The Marine Life Information Network

Global warming (extreme)

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

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

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

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

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

Evidence

Zostera marina is widely distributed on sandy shores and estuaries throughout the northern hemisphere and is known to survive at temperatures of between 0 and 30°C (Zimmerman et al., 2017), although optimum conditions for growth are generally restricted to between 13 and 24°C (Lee et al., 2007). In the UK, sea surface temperatures are currently between 6 and 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, USA, shoots were found to stop growing at 20°C and start to die off at 25°C (Reusch et al., 2005). Numerous studies have reported physiological and biomass and growth rate declines when water temperatures exceed 25°C (Moreno-Marin et al., 2018; Hammer et al., 2018; Mochida et al., 2019; Kim et al., 2024). In addition, temperatures up to 25°C were shown to increase respiration rates in Zostera marina seeds, resulting in oxygen deficits that may hinder seed development and decrease sexual reproductive success (Broderson et al., 2024). When coupled with reduced light intensity and an increase in ammonium concentrations, the negative effect of 25°C water temperature on Zostera marina was shown to intensify (Moreno-Marin et al., 2018). Other combined stressors, such as increased turbidity and decreased pH, have also been observed to amplify negative effects of increased temperature (Perry et al., 2019).

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 Chesapeake Bay, two major Zostera marina die-off events were recorded following two‑week heatwaves in June of 2010 and 2015 (Shields et al., 2019). During these periods, daily mean temperatures reached 28.1 to 28.6°C, triggering rapid declines of 76% (2010) and 64% (2015) in Zostera marina cover, respectively. In response to each collapse, the more heat‑tolerant Ruppia maritima expanded its cover to become the dominant species in the following summer (2011 and 2016). This dominance was temporary after the 2010 die-off as Zostera marina recovered to pre‑die‑off levels within four years, demonstrating its competitive advantage under average temperature conditions (Shields et al., 2019). However, overall Zostera marina extent has declined in this area since 1991 by 54 to 64% (Lefcheck et al., 2017; Richardson et al., 2018; Hensel et al., 2023) and Ruppia maritima has increased by 171% (Hensel et al., 2023). This change in species dominance has been linked to rising temperature, given that Zostera marina begins to decline above 26.5°C, whereas Ruppia maritima benefits over 27.5°C (Richardson et al., 2018). Therefore, as temperatures continue to warm, the growth of Ruppia maritima may be favoured over Zostera marina in this area (Richardson et al., 2018). Zostera marina can form mixed stands with Ruppia maritima in some estuaries and lagoons in the UK, (Davison & Hughes, 1998), which may also be vulnerable to similar competition as temperature rise. .            

In North Carolina where Zostera marina occurs at its southern range limit, water temperatures have increased from 1981 to 2019 (Bartenfelder et al., 2022). The number of days exceeding the stressful threshold of 25°C has nearly doubled, and extreme events above 30°C have become more frequent (Bartenfelder et al., 2022). These climatic changes have coincided with reductions in Zostera marina extent, indicating that eelgrass at the southern range edge is reaching and exceeding its thermal tolerance (Bartenfelder et al., 2022). In the Western Atlantic off the coast of Virginia, a die-off event occurred in June 2015, after water temperatures exceeded 28°C for 13 consecutive days (compared to usual 1 to 5 days), causing shoot density to decline by 90% (Berger et al., 2020). By 2017, shoot density had only returned to half of that present in 2014 before the die off. Short‑term temperature increases linked to El Niño events, combined with two major marine heatwaves, produced anomalies up to 4°C along the Oregon coast and inside the Coos Estuary between 2014 and 2016. Within the estuary, this resulted in more than 100 days per year where temperatures were ≥1.5°C above normal, particularly in shallow areas. These prolonged warm anomalies led to a severe die‑off of Zostera marina. At one long‑term monitoring site, eelgrass shoot density fell from 78 shoots/m² in 2014 to just 5 shoots/m² (94% decline) in 2016, and by 2021, densities showed little sign of recovery (Jarrin et al., 2022).

In James Bay, Canada, a severe reduction in Zostera marina cover occurred between 1995 and 1999. There has been little recovery of eelgrass in this area, and in 2019, there remained a 60% decrease in shoot density and an 80% decrease in shoot biomass compared to pre-decline levels (Leblanc et al., 2023). The exact cause of this decline was not known but environmental factors such as early ice break up and increased water temperatures with extreme variability were thought to have contributed (Leblanc et al., 2023). In other areas of James Bay, hydroelectric development on the La Grande River greatly increased freshwater discharge into the bay, expanding the freshwater plume and driving coastal erosion and sediment stress on eelgrass habitats. When rapid climate change impacts emerged in the late 1990s including earlier sea‑ice breakup, warmer coastal waters, and marine heatwaves, the already stressed eelgrass ecosystem collapsed, losing >90% of Zostera marina biomass along the coast (Fink-Mercier et al., 2024). This subarctic population of Zostera marina was shown to have one of the lowest genetic diversities compared to population in the north-eastern Pacific and north-western Atlantic, potentially making it less resilient to global warming (Jeffery et al., 2024).

Increases in temperature can increase the prevalence and intensity of seagrass wasting disease (Groner et al., 2021; Aoki et al., 2022; Schneck et al., 2023; Graham et al., 2025). For example, Aoki et al. (2022) reported that wasting disease prevalence was three times higher in areas where high temperature anomalies occurred in summer, demonstrating that increasing global temperatures will put Zostera marina meadows at higher risk of the disease (Aoki et al., 2022).

The effect of increased temperature has also been observed to alter reproduction. 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. Conversely, Qin et al. (2020) reported a negative relationship between increasing maximum sea surface temperature and marine heatwave frequency on the flowering frequency, reproductive shoot density, and reproductive energy allocation on Zostera marina, which declined significantly from 2011 to 2018, coinciding with the temperature increases. Sawall et al. (2021) found that Zostera marina subjected to warmer winter and spring temperatures (+3.6°C) flowered approx. 1.5 months earlier in the spring and suffered 40% mortality compared with high survival in ambient conditions. This response was likely a stress response to depleted energy reserves rather than a direct temperature effect. It is not known whether earlier flowering in this species may help or hinder recovery, as while it could increase genetic diversity, earlier seed production could also result in fewer, less viable seeds (Sawall et al., 2021).

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 to 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 seven 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 (e.g., 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 to 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 and 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 and 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. 

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 and 30°C (Zimmerman et al., 2017), although optimum conditions for growth are generally restricted to between 13 and 24°C (Lee et al., 2007). In the UK, sea surface temperatures are currently between 6 and 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, USA, shoots were found to stop growing at 20°C and start to die off at 25°C (Reusch et al., 2005). Numerous studies have reported physiological and biomass and growth rate declines when water temperatures exceed 25°C (Moreno-Marin et al., 2018; Hammer et al., 2018; Mochida et al., 2019; Kim et al., 2024). In addition, temperatures up to 25°C were shown to increase respiration rates in Zostera marina seeds, resulting in oxygen deficits that may hinder seed development and decrease sexual reproductive success (Broderson et al., 2024). When coupled with reduced light intensity and an increase in ammonium concentrations, the negative effect of 25°C water temperature on Zostera marina was shown to intensify (Moreno-Marin et al., 2018). Other combined stressors, such as increased turbidity and decreased pH, have also been observed to amplify negative effects of increased temperature (Perry et al., 2019).

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 Chesapeake Bay, two major Zostera marina die-off events were recorded following two‑week heatwaves in June of 2010 and 2015 (Shields et al., 2019). During these periods, daily mean temperatures reached 28.1 to 28.6°C, triggering rapid declines of 76% (2010) and 64% (2015) in Zostera marina cover, respectively. In response to each collapse, the more heat‑tolerant Ruppia maritima expanded its cover to become the dominant species in the following summer (2011 and 2016). This dominance was temporary after the 2010 die-off as Zostera marina recovered to pre‑die‑off levels within four years, demonstrating its competitive advantage under average temperature conditions (Shields et al., 2019). However, overall Zostera marina extent has declined in this area since 1991 by 54 to 64% (Lefcheck et al., 2017; Richardson et al., 2018; Hensel et al., 2023) and Ruppia maritima has increased by 171% (Hensel et al., 2023). This change in species dominance has been linked to rising temperature, given that Zostera marina begins to decline above 26.5°C, whereas Ruppia maritima benefits over 27.5°C (Richardson et al., 2018). Therefore, as temperatures continue to warm, the growth of Ruppia maritima may be favoured over Zostera marina in this area (Richardson et al., 2018). Zostera marina can form mixed stands with Ruppia maritima in some estuaries and lagoons in the UK, (Davison & Hughes, 1998), which may also be vulnerable to similar competition as temperature rise. .            

In North Carolina where Zostera marina occurs at its southern range limit, water temperatures have increased from 1981 to 2019 (Bartenfelder et al., 2022). The number of days exceeding the stressful threshold of 25°C has nearly doubled, and extreme events above 30°C have become more frequent (Bartenfelder et al., 2022). These climatic changes have coincided with reductions in Zostera marina extent, indicating that eelgrass at the southern range edge is reaching and exceeding its thermal tolerance (Bartenfelder et al., 2022). In the Western Atlantic off the coast of Virginia, a die-off event occurred in June 2015, after water temperatures exceeded 28°C for 13 consecutive days (compared to usual 1 to 5 days), causing shoot density to decline by 90% (Berger et al., 2020). By 2017, shoot density had only returned to half of that present in 2014 before the die off. Short‑term temperature increases linked to El Niño events, combined with two major marine heatwaves, produced anomalies up to 4°C along the Oregon coast and inside the Coos Estuary between 2014 and 2016. Within the estuary, this resulted in more than 100 days per year where temperatures were ≥1.5°C above normal, particularly in shallow areas. These prolonged warm anomalies led to a severe die‑off of Zostera marina. At one long‑term monitoring site, eelgrass shoot density fell from 78 shoots/m² in 2014 to just 5 shoots/m² (94% decline) in 2016, and by 2021, densities showed little sign of recovery (Jarrin et al., 2022).

In James Bay, Canada, a severe reduction in Zostera marina cover occurred between 1995 and 1999. There has been little recovery of eelgrass in this area, and in 2019, there remained a 60% decrease in shoot density and an 80% decrease in shoot biomass compared to pre-decline levels (Leblanc et al., 2023). The exact cause of this decline was not known but environmental factors such as early ice break up and increased water temperatures with extreme variability were thought to have contributed (Leblanc et al., 2023). In other areas of James Bay, hydroelectric development on the La Grande River greatly increased freshwater discharge into the bay, expanding the freshwater plume and driving coastal erosion and sediment stress on eelgrass habitats. When rapid climate change impacts emerged in the late 1990s including earlier sea‑ice breakup, warmer coastal waters, and marine heatwaves, the already stressed eelgrass ecosystem collapsed, losing >90% of Zostera marina biomass along the coast (Fink-Mercier et al., 2024). This subarctic population of Zostera marina was shown to have one of the lowest genetic diversities compared to population in the north-eastern Pacific and north-western Atlantic, potentially making it less resilient to global warming (Jeffery et al., 2024).

Increases in temperature can increase the prevalence and intensity of seagrass wasting disease (Groner et al., 2021; Aoki et al., 2022; Schneck et al., 2023; Graham et al., 2025). For example, Aoki et al. (2022) reported that wasting disease prevalence was three times higher in areas where high temperature anomalies occurred in summer, demonstrating that increasing global temperatures will put Zostera marina meadows at higher risk of the disease (Aoki et al., 2022).

The effect of increased temperature has also been observed to alter reproduction. 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. Conversely, Qin et al. (2020) reported a negative relationship between increasing maximum sea surface temperature and marine heatwave frequency on the flowering frequency, reproductive shoot density, and reproductive energy allocation on Zostera marina, which declined significantly from 2011 to 2018, coinciding with the temperature increases. Sawall et al. (2021) found that Zostera marina subjected to warmer winter and spring temperatures (+3.6°C) flowered approx. 1.5 months earlier in the spring and suffered 40% mortality compared with high survival in ambient conditions. This response was likely a stress response to depleted energy reserves rather than a direct temperature effect. It is not known whether earlier flowering in this species may help or hinder recovery, as while it could increase genetic diversity, earlier seed production could also result in fewer, less viable seeds (Sawall et al., 2021).

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 to 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 seven 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 (e.g., 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 to 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 and 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 and 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. 

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 and 30°C (Zimmerman et al., 2017), although optimum conditions for growth are generally restricted to between 13 and 24°C (Lee et al., 2007). In the UK, sea surface temperatures are currently between 6 and 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, USA, shoots were found to stop growing at 20°C and start to die off at 25°C (Reusch et al., 2005). Numerous studies have reported physiological and biomass and growth rate declines when water temperatures exceed 25°C (Moreno-Marin et al., 2018; Hammer et al., 2018; Mochida et al., 2019; Kim et al., 2024). In addition, temperatures up to 25°C were shown to increase respiration rates in Zostera marina seeds, resulting in oxygen deficits that may hinder seed development and decrease sexual reproductive success (Broderson et al., 2024). When coupled with reduced light intensity and an increase in ammonium concentrations, the negative effect of 25°C water temperature on Zostera marina was shown to intensify (Moreno-Marin et al., 2018). Other combined stressors, such as increased turbidity and decreased pH, have also been observed to amplify negative effects of increased temperature (Perry et al., 2019).

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 Chesapeake Bay, two major Zostera marina die-off events were recorded following two‑week heatwaves in June of 2010 and 2015 (Shields et al., 2019). During these periods, daily mean temperatures reached 28.1 to 28.6°C, triggering rapid declines of 76% (2010) and 64% (2015) in Zostera marina cover, respectively. In response to each collapse, the more heat‑tolerant Ruppia maritima expanded its cover to become the dominant species in the following summer (2011 and 2016). This dominance was temporary after the 2010 die-off as Zostera marina recovered to pre‑die‑off levels within four years, demonstrating its competitive advantage under average temperature conditions (Shields et al., 2019). However, overall Zostera marina extent has declined in this area since 1991 by 54 to 64% (Lefcheck et al., 2017; Richardson et al., 2018; Hensel et al., 2023) and Ruppia maritima has increased by 171% (Hensel et al., 2023). This change in species dominance has been linked to rising temperature, given that Zostera marina begins to decline above 26.5°C, whereas Ruppia maritima benefits over 27.5°C (Richardson et al., 2018). Therefore, as temperatures continue to warm, the growth of Ruppia maritima may be favoured over Zostera marina in this area (Richardson et al., 2018). Zostera marina can form mixed stands with Ruppia maritima in some estuaries and lagoons in the UK, (Davison & Hughes, 1998), which may also be vulnerable to similar competition as temperature rise. .            

In North Carolina where Zostera marina occurs at its southern range limit, water temperatures have increased from 1981 to 2019 (Bartenfelder et al., 2022). The number of days exceeding the stressful threshold of 25°C has nearly doubled, and extreme events above 30°C have become more frequent (Bartenfelder et al., 2022). These climatic changes have coincided with reductions in Zostera marina extent, indicating that eelgrass at the southern range edge is reaching and exceeding its thermal tolerance (Bartenfelder et al., 2022). In the Western Atlantic off the coast of Virginia, a die-off event occurred in June 2015, after water temperatures exceeded 28°C for 13 consecutive days (compared to usual 1 to 5 days), causing shoot density to decline by 90% (Berger et al., 2020). By 2017, shoot density had only returned to half of that present in 2014 before the die off. Short‑term temperature increases linked to El Niño events, combined with two major marine heatwaves, produced anomalies up to 4°C along the Oregon coast and inside the Coos Estuary between 2014 and 2016. Within the estuary, this resulted in more than 100 days per year where temperatures were ≥1.5°C above normal, particularly in shallow areas. These prolonged warm anomalies led to a severe die‑off of Zostera marina. At one long‑term monitoring site, eelgrass shoot density fell from 78 shoots/m² in 2014 to just 5 shoots/m² (94% decline) in 2016, and by 2021, densities showed little sign of recovery (Jarrin et al., 2022).

In James Bay, Canada, a severe reduction in Zostera marina cover occurred between 1995 and 1999. There has been little recovery of eelgrass in this area, and in 2019, there remained a 60% decrease in shoot density and an 80% decrease in shoot biomass compared to pre-decline levels (Leblanc et al., 2023). The exact cause of this decline was not known but environmental factors such as early ice break up and increased water temperatures with extreme variability were thought to have contributed (Leblanc et al., 2023). In other areas of James Bay, hydroelectric development on the La Grande River greatly increased freshwater discharge into the bay, expanding the freshwater plume and driving coastal erosion and sediment stress on eelgrass habitats. When rapid climate change impacts emerged in the late 1990s including earlier sea‑ice breakup, warmer coastal waters, and marine heatwaves, the already stressed eelgrass ecosystem collapsed, losing >90% of Zostera marina biomass along the coast (Fink-Mercier et al., 2024). This subarctic population of Zostera marina was shown to have one of the lowest genetic diversities compared to population in the north-eastern Pacific and north-western Atlantic, potentially making it less resilient to global warming (Jeffery et al., 2024).

Increases in temperature can increase the prevalence and intensity of seagrass wasting disease (Groner et al., 2021; Aoki et al., 2022; Schneck et al., 2023; Graham et al., 2025). For example, Aoki et al. (2022) reported that wasting disease prevalence was three times higher in areas where high temperature anomalies occurred in summer, demonstrating that increasing global temperatures will put Zostera marina meadows at higher risk of the disease (Aoki et al., 2022).

The effect of increased temperature has also been observed to alter reproduction. 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. Conversely, Qin et al. (2020) reported a negative relationship between increasing maximum sea surface temperature and marine heatwave frequency on the flowering frequency, reproductive shoot density, and reproductive energy allocation on Zostera marina, which declined significantly from 2011 to 2018, coinciding with the temperature increases. Sawall et al. (2021) found that Zostera marina subjected to warmer winter and spring temperatures (+3.6°C) flowered approx. 1.5 months earlier in the spring and suffered 40% mortality compared with high survival in ambient conditions. This response was likely a stress response to depleted energy reserves rather than a direct temperature effect. It is not known whether earlier flowering in this species may help or hinder recovery, as while it could increase genetic diversity, earlier seed production could also result in fewer, less viable seeds (Sawall et al., 2021).

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 to 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 seven 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 (e.g., 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 to 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 and 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 and 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 (Marine heatwave pressure definitions).

Evidence

Marine heatwaves due to increased air-sea heat flux are predicted to occur more frequently, last for longer and at increased intensity by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Globally, seagrass populations have declined after severe marine heatwaves (Smith et al., 2024), and it has been suggested that marine heatwaves pose a more immediate and severe threat to temperate seagrass survival than gradual increases in sea surface temperature (Kim et al., 2024). 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 to 27°C, that resulted in mortality of around 66% 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 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 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.

Zostera marina from different latitudes has exhibited different vulnerabilities to heating events, with those found in warmer and more variable conditions showing higher resilience (Nguyen et al., 2021). However, pre‑exposure to marine heatwaves (18°C or 22°C for 28 days) did not increase thermal tolerance in Zostera marina or Zostera noltii from the Western English Channel (King et al., 2024). Although heatwave exposure weakened their physiological performance, Zostera spp. did not become more vulnerable to subsequent heat stress, unlike some kelp species (King et al., 2024). 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.

Marine heatwaves have been associated with declines in Zostera marina extent globally. In the western Atlantic, off the coast of Virginia, USA, in July 2012, temperatures reached over 28°C for 71% of the month (Aoki et al., 2020). Following this heatwave, shoot density in the Zostera marina meadow declined by around 75 to 85% among sites (Aoki et al., 2020). By 2017, some sites within the meadow had recovered to equall or increase in density compared to pre-heatwave levels (Aoki et al., 2020). However, at other sites, Zostera marina did not recover to previous levels, reaching only 40 to 45% recovery by 2017. The most resilient plots were at an intermediate depth (between 1.0 and 1.3 m), suggesting that shallow and deeper meadows did not recover as well (Aoki et al., 2020). In 2015, another marine heatwave occurred in Virginia, with sea surface temperatures reaching between 28°C and 30°C for 10 days in June 2015 (Berger et al., 2024). There was a 90% decrease in Zostera marina shoots in July 2015 (after the heatwave) compared July 2014. Partial recovery of between 68 and 84% shoot density occurred within sites by July 2018 (Berger et al., 2024).

Short‑term temperature increases linked to El Niño events, combined with two major marine heatwaves, produced anomalies up to 4°C along the Oregon coast and inside the Coos Estuary between 2014 and 2016. Within the estuary, this resulted in more than 100 days per year where temperatures were ≥1.5°C above normal, particularly in shallow areas (Jarrin et al., 2022). These prolonged warm anomalies led to a severe die‑off of Zostera marina. At one long‑term monitoring site, eelgrass shoot density fell from 78 shoots/m² in 2014 to just 5 shoots/m² in 2016 (94% decline), and by 2021, densities showed little sign of recovery (Jarrin et al., 2022). The warming effect of these heatwaves spread into other inland estuaries causing ‘estuarine heatwaves’. In some sites, these estuarine heatwaves resulted in significant decreases in biomass from 2015 to 2016 by up to 65%, with continuing declines or little to no recovery recorded in 2019 (Magel et al., 2022). This was especially true for shallower sites, suggesting that deeper sites may be more resilient during such warming events (Magel et al., 2022).

In Chesapeake Bay, USA, two major Zostera marina die-off events were recorded following two‑week heatwaves in June of 2010 and 2015 (Shields et al., 2019). During these periods, daily mean temperatures reached 28.1 to 28.6°C, triggering rapid declines of 76% (2010) and 64% (2015) in Zostera marina cover. In response to each collapse, the more heat‑tolerant angiosperm Ruppia maritima expanded its cover to become the dominant species in the following summer (2011 and 2016). This dominance was temporary after the 2010 die-off as Zostera marina recovered to pre‑die‑off levels within four years, demonstrating its competitive advantage under average temperature conditions (Shields et al., 2019). However, overall Zostera marina extent has declined in this area since 1991 by 54 to 64% (Lefcheck et al., 2017; Richardson et al., 2018; Hensel et al., 2023) and Ruppia maritima has increased by 171% (Hensel et al., 2023). This change in species dominance has been linked to rising temperature, given that Zostera marina begins to decline above approx. 26.5°C, whereas Ruppia maritima benefits over approx. 27.5°C (Richardson et al., 2018). Therefore, as temperatures continue to warm, the growth of Ruppia maritima may be favoured over Zostera marina in this area (Richardson et al., 2018). Zostera marina can form mixed stands with Ruppia maritima in some estuaries and lagoons in the UK, (Davison & Hughes, 1998), which may also be vulnerable to similar competition as temperature rise.

Saha et al. (2020) investigated the effects of simulated marine heatwaves reaching maximum temperatures of 25.2°C on Zostera marina. The experiment compared one summer heatwave (temperature increase of 1.7°C/day for 3 days, peaking at +5.2°C above seasonal averages for 4 days, then cooling over 2 days) with three heatwaves, where the first two involved increases of 1.2°C/day for 3 days, peaking at +3.6°C for 4 days, followed by cooling, and the final heatwave mirroring theone‑heatwave treatment. Photosynthetic rates, respiration, and wasting disease prevalence remained unchanged across treatments within one week after the final heatwave. However, growth rate was reduced by approx. 40% in the three‑heatwave treatment only, suggesting that Zostera marina is generally able to withstand single, short-term heatwave events, however, repeated heatwaves prove more detrimental. In addition, exposure to earlier heatwaves did not confer increased resilience to future heatwaves, indicating that pre-exposure to warming does not buffer eelgrass against subsequent thermal stress (Saha et al., 2020).

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. Marine heatwaves have also been associated with an increase in the prevalence of eelgrass wasting disease, which could suppress a meadow’s ability to recover after a warming event (Groner et al., 2021).

Furthermore, in disturbed meadows, a decrease in sexual reproduction was observed, with beds maintained through vegetative spread, which will lead to decreased genetic diversity and therefore resistance (Potouroglou et al., 2014). The effect of increased temperature has also been observed to alter reproduction. Qin et al. (2020) reported a negative relationship between increasing maximum sea surface temperature and marine heatwave frequency on the flowering frequency, reproductive shoot density, and reproductive energy allocation on Zostera marina, which significantly declined from 2011 to 2018, coinciding with the temperature increases. Sawall et al. (2021) found that Zostera marina subjected to warmer winter and spring temperatures (+3.6°C) flowered approx. 1.5 months earlier in the spring and suffered 40% mortality compared with high survival in ambient conditions. This response was likely a stress response to depleted energy reserves rather than a direct temperature effect. It is not known whether earlier flowering in this species may help or hinder recovery, as while it could increase genetic diversity, earlier seed production could also result in fewer, less viable seeds (Sawall et al., 2021). 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 two to five years from remaining rhizomes, seed set, and/or flowering plants. Under the high emission scenario it is likely that marine heatwaves will lead to an increase in seagrass decline through earlier summer die-offs. In addition, where increased temperatures have allowed more temperature tolerant species, such as Ruppia maritima, to colonize previous Zostera habiat, recovery will be prolonged. 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. 

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). Globally, seagrass populations have declined after severe marine heatwaves (Smith et al., 2024), and it has been suggested that marine heatwaves pose a more immediate and severe threat to temperate seagrass survival than gradual increases in sea surface temperature (Kim et al., 2024). 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 to 27°C, that resulted in mortality of around 66% 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 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 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.

Zostera marina from different latitudes has exhibited different vulnerabilities to heating events, with those found in warmer and more variable conditions showing higher resilience (Nguyen et al., 2021). However, pre‑exposure to marine heatwaves (18°C or 22°C for 28 days) did not increase thermal tolerance in Zostera marina or Zostera noltii from the Western English Channel (King et al., 2024). Although heatwave exposure weakened their physiological performance, Zostera spp. did not become more vulnerable to subsequent heat stress, unlike some kelp species (King et al., 2024). 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.

Marine heatwaves have been associated with declines in Zostera marina extent globally. In the western Atlantic, off the coast of Virginia, USA, in July 2012, temperatures reached over 28°C for 71% of the month (Aoki et al., 2020). Following this heatwave, shoot density in the Zostera marina meadow declined by around 75 to 85% among sites (Aoki et al., 2020). By 2017, some sites within the meadow had recovered to equall or increase in density compared to pre-heatwave levels (Aoki et al., 2020). However, at other sites, Zostera marina did not recover to previous levels, reaching only 40 to 45% recovery by 2017. The most resilient plots were at an intermediate depth (between 1.0 and 1.3 m), suggesting that shallow and deeper meadows did not recover as well (Aoki et al., 2020). In 2015, another marine heatwave occurred in Virginia, with sea surface temperatures reaching between 28°C and 30°C for 10 days in June 2015 (Berger et al., 2024). There was a 90% decrease in Zostera marina shoots in July 2015 (after the heatwave) compared July 2014. Partial recovery of between 68 and 84% shoot density occurred within sites by July 2018 (Berger et al., 2024).

Short‑term temperature increases linked to El Niño events, combined with two major marine heatwaves, produced anomalies up to 4°C along the Oregon coast and inside the Coos Estuary between 2014 and 2016. Within the estuary, this resulted in more than 100 days per year where temperatures were ≥1.5°C above normal, particularly in shallow areas (Jarrin et al., 2022). These prolonged warm anomalies led to a severe die‑off of Zostera marina. At one long‑term monitoring site, eelgrass shoot density fell from 78 shoots/m² in 2014 to just 5 shoots/m² in 2016 (94% decline), and by 2021, densities showed little sign of recovery (Jarrin et al., 2022). The warming effect of these heatwaves spread into other inland estuaries causing ‘estuarine heatwaves’. In some sites, these estuarine heatwaves resulted in significant decreases in biomass from 2015 to 2016 by up to 65%, with continuing declines or little to no recovery recorded in 2019 (Magel et al., 2022). This was especially true for shallower sites, suggesting that deeper sites may be more resilient during such warming events (Magel et al., 2022).

In Chesapeake Bay, USA, two major Zostera marina die-off events were recorded following two‑week heatwaves in June of 2010 and 2015 (Shields et al., 2019). During these periods, daily mean temperatures reached 28.1 to 28.6°C, triggering rapid declines of 76% (2010) and 64% (2015) in Zostera marina cover. In response to each collapse, the more heat‑tolerant angiosperm Ruppia maritima expanded its cover to become the dominant species in the following summer (2011 and 2016). This dominance was temporary after the 2010 die-off as Zostera marina recovered to pre‑die‑off levels within four years, demonstrating its competitive advantage under average temperature conditions (Shields et al., 2019). However, overall Zostera marina extent has declined in this area since 1991 by 54 to 64% (Lefcheck et al., 2017; Richardson et al., 2018; Hensel et al., 2023) and Ruppia maritima has increased by 171% (Hensel et al., 2023). This change in species dominance has been linked to rising temperature, given that Zostera marina begins to decline above approx. 26.5°C, whereas Ruppia maritima benefits over approx. 27.5°C (Richardson et al., 2018). Therefore, as temperatures continue to warm, the growth of Ruppia maritima may be favoured over Zostera marina in this area (Richardson et al., 2018). Zostera marina can form mixed stands with Ruppia maritima in some estuaries and lagoons in the UK, (Davison & Hughes, 1998), which may also be vulnerable to similar competition as temperature rise.

Saha et al. (2020) investigated the effects of simulated marine heatwaves reaching maximum temperatures of 25.2°C on Zostera marina. The experiment compared one summer heatwave (temperature increase of 1.7°C/day for 3 days, peaking at +5.2°C above seasonal averages for 4 days, then cooling over 2 days) with three heatwaves, where the first two involved increases of 1.2°C/day for 3 days, peaking at +3.6°C for 4 days, followed by cooling, and the final heatwave mirroring theone‑heatwave treatment. Photosynthetic rates, respiration, and wasting disease prevalence remained unchanged across treatments within one week after the final heatwave. However, growth rate was reduced by approx. 40% in the three‑heatwave treatment only, suggesting that Zostera marina is generally able to withstand single, short-term heatwave events, however, repeated heatwaves prove more detrimental. In addition, exposure to earlier heatwaves did not confer increased resilience to future heatwaves, indicating that pre-exposure to warming does not buffer eelgrass against subsequent thermal stress (Saha et al., 2020).

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. Marine heatwaves have also been associated with an increase in the prevalence of eelgrass wasting disease, which could suppress a meadow’s ability to recover after a warming event (Groner et al., 2021).

Furthermore, in disturbed meadows, a decrease in sexual reproduction was observed, with beds maintained through vegetative spread, which will lead to decreased genetic diversity and therefore resistance (Potouroglou et al., 2014). The effect of increased temperature has also been observed to alter reproduction. Qin et al. (2020) reported a negative relationship between increasing maximum sea surface temperature and marine heatwave frequency on the flowering frequency, reproductive shoot density, and reproductive energy allocation on Zostera marina, which significantly declined from 2011 to 2018, coinciding with the temperature increases. Sawall et al. (2021) found that Zostera marina subjected to warmer winter and spring temperatures (+3.6°C) flowered approx. 1.5 months earlier in the spring and suffered 40% mortality compared with high survival in ambient conditions. This response was likely a stress response to depleted energy reserves rather than a direct temperature effect. It is not known whether earlier flowering in this species may help or hinder recovery, as while it could increase genetic diversity, earlier seed production could also result in fewer, less viable seeds (Sawall et al., 2021). 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 two to five years from remaining rhizomes, seed set, and/or flowering plants. Under the high emission scenario it is likely that marine heatwaves will lead to an increase in seagrass decline through earlier summer die-offs. In addition, where increased temperatures have allowed more temperature tolerant species, such as Ruppia maritima, to colonize previous Zostera habiat, recovery will be prolonged. 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 (Ocean acidification pressure definitions).

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005). 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 a corresponding 35% increase in H+ ions with no coastal aragonite undersaturation and the aragonite saturation horizon in the NE Atlantic, off the continental shelf, at a depth of 800 m by the end of this century, 2081-2100. 

Evidence

Increasing levels of CO₂ in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005). 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 sea-level by the end of this century (2018-2100) (Sea-level rise pressure definitions).

Evidence

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). 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 sea-level by the end of this century (2018-2100). 

Evidence

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). 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 by the end of this century (2081-2100).

Evidence

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). 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 (Temperature change pressure definition).

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 also encourage the growth of epiphytes, increasing the burden upon seagrass beds and making them more susceptible to disease (Rasmussen, 1977). However, different ecotypes of Zostera marina may be more resilient than others, likely due to genetic and phenotypical differences among populations (Breiter et al., 2024).

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). Numerous studies have reported physiological, biomass, and growth rate declines when water temperatures exceed 25°C (Moreno-Marin et al., 2018; Hammer et al., 2018; Mochida et al., 2019; Kim et al., 2024). 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%). In addition, temperatures up to 25°C were shown to increase respiration rates in Zostera marina seeds, resulting in oxygen deficits that may hinder seed development and decrease sexual reproductive success (Broderson et al., 2024). When coupled with reduced light intensity and an increase in ammonium concentrations, the negative effect of 25°C water temperature on Zostera marina intensified (Moreno-Marin et al., 2018). Other combined stressors, such as increased turbidity and decreased pH, have also been observed to amplify negative effects of increased temperature (Perry et al., 2019). Reduced light levels have also been shown to make Zostera marina more susceptible to the negative effect of higher temperatures. At 24°C under high light (180 μmol photons m²/s), all Zostera marina survived, however, 30% mortality occurred when under low light (60 μmol photons m²/s) (Beca-Carretero et al., 2018, 2021).

Declines in Zostera marina have been observed globally, often attributed to increases in temperature and marine heatwave frequency. High temperatures resulted in a 12-fold increase in mortality of Zostera marina plants (Nejrup & Pedersen, 2008). Moore et al. (2014) found that short-term exposure to a rapid increase of 4 to 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.

In North Carolina where Zostera marina occurs at its southern range limit, water temperatures have increased from 1981 to 2019. The number of days exceeding the stressful threshold of 25°C has nearly doubled, and extreme events above 30°C have become more frequent (Bartenfelder et al., 2022). These climatic changes have coincided with declines in Zostera marina extent, indicating that eelgrass at the southern range edge is reaching and exceeding its thermal tolerance (Bartenfelder et al., 2022). In the Western Atlantic off the coast of Virginia, a die-off event occurred in June 2015, after water temperatures exceeded 28°C for 13 consecutive days (compared to the usual 1 to 5 days), causing shoot density to decline by 90% (Berger et al., 2020). By 2017, shoot density had only returned to half of that present in 2014 before the die off. Short‑term temperature increases linked to El Niño events, combined with two major marine heatwaves, produced anomalies up to 4°C along the Oregon coast and inside the Coos Estuary between 2014 and 2016. This resulted in more than 100 days per year where temperatures were ≥ 1.5°C above normal within the estuary,, particularly in shallow areas. These prolonged warm anomalies led to a severe die‑off of Zostera marina. At one long‑term monitoring site, eelgrass shoot density fell from 78 shoots/m² in 2014 to just 5 shoots/m² in 2016 (94% decline), and by 2021, densities showed little sign of recovery (Jarrin et al., 2022).

In Chesapeake Bay, USA, two major Zostera marina die-off events were recorded following two‑week heatwaves in June of 2010 and 2015 (Shields et al., 2019). During these periods, daily mean temperatures reached 28.1 to 28.6°C, triggering rapid declines of 76% (2010) and 64% (2015) in Zostera marina cover. In response to each collapse, the more heat‑tolerant Ruppia maritima expanded its cover to become the dominant species in the following summer (2011 and 2016). This dominance was temporary after the 2010 die-off as Zostera marina recovered to pre‑die‑off levels within four years, demonstrating its competitive advantage under average temperature conditions (Shields et al., 2019). However, overall, Zostera marina extent has declined in this area since 1991 by 54 to 64% (Lefcheck et al., 2017; Richardson et al., 2018; Hensel et al., 2023) and Ruppia maritima has increased by 171% (Hensel et al., 2023). This change in species dominance has been linked to rising temperature, given that Zostera marina begins to decline above approx. 26.5°C, whereas Ruppia maritima benefits over approx. 27.5°C (Richardson et al., 2018). Therefore, as temperatures continue to warm, the growth of Ruppia maritima may be favoured over Zostera marina in this area (Richardson et al., 2018). Zostera marina can form mixed stands with Ruppia maritima in some estuaries and lagoons in the UK, (Davison & Hughes, 1998), which may also be vulnerable to similar competition as temperature rise.

Increases in temperature can increase the prevalence and intensity of seagrass wasting disease (Groner et al., 2021; Aoki et al., 2022; Schneck et al., 2023; Graham et al., 2025). For example, Aoki et al. (2022) reported that wasting disease prevalence was three times higher in areas where high temperature anomalies occurred in summer, demonstrating that increasing global temperatures may put Zostera marina meadows at higher risk of the disease (Aoki et al., 2022). A 60% decline in Zostera marina shoot density was observed in a meadow in the San Juan Islands (northeast Pacific) between 2013 and 2015. Although the primary driver of this initial decline was not identified, the subsequent 2015 to 2016 marine heatwave was associated with a sharp increase in the prevalence of eelgrass wasting disease, which likely suppressed the meadow’s ability to recover. As a result, shoot densities remained at similarly low levels through 2017 (Groner et al., 2021).

The effect of increased temperature has also been observed to alter reproductive mechanisms. Orth & Moore (1983) reported that the majority (68%) of Zostera marina seeds germinated in the winter months between 0 to 10°C, and that germination was most rapid between 5 to 10°C but virtually no germination was observed when temperatures were above 20°C, in Chesapeake Bay, USA. Qin et al. (2020) reported a negative relationship between increasing maximum sea surface temperature and marine heatwave frequency on the flowering frequency, reproductive shoot density and reproductive energy allocation on Zostera marina, which significantly declined from 2011 to 2018, coinciding with the temperature increases. Sawall et al. (2021) found that Zostera marina subjected to warmer winter and spring temperatures (+3.6°C) flowered approx. 1.5 months earlier in the spring and suffered 40% mortality compared with high survival in ambient conditions. This response was likely a stress response to depleted energy reserves rather than a direct temperature effect. It is not known whether earlier flowering in this species may help or hinder recovery, as while it could increase genetic diversity, earlier seed production could also result in fewer, less viable seeds (Sawall et al., 2021).

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). Areas where increased temperatures have persisted for up to or over two weeks have seen Zostera marina declines of 64 to 94%, particularly when temperatures exceeded 28°C (Shields et al., 2019; Berger et al., 2020; Jarrin et al., 2022). Recovery from these die-offs can range from little or no recovery after five years (e.g., Jarrin et al., 2022) to full recovery after four years (e.g., Shields et al., 2019). Other stressors that can result from elevated temperature, such as increased disease prevalence or reduced light, may hamper recovery of Zostera marina. 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, depending on the maximum temperature reached. In addition, a longer-term or persistent increase in temperature may reduce germination rates and hence reduce recruitment and resilience (Jackson, pers comm., 2019). At the pressure benchmark, resistance is assessed as ‘Medium’ . Recovery will likely be fairly rapid (2 to 10 years) 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 (Temperature change pressure definition).

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. A laboratory experiment that subjected Zostera marina to a simulated marine heatwave (a sudden increase from 20 to 27°C for seven to eight weeks) and simulated upwelling events (a decrease from 27 to 20°C for seven weeks) showed contrasting responses. A sudden and prolonged increase in temperature caused declines in shoot density, biomass and leaf productivity from 40 to 80%. In contrast, a sudden decrease in temperature maintained or slightly increased Zostera marina growth (Kim et al., 2020). These findings suggest that Zostera marina is more vulnerable to high temperature anomalies than to cooling events.

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

Evidence

Around the coast of the USA, Zostera marina has been recorded in salinities up to 51.2 ppt (Lee et al., 2026). However, hypersaline conditions can affect seagrass performance as changes in salinity may increase the energy requirements due to demanding osmotic adjustments (Touchette, 2007). Zostera marina possesses several physiological adaptations that allow it to regulate ions, maintain water balance, and cope with moderate salinity fluctuations, however, it is not among the most hypersaline‑tolerant seagrass species (Sandoval-Gil et al., 2023). Den Hartog (1997) stated that Zostera noltiii has a greater tolerance to extreme salinities compared to Zostera marina due to its intertidal habitat. Vermaat et al. (2000) examined salinity tolerance in Zostera noltiii and found considerable mortalities of plants at a salinity of 35 ppt. Zhang et al. (2023) found that exposure to hypersaline conditions of 40 to 45 psu for durations between 5 and 20 days caused significant declines in Zostera marina growth and marked increases in mortality. Shoots subjected to these high-salinity treatments showed poor recovery even after 30 days under normal salinity conditions (Zhang et al., 2023). These findings suggest that both Zostera species are ill-equipped to withstand high saline conditions

Increased salinity may have been the cause of a major Zostera marina die-off in Lake Grevelingen, Netherlands. In the lake, salinity dropped from marine levels (salinity 30) to brackish (salinity 23) after the lake was closed off from the sea in 1971 (Van Katwijk et al., 2023). During this period, Zostera marina expanded to around 4,600 hectares. When a sluice was opened in 1978, marine salinity rapidly returned, coinciding with a severe winter. Between 1978 and the 1990s, the eelgrass meadows collapsed and eventually disappeared. Examination of historical seeds showed that the Grevelingen population had very low germination at salinities between 20 and 32, unlike neighbouring populations (Van Katwijk et al., 2023). This suggested the population had become adapted to low salinity and lost the ability to recruit under marine conditions, making the rise in salinity a likely driver of its decline (Van Katwijk et al., 2023).

Studies looking to improve germination rates for seedling transplantation in restoration projects have shown that high salinity of >30 ppt can reduce Zostera marina seed germination to less than 16% (Xu et al., 2016). Whereas low salinity (<20 ppt) resulted in a significantly higher germination rate (Xu et al., 2016; Liu et al., 2016).). Salinity had a stronger influence on seed germination than temperature, however, seedlings did not develop well in low salinity, requiring a salinity ≥20 ppt to establish (Xu et al., 2016). These studies demonstrated that Zostera marina is sensitive to increased salinity, particularly during early life stages.

Another potential effect of higher salinity is the increase in susceptibility to disease. Jakobsson-Thor et al. (2018, 2019) reported that salinity had a strong influence on the prevalence and/or intensity of infection by Labyrinthula zosterae, the pathogen known to cause seagrass wasting disease, on natural populations from the west coast of Sweden. They showed that 87 to 100% of Zostera marina shoots had lesions in high-salinity meadows (25 to 29 psu), whereas no infection was detected for those in low salinity meadows (13 to 25 psu) (Jakobsson-Thor et al., 2018). This may suggest that low salinity areas can act as a refuge from infection.

Changes in the 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. Although Zostera plants display a wide tolerance to a range of salinities, an increase from 35 to 40 units for the period of one year may cause mortality in Zostera marina. The subtidal habitat makes the species more vulnerable to salinity extremes compared to the intertidal Zostera noltii, Hence, resistance is assessed as ‘Low’ . Zostera marina may be adversely affected by activities such as brine discharges from seawater desalination plants, especially where the discharges occur in habitats acclimated to variable salinities. Recovery, enabled by recolonization from surrounding communities, may be fairly rapid once conditions return to normal if high salinity tolerant plants remain to reproduce, 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 (Salinity regime change pressure definition detail).

Evidence

In general, seagrass species have a wide salinity tolerance. Nejrup & Pedersen (2008) reported optimum salinities between 10 and 25 ppt. Sola et al. (2020) reported that optimum growth of Zostera marina occurred in a salinity of 19. Den Hartog (1970) reported tolerance to salinities as low as 5 ppt. however, such low salinity was shown to inhibit the growth and photosynthetic rate of Zostera marina, particularly in younger leaves (Sola et al., 2020). 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). Zostera marina showed stronger negative effects in hyposalinity than fully saline conditions (30 to 35) (Sola et al., 2020).

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. Sola et al. (2020) exposed Zostera marina to prolonged (16 days) and sudden (24 to 48 hours) drops down to a salinity of 5. The prolonged exposure led to severely reduced photosynthetic capacity and growth (almost no growth after 16 days), whereas the sudden declines caused rapid reductions in photosynthetic efficiency but did not affect growth (Sola et al., 2020). This suggested that eelgrass has some tolerance to such extreme fluctuations, but persistent low salinity has much more lasting negative effects. Additionally, Namba & Nakaoka (2021) showed that both short but frequent pulses and prolonged drops from high salinity (30 to 32) down to 10 reduced the growth rate of Zostera marina. Low salinity also reduced grazer survival and activity, resulting in higher epiphyte loads on the eelgrass, which likely contributed to further growth decline through shading (Namba & Nakaoka, 2021).

Studies looking to improve germination rates for seedling transplantation in restoration projects have shown that high salinity of >30 ppt can reduce Zostera marina seed germination to less than 16% (Xu et al., 2016). In contrast, low salinity (<20 ppt) resulted in a significantly higher germination rate (Xu et al., 2016; Liu et al., 2016). Salinity had a stronger influence on seed germination than temperature, however, seedlings did not develop well in low salinity, requiring a salinity ≥20 ppt to establish (Xu et al., 2016). These studies demonstrated that Zostera marina is sensitive to low salinity, particularly during early life stages.

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 exaggerated when the seagrass is exposed to multiple stressors at the same time, highlighting the importance to consider negative synergistic effects when conducting 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 and 0.2 m/s for more than one year (Water flow pressure definition). 

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 seagrass’ 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 to 150 cm/s. Krumhansl et al. (2021) demonstrated that Zostera marina existed in mean current speeds of as little as 0.7 cm/s in the Atlantic coast of Nova Scotia, Canada.

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 one 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. (Emergence regime change pressure definition).

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 noltii grows predominantly in the intertidal zone and demonstrates 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 five hours of exposure to air during low tide, leaves of Zostera noltii had lost up to 50% of their water content. Decreasing leaf water content resulted in a reversible reduction in the 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 noltii under a given leaf water content. The experiment confirmed that Zostera marina is most susceptible to local changes in emergence regimes because it is 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. In addition, Park et al. (2016) demonstrated that subtidal Zostera marina were more susceptible to desiccation stress than those in the intertidal. Photosynthetic efficiency declined more rapidly in subtidal plants during air exposure, and longer emersion caused irreversible damage, whereas intertidal plants showed signs of recovery once re-immersed (Park et al., 2016).

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. Therefore, the biotope is 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 (Wave action pressure definition). 

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 (i.e. shoots) 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. 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).

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 as well as 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.

However, in North Carolina, USA, the seed bank density and viability of Zostera marina meadows remained the same before and after hurricanes that occurred in 2018 and 2019 (Baker et al., 2025), suggesting long-term stability despite increased storm-induced wave action. In addition, the storms did not affect seagrass biomass or shoot density, suggesting this species may have some tolerance to increases in wave disturbance.

Sensitivity assessment. No evidence was available to determine the impact of this pressure at the benchmark level. However, a change of 3% may affect the upper shore limits and 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 Transitional metal or organometal (e.g. TBT) contaminants via uncontrolled releases or incidental spills (Transitional metals and organometals pressure definition). 

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Transitional 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 hydrocarbon or polyaromatic hydrocarbon (PAH) contaminants via uncontrolled releases or incidental spills (Hydrocarbon & PAH pressure definition).

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Hydrocarbons and PAH' contaminants on 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 those seen between oil types.  ‘Severe’ and ‘significant’ mortality were reported more often in the tropical species Syringodium filiforme, 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 noltii or the shallow extent of Zostera marina (as syn. Zostera angustifolia), which were more likely to have been in 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 noltii or lower shore intertidal Zostera marina) with premixed oil and dispersant treatment showed a significant decrease in cover within the first week, resulting 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 noltii 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 noltii 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 synthetic compound contaminants via uncontrolled releases or incidental spills (Synthetic compound contamination pressure definition).

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Synthetic compound' contaminants on 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 (Radionuclides contamination pressure definition).

Evidence

No evidence found. 

Introduction of other substances

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

Evidence

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) (deoxygenation pressure definition).

Evidence

The effects of oxygen concentration on the growth and survivability of Zostera marina are little 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 could have an adverse effect. Rasmusson et al. (2020) showed that low oxygen conditions (25 to 30 nmol/ml, about 10% of ambient) after 10 minutes of darkness reduced mitochondrial respiration in Zostera marina, especially at higher temperatures (35 to 40°C). When plants experienced low oxygen during darkness, their photosynthetic rates in next day’s light were also reduced, indicating that short-term hypoxia at night can impair subsequent photosynthetic performance. Low oxygen did not significantly affect the maximum quantum yield, meaning the photosynthetic machinery itself was not directly damaged by hypoxia (Rasmusson et al., 2020).

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

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 (N) and phosphorus (P), can lead to eutrophication which causes 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 noltii 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, algae can 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.

Increases in nutrient pollution are believed to have driven Zostera marina and Zostera noltii declines in the Wadden Sea since the 1930’s (Van Katwijk et al., 2024). Seagrass area (including both Zostera species) declined by 80% from the 1930’s to the 1980’s when highest eutrophication was reported (Van Katwijk et al., 2024). This pollution caused toxic concentrations of ammonium within sediment, increased fouling on seagrass leaves, and increased the abundance of other macroalgae (Reise et al., 2025). Large green algal mats comprised mainly of Ulva sp. prevented sufficient light reaching seagrasses as a result (Reise et al., 2025). Seagrass in this area recovered from 2% in the 1980’s to 13% in the present day, following the reduction of nutrient input during the 1990’s (Van Katwijk et al., 2024).

The effects of nutrient input from point source sewage on Zostera marina and Zostera noltii were examined in the Black Sea (Holmer et al., 2016). Near bottom nutrient concentrations at Station 1 (closest to the sewage source) were: ammonium 4.3 µM, nitrite 0.18 µM, nitrate 0.38 µM and phosphate 0.38 µM, compared to 0.73, 0.06, 0.8, 0.2 µM respectively at Station 2, the next station further from the source. Zostera noltii was absent from Station 1, whereas it was more abundant than Zostera marina at the further three stations. Conversely, Zostera marina biomass was highest at Station 1 compared to the other stations, suggesting that this species was much more tolerant of organic enrichment than Zostera noltii, and may have benefited from increased nutrient concentrations (Holmer et al., 2016).

The positive effect of moderate nitrogen and/or phosphorous enrichment has been reported for many seagrasses (Peralta et al., 2003). Cimon et al. (2021) investigated the effect of nutrient loading in sediment on Zostera marina by placing slow-release nutrient pellets to give 75 g N per m² into sediment and compared the effect after 10 weeks. The nutrient loading significantly increased shoot density, however, this positive effect was negated when the seagrass was shaded by two layers of fiberglass window screen, which reduced light by 68% (Cimon et al., 2021). Jackson et al. (2017) further reported that dissolved nitrogen enrichment of between 68 and 145 µM in porewater positively affected Zostera marina by increasing shoot height, accelerating reproduction and doubling seed output, thus enhancing reproductive success. Increases in leaf growth, shoot density, seed production and reproductive effort in Zostera marina resulting from moderate nutrient enrichment has been observed in numerous other studies (Wang et al., 2020; Qin et al., 2021; Suonan et al., 2022). As such, Vieira et al. (2020) surmised that seagrass meadows globally, including those of Zostera marina and Zostera noltii, can benefit from moderate addition of anthropogenic nutrients, and reported that some of the healthiest meadows in Portugal were close to wastewater treatment plants and food factory effluents.

However, excessive nutrient addition would still lead to ammonium toxicity and water column nitrate inhibition through internal carbon limitation (Touchette & Burkholder, 2000), as well as eutrophication and macroalgal overgrowth. For example, the addition of 100 mg/l of ammonium caused Zostera marina leaf growth rate and photosynthetic rate to decrease significantly (Wang et al., 2021). However, root activity did not decline significantly until 200 mg/l of ammonium, suggesting that root tissue was more tolerant of increased ammonium than leaf tissues.

In addition to nitrate and phosphate enrichment, excessive iron concentrations can also negatively impact Zostera marina survival. Wang et al. (2017) demonstrated that mortality of young plants significantly increased when iron levels reached over 1mg/l, whereas seedlings showed a higher resistance to iron enrichment, with significant increases in mortality occurring at 1.5 mg/l.

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. In addition, physiological stress caused by nutrient enrichment may also leave seagrasses more susceptible to disease. Hughes et al. (2018) demonstrated that increased levels of nitrate (75 to 150 µM) significantly increased infection rate and spread of Labyrinthula in Zostera marina.

Sensitivity assessment. Moderate nutrient enrichment can have positive effects on Zostera marina by increasing shoot growth and density, and enhancing seed production and reproductive success. However, excessive nutrient input has led to the loss of seagrass beds worldwide, due in part to the  smothering by algae and epiphytes, ammonium toxicity, 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. However, in the Wadden Sea, recovery of diminished seagrass beds (loss of 80%) was observedaround 30 years after the reduction in nutrient input, recovering to an area larger than pre-decline levels (Van Katwijk et al., 2024). Therefore, resistance is assessed as ‘None’ based on the severe mortality attributed to nutrient enrichment (e.g. in the Wadden Sea) and loss of seagrasses worldwide. Given the long recovery time in the Wadden Sea and the incomplete recovery of meadows after 20 years in Florida, resilience is assessed as ‘Very low’ and sensitivity is, therefore, assessed as ‘High’.

Organic enrichment

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

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.

The effects of nutrient input from point source sewage on Zostera marina and Zostera noltii were examined in the Black Sea (Holmer et al., 2016). Zostera noltii was absent from Station 1 (closest to the sewage output), whereas it was more abundant than Zostera marina at the further three stations. Conversely, Zostera marina biomass was highest at Station 1 compared to the other stations, suggesting that this species was much more tolerant of organic enrichment than Zostera noltii, and may have benefited from increased nutrient concentrations, although the paper did not quantify the carbon loading caused from the sewage output (Holmer et al., 2016).

Neverauskas (1987) investigated the effects of discharged digested sludge from a sewage treatment on Posidonia spp. and Amphibolis spp. in South Australia. Within five years, the outfall had affected an area of approximately 1,900 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 eight 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 shoots 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 sulphide-oxidizing gill bacteria (Van der Heide et al., 2012). In experiments, the sulphide-oxidizing gill bacteria of Loripes lacteus were shown to reduce sulphide levels in the sediment and enhance the productivity of Zostera noltii, while the oxygen released from the roots of Zostera noltii 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 while moderate nutrient input may benefit the growth and reproduction of some seagrasses (see ‘Nutrient enrichment’ section), excessive 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). The above evidence shows that seagrass beds found in proximity to a source of organic discharge tended to be severely impacted, with important loss of biomass. However, the evidence from the Black Sea suggests that this may not be the case for Zostera marina which was found in highest biomass closest to the sewage outflow, potentially benefitting from the additional nutrients. Although no study was found on the British species, the evidence suggests that Zostera marina may have some tolerance to organic enrichment, particularly when it results in favourable levels of nitrogen and phosphorous. 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’ to represent potential impacts due to indirect anoxia, epiphytic growth, but with ‘Low’ confidence. 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 (Physical loss pressure definition). 

Evidence

All marine habitats and benthic species are considered to have a resistance of 'None' to this pressure and to be unable to recover from a permanent loss of habitat 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 (Physical change in subtratum type pressure definition).

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) (Physical change in sediment type pressure definition). 

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) (Removal of substratum pressure definition). 

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) (Surface abrasion/disturbance pressure definition).

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, sheltered 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. In Chesapeake Bay, USA, boat propellors from commercial seine fishing and crab scraping were shown to dig into the sediment, cutting both leaves and rhizomes of seagrass. This activity created scars of up to 2.8 m in width and over 900 m in length in seagrass meadows dominated by Zostera marina (also containing Ruppia maritima) (Orth et al., 2017). The scars took an average of 2.7 years to fully recover, achieved through both vegetative growth and seed dispersal (Orth et al., 2017).

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. In contrast, deploying and retrieving shrimp pots up to 30 times, and leaving them in the water for up to 9 days while eelgrass was dormant (in September) in Kilkieran Bay, Ireland, did not negatively affect seagrass shoot density or rhizome weight during the growing season (Breen et al., 2024). This may suggest that the impacts of these fishing practices depend on when in the season they occur (Breen et al., 2024).

Grazing: Nacken & Reise (2000) investigated physical disturbance caused by Brent geese (Branta b. bernicla) and widgeon (Anas penelope) feeding on Zostera noltii 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 noltii 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 give rise to this pressure. Given that intertidal habitat is more exposed, making it more readily accessible than subtidal beds. The resilience and recovery of seagrass beds to abrasion of the seabed surface depends on the frequency, persistence, timing, 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 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). Where abrasive pressures remove seagrass shoots but do not penetrate deeply enough to severely impact rhizomes, recovery will be fairly rapid through vegetative growth. Resilience is, therefore, assessed as ‘Medium’ and, sensitivity is assessed as ‘Medium’.

Penetration or disturbance of the substratum subsurface

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

Evidence

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). Barañano et al. (2017) found that disturbance from clam harvesting initially reduced Zostera marina shoot density and biomass, but recovery to levels comparable to the undisturbed site occurred within four months. However, sexual reproductive effort remained markedly lower in the harvested meadow (4%) than in the non‑harvested site (10%), even after full recovery.

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, thereby, causing greater damage. A technical paper by Collins et al. (2010) using SCUBA divers found bare patches (typically 1 to 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 no 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.

In San Francisco Bay, anchoring from boats was shown to have caused between 25 and 41% loss of a Zostera marina meadow, with each boat damaging up to 0.3 hectares of eelgrass due to anchor scour (Kelly et al., 2019). 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 scour 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).

In the English Channel, chains from swing moorings were shown to entirely remove Zostera marina within roughly a 5 m radius of the mooring point, with the extent of damage varying depending on tidal direction and chain movement (Ouisse et al., 2020). Beyond this bare zone, eelgrass persisted, however, canopy height was 60% shorter than unmoored sites up to 15 m from the mooring, suggesting that Zostera marina compensated by producing more shoots, resulting in higher shoot density but lower overall plant biomass (Ouisse et al., 2020). Destruction from swing chain moorings caused at least six hectares of Zostera marina loss overall from other sites along the south-west coast of the UK (and one site in Wales) (Unsworth et al., 2017). Each individual mooring causing an average eelgrass loss of 122 m² (Unsworth et al., 2017).

Conservation mooring systems (CMS), which lift mooring gear off the seabed, aim to reduce chain scouring impact. In Massachusetts, seagrass recovery five years after CMS installation varied strongly with environmental conditions and the initial size of scars (Seto et al., 2024). Recovery was lowest in areas with high wave exposure, large tidal ranges, or very shallow depths where CMS gear could still contact the seabed. Larger scars tended to recover more than small ones, although overall recovery was often incomplete, with persistent unvegetated halos remaining around anchors. Mechanical issues such as fouling or improper installation further reduced recovery, and in some cases small scars expanded rather than healed (Seto et al., 2024).

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 2 to 3%, 46 to 61% and <1% that of the reference sites, respectively. Substantial differences in Zostera marina biomass persisted between disturbed and reference sites up to seven 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.

Recovery after abrasion may depend on when during the growing season of Zostera marina the disturbance occurred. Boardman & Ruesink (2025) found that recovery after dredging was much higher when the disturbance occurred earlier in the growing season (January to April) as seedlings were still able to germinate and establish. Disturbances later in the season resulted in much poorer recovery (Boardman & Ruesink, 2025). In addition, in Kilkieran Bay, Ireland, oyster dredging carried out in early May, before the Zostera marina growing season, showed no significant effect on rhizome weight or shoot density within four months after dredging (Breen et al., 2024).

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 rhizome and 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 six years but 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 (Suspended sediment pressure definition).

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. Nelson (2017) reviewed the effects of reduced light on seagrass species and concluded that significant negative effects (i.e. loss of biomass or shoot density) occurred after more than 60 days of 50 to 90% light reduction. Nelson (2017) demonstrated that Zostera sp. typically showed faster and more severe declines in response to shading compared to other species of seagrass. The critical threshold of light requirements varies among species. In Zostera noltii, minimum light requirement was suggested as 2% in-water irradiance, whereas Zostera marina likely required between 11 to 37% surface irradiance (Erftemeijer & Robin, 2006; Van Katwijk et al., 1997) or between 0.7 and 12.6 mol photons/m²/day (Leger-Diagle et al., 2022; Howarth et al., 2022).

These differences in the light requirement for Zostera are reflected by the position of species along a depth gradient, with Zostera noltii 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.

Increases in nutrient pollution are believed to have driven Zostera marina and Zostera noltii declines in the Wadden Sea since the 1930s (Van Katwijk et al., 2024). Seagrass area (including both Zostera species) declined from 10% in 1930 to 2% during the highest levels of eutrophication in the 1980s (Van Katwijk et al., 2024). Among other impacts, this pollution caused increased fouling on seagrass leaves and increased the abundance of other macroalgae (Reise et al., 2025). Large green algal mats comprised mainly of Ulva sp. prevented sufficient light reaching seagrasses (Reise et al., 2025). Seagrass in this area recovered to 13% in the present day, following the reduction of nutrient input during the 1990s (Van Katwijk et al., 2024).

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. However, shading alone did not cause mortality in the experimental time frame. Cimon et al. (2021) investigated the effect of shading by window screens (reduced light by 68%) on Zostera marina from the St Lawrence Estuary, Canada. After 10 weeks, shading significantly reduced growth rate, shoot density, and carbohydrate reserves, while increasing the abundance of epiphytic algae (Cimon et al., 2021). Bertelli & Unsworth (2018) showed that decreasing light levels to below 20 μmol photons m²/s for seven days significantly reduced Zostera marina growth and photosynthetic rate. After 29 days, leaf size reduced by 41%, and within four to six weeks, shoot mortality occurred (Bertelli & Unsworth, 2018).They concluded that if light levels fall below 20 μmol photons m²/s for multiple weeks, Zostera marina meadows will rapidly decline (Bertelli & Unsworth 2018).

Wong et al. (2021) demonstrated that shading (60 to 80% light reduction) caused major declines in Zostera marina shoot density and biomass whether it was continuous or episodic (12 days shade, 2 days light). Li et al., (2021) showed experimentally that short-term turbidity increases caused substantial declines in Zostera marina; 50 to 100 nephelometric turbidity units (NTU) for 5 to 10 days led to noticeable reductions in survival, with mortality increasing by up 40% during a 30-day recovery period. At 100 to 200 NTU for 10 to 15 days, mortality increased to around 50%. Aboveground productivity decreased by 50% at turbidity levels of 151 to 324 NTU when exposed for 10 to 20 days. The most extreme conditions of 300 to 400 NTU for 15 to 20 days resulted in 100% mortality. Overall, this study showed that relatively brief but intense turbidity spikes can cause 50 to 100% declines in survival and growth, and prolonged high turbidity can become lethal (Li et al., 2021). Note NTUs are measured by light penetration through the fluid medium, while suspended sediment concentration is measured as the dry weight of particulates in suspensions, usually as mg/l.

In a six-month experiment in the Dutch Wadden Sea, Philippart (1995) found that shading induced a 30% decrease in the leaf growth rate, a three-fold increase in the leaf loss rate, and an 80% reduction in the total biomass of Zostera noltii. The decreasing growth rate was 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 noltii 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).

Reduced light levels have also been seen to impact Zostera marina’s physiological tolerance to other stressors. For example, wasting disease prevalence in Zostera marina increased by 35% when exposed to 17% of ambient light (Jakobsson-Thor et al., 2020). In addition, when coupled with 75% reduced light intensity and an increase in ammonium concentrations, the negative effect of 25°C water temperature on Zostera marina intensified (Moreno-Marin et al., 2018). Beca-Carretero et al. (2018, 2021) demonstrated a similar effect, where no mortality occurred at 24°C under high light (180 μmol photons m²/s), but 30% of seagrass died under low light (60 μmol photons m²/s), . 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 (Smothering pressure definition).

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. Nelson (2017) reviewed the effects of reduced light on seagrass species and concluded that significant negative effects (i.e. loss of biomass or shoot density) occurred after more than 60 days of 50 to 90% light reduction. Nelson (2017) demonstrated that Zostera sp. typically showed faster and more severe declines in response to shading compared to other species of seagrass. The critical threshold of light requirements varies among species. In Zostera noltii, minimum light requirement was suggested as 2% in-water irradiance, whereas Zostera marina likely required between 11 to 37% surface irradiance (Erftemeijer & Robin, 2006; Van Katwijk et al., 1997) or between 0.7 and 12.6 mol photons/m²/day (Leger-Diagle et al., 2022; Howarth et al., 2022).

These differences in the light requirement for Zostera are reflected by the position of species along a depth gradient, with Zostera noltii 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.

Increases in nutrient pollution are believed to have driven Zostera marina and Zostera noltii declines in the Wadden Sea since the 1930s (Van Katwijk et al., 2024). Seagrass area (including both Zostera species) declined from 10% in 1930 to 2% during the highest levels of eutrophication in the 1980s (Van Katwijk et al., 2024). Among other impacts, this pollution caused increased fouling on seagrass leaves and increased the abundance of other macroalgae (Reise et al., 2025). Large green algal mats comprised mainly of Ulva sp. prevented sufficient light reaching seagrasses (Reise et al., 2025). Seagrass in this area recovered to 13% in the present day, following the reduction of nutrient input during the 1990s (Van Katwijk et al., 2024).

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. However, shading alone did not cause mortality in the experimental time frame. Cimon et al. (2021) investigated the effect of shading by window screens (reduced light by 68%) on Zostera marina from the St Lawrence Estuary, Canada. After 10 weeks, shading significantly reduced growth rate, shoot density, and carbohydrate reserves, while increasing the abundance of epiphytic algae (Cimon et al., 2021). Bertelli & Unsworth (2018) showed that decreasing light levels to below 20 μmol photons m²/s for seven days significantly reduced Zostera marina growth and photosynthetic rate. After 29 days, leaf size reduced by 41%, and within four to six weeks, shoot mortality occurred (Bertelli & Unsworth, 2018).They concluded that if light levels fall below 20 μmol photons m²/s for multiple weeks, Zostera marina meadows will rapidly decline (Bertelli & Unsworth 2018).

Wong et al. (2021) demonstrated that shading (60 to 80% light reduction) caused major declines in Zostera marina shoot density and biomass whether it was continuous or episodic (12 days shade, 2 days light). Li et al., (2021) showed experimentally that short-term turbidity increases caused substantial declines in Zostera marina; 50 to 100 nephelometric turbidity units (NTU) for 5 to 10 days led to noticeable reductions in survival, with mortality increasing by up 40% during a 30-day recovery period. At 100 to 200 NTU for 10 to 15 days, mortality increased to around 50%. Aboveground productivity decreased by 50% at turbidity levels of 151 to 324 NTU when exposed for 10 to 20 days. The most extreme conditions of 300 to 400 NTU for 15 to 20 days resulted in 100% mortality. Overall, this study showed that relatively brief but intense turbidity spikes can cause 50 to 100% declines in survival and growth, and prolonged high turbidity can become lethal (Li et al., 2021). Note NTUs are measured by light penetration through the fluid medium, while suspended sediment concentration is measured as the dry weight of particulates in suspensions, usually as mg/l.

In a six-month experiment in the Dutch Wadden Sea, Philippart (1995) found that shading induced a 30% decrease in the leaf growth rate, a three-fold increase in the leaf loss rate, and an 80% reduction in the total biomass of Zostera noltii. The decreasing growth rate was 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 noltii 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).

Reduced light levels have also been seen to impact Zostera marina’s physiological tolerance to other stressors. For example, wasting disease prevalence in Zostera marina increased by 35% when exposed to 17% of ambient light (Jakobsson-Thor et al., 2020). In addition, when coupled with 75% reduced light intensity and an increase in ammonium concentrations, the negative effect of 25°C water temperature on Zostera marina intensified (Moreno-Marin et al., 2018). Beca-Carretero et al. (2018, 2021) demonstrated a similar effect, where no mortality occurred at 24°C under high light (180 μmol photons m²/s), but 30% of seagrass died under low light (60 μmol photons m²/s), . 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 (heavy)

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

Evidence

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).

Sensitivity assessment. 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 will be ‘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) (Litter pressure definition). 

Evidence

Not assessed

Electromagnetic changes

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

Evidence

Evidence on the effect of electromagnetic fields (EMFs) on benthic organisms is still severely lacking. Some studies have investigated the effect of anthropogenically induced EMFs on benthic invertebrates at intensities ranging between 2 nT and 40 mT, which is often much higher than in-situ measurements from subsea cables. While some report changes to behaviour, physiology, reproduction, development, immunology, cytotoxicity and orientation, others demonstrate no effect from exposure to the EMF (Albert et al., 2020; Hutchison et al., 2020a), depending on the study species and duration and intensity of exposure. There have been no studies investigating the effect of EMFs at the population or community level for benthic organisms, nor have there been any studies on the effect of EMFs on macrophytes. 

Sensitivity assessment. No studies have examined the effect of EMFs on seagrass, therefore, there is ‘Insufficient evidence’ on which to base an assessment of the likely sensitivity of Zostera marina beds to EMFs.

Underwater noise changes

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

Evidence

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

Introduction of light or shading

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

Evidence

An increase in light might be beneficial while shading by artificial structures will decrease incident light and hence reduce photosynthesis and growth rates. Nelson (2017) reviewed the effects of reduced light on seagrass species and concluded that significant negative effects (i.e. loss of biomass or shoot density) occurred after more than 60 days of 50 to 90% light reduction. Nelson (2017) demonstrated that Zostera sp. typically showed faster and more severe declines in response to shading compared to other species of seagrass. For Zostera marina, the minimum light requirement has been suggested as 11 to 34% surface irradiance (Van Katwijk et al., 1997) or between 0.7 and 12.6 mol photons m² per day (Leger-Diagle et al., 2022; Howarth et al., 2022).

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. However, shading alone did not cause mortality in the experimental time frame. Cimon et al. (2021) investigated the effect of shading (reduced light by 68%) on Zostera marina from the St Lawrence Estuary, Canada. After 10 weeks, shading significantly reduced growth rate, shoot density, and carbohydrate reserves, while increasing the abundance of epiphytic algae (Cimon et al., 2021). Bertelli & Unsworth (2018) showed that decreasing light levels to below 20 μmol photons m²/s for seven days significantly reduced Zostera marina growth and photosynthetic rate. After 29 days, leaf size reduced by 41%, and within four to six weeks, shoot mortality occurred (Bertelli & Unsworth, 2018). Wong et al. (2021) demonstrated that shading (60 to 80% light reduction) caused major declines in Zostera marina shoot density and biomass whether it was continuous or episodic (12 days shade, 2 days light).

Aquaculture and anthropogenic infrastructures are a source of shading in seagrass habitats. Howarth et al. (2022) identified shading and sedimentation as two of the main negative pathways by which shellfish aquaculture affects Zostera marina. Shading was caused by aquaculture structures such as suspended bags, rafts, longlines, and cages. Sedimentation, which reduces light reaching the seabed, resulted from the deposition of faeces and pseudofaeces, physical disturbance during installation or harvesting, and the accumulation of fine sediments beneath structures. Eriander et al. (2017) showed that along the west coast of Sweden, shading from both floating and elevated docks decreased Zostera marina extent by 42 to 64% within 6 m of the dock. Floating docks had a stronger negative impact due to blocking more light, with eelgrass directly below declining by up to 100% compared to docks held up by poles, which led to reductions of up to 70% (Eriander et al., 2017).

Reduced light levels have also been seen to impact Zostera marina’s physiological tolerance to other stressors. For example, wasting disease prevalence in Zostera marina increased by 35% when exposed to 17% of ambient light (Jakobsson-Thor et al., 2020). In addition, when coupled with 75% reduced light intensity and an increase in ammonium concentrations, the negative effect of 25°C water temperature on Zostera marina intensified (Moreno-Marin et al., 2018).

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 tolerate anoxic and sulfidic conditions. The impact of shading on Zostera marina can, therefore, depend on when in the growing season the disturbance occurs. Wong et al. (2020) showed that shading during late summer and autumn causes significantly greater declines in shoot density, growth, and carbohydrate reserves compared with shading during spring or early summer.

Sensitivity assessment. While there is evidence concluding that reduced light impacts the physiology of Zostera marina, records of mortality remain scarce. However, Zostera marina mortality was observed within six weeks under reduced light conditions (Bertelli & Unsworth, 2018), and significant losses were recorded below and around docks, especially underneath floating docks (Eriander et al., 2017). 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 (Barrier to species movement pressure definition).

Evidence

Not relevant–this pressure is considered applicable to mobile species, e.g. fish and marine mammals rather than seabed habitats. Physical and hydrographic barriers may limit 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 (Death for collision pressure definition).

Evidence

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

Visual disturbance

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

Evidence

Not relevant

Genetic modification & translocation of indigenous species

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

Evidence

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 of microbial pathogens

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

Evidence

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 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).

In addition, increases in salinity, temperature, eutrophication, and potentially ocean acidification, have been shown to increase susceptibility of seagrasses to Labyrinthula infection (reviewed by Wang et al., 2024 and Sullivan et al., 2018). However, the combined effect of increased temperature and low salinity was shown to reduce infection rate in eelgrass from the Baltic Sea (Brakel et al., 2019). In addition, Hughes et al. (2018) demonstrated that elevated levels of nitrate (75 to 150 uM) significantly increased infection rate and spread of Labyrinthula in Zostera marina. Graham et al. (2025), showed that deeper subtidal meadows were less susceptible to wasting disease compared to shallow, intertidal meadows,in San Juan Islands, USA, potentially due to the more stable environmental conditions found at depth.

Other pathogens reported to cause disease in Zostera include Phytophthora sp. and Halophytophthora sp. which can infect Zostera marina and Zostera noltii seeds, causing mortality and reducing germination and seed development (Sullivan et al., 2018).

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 (targeted removal pressure definition).

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 (non-targeted removed pressure definition).

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.

Numerous species groups found within the biotope can impact the recovery potential of Zostera marina after disturbance. For example, Gangon et al. (2021) demonstrated that in the northern Baltic Sea, grazing from crustacean and gastropod groups negated the negative effects of nutrient enrichment by controlling algal biomass, preventing overgrowth of Zostera marina. Östman et al. (2016) reported that in North Atlantic Zostera marina systems, top‑down effects are just as important as nutrient enrichment in driving the growth of ephemeral algae. The study showed that when predatory fish such as cod are reduced, mesopredators (e.g., small fish and crabs) increase in abundance, leading to greater consumption of grazers. Grazers play a key role in controlling fast‑growing macroalgae, therefore, their decline allows ephemeral algae to proliferate, producing an effect similar to nutrient-driven eutrophication in Zostera marina systems (Östman et al., 2016).

In addition, the abundance other species may influence Zostera marina sexual reproductive success. Meysick et al. (2019) demonstrated that the presence of bivalves (Magallana gigas and Mytilus edulis) increased the retention of Zostera marina seeds at high flow velocities (18 to 30 cm/s). This was largely due to sediment scouring around the bivalves, which created pits in which seeds became trapped (Meysick et al., 2019). The presence of manila clams, Ruditapes philippinarum, also significantly increased Zostera marina seed burial, thus reducing seed predation, preventing loss of seeds to currents, and facilitating favourable burial depths (Li et al., 2017). However, high densities of the lugworm, Abarenicola pacifica, caused Zostera marina seeds to become buried up to 17 cm, preventing seedlings from establishing (Crow et al., 2023), Infantes et al. (2016) determined that seed loss occurs through predation by hermit crabs, urchins, and the shore crab, Carcinus maenas. The shore crab was shown to cause a 73% loss of Zostera marina seeds in one week (Infantes et al., 2016). The ragworm Hediste diversicolor was also shown to predate on sprouted Zostera marina seeds and seedlings. Therefore, removal of species that feed on Zostera marina seeds may enhance recovery after disturbance.

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. In addition, removal of higher predators that would ultimately cause a decrease in grazing may also negatively affect Zostera habitats through the proliferation of other macroalgae. Therefore, resistance is considered to be 'None', resilience 'Low' and sensitivity 'High'.

The American slipper limpet, Crepidula fornicata

Evidence

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and the Essex coast in 1887-1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 1999, 2018; Hinz et al., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). It ranges from the Baltic Sea, the Kattegat and Skagerrak, the North Sea coasts of the UK, Germany, and Belgium, through the English Channels and into the Irish sea coasts of Ireland and south Wales with records in east and west Scotland, Northern Ireland, northwest France, Spain and south into the Mediterranean (NBN, 2023; OBIS, 2023).

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

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

In Arcachon Bay, France, Crepidula fornicata was limited to muddy sediments that were only ca 8% of the bay 30 years after its introduction (De Montaudouin et al., 2001, 2018). De Montaudouin et al. (2001) attributed its scarcity in Arcachon Bay to the lack of suitable muddy sediments, a lack of bottom trawl fisheries, and large beds of Zostera spp. on intertidal and subtidal. Chains of Crepidula were scarce in the Zostera beds. As a result, De Montaudouin et al. (2018) concluded that Crepidula was not invasive in the Bay of Arcachon. Tillin et al. (2020) suggested that the sediment and the movement of fronds or leaves of seagrass may have reduced its ability to colonize.

Sensitivity assessment. Crepidula has been recorded from areas of strong tidal streams (Hinz et al., 2011), and from the lower intertidal to ca 160 m in depth, but it is most common in the shallow subtidal above 50 m where this biotope is observed (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 2023; Tillin et al., 2020). The sandy, muddy fine sand or muddy sediment characterizing this biotope may besuitable for the colonization by Crepidula fornicata butthe presence of hard substrata (shells, gravel, pebbles and cobbles) that can occur in this biotope could allow Crepidula to gain a foothold. However, Crepidula fornicata was reported as scarce in Zostera spp. in Arcachon Bay, France (De Montaudouin et al., 2001, 2018; Tillin et al., 2020) where they and may have contributed to its failure to colonize areas of the bay. No evidence was found on the effect of Crepidula populations on seagrass-dominated habitats. At present, there is 'Insufficient evidence' to suggest that the Zostera marina biotopes are sensitive to colonization by Crepidula fornicata; further evidence is required.

The carpet sea squirt, Didemnum vexillum

Evidence

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

Although a widespread invader, Didemnum vexillum has a limited ability for natural dispersal since the pelagic larvae remain in the water column for a short time (up to 36 hours). Therefore, it has a short dispersal phase that can allow the species to build localized populations (Herborg et al., 2009; Vercaemer et al., 2015; Holt, 2024). However, Bullard et al. (2007) suggested that Didemnum vexillum can form new colonies asexually by fragmentation. Colonies can produce long tendrils from an encrusting colony, which can fragment, disperse and settle, attaching to suitable hard substrata elsewhere (Bullard et al., 2007; Lambert, 2009; Stefaniak & Whitlatch, 2014). A fragmented colony can spread naturally for up to three weeks transported by ocean currents, attached to floating seaweed, seagrass or other floating biota, or as free-floating spherical colonies (Bullard et al., 2007; Lengyel et al., 2009; Stefaniak & Whitlatch, 2014; Holt, 2024). Fragments can reattach to suitable substrata within six hours of contact. Fragments have the potential to disperse around 20 km before reattachment (Lengyel et al., 2009). Valentine et al. (2007a) reported that colonies of Didemnum vexillum enlarged by 6 to 11 times by asexual budding after 15 days and enlarged 11 to 19 times after 30 days. Valentine et al. (2007a) concluded fragments could successfully grow, survive, and help to spread Didemnum vexillum. While natural fragmentation of tendrils is thought to allow Didemnum vexillum to invade longer distances and increase its dispersal potential, Stefaniak & Whitlatch (2014) found that only 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 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 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. They were not present on the natural seabed under the pontoon, which was composed of silty mud, or on deeper sections of mooring chains 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 bottoms (cited Gittenberger, 2007). In northern Kent, Didemnum vexillum has been recorded covering London clay boulders on Whitstable Flats, West Beach, north Kent, covering tabulate sandstone boulders (0.5 to 2 m across) on the mid-shore and colonizing sediment mounds on the low shore characterized by larger areas of sand, mud and low-lying sediment at Reculver and Bishopstone, north Kent (Hitchin, 2012). It was also recorded from muddy substrata at that site. Hitchin (2012) noted that the site was exposed to enough waves and currents to cause sedimentation. However, Didemnum vexillum grew hanging from on the underside of sandstone boulders nestled on sediment, on consolidated sediment mounds and firm clays, hence burial may prevent colonization and its survival rather than sedimentation alone.

In contrast, Didemnum vexillum’s preference for sheltered conditions, established colonies observed in Georges Bank and Long Island Sound were exposed to moderately strong tidal currents (1 to 2 knots; ca 0.5 to 1 m/s recorded at both sites) that may mobilise sediment (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020). However, Valentine et al. (2007b) describe the substratum as immobile, presumably consolidated, gravel, cobbles, and pebbles. Kleeman (2009) stated that consistent mild wave action or ‘swash zone’ appeared 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, and daily changes in water depth and temperatures (Valentine et al., 2007a). Didemnum vexillum colonies were able to survive exposure to air at low tides for a short time (not exceeding two hours) during rapid colony growth in the summer months of July to September (Valentine et al., 2007a). However, parts of the large established colonies, which were artificially exposed to air for two to three hours in October, were observed desiccated or predated on by grazing periwinkles 30 days later, in the winter month of November (Valentine et al., 2007a). They suggested that the invasive tunicates’ ability to tolerate exposure to air varies with the seasonal growth cycle. Didemnum vexillum also tolerated emersion in Kent, as colonies on the mid-shore at Reculver flourish and survive in air exposure for up to three hours per cycle during 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 six 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 of live or dead eelgrass and on detached pieces of eelgrass Zostera marina, in Lake Tashmoo on Martha’s Vineyard, 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 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 were 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 the 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. The evidence presented shows that Didemnum vexillum can overgrow eelgrass beds. In these biotopes, eelgrass provides suitable substrata and stabilises 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 10 ha 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. Therefore, a resistance of 'Medium' (some mortality, <25%) is suggested as a precaution to reflect the potential reduction in growth and resultant population decline. 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' but with 'Low' confidence.  

The Pacific oyster, Magallana gigas

Evidence

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

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

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

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

Magallana gigas typically occurs in the lower intertidal and shallow subtidal, mostly abundant at depths 1 m to 10 m (Tillin et al., 2020). However, the preferred depth range varies with substratum. On littoral rock in Brittany, the Pacific oyster colonizes all intertidal levels from Mean High Water to Mean Low Water on sheltered (low energy), moderately exposed (moderate energy) and high energy rock shores (Herbert et al., 2012). However, in the Northwest Pacific, Magallana gigas is commonly found on sheltered low energy littoral rock and has less than 10% cover on exposed high energy littoral rock shores (Herbert et al., 2012, 2016). Magallana gigas was recorded in the mid intertidal on hard rock and artificial harbour structures in northern and eastern Adriatic, but absent from beam trawl surveys off the coast of Istria (Ezgeta-Balic et al., 2019). Ezgeta-Balic et al. (2019) noted that the only oyster found in their trawls were large Ostrea edulis, often confused for Magallana gigas by local fishermen. Magallana gigas has not been found at extreme low water levels or subtidally beneath rocky habitats, as it has been in soft sediment areas (Herbert et al., 2012).

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

Pacific oyster aquaculture sometimes co-occurs in areas of seagrass. Agnew et al. (2022) reported that, under laboratory conditions, the presence of Magallana gigas reduced lesion severity and reduced wasting disease infection intensity, likely by filtering the pathogen out of the water. However, oysters previously exposed to the pathogen were also shown to transmit the disease to uninfected seagrass (Agnew et al., 2022). In field trials, oyster presence did not influence wasting disease prevalence or severity, however, effects may have been undetectable due to loss of seagrass tissue (Agnew et al., 2022).

Pacific oysters can coexist with eelgrass on a regional scale. In British Columbia and the American Pacific northwest, where oysters inhabit the high intertidal zone and eelgrass inhabit low intertidal zone to shallow subtidal. But on a finer scale, eelgrass is absent directly adjacent to oysters and the presence of oysters reduces the abundance of eelgrass (Kelly et al., 2007; Padilla, 2010) but it was not clear if this was due to tidal level or exclusion by the oysters (Tillin et al., 2020). Ruesink et al., (2006) reported that the abundance of native oysters (Ostreola conchaphila) and Zostera marina in Willapa Bay, USA, was reduced by the introduction of four non-natives species (cordgrass Spartina alterniflora, Manila clams Ruditapes philippinarum, Japanese eelgrass Zostera japonica, and Pacific oysters Magallana gigas) between ca 1900 and 2000. However, the reduction on Zostera marina cover was due to competition and ecosystem changes attributable to several non-natives rather than Magallana alone.

Sensitivity assessment. No reports were found of Magallana gigas attached to Zostera marina. The sandy, muddy fine sand or muddy sediment found in this biotope would not provide suitable attachment for the species but the presence of hard substrata (shells, gravel, pebbles and cobbles) that can occur in this biotope could allow Magallana to gain a foothold. The evidence from the USA suggests that Zostera marina and Magallana gigas can coexist and overlap but that the Pacific oyster may exclude Zostera marina in shallow water (Kelly et al., 2007; Padilla, 2010). Therefore, resistance is assessed as ‘Low’, resilience and ‘Very low’, and sensitivity is assessed as ‘High’, albeit with ‘Low’ confidence due to the limited available evidence .

Wireweed, Sargassum muticum

Evidence

Sargassum muticum is a circumglobal invasive species (Engelen et al., 2015). It is recorded (2015) from Norway to Morocco and into the Mediterranean in the eastern Atlantic and from Alaska to Baja California in the eastern Pacific and from southern Russia to southern China in the western Pacific (Engelen et al., 2015). It colonizes a variety of habitats and can tolerate -1°C to 30°C and survive salinities below 10 ppt, but has a preference for full salinity ranges, 30 to 34 psu. Although fertilization does not occur below 15 ppt and growth of germlings is limited below 10°C, it can complete its life cycle as long as temperatures are over 8°C for at least four months of the year (Engelen et al., 2015). However, its distribution is limited by the availability of hard substratum (e.g., stones >10 cm) and light (Staehr et al., 2000; Strong & Dring, 2011; Engelen et al., 2015). It is most abundant between 1 and 3 m below mean water. But it has been recorded at 18 m or 30 m in the clear waters of California. However, it is a poor competitor under low light and only develops dense canopies in shallow areas (Engelen et al., 2015).

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 Sargassum 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 sandy and muddy bottoms, whereas Sargassum muticum attaches to hard 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, south-west 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. Firth et al. (2024) showed that Sargassum muticum can also disperse into Zostera marina habitat by attaching to limpet shells. In Devon, south‑west England, surveys found that 5% of Sargassum muticum individuals in seagrass beds were attached to dead limpet shells, and one individual remained attached to a live limpet, indicating that live limpets may act as transport vectors. Sargassum muticum was otherwise attached to rock, gravel, the seagrass matrix itself, and embedded within the sand. The long-term field experiment in this study showed that Zostera marina shoot density was significantly lower in plots where Sargassum muticum co‑occurred (Firth et al., 2024). 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).

Sensitivity assessment. Reports of Sargassum muticum occurring alongside Zostera marina exist for several locations in southwest England, but the ecological consequences of this coexistence remain poorly documented. The sandy, muddy fines and, and muddy sediments characteristic of this biotope may limit the establishment of Sargassum muticum because of the lack of hard surfaces for attachment, although the presence of pebbles, cobbles and shell could allow it to gain a foothold. However, evidence shows that Sargassum muticum can anchor to the seagrass matrix itself, to hardbodied organisms within the biotope, and become embedded in the sediment, and result in reduced Zostera shoot density (Firth et al., 2024). Although these alternative attachment points may allow some Sargassum muticum in soft bottomed Zostera habitats, this does not necessarily imply successful proliferation or sustained colonization. Hhould an expansion occur Sargassum muticum could reduce available habitat for Zostera marina and prevent its recovery (Den Hartog, 1987). Therefore, the resistance is assessed as ‘Low’, resilience as ‘Very low’, and overall sensitivity as ‘High’, but with ‘Low’ confidence.

Wakame, Undaria pinnatifida

Evidence

Undaria pinnatifida (Wakame or Asian kelp) is a large brown seaweed and an Invasive Non-Indigenous Species (INIS) that could out-compete native UK macroalgae species (Farrell & Fletcher, 2006; Thompson & Schiel, 2012; Brodie et al., 2014; Heiser et al., 2014; Arnold et al., 2016; Epstein & Smale, 2017, 2018; Kraan, 2017; Epstein et al., 2019a,b; Tidbury, 2020). Undaria pinnatifida originates from Japan but is currently established on the coastlines of New Zealand, Australia, Northern France, Spain, Italy, the UK, Portugal, Belgium, Holland, Argentina, Mexico, and the USA (De Leij et al., 2017). Undaria pinnatifida was first recorded in the UK in the Hamble Estuary in 1994 (Macleod et al., 2016) and has since proliferated along UK coastlines. One year after its discovery at the Queen Anne Battery marina, Plymouth, it became a major fouling plant on pontoons (Minchin & Nunn, 2014). Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound. In Plymouth Sound, Epstein et al. (2019b) found that within its depth range (+1 to –4 m), Undaria pinnatifida co-existed with seven species of canopy-forming brown macroalgae, including Laminaria hyperborea. Undaria pinnatifida seems to settle better on artificial substrata (e.g., floats, marinas or piers) than on natural rocky shores among local kelps (Vaz-Pinto et al., 2014). It is found predominantly in low intertidal to shallow subtidal habitats (Epstein et al., 2019b) and is significantly more abundant on artificial substrata compared to natural rocky substrata (Heiser et al., 2014; Epstein & Smale, 2018).

Undaria pinnatifida has a wide physiological niche, meaning it can occur in both coastal and estuarine environments, but has a preference for full salinity ranges, 27 to 33 psu, and displays tolerance for varying salinities, turbidity and siltation (Heiser et al., 2014; Epstein & Smale, 2018). Undaria pinnatifida has a greater preference for sites sheltered with low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018). In natural habitats, Undaria pinnatifida was not recorded if the wave fetch was greater than 642 km and increased in abundance and cover in very sheltered sites (Epstein & Smale, 2018).

Sensitivity assessment. Since Undaria pinnatifida prefers fully saline and sheltered conditions, and overlaps with the depth range of this biotope, there is a possibility of it interacting with Zostera marina. However, as Undaria pinnatifida distribution is limited by the availability of hard substrata, the sand, muddy fine sand or mud in this biotope would not provide suitable attachment, and it would be limited to the pebbles, cobbles, or gravel in the sediment, which would probably mitigate its colonization and effect on seagrasses. Therefore, resistance is assessed as ‘High’, resilience and ‘High’, and the biotope is probably ‘Not sensitive’, albeit with ‘Low’ confidence.

Other INIS

Evidence

The effects on native species on seagrass species were reviewed by d’Avack et al. (2014) and Tillin et al. (2020). 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: 

The cord grass Spartina anglica. Spartina anglica is a 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. This species is 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.

Bonnemaison’s hook weed, Bonnemaisonia hamifera. This species can be found in salinities of 14.26 to 37.55 psu, is usually found growing as epifauna on macroalgae in the lower littoral down to 20 m, and has been reported in seagrass habitats within the north-east Atlantic (Tillin et al., 2020), Its ‘Trailliella’ phase prefers very sheltered conditions, such as those in which this biotope can be found. Examples of this species have been seen growing on Zostera marina blades (Johnson et al., 2005, cited in Tillin et al., 2020), making this biotope suitable habitat for Bonnemaisonia hamifera.

A red seaweed, Gracilaria vermiculophylla (syn. Agarophyton vermiculophyllum). This species is known to tolerate a wide salinity range of 5 to 60 psu and inhabit shallow habitats in sheltered areas (Tillin et al., 2020). It is known to form dense mats which could smother and outcompete seagrasses for space and light, though these mats may only be short-lived (Tillin et al., 2020). It has been observed growing amongst Zostera marina beds in Denmark (Tillin et al., 2020) and Zostera noltii beds in Portugal (Marin-Aragón et al., 2024). Agarophyton vermiculophyllum has also been reported to negatively impact Zostera marina through reductions in photosynthesis rate and survival (Martinez-Luscher and Holmer, 2010, as cited in Tillin et al., 2020). High abundances of this red alga was also shown to reduce Zostera noltii biomass, likely through shading and reduced water movement resulting from its thick canopy (Vieira et al., 2020). Examples of this biotope in sheltered areas are suitable habitat for the colonization of Agarophyton vermiculophyllum.

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. This species naturally occurrs on the east coast of North America but was 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.

Chinese mitten crab, Eriocheir sinensis. This species is found in mud and mixed sediment habitats of both marine and estuarine nature, thus tolerating a range of salinities. Eriocheir sinensis favours areas of shallow submerged vegetation and low energy habitats sheltered from wave exposure, making this biotope potentially suitable habitat for this species (Tillin et al., 2020). In estuarine conditions, seagrass beds may be susceptible to the same kind of grazing pressure that this species imposes on freshwater vascular plants, as well as uprooting and burial which disrupts the sediment (Tillin et al., 2020).

Orange striped anemone, Diadumene lineata. This species tends to be found in brackish waters, particularly in bays, estuaries, and marinas where its only requirement is suitable attachment substrata. It can tolerate a large salinity range, from 0.5 to 35 ppt, and is found in shallow waters to depths of a few hundred meters (Cohen, 2011), preferring sheltered areas with low wave exposure (Fofonoff et al., 2003). This species has not been shown to cause negative impacts on the habitats that it colonizes (Fofonoff et al., 2003). The suitable depths, salinity and low energy environments this biotope can be found in, couple with suitable attachment substrata of Zostera blades, makes this biotope potentially suitable habitat for Diadumene lineata (Tillin et al., 2020).

Asian rapa whelk, Rapana venosa. This species colonizes subtidal habitats up to 90 m deep and in salinities of 16 to 35 ppt (Tillin et al., 2020). It can live in a variety of substrata and has been recorded colonizing Zostera beds (Culha et al., 2009, cited in Tillin et al., 2020). Its preferred wave exposure and tidal currents are not known. This species predates on numerous filter feeding species, therefore, reductions in the extent of this functional group within seagrass beds may result in more turbid waters, thus reducing light availability for seagrasses. As such, the impact of colonization of seagrass beds by Rapana venosa has been assessed as ‘Moderate’ but with low confidence (Tillin et al., 2020), and this biotope is considered potentially suitable habitat for Rapana venosa (Tillin et al., 2020).

American lobster, Homarus americanus. This species can be found inhabiting a variety of substratum including muddy sediments as found in this biotope. Adults have an optimal salinity range of 30 to 35 ppt but can survive in salinities over 25 ppt, and are commonly found down to 50 m. Their wave exposure tolerance is not known. This species is known to dig into sediment which could increase turbidity in seagrass beds, as well as cause uprooting of rhizomes and displace the seed bank, however, there is no evidence of such impacts on seagrass beds. Given the suitable depth, salinity, and sediment, as well as offering shelter and camouflage within the seagrass blades, where this biotope is found in the subtidal can be considered suitable habitat for Homarus americanus (Tillin et al., 2020).

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). 

Sensitivity assessment. While this biotope may confer potentially suitable habitat for several INIS, little direct evidence of their occurrence on Zostera marina habitats around the UK and Ireland nor their effects was found. Therefore, there is ‘Insufficient evidence’ from which to assess the sensitivity of this biotope to these species.