteria for good status.  Therefore, at the level of the benchmark resistance of seagrass beds to this pressure is assessed ‘Medium’.  The resilience of seagrass beds this degree of impact is assessed as ‘Medium’ and sensitivity is, therefore, assessed as ‘Medium’.

Organic enrichment

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

Evidence

Organic enrichment may lead to eutrophication with adverse environmental effects including deoxygenation, algal blooms and changes in community structure (see ‘nutrient enrichment’ pressure).  Evidence on the effects of organic enrichment on Zostera species is limited but abundant for other seagrass species.

Neverauskas (1987) investigated the effects of discharged digested sludge from a sewage treatment on Posidonia spp. and Amphibolis spp. in South Australia. Within 5 years, the outfall had affected an area of approximately 1900 ha, 365 ha of which were completely denuded of seagrasses.  The author suggests that the excessive growth of epiphytes on the leaves of seagrasses was a likely cause for reduced abundance.  A subsequent study by Bryars & Neverauskas (2004) determined that 8 years after the cessation of sewage output, total seagrass cover was approximately 28% of its former extent. While these results suggest that seagrasses can return to a severely polluted site if the pollution source is removed, they also suggest that it will take many decades for the seagrass community to recover to its former state.

The effects of organic enrichment from fish farms were investigated on Posidonia oceanica seagrass beds in the Balearic Islands (Delgado et al., 1999).  The fish culture had ceased in 1991; however, seagrass populations were still in decline at the time of sampling.  The site closest to the former fish cages showed a marked reduction in shoot density, shoot size, underground biomass, sucrose concentration and photosynthetic capacities.  The shoot also had high P-concentration in tissues and higher epiphyte biomass compared to the other sites.  Since water conditions had recovered completely by the time of sampling, the authors suggest that the continuous seagrass decline was due to the excess organic matter remaining in the sediment (Delgado et al., 1999).

It should be noted that coastal marine sediments where seagrasses grow are often anoxic and highly reduced due to the high levels of organic matter and slow diffusion of oxygen from the water column to the sediment.  Seagrasses worldwide have been shown to exhibit a three-way symbiotic relationship with the small lucinid bivalves (hatchet-shells, e.g. Loripes and Lucinoma) and their endosymbiotic sulfide-oxidizing gill bacteria (Van der Heide et al., 2012). In experiments, the sulfide-oxidizing gill bacteria of Loripes lacteus were shown to reduce sulfide levels in the sediment and enhance the productivity of Zostera noltei, while the oxygen released from the roots of Zoster noltei was of benefit to Loripes. Nevertheless, the negative effects of the experimental addition of sulphide were not fully prevented by the presence of Loripes (Van der Heide et al., 2012). Therefore, while seagrasses or the Zostera-lucinid symbiosis are adapted to these anoxic sediment conditions if the water column is organically enriched, plants are unable to maintain oxygen supply to the meristem and die fairly quickly.  The enrichment of the water column could, therefore, significantly increase the sensitivity of seagrasses to this pressure. Worldwide evidence suggests that nutrient enrichment is one of the biggest threats to seagrass populations (Jones & Unsworth, 2015).

Sensitivity assessment. The organic enrichment of the marine environment increases turbidity and causes the enrichment of the sediment in organic matter and nutrients (Pergent et al., 1999). Evidence shows that seagrass beds found in proximity to a source of organic discharge were severely impacted with important losses of biomass.  Although no study was found on the British species, the evidence suggests that Zostera marina will be negatively affected by organic enrichment.  No evidence was found addressing the benchmark of this study.  A deposition of 100 gC/m2/year is considerably lower than the amount of organic matter discharged by sewage outlets and fish farms.  Therefore, resistance to this pressure is assessed as ‘Medium’, recovery as ‘Medium’, and sensitivity as ‘Medium’.

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 resulting in 'Very Low' resilience.  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.  Adjacent habitats and species populations may be indirectly affected where meta-population dynamics and trophic networks are disrupted and where the flow of resources e.g. sediments, prey items, loss of nursery habitat etc. is altered.

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

A change to another seabed type (from sediment to hard rock) will result in a permanent loss of suitable habitat for seagrass species. Resistance is thus assessed as ‘None’.  As this pressure represents a permanent change, recovery is impossible as a suitable substratum for seagrasses is lacking. Consequently, resilience is assessed as ‘Very low’.  The habitat, therefore, scores a ‘High’ sensitivity. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

Physical change (to another sediment type)

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

Evidence

Seagrass beds occur almost exclusively in shallow and sheltered coastal waters anchored in sandy and muddy bottoms.  Coarser sediments reduce the vegetative spreading of seagrasses and inhibit seedling colonization (Gray & Elliott, 2009).  Changes in sediment type can, therefore, have wider implications on the distribution of seagrass beds. Hence, change towards a coarser sediment type would inhibit seagrasses from becoming established due to a lack of adequate anchoring substratum.  A more mud dominated habitat, on the other hand, could increase sediment re-suspension and exclude seagrasses due to unfavourable light conditions.  

Sensitivity assessment. The resistance was assessed as ‘Low’. As this pressure represents a permanent change, recovery is impossible without intervention as a suitable substratum for seagrasses is lacking. Consequently, resilience is assessed as ‘Very low’.  The habitat, therefore, scores a ‘High’ sensitivity. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

Habitat structure changes - removal of substratum (extraction)

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

Evidence

The extraction of sediments to 30 cm (the benchmark) will result in the removal of every component of seagrass beds.  Roots and rhizomes are buried no deeper than 20 cm below the surface (see ‘abrasion’ and ‘penetration and/or disturbance of the substratum below the surface of the seabed’ pressures).  Resistance is therefore assessed as ‘None’ and resilience is considered ‘Very Low’ resulting in a ‘High’ sensitivity score.  The confidence assessment for this pressure is high as it is based on the characteristics of the pressure i.e. complete removal of the feature within the pressure footprint.

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

Seagrasses are not physically robust. The leaves and stems of seagrass plants rise above the surface and the roots are shallowly buried so that they are vulnerable to surface abrasion. Activities such as trampling, anchoring, power boating and potting are likely to remove leaves and damage rhizomes. The removal of above-ground biomass would result in a loss of productivity whilst the removal of roots would cause the death of plants.  Seagrasses are limited to shallow, protected waters and soft sediments.  These areas are often open to public access and are widely used in commercial and recreational activities. Evidence for abrasion impacts is summarised below for activities that give rise to this pressure.

Trampling: human wading in shallow coastal waters is a common activity that inherently involves trampling of the substratum.  Trampling may be caused by recreational activities such as walking, horse-riding and off-road driving. These activities are likely to damage rhizomes and cause seeds to be buried too deeply to germinate (Fonseca, 1992).  Negative effects of human trampling on seagrass cover, shoot density, and rhizome biomass, have been reported by Eckrich & Holmquist (2000) for the seagrass Thalassia testudinum. The study found that recovery occurred within a period of seven months after trampling ceased but the reduced cover was still visually distinguishable 14 months after the experiment. A study by Major et al. (2004) found that trampling impact varied depending on substratum type. A significant decrease in shoot density as a result from trampling was only observed at a site with soft muddy substratum with no impact detected on the hard packed sand substratum. Damage from trampling is thus dependent on the substratum type with seagrass beds growing on soft substrata being most vulnerable to this pressure.

Boating activities: boats passing in close proximity to seagrass beds can create waves.  Turbulence from propeller wash and boat wakes can resuspend sediments, break off leaves, dislodge sediments and uproot plants. The re-suspension of sediments is further assessed in ‘changes in suspended sediment’ pressure.  Koch (2002) established that physical damage from boat wakes was greatest at low tide but concluded that negative impacts of boat-generated waves were marginal on seagrass habitats. The physical impact of the engine’s propellers, shearing of leaves and cutting into the bottom, can also have damaging effects on seagrass communities.  In severe cases, propellers cutting into the bottom may completely denude an area resulting in narrow dredged channels through the vegetation called propeller scars.  Scars might expand and merge to form larger denuded areas.  A study in Florida looking at the seagrasses Thalassia testudinum, Syringodium filiforme and Halodule wrightei determined that recovery of seagrass to propeller impact depend on species (Kenworthy et al., 2002).  For Syringodium filiforme recovery was estimated at 1.4 years and for Halodule wrightei at 1.7 years, whilst recovery for Thalassia testudinum was estimated to require 9.5 years.  Variations in recovery time were explained by different growth rates.  However, it is not appropriate to assume that recovery rates are similar from one geographical or climatic region to another and more in-depth research is needed for Zostera species around the British Isles. 

Potting: static gear is commonly deployed in areas where seagrass beds are found, either in the form of pots or as bottom set gill or trammel nets. Damage can be caused during the setting of pots or nets and their associated ground lines and anchors, by their movement over the bottom during rough weather and during recovery. Whilst the potential for damage is lower per unit deployment compared to towed gear (see 'penetration and/or disturbance of the substratum below the surface of the seabed' pressure), there is a risk of cumulative damage if use is intensive. Hall et al. (2008) categorized seagrass beds as being highly sensitive to high intensities of potting (pots lifted daily, with a density of over 5 pots per ha) and medium sensitive to lower levels (pots lifted daily, less than 4 pots per ha).  However, no direct evidence was found to confirm these estimates. 

Grazing: Nacken & Reise (2000) investigated physical disturbance caused by Brent geese (Branta b. bernicla) and widgeon (Anas penelope) feeding on Zostera noltei in the northern Wadden Sea. To graze on leaves and shoots above the sediment and on rhizomes and roots below, birds reworked the entire upper 1 cm layer of sediment and excavated pits by trampling. As a result, birds pitted 12% of the seagrass bed and removed 63% of plant biomass. Plants recovered by the following year with the authors suggesting that seasonal erosion caused by herbivorous wildfowl was necessary for the persistence of Zostera noltei beds (Nacken & Reise, 2000). Similarly, Tubbs & Tubbs (1982, 1983; see Davison & Hughes, 1998) suggested that Zostera sp. can rapidly recover from 'normal' levels of wildfowl grazing. Physical disturbance may, however, be detrimental to seagrass beds as soon as the ‘normal’ level caused by grazing birds is exceeded by human activities. In addition, geese and wigeon do not dive so that shoots below the reach of their necks at low tide are 'safe' from grazing pressure.

Experimental: Boese et al. (2009) examined the recolonization of experimentally created gaps within intertidal perennial and annual Zostera marina beds in the Yaquina River Estuary, USA.  The experiment looked at two zones, the lower intertidal almost continuous seagrass and an upper intertidal transition zone where there were patches of perennial and annual Zostera marina.  The study found that recovery began within a month after a disturbance in the lower intertidal continuous perennial beds and was complete after two years, whereas, plots in the transition zone took almost twice as long to recover.  

Sensitivity assessment. In summary, a wide range of activities gives rise to this pressure with intertidal habitat being more exposed as they are more readily accessible than subtidal beds. The resilience and recovery of seagrass beds to abrasion of the seabed surface depends on the frequency, persistence and extent of the disturbance. Factors such as the size and shape of the impact will also influence the sensitivity of seagrass. There is also considerable evidence that the type of substratum plays a role in determining the magnitude of impact. Soft and muddy substratum is thought to be more easily damaged than harder more compact ground. Finally, temporal effects should also be taken into account. The state of the tide will influence the magnitude of damage as will seasonal effects with damage in winter likely to have less impact than the damage that occurs during the growing season. Overall, studies suggest little resistance to abrasion resulting in an assessment of ‘Low’ resistance. Physical disturbance and removal of plants can lead to increased patchiness and destabilisation of the seagrass bed, which in turn can lead to reduced sedimentation within the seagrass bed, increased erosion, and loss of larger areas of plants (Davison & Hughes, 1998). Recovery will, however, be fairly rapid and resilience is assessed as ‘Medium’. Therefore, sensitivity is assessed as ‘Medium’ to this pressure.

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

Seagrass species are vulnerable to physical damage. The leaves and stems of seagrass plants rise above the surface and the roots are shallowly buried. Activities such as digging and raking for clams, anchoring and mooring will penetrate the substratum to an average depth of 5 cm removing plant biomass above and below ground. Abrasion to the substratum to a depth greater than 5 cm will directly impact seagrass habitats and all biomass (leaves, rhizomes) will be completely removed leading to the death of the plant in the area impacted.  Seagrass beds are often associated with commercially important bivalves.  Fisheries targeting these species are therefore likely to impact seagrass habitats and are the most widespread (and best studied) activities giving rise to this pressure on this habitat. The extent of the damage on seagrass beds depends on the activity.

Clam digging and clam raking: Boese (2002) investigated the effects of manual clam harvesting on Zostera marina by raking and digging for clams in experimental plots in Yaquina Bay, USA.  After three monthly treatments, measures of biomass, primary production (leaf elongation), and percent cover were compared between disturbed and undisturbed plots. The study found that clam raking treatments visibly removed large numbers of seagrass leaves and some below-ground rhizomes. However, two weeks after the end of the experiment, no statistical difference in percentage cover was observed between disturbed and control plots indicating a fast recovery rate. Clam digging, on the other hand, caused visual differences in percentage cover for 10 months after the end of the experiment, although differences were not statistically significant. Boese (2002) concluded that recreational clamming is unlikely to have a major impact on seagrass beds in the Yaquina estuary. The author calls, however, to view the results with caution as multi-year disturbances were not investigated and differences in sediment characteristics are likely to influence the resistance and resilience of seagrasses to this pressure. Similarly, Peterson et al. (1987) found that hand raking and moderate clam-kicking (a commercial harvesting method in which propeller wash is used to dislodge hard clams) resulted in a reduction in Zostera marina biomass by approximately 25%. No differences between control and experimental areas were apparent one year after the experiment. However, at a higher intensity, clam-kicking reduced seagrass biomass to about half of control levels and recovery remained incomplete four years after the end of the experiment (Peterson et al., 1987).

Anchoring and mooring: an anchor landing on a patch of seagrass can bend, damage and break seagrass shoots (Montefalcone et al., 2006) and an anchor being dragged as the boat moves driven by wind or tide causes abrasion of the seabed. Milazzo et al. (2004) found that the extent of damage depended on the type of anchor with the folding grapnel having the greatest impact. The study further determined that heavier anchors (often associated with larger boats) will sink deeper into the substratum and thereby causing greater damage. A technical paper by Collins et al. (2010) using SCUBA divers found bare patches (typically 1–4 m2) were caused by anchoring by leisure boats in Studland Bay, UK. The study further determined that average shear vane stress was significantly higher in intact seagrass beds compared to scars indicating a less cohesive and more mobile substratum caused by anchors. Axelsson et al. (2012) also investigated anchor damage in Studland Bay. The study did not provide consistent evidence of boat anchoring impacting the seagrass habitat in this location. The study did, however, observe higher shoot density and percentage cover of seagrass in a voluntary anchor zone compared to a control area where anchoring occurred. The authors recommended longer monitoring in order to determine whether the trend was caused by natural variations or the effects of anchor exclusion. Traditional mooring further contributes to the degradation of seagrass habitats.  A traditional swing mooring is a buoy on a chain attached to a static anchoring block fixed on the seabed, to buffer any direct force on the permanent block, the chain lies on the seabed where it moves around with wind and tides, as the chain pivots on the block it scours the seabed.  In proximity to seagrass beds, the chain usually removes not only the seagrass above ground parts such as leaves and shoots but also the roots anchored in the sediment.  Further sediment abrasion may occur in the vicinity to the anchoring blocks due to eddying of currents.  The blocks themselves may increase the competition of seagrass with other algae as they provide ideal settlement surfaces. Boats might also moor on intertidal sediments.  When the tide goes out, the boat sits directly on top of the soft sediment.  Walker et al. (1989) found that boat moorings caused circular or semi-circular depressions of bare sand within seagrass beds between 3 to 300 m2 causing important habitat fragmentation.  The scours created by moorings in the seagrass canopy interfere with the physical integrity of the meadow.  Though relatively small areas of seagrass are damaged by moorings, the effect is much greater than if an equivalent area was lost from the edge of a meadow.  Such mooring scars have been observed for Zostera marina around the UK such as in Porth Dinllaen in the Pen Llyna’r Sarnau Special Area of Conservation, Wales (Egerton, 2011) and at Studland Bay (Jackson et al., 2013).

Trawling: bottom trawling and dragging are industrial fishing methods which scour the seabed to collect target species.  Neckles et al. (2005) investigated the effects of trawling for the blue mussels Mytilus edulis on Zostera marina beds in Maquoit Bay, USA.  Impacted sites ranged from 3.4 to 31.8 ha in size and were characterized by the removal of above- and belowground plant material from the majority of the bottom.  The study found that one year after the last trawl, Zostera marina shoot density, shoot height and total biomass averaged respectively to 2-3%, 46-61% and < 1% that of the reference sites.  Substantial differences in Zostera marina biomass persisted between disturbed and reference sites up to 7 years after trawling.  Rates of recovery depended on initial fishing intensity but the authors estimated that an average of 10.6 years was required for Zostera marina shoot density to match pre-trawling standards.

Dredging and suction dredging: the effects of dredging for scallops on Zostera marina beds were investigated by Fonseca et al. (1984) in Nova Scotia, USA.  Dredging was carried out when Zostera marina was in its vegetative stage on hard sand and on soft mud substrata.  The damage was assessed by analysing the effects of scallop harvesting on seagrass foliar dry weight and on the number of shoots. Lower levels of dredging (15 dredges) had a different impact depending on substrata, with the hard bottom retaining a significantly greater overall biomass than the soft bottom.  However, an increase in dredging effort (30 dredges) led to a significant reduction in Zostera marina biomass and shoot number on both hard and soft bottoms. Solway Firth is a British example of the detrimental effects of dredging on seagrass habitats. In the area, where harvesting for cockles by hand is a traditional practice, suction dredging was introduced in the 1980s to increase the yield.  A study by Perkins (1988) found that where suction dredging occurred, the sediment was smoothened and characterized by a total absence of Zostera plants.  The study concluded that the fishery was causing widespread damage and could even completely eradicate Zostera from affected areas.  Due to concerns over the sustainability of this fishing activity, the impacts on cockle and Zostera stocks, and the effects on overwintering wildfowl, the fishery was closed to all forms of mechanical harvesting in 1994.

Sensitivity assessment. The deployment of fishing gears on seagrass beds results in physical damage to the above surface part of the plants as well as to the root systems. Seagrasses do not have an avoidance mechanism; resistance to this pressure is therefore assessed as ‘None’. The recovery of seagrass beds after disturbance to the sub-surface of the sediment will be slow with the speed depending on the extent of removal.  Rates may be accelerated where adjacent seed sources and viable seagrass beds are present but can be considerably longer where rhizomes and seed banks were removed.  Using a model simulation, it has been suggested that with favourable environmental conditions, seagrass beds might recover from dragging disturbance in 6 years but, conversely, recovery under conditions less favourable to seagrass growth could require 20 years or longer (Neckles et al., 2005). Resilience is thus assessed as ‘Low’. The mechanical harvest of shellfish damaging the sub-surface of the sediments poses a very severe threat to seagrass habitats, yielding a ‘High’ sensitivity score.

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

Irradiance decreases exponentially with increasing depth, and the suspended sediment concentration has a direct linear effect on light attenuation (van Duin et al., 2001).  Changes in suspended solids will thus reduce the light available for seagrass plants necessary for photosynthesis.  Impaired productivity due to a decrease in photosynthesis will affect the growth and reproductive abilities of plants.  Turbidity also results in a reduction of the amount of oxygen available for respiration by the roots and rhizomes thus lowering nutrient uptake.  The resulting hypoxic conditions will lead to a build-up of sulphides and ammonium, which can be toxic to seagrass at high concentrations (Mateo et al., 2006).  Giesen et al. (1990a,b; Davison & Hughes, 1998) suggested that considerable declines in seagrass populations in the Wadden Sea were related to increases in turbidity from dredging and deposit extraction.

Water clarity is a vital component for seagrass beds as it determines the depth-penetration of photosynthetically active radiation of sunlight.  Seagrasses have light requirements an order of magnitude higher than other marine macrophytes making water clarity a primary factor in determining the maximum depth at which seagrasses can occur.  The critical threshold of light requirements varies among species ranging from 2% in-water irradiance for Zostera noltei, to 11 to 37% for Zostera marina (Erftemeijer & Robin, 2006). These differences in the light requirement for Zostera are reflected by the position of species along a depth gradient with Zostera noltei occurring predominantly in the intertidal and Zostera marina found at greater depth in the subtidal.  However, differences in light requirements also vary within species.  For example, the minimum light requirement for Zostera marina in a Danish embayment was 11% in-water irradiance, whereas the estimated light requirement for the same species in the Netherlands was 29.4% in-water irradiance (Olesen, 1993).  This variability within species is likely attributed to photo-acclimation to local light regimes.  In a six month long experiment in the Dutch Wadden Sea, Philippart (1995) found that shading induced a 30% decrease in the leaf growth rate, a 3-fold increase in the leaf loss rate, and an 80% reduction in the total biomass of Zostera noltei.  The decreasing growth rate is most probably due to reduced photosynthesis caused by shading.  The increased leaf loss may have been the result of enhanced deterioration of leaf material under low light conditions.  The study also established that during the summer period, the maximum biomass of Zostera noltei under the control light conditions was almost 10 times higher than those under the low light conditions (incident light reduced to 45% of natural light conditions).  The summer is a critical period for maintenance and growth of vegetative shoots.  The effects of shading may, therefore, be most severe during the summer months.  A similar response to reduced light availability for Zostera marina was observed by Moore & Wetzel (2000).

Increases in turbidity over a prolonged period of time are therefore highly likely to impact seagrass species.  Sensitivity will depend on individual seagrass beds.  Older, more established perennial meadows have greater carbohydrate reserves and are thus more able to resist changes in light penetration than annual plants (Alcoverro et al., 2001).  Seagrass plants found in clear waters may be able to tolerate sporadic high turbidity (Newell & Koch, 2004).  However, where seagrass beds are already exposed to low light conditions, then losses may result from even short-term events (Williams, 1988).  The growth of both Zostera marina and its associated epiphytes are reduced by increased shading due to turbidity (reduction of light penetration by 42, 28 and 9%).  Backman & Barilotti (1976) further established that intensive shading (reduction of light penetration by 63%) inhibited flowering in Zostera marina plants.

Sensitivity assessment. Turbidity is an important factor controlling production and ultimately survival and recruitment of seagrasses.  Seagrass populations are likely to survive short-term increases in turbidity, however, a prolonged increase in light attenuation, especially at the lower depths of its distribution, will probably result in loss or damage of the population. Therefore, resistance is assessed as ‘Low’.  A loss of seagrass beds will promote the re-suspension of sediments, making recovery unlikely as seagrass beds are required to initially stabilise the sediment and reduce turbidity levels (Van der Heide et al., 2007).  A high turbidity state appears to be a highly resilient alternative stable state; hence return to the seagrass biotope is unlikely resulting in ‘Low’ resilience.  Zostera marina should be considered intolerant of any activity that changes the sediment regime where the change is greater than expected due to natural events, and sensitivity is assessed as  ‘High’.

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

Several studies have documented the deterioration of seagrass meadows by smothering due to excessive sedimentation.  The consequences of enhanced sedimentation for seagrass beds depend on several factors such as the life history stage as well as the depth and timing of burial. 

Early life stages of seagrass, smaller in size than adult plants, are most vulnerable to this pressure as even a small load of added sediment will lead to the complete burial. Wang et al. (2016) reported that seed germination and seedling establishment varied with burial depth and sediment type, in Zostera marina from Lidao Bay, Shandong Peninsula, China. High germination rates (76-90%) were found in seeds on the sediment surface or buried to 1 cm after seven weeks in the laboratory, regardless of sediment type but germination rates fell to below 40% with increasing burial. Only those seeds in a sand:silt mixture (2:1) were able to germinate when buried to 5 cm. seedling establishment was similar and seeds buried to shallow depths had the highest establishment rates. A maximum of seedling establishment of ca 30% was recorded buried to 1 cm in sand:silt (2:1) (Wang et al., 2016). Wang et al. (2016) also reported that the burial limit of Zostera marina seeds varied between 1 and 9 cm in various studies but noted that survival of seeds probably depended on a range of factors including depth, sediment type and oxygen levels in the sediment.

Vermaat et al. (1997) found that adult Zostera marina in the Dutch Wadden Sea was able to cope with sedimentation rates between 2 and 13 cm per year as the plant has the capacity to elongate vertical stems enabling it to raise the leaf canopy above the sediment load.  However, a study in the USA observed over 50% mortality of plants of Zostera marina in field burial treatments of 4 cm (corresponding to 25% of plant height) for 24 days (Mills & Fonseca, 2003).  Plants buried 75% or more of their height (16 cm) experienced 100% mortality, which indicated a low resistance of Zostera marina to burial.  Munkes et al. (2015) noted that the Zostera marina in the Mill & Fonseca (2003) study were smaller than the normal ca 50 cm leaf length. Munke et al. (2015) noted that the effect of burial depended on the actual leaf length rather than species size.  In their experimental field study in Kiel Bight, Munke et al. (2015) found negative effects on shoot mortality, delayed growth and flowering and reduced carbohydrate storage even after burial under the sand at 5 cm (ca 10% of plant height) and four weeks, the lowest burial depth and shortest duration examined. The effects were significant enough to affect the next year's growth. Burial by greater than 5 cm resulted in shoot mortality but burial by 5 cm reduced maximum biomass, leaf length, starch storage and flowering capacity. Burial by 20 cm (ca 40% of plant height) resulted in high shoot mortality (ca 97%) after 10 weeks. Munke et al. (2015) also noted that their experimental seagrass meadow was in good condition and that stressed populations could exhibit more adverse effects.

The timing of the siltation event also plays a role in particular for intertidal beds.  At low tide, the seagrass bed is exposed with plants lying flat on the substratum. The addition of material would immediately smother the entire plant and have a greater impact on leaves and stem than if added on plants standing upright.  The resistance of intertidal beds to this pressure may thus vary with time of day.

Sensitivity assessment. Above studies suggest that Zostera marina is not resistant to smothering with some variation between the critical threshold depths of burial and sediment type. All studies, however, indicate that at the level of the benchmark (5 cm of fine material added to the seabed) some mortality and loss of biomass may occur. Therefore, resistance is assessed as  'Low'. Some plants will survive by successfully relocating rhizomes closer to the sediment surface. With the benchmark set at ‘material added to the seabed in a single event’, the sensitivity will be greater than if burial occurred in a continuous way. In addition, seagrass beds are restricted to low energy environments, suggesting that once the silt is deposited, it will remain in place for a long period of time so habitat conditions will not reduce exposure. Resilience is therefore assessed as 'Medium' and sensitivity as ‘Medium’ siltation at the pressure benchmark.

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

Zostera marina is intolerant of smothering by excessive siltation (see above). Seagrasses can cope with small rates of sedimentation by relocating their rhizomes closer to the sediment surface (Vermaat et al., 1997). Mills & Fonseca (2003) however observed 100% mortality in Zostera marina plants buried at a depth of 16 cm. Burial by 20 cm (ca 40% of plant height) resulted in high shoot mortality (ca 97%) after 10 weeks (Munke et al., 2015).

Resistance to sedimentation at the pressure benchmark (30 cm of added material) is therefore assessed as ‘None’ as all individuals exposed to siltation are predicted to die and consequent resilience as ‘Low’ to ‘Very Low’. In addition, seagrass beds are restricted to low energy environments, suggesting that once the silt is deposited, it will remain in place for a long period of time so habitat conditions will not reduce exposure. Sensitivity based on combined resistance and resilience is therefore assessed as ‘High’.

Litter

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

Evidence

Not assessed

Electromagnetic changes

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

Evidence

No evidence

Underwater noise changes

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

Evidence

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

Introduction of light or shading

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

Evidence

An increase in light might be beneficial while shading by artificial structures will decrease incident light and hence reduce photosynthesis and growth rates.  For example, in mesocosm experiments, Frederick et al. (1995) noted that shading (at 11, 21, 41, 61, and 94% of incident surface light for one week) resulted in a reduction in shoot density and an increase in shoot height. But shading alone did not cause mortality in the experimental time frame.  Holmer & Laursen (2002) noted that shading affected Zostera marina from a low-light, organic rich sediment population more than light saturated, low-organic sediment population.  However, the effects were significant in spring but not in autumn, and were also related to the plant's ability to tolerant anoxic and sulfidic conditions. 

Overall, there is little evidence of seagrass mortality resulting from shading directly but the effects of shading and smothering from epiphytes and macroalgae are discussed under nutrient enrichment and the effects of light attenuation under 'water clarity' above. However, the effects of shading could mirror those of reduced water clarity (increased turbidity) depending on the scale of the artificial structure. Therefore, a resistance of 'Low', with a resilience of 'Low' and sensitivity of 'High' is suggested, albeit with low confidence. 

Barrier to species movement

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

Evidence

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

Death or injury by collision

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

Evidence

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

Visual disturbance

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

Evidence

Not relevant

Genetic modification & translocation of indigenous species

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

Evidence

Translocation of seagrass seeds, rhizomes and seedlings is a common practice globally to counter the trend of decline of seagrass beds.  However, Williams & Davis (1996) found that levels of genetic diversity of restored Zostera marina beds in Baja California, USA, were significantly lower than in natural populations.  A subsequent study by Williams (2001) determined that the observed genetic bottleneck was a consequence of the collection protocol of source material (i.e. founder effect).  Founder effects are likely to occur if seeds used to revegetate restoration sites are collected from a limited number of sources.  Similar to episodes of colonization, the ‘founding’ propagules can represent only a portion of the genetic diversity present in the source populations, and they might hybridize with local genotypes (Hufford & Mazer, 2003).  The loss of genetic variation can lead to lower rates of seed germination and fewer reproductive shoots, suggesting that there might be long-term detrimental effects for population fitness.  Williams (2001) affirms that genetic variation is essential in determining the potential of seagrass to rapidly adapt to a changing environment.  Transplanted populations are therefore more sensitive to external stressors such as eutrophication and habitat fragmentation, with a markedly reduced community resilience than natural populations (Hughes & Stachowicz, 2004).

Translocation also has the potential to transport pathogens to uninfected areas (see 'introduction of microbial pathogens' pressure).  The sensitivity of the ‘donor’ population to harvesting to supply stock for translocation is assessed for the pressure ‘removal of target species’. No evidence was found for the impacts of translocated beds on adjacent natural seagrass beds.  However, it has been suggested that translocation of plants and propagules may lead to hybridization with local wild populations. If this leads to loss of genetic variation there may be long-term effects on the potential to adapt to changing environments and other stressors.

Sensitivity assessment. Presently, there is no evidence of loss of habitat due to genetic modification and translocation of seagrass species, resistance and resilience to this pressure are thus considered to be ‘High’ (no impact to recover from). Overall the biotope is therefore 'Not Sensitive' to this pressure. However, if hybridization occurred, recovery would not be considered possible unless the population is eradicated and replaced.  In this case, resilience is thus deemed ‘Very Low’ resulting in an overall ‘Low’ sensitivity score.  As there is no direct evidence to support assessments, these are based on expert judgement. 

Introduction or spread of invasive non-indigenous species

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

Evidence

The effects on native species on seagrass species were reviewed by d’Avack et al. (2014). The review reported several non-native invasive plants as well as invertebrate species negatively impacting British seagrass beds. The (potential) impact of each invasive non-indigenous species (INIS) is reported below. 

Non-native invasive plants: among the INIS currently present in the UK, the large brown seaweed Sargassum muticum has the most direct impact on Zostera species. Druehl (1973) was the first to raise concern about the potential negative effects of S. muticum on Zostera beds in British waters.  Zostera and Sargassum muticum were thought to be spatially separated due to their preferred habitat.  Zostera species grow on sand and muddy bottoms, whereas Sargassum muticum attaches to solid substratum.  However, when the seabed consists of a mixed substratum of sand, gravel and stones both species may occur together.  Even though there are no indications of direct competition between the two species (Den Hartog, 1997), Sargassum muticum establishes itself within seagrass habitats where beds are retreating due to natural or anthropogenic causes.  The invasive seaweed almost immediately occupies the empty spaces thereby interfering with the natural regeneration cycle of the bed.  In addition, a study in Salcombe, SW England by Tweedley et al. (2008) demonstrated that the presence of Zostera marina may help the attachment of Sargassum muticum on soft substrata by trapping drifting fragments thereby allowing viable algae spores to settle on the seagrass matrix in an otherwise unfavourable environment.  Once the invasive seaweed establishes itself, Zostera marina is unable to regain the lost territory indicating that eventually, Sargassum muticum is able to replace seagrass beds particularly on mixed substratum (Den Hartog, 1997).

The cord grass Spartina anglica is non-native grass, which was recorded to have negative effects on seagrass beds.  This hybrid species of native (Spartina alterniflora) and an introduced cord grass species (Spartina maritima) colonizes the upper part of mud flats, where due to its extensive root system, it effectively traps and retains sediments. Spartina anglica has rapidly colonized mudflats in England and Wales due to its fast growth rate and high fecundity.  Deliberate planting to stabilise sediments accelerated its spread throughout Britain (Hubbard & Stebbings, 1967).  By consolidating the sediments the plant is responsible for raising mud flats as well as reducing sediment availability elsewhere.  Butcher (1934) raised concerns that its pioneering consolidation may result in the removal of sediments from Zostera beds.  Declines in Zostera noltei due to the encroachment of Spartina anglica were observed in Lindisfarne National Reserve in north-east England (Percival et al., 1998). The reduction in Zostera noltei beds had a direct impact on wildfowl populations as the food availability for the wildfowl was reduced on the top of the shore.  This pressure will affect the upper limits of the intertidal rather than subtidal biotopes.

The invasive green algae Codium fragile ssp. tomentosoides, now found throughout Britain has been reported to occur in habitats dominated by Zostera marina (Gabary et al.,1997).  It was initially thought that Zostera out-competes Codium at high Zostera densities (Malinowski & Ramus, 1973).  But a study by Gabary et al. (2004) in Canada found that the invasive alga has morphological adaptations that allow it to compete with Zostera even in healthy eelgrass beds.  Codium fragile ssp. tomentosoides have a wide salinity tolerance 12 to 40 ppt and are thus a concern to biotopes in full as well as in reduced salinity.  However, direct ecological impacts remain unknown and no quantitative evidence is available to assess resistance at the benchmark.

Non-native invasive invertebrates: benthic macroinvertebrates can have a significant impact on seagrass beds, by either influencing abundance through seed herbivory (Fishman & Orth, 1996) or by influencing seed germination and seedling development by affecting vertical distribution of seeds.  Some species have a positive effect by burying seeds to shallow depths and thereby reducing seed predation and facilitating seed germination whilst other species bury seeds too deep to allow germination.  The invasive polychaete Marenzelleria viridis, a species naturally occurring on the east coast of North America but introduced Europe via transport in ballast waters, was recorded to directly impact seed banks of Zostera marina beds in its new territory (Delefosse & Kristensen, 2012). The study carried out on the island of Fyn, Denmark, determined that the impact of Marenzelleria viridis on seagrass beds depended on the abundance of worms within a bed.  Negative effects were only observed at high abundances (1600 individual per m²) causing seeds to be buried too deep to germinate.  However, the study by Delefosse & Kristensen (2012) is the only publication on the impact of this particular invasive species on seagrass beds, and more evidence is needed in order to determine the ecological implications of this introduced polychaete in UK waters.

Other invasive species could affect seagrass beds via indirect pathways.  For instance, the Atlantic oyster drill Urosalpinx cinerea, a small predatory sea snail is unlikely to have a direct effect on seagrass beds but by preying on mussels and other bivalves, the sea snail could be responsible for a drop in water clarity which in turn will affect Zostera species (see sections below on changes in suspended solids).  The invasive Pacific oyster Magallana gigas can also have negative effects. Oysters physically alter their environment by increasing habitat complexity and altering water flow and causing sulphide to accumulate in the sediment.  Sulphide is toxic to eelgrass and a decline in Zostera marina as a consequence of invasive oyster growth was observed in British Columbia, Canada (Kelly & Volpe, 2007).  The authors did not state the level of effect quantitatively and therefore the level of impact in terms of the resistance benchmarks used in this study is not clear.

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009). Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024). In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Michin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

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

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

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

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

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

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

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

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

However, some studies on Georges Bank, USA and Sandwich, Massachusetts observed colonies were able to survive partial covering by sand (Bullard et al., 2007; Valentine et al., 2007a). Gittenberger et al. (2015) reported that Didemnum vexillum was able to overgrow sandy bottom (cited Gittenberger, 2007). In northern Kent, Didemnum vexillum has been recorded covering London clay boulders on Whitestable Flats, West Beach, north Kent, covering tabulate sandstone boulders (0.5 to 2 m across) on the mid-shore and colonizing sediment mounds on the low shore characterized by larger areas of sand, mud and low-lying sediment at Reculver and Bishopstone, north Kent (Hitchin, 2012). It was also recorded from muddy substrata at that site. Hitchin (2012) noted that the site was exposed to enough waves and currents to cause sedimentation. However, Didemnum vexillum grew hanging from on the underside of sandstone boulders nestled on sediment, on consolidated sediment mounds and firm clays, hence burial may prevent colonization and its survival rather than sedimentation alone.

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

The Sandwich tide pools were subject to air exposure at low tide, daily changes in water depth and temperatures (Valentine et al., 2007a). Didemnum vexillum colonies are able to survive exposure to air at low tides for a short time (not exceeding two hours) during rapid colony growth in summer months July to September (Valentine et al., 2007a). However, parts of the large established colonies, which were artificially exposed to air for two to three hours in October, were observed desiccated or predated on by grazing periwinkles 30 days later, in the winter month November (Valentine et al., 2007a). They suggested that the invasive tunicates’ ability to tolerate exposure to air varies with the seasonal growth cycle. Didemnum vexillum also tolerated emersion in Kent, as colonies on the mid-shore at Reculver flourish and survive in air exposure for up to three hours per cycle during springs (Hitchin, 2012). Hitchin (2012) suggested the porous nature of the sandstone boulders the species colonized retained water. The Kent shore was sheltered but held water due to its shallow slope and flats, which may allow Didemnum sp. to survive in the low to mid-shore. There is evidence that Didemnum vexillum died when exposed to air for more than 6 hours (Laing et al., 2010).

Limited evidence was found on Didemnum vexillum populations established and growing on eelgrass, and what ecological impacts this may cause, but most reported evidence of other tunicates overgrowing eelgrass and macroalgae. Didemnum vexillum was first reported growing on the stalk and blade on live or dead eelgrass and on detached pieces of eelgrass Zostera marina, in Lake Tashmoo on Martha’s Vineayard, New England, which is described as a marine pond with an expansive eelgrass meadow and shellfish aquaculture site, and a seabed composed of a fine-grained sediment (Carman & Grunden, 2010; Carman et al., 2014). The colonies of Didemnum vexillum were mainly found growing on the bottom of a dingy for public landing (eastern shore) and on an aquaculture float (western shore). Here, pieces of eelgrass were growing and incorporated into the Didemnum vexillum colonies. Didemnum vexillum was not found near the north or south shore end of the pond. This suggested that the little artificial hard substrata available has allowed Didemnum to colonize the natural substratum that surrounded the artificial substrata (Carman & Grunden, 2010). Didemnum vexillum was not observed attached to the fine sediment (Carman & Grunden, 2010).

There is little direct evidence on how the invasive species may impact eelgrass beds. However, it was suggested that as Didemnum vexillum smothers bivalves and other sessile organisms, it can probably smother plants too (Carman & Grunden, 2010). Based on evidence from other invasive tunicates, it is also suggested that fouling by Didemnum vexillum and other invasive tunicates may block light, reducing photosynthesis and eelgrass shoot growth and survival (Wong & Vercaemer, 2012; Long & Grosholz, 2015; Tillin et al., 2020). This may also affect the other epifauna associated with eelgrass and eelgrass beds (Long & Grosholz, 2015).

In the field, Long & Groscholz (2015), found a negative effect of Didemnum vexillum overgrowth on eelgrass when it covers up to around 20% of the length of an individual eelgrass shoot. The eelgrass aboveground growth rate and biomass production was lower for eelgrass overgrown by Didemnum vexillum. Where Didemnum vexillum occurred on intertidal eelgrass the invasive species can grow in large clumps and ‘glue’ together multiple eelgrass shoots (Long & Grosholz, 2015). In mesocosm experiments, a significant decrease in the aboveground biomass in eelgrass was observed due to overgrowth by Didemnum vexillum, even though mesocosms had relatively lower cover of Didmenum vexillum compared to the field. However, there was no significant difference in the effect of overgrowth on eelgrass length production index (Long & Groscholz, 2015). Overall, the overgrowth did not have significant effects on biomass or morphology metrics in the experiment. However, Long & Groscholz (2015) suggested that more overgrowth on the terminal shoot, rather than on its rhizomes or other parts of the eelgrass may reveal trends in the growth rate.

Sensitivity assessment. Invasive species are affecting seagrass habitats around the UK with invasive flora having the greatest impact on seagrass beds so far recorded. However, there are extensive knowledge gaps on how invasive species influence the health of Zostera beds in UK waters. More research is needed in order to fully comprehend this pressure. Resistance is assessed as 'Low'. Return to ‘normal’ conditions is highly unlikely if an invasive species would come to dominate the biotope. Indeed recovery would only be possible if the majority of the INIS were removed (through either natural or unnatural process) to allow the re-establishment of other species. Therefore, actual resilience is assessed as ‘Low’ resulting in an overall ‘High’ sensitivity score.

The evidence presented shows Didemnum vexillum can overgrow eelgrass beds. In these biotopes, eelgrass provides suitable substrata and stabilise the sediment for successful colonization of Didemnum vexillum, which may otherwise not colonize sandy and muddy sediments. There is no direct evidence of Didemnum vexillum causing mortality amongst Zostera beds, however, fouling of Didemnum vexillum could potentially contribute to the population decline of Zostera, as it is likely to smother the eelgrass. Evidence has suggested that smothering of eelgrass causes negative effects on the population. For example, Den Hartog (1994) reported the growth of a dense blanket of Ulva radiata in Langstone Harbour in 1991 that resulted in the loss of 10ha of Zostera marina and Zostera noltei. Subsequently, by summer 1992, the Zostera sp. were absent, however this may have been exacerbated by grazing by Brent geese. The mechanisms responsible for seagrass decline under eutrophication are complex and involve direct and indirect effects relating to changes in water quality, smothering by macroalgal blooms (Den Hartog & Phillips, 2000), and competition for light and nutrients with epiphytic microalgae and with phytoplankton (Nienhuis, 1996). Short & Burdick (1996) found that excessive nitrogen loading stimulated the proliferation of algal competitors that caused shading and thereby stressing Zostera plants. Therefore, a resistance of 'Medium' (some mortality, <25%) is suggested as a precaution. Resilience is likely to be 'Very low' as Didemnum vexillum would need to be physically removed for recovery to occur. Hence, sensitivity to invasion by Didemnum is assessed as 'Medium'. 

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

Historic records show that seagrass species, in particular, Zostera marina, are highly susceptible to microbial pathogens.  During the 1930s, a so-called ‘wasting disease’ decimated the eelgrass Zostera marina in Europe and along the Atlantic Coast of North America with over 90% loss (Muehlstein, 1989).  Wasting disease resulted in black lesions on the leaf blades which potentially lead to loss of productivity, degradation of shoots and roots, eventually leading to the loss of large areas of seagrass (Den Hartog, 1987). Wasting disease is caused by infection with a marine slime mould-like protist, called Labyrinthula zosterae (Short et al., 1987; Muehlstein et al., 1991).  Recovery of seagrass beds after the epidemic has been extremely slow or more or less absent in some areas such as the Wadden Sea (Van der Heide et al., 2007).  The disease continues to affect Zostera marina in temperate regions with variable degrees of losses but not to the extent of an epidemic (Short et al., 1988).  The exact conditions responsible for an outbreak are still unknown but it has been shown that already weakened plants are more susceptible to infection (Tutin, 1938; Rasmussen, 1977) and that salinity plays a role the pathogen activity (Muehlstein et al., 1988).

Sensitivity assessment. Zostera marina is highly susceptible to microbial pathogens, which were in the past responsible for important reductions in seagrass populations. A sensitivity of ‘High’ has been recorded (‘Low’ resistance, ‘Low’ resilience)

Removal of target species

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

Evidence

Seagrass is not targeted by commercial fishery in the UK at present.  Seeds and shoots are, however, harvested for extensive transplantation project aimed at promoting seagrass populations in areas denuded by natural or anthropogenic causes. Divers are most commonly employed to remove material from the source population, an activity with a low overall impact on seagrass habitats.  However, in the USA, a mechanical seed harvesting technique was invented and put into practice (Orth & Marion, 2007).  The mechanised harvester is able to drastically increase the number of Zostera seed collected from a source population (1.68 million seeds in one day compared to 2.5 million seeds collected by divers in one year).  However, the large-scale removal of seeds, the productive output of seagrasses, can affect the integrity of the natural seagrass beds.  To date, no mechanical harvesting has been employed in the UK.  The ecological impact of seed collection by divers is low; the harvesting of Zostera in British waters has, therefore, a minimal effect on natural seagrass habitats.  The effect of the translocation of species is covered in the pressure 'genetic modification and translocation of indigenous species’.  

Harvesting of seagrasses as a craft material is a small but growing, industry.  The present legislation for the conservation of seagrasses will discourage the expansion of this industry (see Jackson et al. 2013 for a full list of the political framework for seagrass protection in the UK). Seagrass beds are not considered dependent on any of the organisms that may be targeted for direct removal e.g. oysters, clams and mussels.  However, an indirect effect of fisheries targeting bivalves is a change in the water clarity, crucial for the growth and development of Zostera species. Indeed bivalves have been shown to significantly contribute to the clearance of the water column which subsequently increases light penetration, facilitating the growth and reproduction of Zostera species (Wall et al., 2008).  Newell & Koch (2004) using modelling, predicted that when sediments were resuspended, the presence of even low numbers of oysters (25 g dry tissue weight/ m²) distributed uniformly throughout the domain, reduced suspended sediment concentrations by nearly an order of magnitude.  A healthy population of suspension-feeding bivalves thus improves habitat quality and promotes seagrass productivity by mitigating the effects of increased water turbidity in degraded, light-limited habitats (see, changes in suspended solids).  Bivalves also contribute pseudofaeces to fertilize seagrass sediments (Bradley & Heck Jr, 1999).

Seagrass plants may be directly removed or damaged by static or mobile gears that target other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of target species on this biotope.

Sensitivity assessment. Seagrass beds have no avoidance mechanisms to escape targeted harvesting of leaves, shoots and rhizomes. Resistance to this pressure is therefore assessed as ‘None’.  Studies of the effects of wildfowl grazing (see resilience and recovery above) suggest that recovery from the removal of target species will be rapid resulting in a 'Medium' resilience score. Added anthropogenic disturbance may, however, be detrimental to seagrass beds as soon as the ‘normal’ level caused by grazing birds is exceeded by human activities. Overall the sensitivity of this biotope is deemed ‘Medium’ to this pressure.

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

Filter-feeders such as mussels, clams and scallops are often associated with seagrass beds.  Fisheries targeting these bivalves employ methods such as trawling, dredging, digging and raking which all result in the non-targeted removal of seagrass species.  The direct physical effects of such fishing methods on seagrass are described in detail for the pressure ‘penetration and/or disturbance of the substratum’. Seagrass plants and the sedimentary habitat may be directly removed or damaged by static or mobile gears that are targeting other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of non-target species in this biotope.

Incidental removal of the key characterizing seagrass species and associated species would alter the character of the biotope. The biotope is characterized by the presence of beds of seagrass, these provide habitat structure and attachment surfaces for epiphytic species. These may also modify local habitats through changes in water flow and the trapping of sediments. The loss of the turf due to incidental removal as by-catch would, therefore, alter the character of the habitat and result in the loss of habitat structure and species richness. The ecological services such as primary and secondary production and habitat engineering provided by seagrass and the associated species would also be lost.

Sensitivity assessment. Incidental removal of seagrass as by-catch would be detrimental, altering the character of the biotope and removing the habitat structure, and could lead to reclassification of the biotope where extensive removal occurs.  Therefore, resistance is considered to be 'None', resilience 'Low' and sensitivity 'High'.