The Marine Life Information Network

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 angiosperms. 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 marine plants, temperature affects biological processes by increasing the 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). 

Ruppia maritima occupies shallow waters, creating meadows at depths of 0.5 to 2 m (Halabowski, Sowa & Krodkiewska, 2018). Verhoeven (1979) noted that Ruppia maritima plants survived between 0 and 38°C, grew exponentially between 10 and 30°C and withstood fluctuations of 15°C in laboratory experiments. For example, Ruppia maritima has been observed in an isolated inland coal mine settling pond in Poland that experiences a temperature range of 7.0 to 20.1°C (Halabowski, Sowa & Krodkiewska, 2018), in the coastal and estuarine waters of Florida Bay, which experience low temperatures of 3.0 to 6.6°C and highs >30°C (Strazisar et al., 2015), and off the coast of Thailand between 27 and 33°C (Rasmusson et al., 2021). However, temperatures above 30°C were harmful if sustained for prolonged periods of time, and Ruppia maritima was replaced by 

Stuckenia (syn. Potamogeton) pectinatus grows in high-temperature environments, such as in the vicinity of thermal effluent (Kantrud, 1991). Anderson (1969), cited in Arnold et al. (2017), sampled Ruppia maritima from an industrial thermal plume on the Patuxent River (Maryland, USA) and found that the upper lethal temperature for the species was 45°C. However, temperatures above 23 to 25°C had a negative influence on photosynthesis.

In a lab experiment using Ruppia maritima samples from Thailand, where plants were incubated for five days at 43°C (a 15°C increase from their reported ambient temperature of 28°C), Ruppia maritima displayed sensitivity to both direct and prolonged temperature stress with a drastic decrease in the photosynthetic parameters leading to mortality (Rasmusson et al., 2021). Therefore, populations of similar Ruppia maritima in tropical environments may suffer a decline as temperatures increase and their productivity is highly reduced. High temperatures have also been documented to affect Ruppia maritima reproduction. Optimal seed germination occurs at 15 to 20°C, and in Europe, seed germination is observed to occur at temperatures beginning at 16°C, but only after a period of cold stratification at 2 to 4°C (Arnold et al., 2017). If colder climates, such as Chesapeake Bay, USA, become more subtropical, it may not be cold enough for Ruppia spp. to reproduce by seed, reducing overall population resilience (Arnold et al., 2017).

As temperatures increase alongside salinity, Ruppia maritima seedling and adult genet survival was documented decreasing, with few genets surviving in the hot, salty water of the western region of Florida Bay, U.S.A. (>36 psu and 35°C) (Strazisar et al., 2021). A similar effect was also observed in another seagrass, Ruppia sinensis (collected from Northern China), where tests were carried out on the plant alongside increasing temperature (0 to 30°C) and increasing salinity (0 to 50 psu) (Gu et al.,2018). The authors reported that optimum seed germination occurred at 30°C and a salinity of 5 psu, while low temperature (0°C) and high salinity (salinity 40 to 50) significantly inhibited seed germination.  Seed germination showed a bimodal pattern at suitable salinities (5 to 10) with increasing temperature, and storing seeds at low temperatures (0°C) or high salinities (40 to 50) promoted germination after transferal to optimal germination conditions (Gu et al., 2018). 

However, it has been suggested that Ruppia’s wide temperature tolerance may make it a ‘winner’ in a warmer climate, and increases in temperature may provide opportunities for Ruppia maritima to expand current populations over competitor seagrasses (Arnold et al., 2017). For example, over a seven-year period, the York River in Chesapeake Bay was studied to understand the effect of short-term temperature stress on seagrasses. Shields, Parrish & Moore (2019) observed that the more heat-tolerant Ruppia maritima replaced Zostera marina in response to increasing temperatures after Zostera marina die-off events. Yet Ruppia maritima was also observed retreating again once Zostera marina began to recover (Shields, Parrish & Moore, 2019).

Hensel et al. (2023), also observing the changes in seagrass communities in Chesapeake Bay, documented over a 35-year period (combining survey and modelling data) that, due to repeated marine heatwaves, since 1991, there has been a 54% retraction of the formerly dominant eelgrass, Zostera marina, allowing for a 171% expansion of the temperature-tolerant widgeongrass, Ruppia maritima. Ruppia maritima has a 7.4°C higher optimal growing temperature than Zostera marina, which aids it in outcompeting Zostera marina during the summer months when heatwaves are at their peak (Hensel et al., 2023). They concluded that Ruppia maritima is responsible for rapid, extensive meadow recoveries (Hensel et al., 2023). This expansion and retraction of Ruppia maritima populations have been observed elsewhere, such as in San Diego, California, USA, during an El Niño Southern Oscillation period in which daily maximum water temperatures increased between 1.5 and 2.5°C, in which Zostera marina declined, and Ruppia maritima expanded into areas that were previously dominated by Zostera marina (Johnson et al., 2003 cited in Shields, Parrish & Moore, 2019). Richardson, Lefcheck & Orth (2018) modelled the abundance of foundational seagrass species in response to warming in Chesapeake Bay, USA. They found that the mean cover of Ruppia maritima increased over the last three decades, from 6.8% in the 1990s to 7.5% in the 2000s, and then to 11.4% in the 2010s. In addition, the response of Ruppia maritima to temperature was curvilinear (having a non-linear relationship with water temperature), at first slightly increasing (from 25°C) and then decreasing steeply (from 26°C) with temperature, but then increasing steeply and becoming positive when temperatures exceeded 27.5°C (Richardson, Lefcheck & Orth, 2018).

The presence of Ruppia cirrhosa has been reported from Europe to Central and South America, including bordering most countries in the Mediterranean Sea, where Ruppia maritima and Ruppia drepanensis are much less common (Mannino et al., 2015). Ruppia cirrhosa has a temperature range of 5 to 30°C (Mannino et al., 2015), with an upper thermal tolerance of 34 to 35°C and a growth optimum of 20 to 30°C (Tsioli et al., 2019). Ruppia cirrhosa specifically from Sweden have been observed in temperatures between −4 and 24°C (Rasmusson et al., 2021). In a lab experiment using Ruppia cirrhosa samples from Sweden, plants were incubated for five days at 33°C (a 15°C increase from their reported ambient temperature18°C), Ruppia cirrhosa displayed a minor response to increases in temperature (Rasmusson et al., 2021). Thus, indicating that in predicted warming scenarios, Swedish Ruppia cirrhosa exhibited high tolerance to increasing temperature, may sustain efficient photosynthesis and potentially outcompete more heat-sensitive species.

Species living among Ruppia found in lagoons and shallow lochs are probably adapted to fluctuating temperatures, while mobile species are likely to move to deeper waters. Benthic infauna are likely to be protected from temperature extremes by their benthic habit. However, a proportion of the Arenicola marina population may be lost at temperatures above 20°C and excluded from habitats suffering from more extreme fluctuations in temperature.

Sensitivity assessment. A 5°C change in temperature for one month or a 2°C change in temperature for one year is unlikely to severely affect Ruppia plants. Therefore, resistance is assessed as 'High', resulting in a 'High' resilience score (no impact to recover from). Ruppia maritima has a very wide temperature tolerance and is deemed ‘Not sensitive’ to this pressure at the benchmark level.

Temperature decrease (local)

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

Evidence

A decrease in temperature is likely to delay the onset of budding and germination and subsequent reproduction in Ruppia spp., which may be of particular importance for annual species (see above). Verhoeven (1979) noted that all Ruppia taxa survive between 0-38°C, grow exponentially at 10-30°C and survive daily fluctuations of 15°C in culture. For example, Ruppia maritima has been observed in an isolated inland coal mine settling pond in Poland that experiences a temperature range of 7.0 to 20.1°C (Halabowski, Sowa & Krodkiewska, 2018), in the coastal and estuarine waters of Florida Bay, which experience low temperatures of 3.0 to 6.6°C and highs >30°C (Strazisar et al., 2015), and off the coast of Thailand between 27 and 33°C (Rasmusson et al., 2021). Kantrud (1991) reported that in North American wetlands that freeze in winter, Ruppia spp. behaved as annuals. Verhoeven (1979) reported that the distribution of Ruppia maritima and Ruppia cirrhosa extended north to Norway (ca 69 deg N and 68 deg N, respectively), suggesting that these species could be tolerant of the average winter temperatures encountered in the UK.

In Ruppia maritima, optimal seed germination occurs at 15 to 20°C, and in Europe, seed germination is observed to occur at temperatures beginning at 16°C, but only after a period of cold stratification at 2 to 4°C (Arnold et al., 2017). If colder climates become colder, it may not be warm enough for Ruppia spp. to reproduce by seed, reducing overall population resilience (Arnold et al., 2017). Strazisar et al. (2015) studied the life history of Ruppia maritima in the costal and estuarine waters of Florida Bay, USA, and found that even very short durations of low temperatures (minimum of 3.0 to 6.6°C and average of 9.2 to 9.3 °C for 3 days) or salinity variability at all sites was adequate to induce seed germination, and did not inhibit it. In addition, Gu et al. (2018) found that in the seagrass Ruppia sinensis, low temperature (0°C) and high salinity (salinity 40 to 50) significantly inhibited seed germination, while storing seeds at low temperatures (0°C) or high salinities (40 to 50) promoted germination after transfer to optimal germination conditions. 

The presence of Ruppia cirrhosa has been reported from Europe to Central and South America, including bordering most countries in the Mediterranean Sea, where Ruppia maritima and Ruppia drepanensis are much less common (Mannino et al., 2015). Ruppia cirrhosa have a temperature range of 5 to 30°C (Mannino et al., 2015), with an upper thermal tolerance of 34 to 35°C and a growth optimum of 20 to 30°C (Tsioli et al., 2019), and Ruppia cirrhosa specifically from Sweden have been observed in temperatures between −4 and 24°C (Rasmusson et al., 2021).

Many of the species found within the Ruppia spp. communities are typically lagoonal or shallow water species, adapted to fluctuating temperatures. Infaunal polychaetes are protected from temperature extremes by their burrowing habit, however, a proportion of the Arenicola marina population may be lost below 5°C in areas subject to extreme fluctuations in temperature.

Sensitivity assessment. Overall, the Ruppia spp. stand will not be damaged by a decrease in temperature at the benchmark level, but some species will decrease in abundance, while mobile species may move to deeper water, resulting in a reduced species richness. A 5°C change in temperature for one month or a 2°C change in temperature for one year is unlikely to severely affect Ruppia plants. Resistance is therefore assessed as 'High', resulting in a 'High' resilience score (no impact to recover from). Ruppia maritima has a very wide temperature tolerance and is deemed ‘Not sensitive’ at the benchmark level.

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

Ruppia maritima has a wide salinity tolerance and is found growing in full saline, brackish, and freshwater environments. Ruppia spp. tolerate a wider range of ionic strengths and salinities than any other aquatic angiosperm, occurring between 0.6 and 390 g/l (Kantrud, 1991; Hillmann, DeMarco & La Peyre, 2019; Lee et al., 2026). However, the reported salinity tolerances vary with region and with species. Ruppia maritima was reported to be abundant at salinities between 15 and >100 g/l in North American wetlands and between 0.57 and 27 g/l in European sites and has a general salinity range between 0 and >100 psu (Verhoeven, 1979; Kantrud, 1991; Hillmann, DeMarco & La Peyre, 2019; Lee et al., 2026). For example, the Florida Everglades, USA, experiences a range of salinities from fully freshwater to fully marine, yet Ruppia maritima is found at the ecotone (hydrological stochasticity and steep salinity gradients, particularly at the interface of freshwater-marine transitions) between the freshwater Everglades marshes and hypersaline Florida Bay, experiencing a salinity range of 0 to 70 psu (Strazisar et al., 2015). In Southeast Bulgaria, in Atanasovsko Lake off the Black Sea coast, Ruppia maritima was observed in hypersaline conditions (>40‰) in shallow (<1 m) polymictic waters, with eutrophic to polytrophic conditions (Gecheva et al., 2017). Similarly, in Indonesia, Ruppia spp. are observed in the intertidal zone of the Danau Laut Mati, or Dead Sea Lake, where salinity ranges between 40 and 42 ppt, both vertically in-depth and temporally (Kurniawan et al., 2024). In some sites, the high salinity tolerance of Ruppia maritima excluded other low-saline-tolerant plant species, allowing a dense population to develop, as was the case in an inland coal mine setting pond in Poland, where salinity ranged between 14.2 and 26.5 psu (Halabowski, Sowa & Krodkiewska, 2018). Despite this tolerance to high salinities, Ruppia maritima exhibited its maximum biomass in freshwater and a secondary biomass peak in mesohaline water (Rodríguez-Gallego et al., 2015).

Strazisar et al. (2015) found that seedling survival in Ruppia maritima fluctuated with salinity, with only a 13% survival towards more saline/marine waters under rapid salinity changes. Adult survival also decreased from less saline (93%) to higher saline (25%) sites, resulting in a significant decline in per capita clonal reproduction rates. The low survival rate observed was associated with rapid salinity fluctuations (2.5 to 20 psu) with short periodicities (<24 hours). Adult survival was mostly negatively associated with repeated fluctuations of <72 hours, and, in addition to the low seedling survival observed, no sexual reproduction occurred at any of the ecotone sites (Strazisar et al., 2015). Yet, germination rates were quite high at all sites (≥33%), indicating that no site inhibited germination of Ruppia maritima seeds across the ecotone in this study. Furthermore, total biomass (shoot, root, and rhizome) was over two times greater at the upper and central sites, where salinity fluctuated less, compared with the lower site (12, 12, and 5 mg/shoot, respectively), which indicated that vegetative adults at the upper sites had also increased their growth relative to the lower ecotone site (Strazisar et al., 2015). Based on the data from Strazisar et al. (2015), Ruppia maritima seedling and adult survival are negatively associated with high frequencies of salinity pulses (2.5 to 20 psu), particularly on the overall development of seeds to vegetatively reproductive adults between 12 and 48 hours. Conversely, less extreme salinity fluctuations correlated positively with the overall transition from seed to vegetatively reproductive adult (a 2.5 to 5 psu change for >48 to 72 hours, and a >5 to10 psu change for >72 to 96 hours). Modest salinity fluctuations did not inhibit Ruppia maritima life history development and likely explained this species’ relative dominance in transitional waters (Strazisar et al., 2015).

As temperatures increase alongside salinity, Ruppia maritima seedling and adult genet survival was documented to decrease, with few genets surviving in the hot, salty water of the western region of Florida Bay, USA (>36 psu and 35°C) (Strazisar et al., 2021). A similar effect was also observed in Ruppia sinensis (collected from Northern China), where tests were carried out on the plant alongside increasing temperature (0 to 30°C) and increasing salinity (0 to 50 psu). Gu et al. (2018) found that optimum seed germination occurred at 30°C and a salinity of 5 psu, and that low temperature (0°C) and high salinity (salinity 40 to 50) significantly inhibited seed germination. Seed germination showed a bimodal pattern at suitable salinities (5 to 10) with increasing temperature, and storing seeds at low temperatures (0°C) or high salinities (40 to 50) promoted germination after transferal to optimal germination conditions (Gu et al., 2018). 

Ruppia cirrhosa tolerates a wide range of salinity (1.5 to 60 psu) and has a wider optimum salinity range than that of Ruppia maritima (up to 60 psu compared to 0.3 to 15 psu) (Mannino et al., 2015). In European sites, Ruppia cirrhosa is documented to tolerate salinity levels of 2.7 to 108.3 g/l (Verhoeven, 1979). For example, Ruppia cirrhosa found in the Eastern Macedonian and Thrace Region of Greece were euryhaline, found in waters ranging from 5 to 55 or 60 psu (Tsioli et al., 2019).

Kantrud (1991) concluded that the optimum salinity for Ruppia spp. growth was 5 to 20 g/l, while slightly lower salinities early in spring may enhance germination and seed formation. However, high salinity alongside a high pH was demonstrated to sharply reduce the relative growth rate of Ruppia maritima as well as reduce enzyme activity and its nutrient removal efficiency (Chen et al., 2017). Rapid fluctuations were found to kill Ruppia spp. when salinities rise ca >18 g/l in a few weeks (Verhoeven, 1979). However, Ruppia spp. was also reported to survive a drop of at least 14 g/l in 24 hrs (Kantrud, 1991). Overall, Ruppia spp. are probably not directly sensitive to changes in salinity at the benchmark level. Their exclusion from very low to freshwater, or nearly full seawater, is probably due to competitive exclusion by other aquatic plants, e.g. Stuckenia (syn. Potamogeton) pectinatus.

As the salinity increases, low salinity species are likely to be replaced by comparable marine forms. Typically, lagoonal species (e.g. the hydrobids, some gammarids, and Cerastoderma glaucum) are adapted to a wide range of salinities and are unlikely to be affected. Estuarine and low salinity polychaetes present in the benthos are likely to be replaced by more marine species as the salinity increases, e.g. the abundance of Hediste diversicolor is likely to fall while the abundance of Arenicola marina may increase. Sticklebacks are found in marine and freshwater habitats, and the sand goby tolerates a wide range of salinities. Therefore, the biotope as a whole will probably be little affected by increases in salinity at the benchmark level, although some species may be replaced by more marine members of the same group.

Sensitivity assessment. Ruppia maritima exclusion from freshwater or nearly full seawater habitats is most probably due to competitive exclusion by other aquatic plants. An increase in salinity from 35 to >40 units for one year is unlikely to severely affect Ruppia plants. Resistance is therefore assessed as 'High', resulting in a 'High' resilience score (no impact to recover from). Overall, Ruppia maritima has a very wide salinity tolerance and is assessed as ‘Not sensitive’ to an increase in salinity at the benchmark level.

Salinity decrease (local)

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

Evidence

Ruppia maritima has a wide salinity tolerance and is found growing in full saline, brackish, and freshwater environments. Ruppia spp. tolerate a wider range of ionic strengths and salinities than any other aquatic angiosperm, occurring between 0.6 and 390g/l (Kantrud, 1991; Hillmann, DeMarco & La Peyre, 2019; Lee et al., 2026). However, the reported salinity tolerances vary with region and with species. Ruppia maritima was reported to be abundant at salinities between 15 and >100g/l in North American wetlands and between 0.57 and 27g/l in European sites and has a general salinity range between 0 and >100 psu (Verhoeven, 1979; Kantrud, 1991; Hillmann, DeMarco & La Peyre, 2019; Lee et al., 2026). For example, the Florida Everglades, USA, experiences a range of salinities from fully freshwater to fully marine, yet Ruppia maritima is found at the ecotone between the freshwater Everglades marshes and hypersaline Florida Bay, experiencing a salinity range of 0 to 70 psu (Strazisar et al., 2015). In Southeast Bulgaria, in Atanasovsko Lake off the Black Sea coast, Ruppia maritima was observed in hypersaline conditions (>40‰) in shallow (<1 m) polymictic waters, with eutrophic to polytrophic conditions (Gecheva et al., 2017). Similarly, in Indonesia, Ruppia spp. are observed in the intertidal zone of the Danau Laut Mati, or Dead Sea Lake, where salinity ranges between 40 and 42 ppt, both vertically in-depth and temporally (Kurniawan et al., 2024). In some sites, the high salinity tolerance of Ruppia maritima excluded other low-saline-tolerant plant species, allowing a dense population to develop, as was the case in an inland coal mine setting pond in Poland, where salinity ranged between 14.2 and 26.5 psu (Halabowski, Sowa & Krodkiewska, 2018). Despite this tolerance to high salinities, Ruppia maritima exhibited its maximum biomass in freshwater and a secondary biomass peak in mesohaline water (Rodríguez-Gallego et al., 2015).

Strazisar et al. (2015) found that seedling survival in Ruppia maritima fluctuated with salinity, with only a 13% survival towards more saline/marine waters under rapid salinity changes. Adult survival also decreased from less saline (93%) to higher saline (25%) sites, resulting in a significant decline in per capita clonal reproduction rates. The low survival rate observed was associated with rapid salinity fluctuations (2.5 to 20 psu) with short periodicities (<24 hours). Adult survival was mostly negatively associated with repeated fluctuations of <72 hours, and, in addition to the low seedling survival observed, no sexual reproduction occurred at any of the ecotone sites (Strazisar et al., 2015). Yet, germination rates were quite high at all sites (≥33%), indicating that no site inhibited germination of Ruppia maritima seeds across the ecotone in this study. Furthermore, total biomass (shoot, root, and rhizome) was over two times greater at the upper and central sites, where salinity fluctuated less, compared with the lower site (12, 12, and 5 mg/shoot, respectively), which indicated that vegetative adults at the upper sites had also increased their growth relative to the lower ecotone site (Strazisar et al., 2015). Based on the data from Strazisar et al. (2015), Ruppia maritima seedling and adult survival are negatively associated with high frequencies of salinity pulses (2.5 to 20 psu), particularly on the overall development of seeds to vegetatively reproductive adults between 12 and 48 hours. Conversely, less extreme salinity fluctuations correlated positively with the overall transition from seed to vegetatively reproductive adult (a 2.5 to 5 psu change for >48 to 72 hours, and a >5 to 10 psu change for >72 to 96 hours). Modest salinity fluctuations did not inhibit Ruppia maritima life history development and likely explained this species’ relative dominance in transitional waters (Strazisar et al., 2015). However, La Peyre & Rowe (2003) found that the relative growth rate of Ruppia maritima was significantly lowered during a short experimental freshwater pulse but did not report any mortality. 

Kantrud (1991) concluded that the optimum salinity for Ruppia spp. growth was 5 to 20 g/l, while slightly lower salinities early in spring may enhance germination and seed formation. Rapid fluctuations were found to kill Ruppia spp. when salinities rise ca >18 g/l in a few weeks (Verhoeven, 1979). However, Ruppia spp. was also reported to survive a drop of at least 14 g/l in 24 hrs (Kantrud, 1991). Overall, Ruppia spp. are probably not directly sensitive to changes in salinity at the benchmark level. Their exclusion from very low to freshwater, or nearly full seawater, is probably due to competitive exclusion by other aquatic plants, e.g. Stuckenia (syn. Potamogeton) pectinatus.

Ruppia cirrhosa tolerates a wide range of salinity, 1.5 to 60 psu, and has a wider optimum salinity range than that of Ruppia maritima (up to 60 psu compared to 0.3 to 15 psu) (Mannino et al., 2015). In European sites, Ruppia cirrhosa is documented to tolerate salinity levels of 2.7 to 108.3 g/l (Verhoeven, 1979). For example, Ruppia cirrhosa found in the Eastern Macedonian and Thrace Region of Greece were euryhaline, found in waters ranging from 5 to 55/60 psu (Tsioli et al., 2019).

Most of the typically lagoonal species (e.g. Cerastoderma glaucum, Gammarus insensibilis and hydrobids) will be little affected by changes in salinity. However, Gammarus insensibilis was reported to disappear in areas affected by prolonged exposure to freshwater. Similarly, Arenicola marina does not tolerate salinities below 24 psu and is likely to be replaced by Hediste diversicolor. Sticklebacks are found in marine and freshwater habitats, and the sand goby tolerates a wide range of salinities. As the salinity decreases, the species composition is likely to change towards more freshwater-tolerant species, including insects, although the functional groups will probably remain, and the species richness may increase.

A short-term decrease in salinity is unlikely to affect the biotope adversely. However, prolonged exposure to low salinities or freshwater is likely to result in the replacement of the Ruppia community by other aquatic plant communities, e.g. Stuckenia (syn. Potamogeton) pectinatus. Once prior conditions return, recovery is likely to be rapid.

Sensitivity assessment. Ruppia maritima exclusion from freshwater or nearly full seawater habitats is most probably due to competitive exclusion by other aquatic plants. A decrease in salinity from 'reduced' to 'low' salinity for one year is unlikely to severely affect Ruppia plants due to competition. Resistance is therefore assessed as 'Medium' to represent the potential loss of extent or abundance, but once the salinity returns to prior levels, recovery is likely to be rapid, and resilience is probably 'High'. Overall, Ruppia maritima beds are assessed as ‘Low’ sensitivity to a decrease in salinity at the benchmark level.

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

The SS.SMp.SSGr.Rup biotope is found in extremely sheltered conditions in very weak tidal streams. An increase in water flow at the benchmark level  is likely to damage leaves and shoots and probably remove the vegetation and a proportion of the root system. The root system of Ruppia spp. is poorly developed consisting of horizontal runners a few millimetres below the sediment surface and only 1-2 thin roots per 10-20 cm along the rhizome. Therefore, Ruppia spp. are not very resistant of water flow and are limited to still, sheltered waters such as lagoons and bays where current flow is less than in adjacent channels and tidal rivers (Verhoeven, 1979; Kantrud, 1991). Verhoeven (1979) suggested that Ruppia maritima was particularly intolerant while Ruppia cirrhosa occurred in larger waters at more exposed but still sheltered sites. In addition, turbulent water flow resulting in resuspension of sediment can indirectly reduce Ruppia productivity due to increased turbidity (see below). Kantrud (1991) reported that Ruppia spp. can occur in areas of 'considerable' current flow, e.g. Ruppia beds fertilized in situ with phosphorus were found to grow well in currents up to 4cm/s. However, 4cm/s is considered to be negligible (see benchmark). Epiphytes and algal mats would also be lost. 

Most of the benthic infauna are found in areas of stronger currents (e.g. Arenicola marina), and many of the mobile species (e.g. amphipods, isopods, shrimp, crabs and fish) would migrate to other suitable substrata or habitats. However, where present Cerastoderma glaucum is only found in areas of weak water flow and may be lost.

Sensitivity assessment. Any change in water movement will have a considerable impact on the integrity of seagrass habitat.  A change in water flow at the level of the benchmark of 10 to 20 cm/s for more than 1 year would remove of the Ruppia maritima plants by removal of the muddy substratum and drag on the plants. Resistance is thus assessed as ‘Low’.  Once the water flow regime returned to prior conditions then recovery would probably take 2-10 years, a ‘Medium’ resilience. Overall, Ruppia maritima has a ‘Medium’ sensitivity to this pressure. 

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

Ruppia maritima occurs in tidal areas, from mean high water (MHW) to mean low water (MLW). Kantrud (1991) reported that the grass is restricted to areas exposed for a maximum of four hours daily or approximately seven hours per low tide, but quickly disappears from areas emersed for extended periods. Verhoeven (1979) stated that the Ruppia spp. had a low tolerance to drought as plants (except ripe seeds) died within a few days of being exposed to aerial conditions. Changes in emergence regime and increased aerial exposure are thus likely to result in reduced growth, productivity and the loss of the upper portion of the population.

Gu et al. (2019) conducted an in-situ investigation of the influence of desiccation on sediment seed banks and population recruitment of the seagrass Ruppia sinensis in the Yellow River Delta, China. Although Ruppia sinensis was significantly impacted by the desiccation of its habitat, many sediment seeds survived the long-term desiccation of their environment, and desiccation did not completely degrade the population. However, desiccation markedly delayed seed germination and seedling establishment. Gu et al. (2019) concluded that, unlike most seagrass species, Ruppia sinensis seeds in sediment exhibited considerable tolerance and adaptability to desiccation of their habitat. 

Other components of the community might also be affected. For instance, Cerastoderma glaucum is thought to be intolerant of changes in emergence and may be lost. Hydrobia spp. on the other hand, inhabit salt marshes and are tolerant of emersion. Gammarids and isopods can either migrate to deeper water, burrow in the sediment, or find shelter in damp weed to avoid aerial exposure. Algal mats retain water, and while their surface may be bleached or desiccated in hot weather, they are likely to recover quickly.

Sensitivity assessment. Seagrass growing in intertidal habitats have greater tolerance to exposure to air than species inhabiting subtidal beds. Changes in emergence regime may, however, cause some mortality in Ruppia plants, resulting in a reduction in the upper shore extent. Resistance is therefore assessed as ‘Medium’. Recovery will be enabled by recolonization from surrounding communities located further down the shore and via the remaining seed bank. Recovery is therefore considered to be fairly rapid, resulting in a ‘Medium’ resilience score. The biotope has a ‘Medium’ sensitivity to this pressure.

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

McCann (1945) noted that waves caused injury to Ruppia branches, leaving broken tips incapable of survival, and Verhoeven (1979) observed that the base of leaves detached easily in turbulent water to avoid damage to the root system. However, the root system is weak (see water flow), and Ruppia beds are restricted to areas protected from wave action and with little fetch and wind-induced water turbulence. Wave action also resuspends sediment, increasing turbidity and hence reducing productivity. This biotope (SS.SMp.SSgr.Rup) is found in extremely sheltered areas, therefore, an increase in wave action is likely to remove the surface vegetation and the majority of the root system. Ruppia maritima has a less extensive root and rhizome system than, for example, Zostera marina, increasing its susceptibility to mechanical disturbance and increased wave energy (Hensel et al., 2023). However, Ruppia maritima beds along the mainland coastline of eastern Mississippi Sound, USA, did not show any short-term declines in their growth or abundance following 2005 hurricanes Katrina and Rita (Cho et al., 2017).

Most lagoonal species are adapted to sheltered conditions and are likely to be adversely affected by increases in wave exposure, e.g. Gammarus insensibilis and Cerastoderma edule, resulting in loss of a proportion of the population. The resident gastropods, e.g. Hydrobia ulvae, are unlikely to be directly affected and will switch to alternative food supplies, however, should the increase in wave exposure be significant enough to change the sediment type, e.g. to coarse sands, they are likely to be lost. Benthic species, such as Arenicola marina, can tolerate sheltered to moderately exposed conditions and would probably be little affected at the benchmark level.

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

Sensitivity assessment. Exposure models from Studland Bay and Salcombe, where seagrass beds are limited to low wave exposure, show that even a change of 3% (in significant wave height) 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). The root system of Ruppia is more fragile than that of seagrass and, therefore, more likely to be damaged. At the benchmark level, an increase in wave exposure is likely to remove surface vegetation and the majority of the root system, causing mortality. Hencce, resistance is assessed as ‘Low’. 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

This pressure is Not assessed but evidence is presented where available.

Little information concerning the effects of non-synthetic compound contamination on Ruppia maritima was found. However, Ruppia maritima have been found growing in a polluted coal mine settling pond in Poland, which contained heavy metal contaminants (Halabowski, Sowa & Krodkiewska, 2018), and it was noted how heavy metals can cause some deformation in Ruppia maritima, with nickel, for example, reducing plant growth (Bhalerao et al., 2015 cited in Halabowski, Sowa & Krodkiewska, 2018). Yet, this may not be usual, as compared to previous studies, Halabowski, Sowa & Krodkiewska (2018) did not find the presence of Ruppia maritima in sediments with similar levels of heavy metal contamination.

Other seagrasses, such as Zostera marina, are known to accumulate TBT, but no detrimental effects were observed in the field (Williams et al., 1994). Naphthalene, Pentachlorophenol, Aldicarb and Kepone reduce nitrogen fixation and may affect Z. marina viability. TBT contamination is likely to adversely affect grazing gastropods, resulting in increased algal growth, reduced primary productivity and potential smothering of the biotope. Bryan & Gibbs (1991) suggested that TBT may cause reproductive failure or larval mortality in bivalve molluscs, e.g. Pecten maximus at ca. 50 ng/l TBT.

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

This pressure is Not assessed but evidence is presented where available.

Little information was found on the effects of hydrocarbon contamination on Ruppia plants. However, the implications of oil exposure on Ruppia maritima were studied following the 2010 Deepwater Horizon oil spill. Martin, Hollis & Turner (2015) grew Ruppia maritima in a range of manipulated sediment oil concentrations (0, 0.26, 0.53, and 1.05 ml oil/l tank volume) and found that while no statistical differences were detected in growth, plants exhibited significant changes to reproductive output, root morphology, and uprooting force. Significant reductions in inflorescences and fruiting bodies at higher oil concentrations were observed in addition to roots growing in high oil being shorter and wider, and plants in medium and high oil required less force to uproot (Martin, Hollis & Turner, 2015). Martin, Hollis & Turner (2015) concluded that, given the importance of sexual reproduction for Ruppia maritima, oil contamination may have substantial population-level effects, and areas containing buried oil may be more susceptible to high-energy storm events due to the reduction in uprooting force. However, Ruppia maritima beds along the mainland coastline of eastern Mississippi Sound, USA, did not show any short-term declines in their growth or abundance following the 2010 Deepwater Horizon oil spill. (Cho et al., 2017). Furthermore, Kenworthy et al. (2017) found that Ruppia maritima in the Chandeleur Islands, Louisiana, USA, was exposed to Deepwater Horizon oil, with evidence of oil-related compounds detected in its tissues and surrounding sediments. Despite this exposure, Ruppia maritima demonstrated resilience and was the most frequently encountered species in the Chandeleur Islands during post-exposure sampling in June 2011. It showed high densities and widespread distribution, with flowering plants observed and canopies occupying the entire water column up to the surface.

Healthy populations of Zostera spp. have been observed in the presence of long-term, low-level, hydrocarbon effluent, for example, in Milford Haven, Wales (Hiscock, 1987). The Amoco Cadiz oil spill off Roscoff caused Zostera marina leaves to blacken for 1-2 weeks but had little effect on growth, production or reproduction after the leaves were covered in oil for six hours (Jacobs, 1980). The Amoco Cadiz oil spill did, however, result in the virtual disappearance of Amphipods, Tanaidacea and Echinodermata from Zostera marina beds and caused a decrease in numbers of Gastropoda, sedentary Polychaeta and Bivalvia. The numbers of most groups returned to normal within a year, except Echinoderms, which recovered more slowly and Amphipods, which did not show any signs of recovery (Jacobs, 1980).

Removal of oil-intolerant gastropod grazers may result in smothering of seagrasses by epiphytes (Davison & Hughes, 1998). 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.

Experimental treatment of Zostera spp. with crude oil and dispersants halted growth but had little effect on cover, whereas pre-mixed oil and dispersant caused rapid death and significant decline in cover within 1 week, suggesting that dispersant treatments should be avoided (Davison & Hughes, 1998).

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

This pressure is Not assessed but evidence is presented where available.

Johnston & Bird (1995) determined that Ruppia maritima was able to better tolerate the effects of herbicides compared to other aquatic plants. Although photosynthesis was reduced by 0.05 mg/l atrazine after 35 days exposure, growth was reduced at atrazine concentrations < 5mg/l but continued at 10 mg/l. However, Kantrud (1991) reported that herbicides (atrazine and alachlor) in agricultural runoff reduced Ruppia growth and biomass in Chesapeake Bay and noted that 1.0 ppm of atrazine had been used to control Ruppia growth in wetlands. Only small numbers of plants survived 4 years after treatment with the herbicide 2,4 D-ester at 112 kg/ha (Kantrud, 1991). Cole et al. (1999) suggested that herbicides were, not surprisingly, very toxic to algae and macrophytes. Similarly, most pesticides and herbicides were suggested to be very toxic for invertebrates, especially crustaceans (amphipods, isopods, mysids, shrimp and crabs) as well as fish (Cole et al., 1999). For example, Lindane was shown to be very toxic to gobies (Gobius spp.: see Pomatoschistus minutus) (Ebere & Akintonwa, 1992) .

Synthetic chemicals found in agricultural, urban and industrial discharges are likely to adversely affect the biotope. Herbicides in particular are likely to reduce growth and productivity of Ruppia beds, and may result in mortality. In addition, loss of particularly intolerant crustaceans may result in unchecked growth of epiphytes, which would again reduce photosynthesis and productivity of Ruppia plants.

Radionuclide contamination

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

Evidence

No evidence 

Introduction of other substances

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

Evidence

This pressure is Not assessed.

De-oxygenation

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

Evidence

Limited information on the effects of oxygen concentration on the growth and survivability of Ruppia maritima is reported in the literature. Yet Ruppia maritima have been observed in waters with dissolved oxygen concentrations of 0.3 to 17.6 mg/l (Strazisar et al., 2015; Hillmann, DeMarco & La Peyre, 2019). Ruppia spp. favour aerobic sediments with low levels of sulphides and free H₂S, but will grow in reduced conditions, since the leaves supply oxygen to the roots. Roots of Ruppia maritima also show a higher threshold towards sulphides, 1.8 to5.9 mM H₂S (Pedersen & Kristensen, 2015). Senescence and loss of stems can coincide with increases in H₂S in the sediment and may be a factor regulating the decrease in Ruppia species in hot summer months (Kantrud, 1991). Germination may also be affected by oxygen levels, and seeds in poorly oxygenated sediments lie dormant until the next year (Kantrud, 1991). However, the presence of Ruppia in reduced sediment suggests that it would tolerate low oxygen levels comparable to the benchmark, especially since photosynthesis produces oxygen. For example, Ruppia maritima have been found growing in a polluted coal mine settling pond in Poland, which contained a dissolved oxygen concentration of 5.7 to 17.6 mg O₂/l (Halabowski, Sowa & Krodkiewska, 2018). In Florida Bay, USA, Ruppia maritima sites were monitored for evidence of hypoxia, but oxygen levels remained >6 mg/l and did not appear to be controlling species community shifts (Strazisar et al., 2015). However, the sediments Ruppia maritima occupies may be more aerobic due to the sediments' larger grain size compared to the sediments of other seagrasses, such as Zostera marina, and have been found in coarser sediments than Zostera in Goodwin Island in the Chesapeake Bay, USA (French et al., 2024).

In Ruppia spiralis, the accumulation of organic matter and the dominance of anaerobic bacterial processes in nutrient-rich environments can cause dystrophic events/crises (rapid environmental degradation), leading to hypoxia/anoxia, which further harms Ruppia spiralis by causing the detachment of fronds and stems and contributing to the decay of the plant (Lenzi & Cianchi, 2022).

Mud snails are relatively tolerant of reduced hypoxic muds and can tolerate aerial exposure for over a week, suggesting that they are capable of anaerobic respiration. Benthic infaunal species are probably tolerant of hypoxia, e.g. Arenicola marina, which can tolerate 9 days without oxygen (Hayward, 1994) and Cerastoderma glaucum, which tolerates 84 hrs in the absence of oxygen (Boyden, 1972). Most polychaetes are capable of anaerobic metabolism, while mobile fish and gobies migrate out of the affected area in response to decreasing oxygen levels (Diaz & Rosenberg, 1995). Small mobile shrimp, amphipods and isopods will probably also migrate out of the affected area. 

Sensitivity assessment. Therefore, the Ruppia stands, and benthic infauna will probably tolerate hypoxia at the level of the benchmark, but increased epiphyte growth due to reduced numbers, but not loss of grazers, may reduce Ruppia spp. productivity. However, species richness is likely to decline. Therefore, a resistance of 'High' is recorded, with a resilience of 'High', and resultant sensitivity of 'Not sensitive' at the benchmark level.

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

Ruppia maritima seems to display some tolerance to nutrient enrichment. For example, Ruppia maritima has been found growing in a polluted coal mine settling pond in Poland that contained a nitrate and phosphate concentration of 3.1 to 12.0 mg NO₃/l and 0.001 to 1.2 mg PO₄³/l, respectively, with the maximum concentrations of nitrates higher than those that have been reported in previous research (0 to 1.0 mg NO₂/l) (Halabowski, Sowa & Krodkiewska, 2018). In mesocosm experiments, Ruppia maritima was shown to increase shoot production by >300% over controls after the addition of 10 µM water column NO₃-N /day (Burkholder et al., 1994). Burkholder et al. (1994) went on to suggest that Ruppia maritima and Halodule wrightii effectively control nitrate uptake and could be transplanted to replace Zostera marina in nitrate-enriched waters where the eelgrass had disappeared. Therefore, it appears that Ruppia spp. will benefit from low nutrient enrichment. However, rapid recruitment and biomass development for Ruppia maritima requires high nutrient availability of both nitrogen (for vegetative growth) and phosphorus (for reproductive shoot development) (Strazisar et al., 2016). In addition, Strazisar et al. (2021) found that seeding survival of Ruppia maritima decreased under phosphorus-limited conditions. Ahmadi et al. (2017) suggested using Ruppia maritima as an advanced coastal treatment of saline municipal wastewater. Through an in-situ pilot study where plants were fed by activated sludge effluent (at Chobeineh wastewater treatment plant in Ahvaz city, Iran), Ruppia maritime exhibited a total nitrogen and phosphorus uptake of 44.48 to 45.78% and 63.6 to 73.01%, respectively, dependent on the level of electrical conductivity, 10, 15, and 20 ms/cm, with the highest growth and uptake rates observed at 10 ms/cm. Ahmadi et al. (2017) concluded that Ruppia maritima should be considered a promising saline-tolerant plant in the advanced treatment of saline wastewater.

In addition, Ruppia maritima may be more tolerable to fluctuations in nutrients compared to competing seagrasses like Zostera marina, rebounding after stressful water quality events and more suited to surviving in degraded conditions (Hensel et al., 2023; French et al., 2024). However, Ruppia maritima was a significant contributor to large seagrass population crashes in the Chesapeake Bay area since the 1980s due to nutrient pulses driven by springtime runoff, such as its population halving in 2019 (Hensel et al., 2023). Since Ruppia maritima became dominant in Chesapeake Bay in 1999, seagrass cover has become increasingly variable. Between 1999 and 2019, Chesapeake Bay has either gained or lost at least 20% total seagrass area from the previous year, and Ruppia maritima cover has been approx. three-times more variable than historical Zostera marina cover (Hensel et al., 2023). Therefore, Hensel et al. (2023) stated that Ruppia maritima is most negatively affected by increased watershed nutrient pollution, and at current nutrient levels in Chesapeake Bay, predicted a future amplification of boom–bust cycles in Ruppia maritima cover, alternating between high in dry years and low in wet years.

Similar to Ruppia maritima, Ruppia cirrhosa has been documented recovering from nutrient fluctuation over a relatively short period of time. A significant increase in aquatic angiosperms, such as Ruppia cirrhosa and Stuckenia pectinata, was observed four to five years after sewage improvements (changes in hydrological management and improved sewage treatment) reduced the pollution in a shallow, brackish coastal lagoon in Corsica, France (Biguglia lagoon) (Pasqualini et al., 2017). The Biguglia lagoon remains a major site of agricultural and urban activities, favouring large nitrogen inputs. However, the restored communities closely resemble pristine communities, which suggests that both Ruppia cirrhosa and lagoon habitats are resilient and able to recover from natural and anthropogenic disturbances (Pasqualini et al., 2017).

Nutrient enrichment is known to have indirect adverse effects. Nutrient enrichment stimulates epiphyte growth, which interferes with the nutrient exchange across the Ruppia leaves and shades out light, reducing primary productivity, growth and reproduction. Similarly, nutrients stimulate phytoplankton blooms that compete for nutrients, but more importantly, increase the turbidity (see water clarity). and absorb light, reducing Ruppia productivity. Twilley et al. (1985) found that epiphyte growth in nutrient-enriched conditions reduced the light incident on Ruppia leaves by >80%, resulting in significant decreases in macrophyte biomass at medium to high levels of enrichment (0.86 and 1.68 g N /m²/day, respectively). Ruppia maritima production collapsed after six weeks at high nitrogen levels. However, the epiphytic growth only resulted in loss of macrophytes due to the additional turbidity caused by the phytoplankton (Twilley et al., 1985). Kantrud (1991) concluded that while nutrients can stimulate growth, growth is severely limited by phytoplankton and epiphytes in eutrophic conditions.

Ruppia spp. can tolerate organic-rich sediments and has been reported to grow in extremely reduced sediments since leaves supply oxygen to the root system (Kantrud, 1991). However, Azzoni et al. (2001) noted that the oxygen supply to the roots detoxified the sulphide levels around the root system, but that once this capacity was exhausted, perhaps due to additional nutrients or reduction in plant productivity, sulphide rapidly built up and killed the root system and hence the plant.

The nationally scarce foxtail stonewort Lamprothamnium papulosum was reported to be absent where the total phosphate concentration was greater than 100 µg/l and may be lost due to nutrient enrichment (Bamber et al., 2001). Most grazing, suspension, and deposit feeding members of the community will probably benefit from the increased epiphyte and phytoplankton productivity, as would their predators.

Sensitivity assessment. The above evidence demonstrates that Ruppia can survive nutrient-enriched habitats and that growth can be stimulated by nutrient enrichment, while seed germination and seedling development require ‘high’ nutrient availability. Ruppia has also been suggested for the ‘treatment’ or mitigation of saline wastewater. This evidence suggests that Ruppia is resistant to nutrient enrichment. However, Ruppia beds were also shown to fluctuate and experience boom-bust cycles due to cycles of nutrient levels in Chesapeake Bay. Therefore, its resistance is probably site-specific and probably influenced by indirect effects (e.g. epiphyte or phytoplankton growth) more than just nutrient concentration alone. Therefore, a significant increase in nutrients and subsequent eutrophication may result in loss of the biotope. Therefore, the worst-case resistance to nutrient enrichment is assessed as ‘Low’. Hence, resilience is assessed as 'Medium', and the biotope is considered to have a ‘Medium’ sensitivity to nutrient enrichment.

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). Yet, Ruppia maritima seems to display some tolerance to organic enrichment. Ahmadi et al. (2017) suggested using Ruppia maritima as an advanced coastal treatment of saline municipal wastewater. Through an in-situ pilot study where plants were fed by activated sludge effluent (at Chobeineh wastewater treatment plant in Ahvaz city, Iran), Ruppia maritime exhibited a total nitrogen and phosphorus uptake of 44.48 to 45.78% and 63.6 to 73.01%, respectively, dependent on the level of electrical conductivity, 10, 15, and 20 ms/cm, with the highest growth and uptake rates observed at 10 ms/cm. Ahmadi et al. (2017) concluded that Ruppia maritima should be considered a promising saline-tolerant plant in the advanced treatment of saline wastewater.

In addition, Ruppia maritima may be more tolerable to fluctuations in nutrients, rebounding after stressful water quality events and more suited to surviving in degraded conditions, compared to competing seagrasses like Zostera marina (Hensel et al., 2023; French et al., 2024). However, Ruppia maritima was a significant contributor to large seagrass population crashes in the Chesapeake Bay area since the 1980s due to nutrient pulses driven by springtime runoff, such as its population halving in 2019 (Hensel et al., 2023). Since Ruppia maritima became dominant in Chesapeake Bay in 1999, seagrass cover has become increasingly variable. Between 1999 and 2019, Chesapeake Bay has either gained or lost at least 20% total seagrass area from the previous year, and Ruppia maritima cover has been approx. three times more variable than historical Zostera marina cover (Hensel et al., 2023). Therefore, Hensel et al. (2023) stated that Ruppia maritima is most negatively affected by increased watershed nutrient pollution, and at current nutrient levels in Chesapeake Bay, predicted a future amplification of boom–bust cycles in Ruppia maritima cover, alternating between high in dry years and low in wet years.

However, Lanari et al. (2018) studied the impact of short-term depositions of macroalgal blooms on Ruppia maritima in the Patos Lagoon estuary, southern Brazil, and found that unstable algal depositions promoted significant reductions in Ruppia maritima biomass, by reducing their shoot height and density, and rhizome length. Lanari et al. (2018) concluded that this process may diminish the resilience of Ruppia maritima meadows, with impacts on estuarine nutrient cycling and secondary production. Previously recorded reductions of Ruppia maritima abundance and distribution in the Patos Lagoon estuary resulted from anomalous precipitations and high fluvial discharge associated with El Niño events (Odebrecht et al., 2010 cited in Lanari et al., 2018). However, high runoff would also be detrimental due to high suspended sediment load and resultant light attenuation. 

Similar to Ruppia maritima, Ruppia cirrhosa has been documented recovering from nutrient fluctuation over a relatively short period of time. Since 2007, and after changes in hydrological management and improvements in sewage treatment, which polluted a shallow, brackish coastal lagoon in Corsica, France (Biguglia lagoon), a significant increase in Ruppia cirrhosa aquatic angiosperms was observed four to five years after sewage improvements (Pasqualini et al., 2017). The Biguglia lagoon remains a major site of agricultural and urban activities, favouring large nitrogen inputs, however, the restored seagrass communities closely resemble pristine communities, which suggests that both Ruppia cirrhosa and lagoon habitats are resilient and able to recover from natural and anthropogenic disturbances (Pasqualini et al., 2017).

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 are adapted to these conditions, but 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.

Sensitivity assessment. Ruppia has been suggested for the ‘treatment’ or mitigation of saline wastewater. This evidence suggests that Ruppia is resistant to nutrient and organic enrichment. However, Ruppia beds were also shown to fluctuate and experience boom-bust cycles due to cycles of nutrient levels in Chesapeake Bay and suffer significant reductions in biomass due to the deposition of dead algae after algal blooms. Therefore, its resistance is probably site-specific and probably influenced by indirect effects (e.g. hypoxia and sediment) more than just organic enrichment alone.  Therefore, the worst-case resistance to nutrient enrichment is assessed as ‘Low’. Hence, resilience is assessed as 'Medium', and the biotope is considered to have a ‘Medium’ sensitivity to organic enrichment.   

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. Ruppia maritima will be unable to recover from a permanent loss of habitat resulting in a ‘Very Low’ resilience score.  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

Ruppia maritima occurs almost exclusively in shallow and sheltered coastal waters anchored in sandy and muddy bottoms.  A physical change to another seabed type, i.e. from sedimentary to hard rock substratum would result in loss of the Ruppia bed and its associated community. As a permanent change, there is no opportunity for recovery without intervention.  Therefore, the resistance is 'None', and resilience is 'Very Low', resulting in a sensitivity of 'High'. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

Physical change (to another sediment type)

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

Evidence

Ruppia maritima occurs almost exclusively in shallow and sheltered coastal waters anchored in sandy, muddy, and clay bottoms (Verhoeven, 1979; Kantrud, 1991; Halabowski, Sowa & Krodkiewska, 2018). However, Ruppia maritima have also been observed in coarse sediments (grain size >20 mm), which is not consistent with previous studies of the species (Halabowski, Sowa & Krodkiewska, 2018). A physical change to another seabed type, i.e. from sedimentary to hard rock substratum, would result in loss of the Ruppia bed and its associated community. As a permanent change, there is no opportunity for recovery without intervention. Therefore, the resistance is 'None', and resilience is 'Very Low', resulting in a sensitivity of 'High'. 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 Ruppia beds.  Roots and rhizomes are buried no deeper than 20 cm below the surface (see ‘abrasion’ pressures).  Resistance is therefore assessed as ‘None’ for all seagrass biotopes and resilience is considered ‘Very Low’ resulting in a ‘High’ sensitivity score. 

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

d'Avack et al. (2014) reviewed the impacts of physical damage (abrasion and penetration) to seagrasses to the pressure benchmark. The report found that a large amount of research had been conducted, however, with the majority of studies focusing on Zostera species. The sensitivity of Ruppia to this pressure is thus largely based on expert judgement, but with a high level of confidence due to the morphology of the plant. Similar to Zostera species, Ruppia maritima has a shallow and weak root system and is incapable of vertical rhizome growth. However, Ruppia maritima has a less extensive root and rhizome system than Zostera marina, increasing its susceptibility to mechanical disturbance and increased wave energy (Hensel et al., 2023). Seagrasses are not physically robust. Ruppia stems and leaves are damaged by wave action or water turbulence, and the root system is shallow and weak (Verhoeven, 1979; Kantrud, 1991). Therefore, it is likely that Ruppia spp. are intolerant of physical disturbance and that a proportion of the vegetation may be removed and rhizomes broken by any physical disturbance, e.g. trampling, anchoring, power boating and potting. However, in subtropical areas, wintering wildfowl were reported to consume entire stands of Ruppia spp., which grew back in a few weeks (Kantrud, 1991). Similarly, Steiglitz (1966, cited in Kantrud, 1991) suggested that wildfowl could consume 50% of the standing crop without damaging the Ruppia bed. 

In Chesapeake Bay, USA, the recovery of seagrass beds was observed after boat propeller scaring. Orth, Lefcheck & Wilcox (2017) observed considerable scaring during a 26-year monitoring period (1987 to 2015), where scars averaged 5,575 and 3,206 m long in the two bays with the most intense scaring. However, on average, individual scars only persisted for 2.7 years, with a recovery time of 6.2 years for an entire set of new scars, implying quick recovery, aided by the diverse reproductive habits of the two seagrasses found in this region, Zostera marina and Ruppia maritima (Orth, Lefcheck & Wilcox, 2017). Both Ruppia maritima and Zostera marina reproduce asexually and sexually in this region, and their clonal growth is reported up to 25 cm/year (Orth & Moore, 1982). As scars were formed inside an existing meadow, clonal growth from the edges of the scars can cover a portion of the scar in the first year and is likely responsible for much of the revegetation noted in Chesapeake Bay (Orth, Lefcheck & Wilcox, 2017).

Benthic infauna such as polychaetes (e.g. Arenicola marina or Pygospio elegans) are partly protected from abrasion due to their infaunal habit, but a proportion are likely to be killed by any mechanical disturbance that penetrates the sediment (e.g. anchors). Similarly, the shell of Cerastoderma glaucum is relatively thin, and individuals are likely to be damaged or killed by abrasion. Macroalgae are relatively flexible and unlikely to be damaged. However, resident grazers (e.g. gammarid amphipods, isopods, or gastropods) are likely to be killed by direct physical contact, although they are generally small enough to be swept aside, or able swimmers, and most will probably escape.

Sensitivity assessment. Ruppia beds are particularly fragile and likely to be damaged by physical disturbance, so that a resistance of 'Low' is suggested. However, rhizomes and the seed bank remain, recovery will be rapid, so that their resilience is probably 'High', resulting in a sensitivity of 'Low’.

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

d'Avack et al. (2014) reviewed the impacts of physical damage (abrasion and penetration) seagrasses to the pressure benchmark. The report found that a large amount of research had been conducted with however the majority of studies focusing on Zostera species. The sensitivity of Ruppia to this pressure is thus largely based on expert judgement but with a high level of confidence due to the morphology of the plant. Similar to Zostera species, Ruppia maritima has a shallow and weak root system and is incapable of vertical rhizome growth. Seagrasses are not physically robust. Ruppia stems and leaves are damaged by wave action or water turbulence and the root system is shallow and weak (Verhoeven, 1979; Kantrud, 1991). Therefore, it is likely that Ruppia spp. are intolerant of physical disturbance and that a proportion of the vegetation may be removed and rhizomes broken by any physical disturbance. However, in subtropical areas wintering wildfowl were reported to consume entire stands of Ruppia spp. which grew back in a few weeks (Kantrud, 1991). Similarly, Steiglitz (1966, cited in Kantrud, 1991) suggested that wildfowl could consume 50% of the standing crop without damaging the Ruppia bed. 

Benthic infauna such as polychaetes (e.g. Arenicola marina or Pygospio elegans) are partly protected from abrasion due to their infaunal habit but a proportion are likely to be killed by any mechanical disturbance that penetrates the sediment (e.g. anchors). Similarly, the shell of Cerastoderma glaucum is relatively thin and individuals are likely to be damaged or killed by abrasion. Macroalgae and relatively flexible and unlikely to be damaged. However, resident grazers (e.g. gammarid amphipods, isopods, or gastropods) are likely to be killed by direct physical contact, although they are generally small enough to be swept aside, or able swimmers and most will probably escape.

Sensitivity assessment.  Ruppia beds are particularly fragile and likely to be damaged by physical disturbance.  Penetrative activities (e.g. demersal trawls and dredges) are likely to damage the rhizomes directly (as they lie within the top 5 cm of sediment) so that a resistance of 'Low' is suggested. However, where rhizomes and the seed bank remain, recovery will be rapid, so that their resilience is probably 'High', resulting in a sensitivity of 'Low'.

Changes in suspended solids (water clarity)

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

Evidence

Water clarity is a vital component for seagrass beds as it determines the depth penetration of photosynthetically active radiation of sunlight. Increased turbidity results from increases in dissolved organics (e.g. humic acids or gelbstoff), organic particulates and suspended sediment, or blooms of phytoplankton and zooplankton. 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 plants can occur. Ruppia spp. require high light levels and only normally develop well in clear water, and are always reduced or absent from turbid waters (Verhoeven, 1979). Joanen & Glasgow (1965) found that plants preferred turbidity levels less than 25 to 55 ppm (equivalent to 25 to 55 mg/l). However, Hillmann, DeMarco & La Peyre (2019) observed Ruppia maritima growing across the Barataria Basin, an estuary in Louisiana, USA, in turbidity ranges of 1.4 to 126.0 NTU (nephelometric turbidity units, or roughly 0.46 to 42 mg/l), with fresh and intermediate waters having higher turbidity values than those of brackish and saline waters. Wetzel & Penhale (1983) compared the photosynthetic parameters of Ruppia maritima and Zostera marina. Ruppia maritima was found to be photosynthetically less efficient in low levels of underwater light compared to Zostera marina. Ruppia maritima also has a relatively high ratio of chlorophyll α to chlorophyll β, suggesting that it is less adapted to low-light environments than other seagrasses (Evans et al., 1986). A shading experiment by Congdon & McComb (1979) on Ruppia maritima determined that a 40% reduction in light availability resulted in a 50% reduction in standing crop. While Ruppia beds can survive short-term changes in turbidity, Kantrud (1991) concluded that the control of turbidity levels was crucial for the management of Ruppia beds.

Sensitivity assessment. Turbidity is an important factor controlling production and ultimately survival and recruitment of Ruppia plants. Populations are likely to survive short-term increases in turbidity, however, a prolonged increase in light attenuation, e.g. a change from clear (<10 mg/l) to intermediate (10-100 mg/l) water clarity at the benchmark level for a year, especially at the lower depths of distribution, will probably result in loss or damage of the population. Therefore, resistance is assessed as ‘Low’. However, while rhizomes and the seed bank remain, recovery will be rapid, so that their resilience is probably 'High', resulting in a sensitivity of 'Low’.

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

Water clarity is a vital component for seagrass beds as it determines the depth penetration of photosynthetically active radiation of sunlight. Increased turbidity results from increases in dissolved organics (e.g. humic acids or gelbstoff), organic particulates and suspended sediment, or blooms of phytoplankton and zooplankton. 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 plants can occur. Ruppia spp. require high light levels and only normally develop well in clear water, and are always reduced or absent from turbid waters (Verhoeven, 1979). Joanen & Glasgow (1965) found that plants preferred turbidity levels less than 25 to 55 ppm (equivalent to 25 to 55 mg/l). However, Hillmann, DeMarco & La Peyre (2019) observed Ruppia maritima growing across the Barataria Basin, an estuary in Louisiana, USA, in turbidity ranges of 1.4 to 126.0 NTU (nephelometric turbidity units, or roughly 0.46 to 42 mg/l), with fresh and intermediate waters having higher turbidity values than those of brackish and saline waters. Wetzel & Penhale (1983) compared the photosynthetic parameters of Ruppia maritima and Zostera marina. Ruppia maritima was found to be photosynthetically less efficient in low levels of underwater light compared to Zostera marina. Ruppia maritima also has a relatively high ratio of chlorophyll α to chlorophyll β, suggesting that it is less adapted to low-light environments than other seagrasses (Evans et al., 1986). A shading experiment by Congdon & McComb (1979) on Ruppia maritima determined that a 40% reduction in light availability resulted in a 50% reduction in standing crop. While Ruppia beds can survive short-term changes in turbidity, Kantrud (1991) concluded that the control of turbidity levels was crucial for the management of Ruppia beds.

Sensitivity assessment. Turbidity is an important factor controlling production and ultimately survival and recruitment of Ruppia plants. Populations are likely to survive short-term increases in turbidity, however, a prolonged increase in light attenuation, e.g. a change from clear (<10 mg/l) to intermediate (10-100 mg/l) water clarity at the benchmark level for a year, especially at the lower depths of distribution, will probably result in loss or damage of the population. Therefore, resistance is assessed as ‘Low’. However, while rhizomes and the seed bank remain, recovery will be rapid, so that their resilience is probably 'High', resulting in a sensitivity of 'Low’.

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

Ruppia maritima is intolerant of smothering by excessive siltation. In addition, Ruppia 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.  Resistance is assessed as ‘None’ as all individuals exposed to siltation at the benchmark level are predicted to die and consequent resilience as ‘Low’ to ‘Very Low’.  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., 2020), 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. 

No studies have examined the effect of EMFs on any of the characterizing species. However, one study was performed on the reef-forming annelid, Ficopomatus enigmaticus (Oliva et al., 2023). Sperm cells from this species were exposed to 0.5 and 1.0 mT of static magnetic field. After only three hours of exposure, sperm fertilization rate was reduced, and significant increases in DNA damage and mitochondrial activity, indicative of a stress response, were reported. However, there is ‘Insufficient evidence’ on which to base an assessment of the likely sensitivity of this biotope to EMFs.

Underwater noise changes

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

Evidence

Not relevant

Introduction of light or shading

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

Evidence

Ruppia spp. require high light levels and only normally develop well in clear water, and are always reduced or absent from turbid waters (Verhoeven, 1979). Joanen & Glasgow (1965) found that plants preferred turbidity levels less than 25 to 55 ppm (equivalent to 25 to 55 mg/l). However, Hillmann, DeMarco & La Peyre (2019) observed Ruppia maritima growing across the Barataria Basin, an estuary in Louisiana, USA, in turbidity ranges of 1.4 to 126.0 NTU (nephelometric turbidity units, or roughly 0.46 to 42 mg/l), with fresh and intermediate waters having higher turbidity values than those of brackish and saline waters. Bittner, Roesler & Barnes (2020) estimated seagrass populations along the Texas portion of the Gulf Coast, USA, using a species distribution model, and found that Ruppia maritima (alongside Syringodium filiforme and Halophila engelmannii) were most influenced by benthic light availability. In addition, the distribution model predicted a lack of suitable habitat near sites characterized by abundant human development, due to human disturbances often leading to elevated nitrate concentrations and decreased benthic light availability. However, Ruppia maritima has a relatively low minimum light requirement compared to some seagrasses, approximately 8 to 16% surface irradiance, compared to Halodule wrightii and Thalassia testudinum, 25 to 27% and 18 to 23%, respectively (Choice et al., 2014 cited in Bittner, Roesler & Barnes, 2020).

Seed survival for Ruppia maritima is particularly limited by phosphorus and light, with seedling survival and development limited to ≤3 days in light and 0% survival in the dark, and a preference for asexual reproduction to compensate for seedling death (Strazisar et al., 2021). At phosphorus-limited sites, adequate light for Ruppia maritima is >20,000 lumens /m², although Ruppia maritima has also been observed growing and reproducing in approx. 14,000 lumens m² when phosphorus nutrients were sufficient, and there was an absence of competing plants, such as Chara hornemannii (Strazisar et al., 2021).

While light availability limits all seagrass species, Ruppia maritima meadows are more likely to totally collapse in a poor light environment shared with Zostera marina, as compared to competing Zostera, Ruppia has higher light requirements (i.e., light compensation point of 49.5 μmol/m²/s compared to 17.95 μmol/m²/s) and a shorter canopy (5 to 20 cm vs. 30 to 50 cm) that does not reach high into the water column until midsummer (Batiuk et al., 2000 and Moore, 2004 cited in Hensel et al., 2023). Wetzel & Penhale (1983) compared the photosynthetic parameters of Ruppia maritima and Zostera marina. Ruppia maritima was found to be photosynthetically less efficient in low levels of underwater light compared to Zostera marina. Ruppia maritima also has a relatively high ratio of chlorophyll α to chlorophyll β, suggesting that it is less adapted to low-light environments than other seagrasses (Evans et al., 1986). A shading experiment by Congdon & McComb (1979) on Ruppia maritima determined that a 40% reduction in light availability resulted in a 50% reduction in standing crop. Kantrud (1991) noted that poor insolation due to fog, mountains, and short days reduced Ruppia spp. productivity. While Ruppia beds can survive short-term changes in turbidity, Kantrud (1991) concluded that the control of turbidity levels was crucial for the management of Ruppia beds.

Sensitivity assessment. It is likely that 'shading' of a Ruppia bed would result in reduced growth and productivity, potential competition with shade-tolerant species and, if prolonged, loss of the bed in a 'shaded' area. Therefore, a resistance of 'Low' is given, with a resilience of 'High', resulting in a sensitivity of 'Low'. However, if the 'shading' was caused by a permanent structure, then the resilience would be 'Very low' and sensitivity 'High'.

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. The potential effects of vessel grounding are covered under abrasion above. 

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.  Zostera marina is the seagrass species most commonly translocated. Williams and Davis (1996) found that levels of genetic diversity of restored eelgrass Zostera marina beds. 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 reduced community resilience, compared to natural populations (Hughes & Stachowicz, 2004).

Even though restoration efforts tend to focus on Zostera marina, transplantations of Ruppia maritima (Bird et al., 1994) have also been undertaken and large areas of wetland planted with Ruppia to feed wildfowl in the USA (Kantrud, 1991). Similar reductions in genetic diversity are possible. 

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 hybridisation 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 Ruppia species. 

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

Kantrud (1991) reported possible pathogenic effects of fungi, that produce 'tubercles' on the Ruppia leaves. Kantrud (1991) also states that 'vegetative reproduction usually allows Ruppia spp. to survive Rhizoctonia infestations' and that Ruppia spp. probably suffer less from diseases than other aquatic angiosperms.

Sensitivity assessment. Therefore, the above evidence suggests a resistance of 'High', a resilience of 'High' and, hence, a sensitivity of 'Not sensitive'. 

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

In accessible areas, extraction of Arenicola marina for bait is likely to disturb the sediment and benthic infauna, although the Ruppia stands themselves would probably recover quickly. Similarly, Arenicola marina populations are thought to recover rapidly, although in isolated areas recovery may take longer due to the lack of a pelagic larvae.

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

In accessible areas, extraction of Arenicola marina for bait is likely to disturb the sediment and benthic infauna, although the Ruppia maritima stands themselves would probably recover quickly. Similarly, Arenicola marina populations are thought to recover rapidly, although in isolated areas recovery may take longer due to the lack of pelagic larvae. Direct, physical impacts from harvesting 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 on this biotope. 

Incidental removal of the key characterizing seagrass species and associated species would alter the character of the biotope. The biotope is characterized by the presence of beds of Ruppia maritima, these provide habitat structure and 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 Ruppia maritima and the associated species would also be lost.

Sensitivity assessment. Incidental removal of Ruppia spp. as by-catch would be detrimental, altering the character of the biotope and removing the habitat structure, and could lead to reclassification of the biotope where extensive removal occurs.  Therefore, resistance is considered to be 'None', resilience 'High'  (but see caveats in the resilience section) and sensitivity 'Medium'.

The American slipper limpet, Crepidula fornicata

Evidence

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 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). Close examination of the literature (2023) shows that evidence of its colonization and density on bedrock in the infralittoral or circalittoral was lacking. Tillin et al. (2020) suggested that Crepidula could colonize circalittoral rock due to its presence on tide-swept rough grounds at 60 metres in the English Channel (Hinz et al., 2011). However, Hinz et al. (2011) reported that Crepidula fornicata only dominated one assemblage (with an average of 181 individuals per trawl) on a gravel substratum with boulders. Bohn et al. (2015) noted that Crepidula occurred at low density or was absent in areas dominated by boulders. Bohn et al. (2013a, 2013b, 2015) and Preston et al. (2020) showed that while Crepidula could settle on slate panels or ‘stone’, it preferred shell, especially that of conspecifics.

Sensitivity assessment. The infralittoral muddy sand characterizing this biotope and the wide salinity range of Ruppia maritima are likely to be unsuitable for the colonization by Crepidula fornicata, although the lack of wave action might allow limited colonization than more exposed sites. Crepidula has been recorded from areas of strong tidal streams (Hinz et al., 2011), and has been recorded 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). Although Crepidula fornicate has been observed in seagrass beds in Arcachon Bay, France (Tillin et al., 2020), no evidence was found on the effect of Crepidula populations on seagrass-dominated habitats or infralittoral habitats. At present, there is 'Insufficient evidence' to suggest that the infralittoral muddy sand biotopes are sensitive to colonization by Crepidula fornicata or other invasive species; 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; Minchin & 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-mediated transport via aquaculture facilities, boat hulls, commercial fishing vessels, and ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijkstra et al., 2007; Griffith et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009). For example, in oyster farms in British Columbia, large fragments of Didemnum sp. come off oyster strings when they are pulled out of water, and other fragments can be pulled off oysters and mussels and thrown back into the water, which is likely to aid dispersal of the invasive species (Bullard et al., 2007). Dijkstra et al. (2007) hypothesised that Didemnum sp. was introduced to the Gulf of Maine with oyster aquaculture in the Damariscotta River and transported via Pacific oysters.

Didemnum vexillum was likely introduced into the UK from northern Europe or Ireland via poorly maintained or not antifouled vessels, movement of contaminated shellfish stock and aquaculture equipment, or via marine industries such as oil, gas, renewables, and dredging (Holt, 2024). Recent evidence from genetic material suggests 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, 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°C and 20°C and slow or cease below 9°C 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°C to 22°C) and the lowest average temperatures were recorded in July (9°C 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., 2007a&b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012; Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent, and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020).

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

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

Didemnum vexillum has not been reported from Ruppia beds but has been reported to colonize seagrass communities to a limited extent.   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.  In both the northeast and northwest United States, overgrowth of Zostera marina blades by Didemnum vexillum reduced the above-ground growth of the seagrass, and a reduction in eelgrass growth, due to light reduction, has been observed in the past as a result of fouling by other invasive ascidians (Tillin et al., 2020). In addition to reduced growth, seagrasses may be vulnerable to uprooting, especially if combined with burrowing, contributing to erosion (Tillin et al., 2020). 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.  This biotope experiences very weak water flow (negligible) and is extremely sheltered from wave exposure. Didemnum vexillum regresses as temperatures decline in winter, so shallow examples may be able to recover their condition in winter (Gittenberger, 2007; Valentine et al., 2007a; Herborg et al., 2009). 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 due to the lack of direct evidence within Ruppia beds.  

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 northeastern Adriatic Sea, where it was not introduced for aquaculture (Ezgeta-Balic et al., 2019). It was also suggested that some Magallana gigas populations were established in southwest England from France, possibly via fouling on ships (GBNNSIP, 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).  

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).

Sensitivity assessment. No reports were found of Magallana gigas attached to macroalgae or seagrass (Tillin et al., 2020), and the muddy sand in this biotope would not provide suitable attachment for the species. However, a study in British Columbia, Canada, found that while oysters and Zostera marina coexist at a regional scale, eelgrass is typically absent directly seaward of oyster beds, it is not clear if this was due to tidal level or exclusion by the oysters (Tillin et al., 2020). In addition, as Magallana gigas prefers salinities of 20 psu or above, it would likely only be observed in the most marine areas of this biotope, such as the mouths of estuaries or coastal shores, 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.

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). 

Sensitivity assessment. Since Sargassum muticum can survive in estuarine conditions and overlaps with the depth range of this biotope, there is a possibility of it interacting with Ruppia maritima. However, as Sargassum muticum distribution is limited by the availability of hard substrata, the muddy sand in this biotope would not provide suitable attachment, 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.

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 can survive in estuarine conditions and overlaps with the depth range of this biotope, there is a possibility of it interacting with Ruppia maritima. However, as Undaria pinnatifida distribution is limited by the availability of hard substrata, the muddy sand in this biotope would not provide suitable attachment, 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

Tillin et al. (2020) noted that the red algae, Bonnemaisonia hamifera, grows predominantly epiphytically on macroalgae in lower littoral and shallow subtidal habitats, and has been documented growing on Zostera marina and Ruppia maritima. Bonnemaisonia hamifera may also occur on rocks or commonly as a key element within the ‘infralittoral muddy gravel’ biotope, widely found in lagoons and sea lochs, where the mud is often gravelly with cobbles and can be black and anoxic close to the surface (Tillin et al., 2020). Bonnemaisonia hamifera has a preference for salinities between 14.26 and 37.55 psu, depths to 20 m, and very sheltered conditions (Tillin et al., 2020).

Other invasive species which could have a potential impact on seagrasses and the biotope include the Chinese mitten crab, Eriocheir sinensis, which may graze on seagrass; the red seaweed, Agarophyton vermiculophyllum, which may alter the trophic dynamics and nutrient cycling of seagrasses and compete with native seagrass beds for light and oxygen; the orange-striped green sea anemone, Diadumene lineata, which may alter food webs and nutrient cycling within the ecosystem at high abundances; and the American lobster, Homarus americanus, which have been recorded living in sublittoral macrophyte-dominated sediment and eelgrass habitats, which provides a suitable burrowing substratum as well as some shelter camouflage (Tillin et al., 2020).

d’Avack et al. (2014) found that invasive flora had the greatest impact on seagrass beds but pointed out extensive knowledge gaps on how invasive species influence the health of Zostera beds in UK waters. Indeed, recovery would only be possible if the majority of the invasive species were removed (through either natural or unnatural processes) to allow the re-establishment of the characterizing species. More research is needed in order to fully comprehend this pressure on seagrasses. Therefore, there is ‘Insufficient evidence’ on the effect of other invasive species on Ruppia spp.