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

Fucus serratus is a cold-temperate species with a broad North Atlantic distribution extending from Svalbard to Portugal and across Northeast America. Within the British Isles, it occurs centrally within its geographic range (Yesson et al., 2015). Broad-scale habitat suitability modelling indicates that temperature is an important environmental predictor for this species across Europe (Westmeijer et al., 2019), but around the British Isles, where Fucus serratus does not exhibit a negative association with warmer conditions, temperature does not generally appear to limit its distribution (Yesson et al., 2015).

Laboratory and mesocosm studies indicate that adult thalli tolerate temperatures exceeding typical UK summer maxima. Nielsen et al. (2014) reported no reduction in the growth of adult Fucus serratus specimens from Scotland at 22°C, and Arrontes (1993) recorded survival after one week at 25°C. Figueroa et al. (2019), using individuals from the species’ southern distribution limit, identified an optimal growth range of 14 to 20°C, with photosynthetic adjustment occurring across the 8 to 28°C range. Migné et al. (2021) recorded canopy-level aerial production rates exceeding 1 g C/m²/hour during emersion under high light and temperature conditions in the southern English Channel. Although photoinhibition occurred in thalli at the top of the canopy under harsh conditions, photosystem II performance remained high within lower canopy layers, with temperatures recorded between 12 and 23.3°C below the canopy and 14 to 32.9°C above it. This vertical thermal buffering demonstrated in situ moderation of heat stress within intact stands.

Short-term atmospheric heatwave simulations also indicate buffering within canopies. Harris et al. (2025) simulated a three-day atmospheric heatwave, with heated treatments reaching 18.0 to 31.2°C above ambient air temperatures. The upper canopy fronds substantially reduced the thermal stress experienced by the lower canopy fronds within the Fucus serratus canopy, even when air temperatures exceeded previously estimated critical thermal maxima. However, repeated exposures reduced resilience.

In contrast, studies conducted at the southern distribution limit (NW Iberian Peninsula) of Fucus serratus consistently report reduced performance and range contraction associated with warming. Casado-Amezúa et al. (2019) estimated a contraction of approximately 45% in the extent of occurrence for Fucus serratus over three decades. Fernández (2016) reported large-scale replacement of cold-temperate canopy species, including Fucus serratus, by warm-temperate species along northern Spain. Álvarez-Losada et al. (2020) described localised loss near former distributional limits, with subsequent shifts to turf-dominated assemblages driven primarily by canopy loss.

Duarte & Viejo (2018) identified marked differences in recruitment and adult performance between exposed coasts and upwelling-influenced rias, suggesting the latter act as climatic refugia. Viejo et al. (2024) reported absence at sites where autumn maximum temperatures exceeded 16.6°C and projected potential loss in rias with an increase of 1.5°C in maximum autumn temperature. García et al. (2021) and Pereira et al. (2025) demonstrated elevated mortality and reduced germling survival at high temperatures in southern populations, with mortality in Fucus serratus beginning at 36°C under experimental emersion heat stress and reaching 50% after prolonged exposure (Pereira et al., 2025). Pedersen et al. (2025) found negative effects of 25°C exposure over five weeks, particularly when combined with reduced salinity. Earlier experimental and field studies from Spain and Portugal also documented the adverse effects of elevated temperature on growth, physiological performance, and reproductive output (Pearson et al., 2009; Viejo et al., 2011; Martínez et al., 2012). Jueterbock et al. (2014) attributed reduced performance in southern populations in part to restricted within-population genetic diversity, suggesting limited adaptive capacity at the warm range edge. In contrast, southwest Ireland and Brittany have been identified as hotspots of genetic diversity (Coyer et al., 2003; Hoarau et al., 2007), which may enhance resilience to warming through greater genetic variation. These findings indicate that phenotypic plasticity and genetic diversity are important mediators of population-level sensitivity to temperature stress.

However, these southern-edge studies reflect populations already near upper thermal limits and frequently incorporate interacting stressors such as emersion heat, low salinity, or chronic warming. In contrast, monitoring in the Barents Sea (Kolbeeva et al., 2025) recorded increased biomass and stable Fucus serratus populations between 2021 and 2024, potentially associated with relatively warmer summers and higher salinity in the region. Similarly, Armitage et al. (2017) found that under relatively warm summer conditions in Norway (mean 15.1°C), Fucus serratus exhibited greater weight gain than competing canopy species.

Marine heatwave mesocosm experiments (Atkinson et al., 2020) applying +1.5, +2.0, and +3.5°C increases for 14 days recorded reductions in growth and photosynthetic performance in Fucus serratus, while enhancing the performance of Sargassum muticum, a non-native competitor. Saha et al. (2025) reported altered bacterial settlement patterns on Fucus serratus under +5°C laboratory exposure (21°C vs 16°C), indicating temperature-mediated shifts in algae-bacteria interactions. Provera et al. (2021) observed lower photosynthetic responsiveness to warming in Fucus serratus compared to turf-forming species. Experimental work along a controlled temperature gradient (7 to 31°C) in Northwest Iberia further demonstrated differential physiological sensitivity between cold- and warm-affinity seaweeds (Díaz-Acosta et al., 2021). In Fucus serratus, net photosynthesis showed lower responsiveness to warming than respiration, indicating a disproportionate increase in respiratory demand with temperature. This imbalance suggests potential reductions in carbon balance under sustained warming, supporting evidence that cold-affinity species are more vulnerable to elevated temperatures than co-occurring warm-affinity taxa.

Genetic analyses provide additional context. Jueterbock et al. (2018) found temporal stability in genetic structure across central populations over five to ten generations, although diversity declined at the southern range edge between 1999 and 2010. Bioclimate envelope modelling for North America projected northward expansion of approximately 500 km and southern retraction of similar magnitude by 2100 under high-emission scenarios (Khan et al., 2018).

Algal reproduction appears relatively insensitive to small temperature variations. A systematic review of algal reproductive phenology (de Bettignies et al., 2018) identified only two studies in which temperature strongly predicted gamete release, both involving temperature differences exceeding those typically observed between comparable UK sites (Norton, 1981; Bacon & Vadas, 1991). However, Nielsen et al. (2014) reported that germlings were negatively affected by increased temperature, indicating that early life stages are more vulnerable than mature algae to this pressure.

Little direct evidence was found to assess the effects of increased temperature on piddocks, and the assessment is based on distribution records and evidence for spawning in response to temperature changes. Pholas dactylus occurs in the Mediterranean and the East Atlantic, from Norway to Cape Verde Islands (Micu, 2007). Temperature influences the timing of reproduction in Pholas dactylus, which usually spawns between July and August. Increased summer temperatures in 1982 induced spawning in July on the south coast of England (Knight, 1984). Species distribution models suggested that the distribution of Pholas dactylus could expand northward in the next century due to ocean warming (Schultz et al., 2024).

Similar observations have been made for other piddock species. Spawning of the piddock Petricolaria pholadiformis was initiated by increasing water temperature (>18°C) (Duval, 1963a), so elevated temperatures outside of usual seasons may disrupt normal spawning periods. The spawning of Barnea candida was also reported to be disrupted by temperature changes. Barnea candida normally spawns in September when temperatures are dropping (El-Maghraby, 1955). However, a rise in temperature in late June of 1956 induced spawning in some specimens of Barnea candida (Duval, 1963b). Disruption from established spawning periods, caused by temperature changes, may be detrimental to the survival of recruits, as other factors influencing their survival may not be optimal, and some mortality may result. Established populations may otherwise remain unaffected by elevated temperatures.

The current distribution of Hiatella arctica is predominantly arctic and boreal (Gordillo, 2001; Sejr et al., 2004), and palaeoecological reviews describe the genus as ‘consistently linked to cool temperate and polar regions’ (Gordillo, 2001). However, populations of Hiatella arctica occur in the Mediterranean and have clearly acclimated to the warmer temperatures (Oberlechner, 2008). Laboratory experiments on filtration rates of Hiatella arctica found that activity was strongly linked to temperature (Ali, 1970). Activity rates rose steadily between 0°C and a maximum between 15°C and 17°C and fell sharply to almost no activity at 25°C (Ali, 1970). Although activity may be reduced, Hiatella arctica have very low metabolic rates and may be able to sustain a period of reduced activity. Regression models developed by Bourget et al. (2003) found that temperature and water transparency (measured in metres and indicating the level of inorganic suspended solids) explained only 40% of the variation in biomass of Hiatella arctica fouling navigation buoys in the Gulf of St Lawrence system (Canada). These findings suggest that other variables play a more significant role in determining settlement, survival, and growth over a year in this system. However, the models indicated that biomass was higher when temperatures were higher (around 14°C), although a causal link was not identified (Bourget et al., 2003).

Sensitivity assessment. Fucus serratus is found in the middle of its natural temperature range in the British Isles and is therefore not likely to be affected by an increase in temperature at the pressure benchmark. An increase in acute or chronic temperature above average British and Irish temperatures is not likely to have a detrimental effect on Fucus serratus and its associated communities, based on global distribution. However, phenotypic plasticity would influence the tolerance of an individual population. The global distribution of the piddock species, Petricolaria pholadiformis, Pholas dactylus and Barnea candida, suggests that these species can tolerate warmer waters than currently experienced in the UK and may therefore be tolerant of a chronic increase in temperature. Experiments by Ali (1970) suggest that Hiatella arctica could tolerate an acute or chronic increase in temperature at the pressure benchmark. Although an acute or chronic increase may result in sub-lethal effects on feeding and hence a reduction in growth and potentially reproduction. Therefore, resistance to an acute change in temperature is assessed as ‘High’, resilience as ‘High’ (no impact to recover from), and sensitivity as ‘Not Sensitive’ for all characteristic species.  For all species, it should be noted that the timing of acute changes may lead to greater impacts.  Temperature increases in the warmest months may exceed thermal tolerances, whilst changes in colder periods may stress individuals acclimated to the lower temperatures.

Temperature decrease (local)

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

Evidence

Many intertidal species are tolerant of freezing conditions as they are exposed to extremes of low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period of time during the tidal cycle. In winter, air temperatures are colder than the sea; conversely, in summer, air temperatures are much warmer than the sea. Species that occur in the intertidal are therefore generally adapted to tolerate a range of temperatures, with the width of the thermal niche positively correlated with the height of the shore (Davenport & Davenport, 2005). Local populations may be acclimated to the prevailing temperature regime and may therefore exhibit different tolerances than other populations subject to different salinity conditions. Therefore, caution should be used when inferring tolerances from populations in different regions. 

Lüning (1984) reported that Fucus serratus survived in the laboratory for a week at temperatures between 0°C and 25°C. Fucus serratus is found along the Atlantic coast of Europe from Svalbard to Portugal and on the shores of north-east America; it is one of three non-indigenous macroalgae identified as successfully established in Iceland (Micael et al., 2021). Hence, the seaweed is within its thermal range in the British Isles. Lüning (1984) placed this species in his ‘Cold temperature North Atlantic group’.

Little empirical evidence was found to assess the effects of decreased temperature on piddocks, and the assessment is based on distribution records and evidence for spawning in response to temperature changes. The American piddock Petricolaria pholadiformis has a wide distribution and is found as far north as the Skaggerak, Kattegat and Limfjord (Jensen, 2010; Huber & Gofas, 2015). Pholas dactylus occurs in the Mediterranean and the East Atlantic, from Norway to Cape Verde Islands (Micu, 2007). Barnea candida is distributed from Norway to the Mediterranean and West Africa (Gofas, 2015). Temperature changes have been observed to initiate spawning by Pholas dactylus, which usually spawns between July and August. Increased summer temperatures in 1982 induced spawning in July on the south coast of England (Knight, 1984). Spawning of Petricolaria pholadiformis is initiated by increasing water temperature (>18°C) (Duval, 1963a), so decreased temperatures may disrupt normal spawning periods when this coincides with the reproductive season. The spawning of Barnea candida was also reported to be disrupted by temperature change. Barnea candida normally spawns in September when temperatures are dropping (El-Maghraby, 1955). Disruption from established spawning periods, caused by decreased temperatures, may be detrimental to the survival of recruits, as other factors influencing their survival may not be optimal, and some mortality may result. Established populations may otherwise remain unaffected by decreased temperatures.

Gordillo & Aitken (2000), in a review of environmental factors relevant to re-interpreting Late Quaternary environments from fossil collections, suggest that Hiatella arctica is eurythermal, based on Aitken (1990) and Peacock (1993). The current distribution of Hiatella arctica is predominantly arctic and boreal (Sejr et al., 2004; Gordillo, 2001) and palaeoecological reviews describe the genus as ‘consistently linked to cool temperate and polar regions’ (Gordillo, 2001), suggesting that this species would not be sensitive to a decrease in temperature at the pressure benchmark, within temperate regions. Regression models developed by Bourget et al. (2003) found that temperature and water transparency (measured in metres and indicating the level of inorganic suspended solids) explained only 40% of the variation in biomass of Hiatella arctica fouling navigation buoys in the Gulf of St Lawrence system (Canada). These findings suggest that other variables play a more significant role in determining settlement, survival, and growth over a year in this system. However, the models indicated that biomass was higher where temperatures were higher (around 14°C), although a causal link was not identified (Bourget et al., 2003). 

Sensitivity assessment. A decrease in acute or chronic temperature above average British and Irish temperatures is not likely to adversely effect Fucus serratus and associated communities, based on global distribution. However, it should be noted that phenotypic plasticity will influence the tolerance of the individual population. The global distribution of the piddock species and Hiatella arctica suggests that these species can tolerate cooler waters than currently experienced in the UK and may therefore be tolerant of a chronic decrease in temperature at the benchmark level. Decreased temperatures may, depending on timing, interfere with spawning cues, which appear to be temperature driven. The effects will depend on the seasonality of occurrence and the species affected. Adult populations may be unaffected and, in these relatively long-lived species, an unfavourable recruitment may be compensated for in a following year. Based on the characterizing species, resistance to an acute and chronic decrease in temperature at the pressure benchmark is therefore assessed as ‘High’ and resilience as ‘High’ (within two years) and the biotope is considered ‘Not sensitive’.

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

Little direct evidence was found to assess sensitivity to this pressure. Local populations may be acclimated to the prevailing salinity regime and may therefore exhibit different tolerances to other populations subject to different salinity conditions, and therefore caution should be used when inferring tolerances from populations in different regions. Biotopes found in the intertidal will naturally experience fluctuations in salinity, where evaporation increases salinity and inputs of rainwater expose individuals to freshwater. Species found in the intertidal are therefore likely to have some form of behavioural or physiological adaptations to changes in salinity. Fucoids are able to compensate for changes in salinity by adjusting internal ion concentrations. However, this will occur at a cost, reducing photosynthetic rate and hence affecting the growth rate of the seaweed. Fucus serratus commonly inhabits narrow fjords where salinity can vary widely along a spatial (kms) and/or temporal (hours to daily) scale. Growth rates for Fucus serratus are maximal at a salinity of 20 PSU, with the critical limit for recruitment set at 7 PSU (Malm et al., 2001). An increase in salinity at the pressure benchmark could therefore impact growth.

Field monitoring from the Barents Sea recorded increased Fucus biomass associated with higher summer salinity, with Fucus serratus populations remaining stable between 2021 and 2024 (Kolbeeva et al., 2025). Cover increased across all research sites, and elevated salinity combined with warm conditions was suggested to stimulate growth. At the southern distribution limit in Northwest Iberia, fine-scale modelling identified mean salinity as an important predictor of Fucus serratus, with a higher probability of occurrence at salinities between 33 and 35 PSU and absence recorded below a mean salinity of 33 PSU (Viejo et al., 2024). These findings indicate that Fucus serratus tolerates and may benefit from salinity within or slightly above average full marine conditions, although salinity interacts with temperature and grazing pressure at range edges.

No direct empirical evidence was found to assess this pressure for Pholas dactylus, so the assessment is based on the reported distribution of the biotope and other piddocks. Pholas dactylus has been recorded in salinities of 30 to 35 PSU, with a small number of records from 35 to 40 PSU (OBIS, 2025). Barnea candida is reported to extend into estuarine environments in salinities down to 20 PSU (Fish & Fish, 1996). Petricolaria pholadiformis is particularly common off the Essex and Thames estuary, e.g. the River Medway (Bamber, 1985), suggesting tolerance of brackish waters. Zenetos et al. (2009) suggest that at all sites where Petricolaria pholadiformis has been found has some freshwater inflow into the sea. According to the literature, the species in its native range inhabits environments with salinities between 29 and 35 PSU, while in the Baltic Sea it is reported from salinities 10 to 30 PSU (Gollasch & Mecke, 1996, cited from Zenetos et al. 2009). According to Castagna & Chanley (1973, cited from Zenetos et al. 2009), the lower salinity tolerance of Petricolaria pholadiformis is 7.5 to 10 PSU. It appears that reduced salinity facilitates its establishment (Zenetos et al., 2009).

Filippov et al. (2003, abstract only) tested the salinity tolerances of Hiatella arctica obtained from the White Sea. The salinity tolerance of individuals kept at 25 PSU was 17 to 36 PSU. Acclimation of Hiatella arctica allowed them to adapt to higher or lower salinities, with the potential tolerance range of acclimated individuals assessed as 13 to 42 PSU. 

Sensitivity assessment. Although some increases in salinity may be tolerated by the species present, these are generally short-term and mitigated during tidal inundation. Fucus serratus occurs in fully marine conditions and is regularly exposed to short-term fluctuations in salinity in fjords and estuarine-influenced coasts. Hiatella arctica are relatively euryhaline and may acclimate to increases in salinity >40 PSU. However, no evidence was found to suggest that the characteristic piddock species could tolerate a salinity increase at the pressure benchmark level. Therefore, there is Insufficient Evidence to assess the sensitivity of the whole biotope to this pressure.

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

Biotopes found in the intertidal zone will naturally experience fluctuations in salinity, where evaporation increases salinity, and rainwater exposes individuals to freshwater. Species found in the intertidal are therefore likely to have some form of behavioural or physiological adaptations to changes in salinity.

Fucoids are able to compensate for changes in salinity by adjusting internal ion concentrations. However, this will occur at a cost, reducing the photosynthetic rate and hence affecting the growth rate of the seaweed. Growth rates for Fucus serratus are maximal at a salinity of 20 PSU, with the critical limit for recruitment set at 7 PSU (Malm et al., 2001). While low salinity has long been considered a limiting factor for fucoid reproduction, recent work suggests that typical salinity variation in UK populations is well above thresholds causing reproductive failure (Ardehed et al., 2016; Kinnby et al., 2019). Fertilization and germling survival are only substantially reduced under strongly reduced salinity (<5 to 6 PSU; Brawley, 1992; Serrão et al., 1996, 1999; Malm et al., 2001).

Experimental exposure of Fucus serratus to very low salinity (5 PSU) for five weeks caused negative effects on survival, growth, photosynthesis, and oxidative stress, particularly when combined with elevated temperature (Pedersen et al., 2025). At the southern edge of its distribution in Northwest Iberia, mean salinity below 33 PSU was associated with the absence of Fucus serratus, while sites with salinity between 33 and 34 PSU showed the greatest probability of occurrence (Viejo et al., 2024). These findings indicate that Fucus serratus is generally tolerant of small reductions in salinity but may be constrained by very low salinity or interacting stressors, particularly in combination with high temperatures at range edges.

No direct empirical evidence was found to assess this pressure for the characteristic piddock species, and the assessment is largely based on the reported distribution of the biotope and piddocks, which are recorded in full salinity habitats (Connor et al., 2004). Pholas dactylus has been recorded in salinities ranging from 30 to 35 PSU, with a small number of records from 35 to 40 PSU (OBIS, 2025). No information was found for the salinity tolerance of Pholas dactylus. Barnea candida is reported to extend into estuarine environments at salinities down to 20 PSU (Fish & Fish, 1996). Barnea candida is reported to extend into estuarine environments in salinities down to 20 PSU (Fish & Fish, 1996). Petricolaria pholadiformis is particularly common off the Essex and Thames estuary, e.g. the River Medway (Bamber, 1985), suggesting tolerance of brackish waters. Zenetos et al. (2009) suggested that there was some freshwater inflow into the sea at all sites where Petricolaria pholadiformis was found.  According to the literature, the species in its native range inhabits environments with salinities between 29 and 35 PSU, while in the Baltic Sea it is reported from salinities 10 to 30 PSU (Gollasch & Mecke, 1996, cited from Zenetos et al. 2009). According to Castagna & Chanley (1973, cited from Zenetos et al. 2009), the lower salinity tolerance of Petricolaria pholadiformis is 7.5 to 10 PSU. It appears that reduced salinity facilitates its establishment (Zenetos et al., 2009).

Filippov et al. (2003, abstract only) tested the salinity tolerances of Hiatella arctica obtained from the White Sea. The salinity tolerance of individuals kept at 25 PSU was 17 to 36 PSU. Acclimation of Hiatella arctica allowed them to adapt to higher or lower salinities, with the potential tolerance range of acclimated individuals assessed as 13 to 42 PSU. Gordillo & Aitken (2000) suggested that the normal minimum salinity tolerance of Hiatella arctica was 20 PSU (based on Aitken (1990) and Peacock (1993), in a review of environmental factors relevant to re-interpreting Late Quaternary environments from fossil collections. 

Sensitivity assessment. Hiatella arctica is relatively euryhaline and may acclimate to decreases in salinity from full to reduced (18 to 30 PSU) or variable (18 to 35 PSU). The impact will be mediated by the length of exposure to lower salinities, with the evidence suggesting that long-term exposure to salinities <20 PSU is harmful. Reductions in salinity at the lower end of the pressure benchmark are likely to result in a reduction in species abundance and richness as less tolerant species either move away or perish. In addition, based on reported distributions of the biotope and piddocks, a benchmark-level decrease in salinity from full to reduced would result in decreased abundance of Pholas dactylus in biotopes that were previously fully marine. Therefore, resistance is assessed as ‘Low’, and resilience (following return to full salinity) as ‘Medium’, and sensitivity is assessed as ‘Medium’ 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

Moderate water movement is beneficial to seaweeds as it carries a supply of nutrients and gases to the plants and removes waste products. Propagule dispersal, fertilization, settlement, and recruitment are also influenced by water movement (Pearson & Brawley, 1996). Increased water flow can increase scour through increased sediment movement. Small life stages of macroalgae are likely to be affected by the removal of recruits from the substratum, reducing successful recruitment (Devinny & Volse, 1978) (see ‘siltation’ pressures). A reduction in water flow can cause a thicker boundary layer, resulting in lower absorption of nutrients and CO₂ by the macroalgae. Slower water movement can also cause oxygen deficiency, directly impacting the fitness of algae (Olsenz, 2011). 

Higher water flow rates increase mechanical stress on macroalgae by increasing drag. This can result in individuals being torn off the substratum. Jonsson et al. (2006) found that a flow speed of 7 to 8 m/s completely dislodged Fucus vesiculosus individuals larger than 10 cm. Smaller individuals are likely to better withstand increased water flow as they experience less drag. Biogenic habitat structures reduce the effects of water flows on individuals by slowing and disrupting the flow. The fronds of Fucus serratus and the red algal turf will reduce the flow experienced by the turf. Boller & Carrington (2006), for example, found that the canopy created by the taller turf of Chondrus cripsus reduced drag forces on individual plants by 15 to 65%. The crustose holdfasts of Osmundea pinnatifida, Corallina officinalis and the coralline crusts are securely attached and, as these are relatively flat, are subject to less drag than upright fronds and are likely to tolerate changes in water flows at the pressure benchmark. Moderate water movement is beneficial to seaweeds as it carries nutrients and gases to the plants and removes waste products. However, if the flow becomes too strong, plants may become dislodged.

Established adult piddocks are, to a large extent, protected from the direct effects of increased water flow, owing to their environmental position within the substratum. Increases or decreases in flow rates may affect suspension feeding by altering the delivery of suspended particles or the efficiency of filter feeding. However, no evidence was found to inform the sensitivity assessment. However, other biotopes characterized by piddocks (IR.MIR.KR.Ldig.Pid and CR.MCR.SfR.Pid) have been found in areas where tidal flows vary between 0.5 and 1.5 m/s (Connor et al., 2004), suggesting that changes in flow rates within this range will not negatively impact piddocks. Adult piddocks may become exposed should physical erosion occur at a greater rate than burrowing and be lost from the substratum. At higher densities, bioerosion by piddocks may destabilise the substratum, increasing vulnerability to erosion and resulting in loss of habitat, Fucus serratus and red algal turf.

Growth and reproduction of Semibalanus balanoides are influenced by food supply and water velocity (Bertness et al., 1991). Laboratory experiments demonstrate that barnacle feeding behaviour alters over different flow rates but that barnacles can feed at a variety of flow speeds (Sanford et al., 1994). Flow tank experiments using velocities of 0.03, 0.07 and 0.2 m/s showed that a higher proportion of barnacles fed at higher flow rates (Sanford et al., 1994). Feeding was passive, meaning the cirri were held out to the flow to catch particles; although active beating of the cirri to generate feeding currents occurs in still water (Crisp & Southward, 1961). Field observations at sites in southern New England (USA) that experience a number of different measured flow speeds found that Semibalanus balanoides from all sites responded quickly to higher flow speeds, with a higher proportion of individuals feeding when current speeds were higher. Barnacles were present at a range of sites, varying from sheltered sites with lower flow rates (maximum observed flow rates <0.06 to 0.1 m/s), a bay site with higher flow rates (maximum observed flows 0.2 to 0.3 m/s) and open coast sites (maximum observed flows 0.2 to 0.4 m/s). Recruitment was higher at the site with flow rates of 0.2 to 0.3 m/s (although this may be influenced by supply) and at higher flow microhabitats within all sites. Both laboratory and field observations indicate that flow is an important factor with effects on feeding, growth and recruitment in Semibalanus balanoides (Sanford et al., 1994; Leonard et al., 1998). However, the results suggest that flow is not a limiting factor determining the overall distribution of barnacles, as they can adapt to a variety of flow speeds.

Patella vulgata inhabits a range of tidal conditions and is therefore likely to tolerate a change in water flow rate. The streamlined profile of limpet shells is of importance in increasing their tolerance of water movement, and this is undoubtedly one factor in determining the different shapes of limpets at different exposures. With increasing exposure to wave action, the shell develops into a low profile, reducing the risk of being swept away. The strong muscular foot and a thin film of mucus between the foot and the rock enable Patella vulgata to grip very strongly to the substratum (Fretter & Graham, 1994). The ability of limpets to resist accelerating, as distinct from constant currents, may set a limit to the kind of habitat which they can occupy and limit the size to which they can grow.

Shell morphology within littorinids varies according to environmental conditions. In sheltered areas, where Carcinus maenas is more prevalent, shell apertures are small to inhibit predation. In exposed areas, the foot surface is larger to allow greater attachment, and the shell spire is lower to reduce drag (Raffaelli 1982, Crothers, 1992). Steromphala (syn. Gibbula) cineraria also appears to tolerate a range of wave exposures from exposed sites to those that are very sheltered (Frid & Fordham, 1994), suggesting it can also adapt to different water velocities. As with Littorina littorea, the morphology of the shell varies according to wave exposure, allowing individuals to adapt to different conditions in the habitat in which the larvae settle (Frid & Fordham, 1994). Steromphala (syn. Gibbula) cineraria is, however, absent from areas with very strong and turbulent flow.

Sensitivity assessment. Based on the exposure of Fucus serratus and red algal turfs and piddocks to water flows between 0.5 and 1.5 m/s in other biotopes (Connor et al., 2004), the biotope is considered to be unimpacted by changes within this range as long as these do not lead to increased erosion of the substratum. Resistance is therefore assessed as ‘High’ and resilience as ‘High’ (based on no impact to recover from), so that the biotope is considered to be ‘Not sensitive’.

Emergence regime changes

Benchmark.  1) A change in the time covered or not covered by the sea for a period of ≥1 year, or 2) an increase in relative sea level or decrease in high water level for ≥1 year. (Emergence regime change pressure definition).

Evidence

Emergence regime is a key factor structuring intertidal biotopes. Changes in emergence can lead to greater exposure to desiccation, temperature and salinity variation, reduced levels of time for filter feeding and nutrient uptake and photosynthesising opportunities for the characterizing species. Changes in emergence can also alter competitive interactions and trophic interactions such as grazing and predation. This biotope occurs in the mid shore (Connor et al., 2004) and usually occurs immediately below either a Fucus vesiculosus-barnacle mosaic (FvesB) or a Mytilus edulis and piddocks-dominated biotope (MytPid) on moderately exposed shores or a dense canopy of Fucus vesiculosus (Fves) or Ascophyllum nodosum (Asc.FS) on sheltered shores. The littoral fringe below on moderately exposed shores are dominated by the kelp Laminaria digitata (Ldig.Pid), while the kelp Laminaria saccharina may co-dominate on sheltered shores (Slat.Ldig; Slat.Ft) (JNCC, 2022). 

Fucoids can tolerate periodic desiccation but only to a limited extent. Fucus serratus is more susceptible to desiccation than other Fucus species that are located further up the shore and subjected more frequently to aerial exposure (Schonbeck & Norton, 1978). In experiments, Fucus serratus did not survive transplantation further up the shore, e.g. into the Fucus spiralis belt (Schonbeck & Norton, 1978). The critical water content for Fucus serratus was estimated at 40%, with water losses past this point causing irreversible damage. Beer et al. (2014) found that Fucus serratus could not regain any positive photosynthetic rates after rehydrating from 10% water content. The upper shore extent of Fucus serratus populations may be replaced by species more tolerant of desiccation and more characteristic of the mid-eulittoral, such as Fucus vesiculosus or Ascophyllum nodosum. 

Early life history stages will be more susceptible to this pressure (Henry & Van Alstyne, 2004). However, germlings are protected from desiccation by the canopy of adults. A study by Brawley & Jonhnson (1991) showed that germling survival under adult canopy was close to 100%, whereas survival on the adjacent bare rock was close to 0% during exposure to aerial conditions. The Fucus canopy is also likely to protect other underlying species to a great extent. Mortalities of other components of the community will, however, occur if the canopy is removed (see ‘abrasion’ pressure).

Physiological studies confirm that Fucus serratus is particularly susceptible to emersion stress compared to upper shore fucoids. During low tide, oxidative stress and photoinhibition increase in Fucus serratus, with reactive oxygen species and lipid peroxidation rising in summer conditions, whereas mid- and upper-shore species show greater resilience (Martins et al., 2021). Canopy shading and frond overlap can mitigate desiccation and light stress for understory layers, maintaining high photosynthetic activity even during emersion (Bordeyne et al., 2017; Fernández et al., 2015). Additive and interacting effects of air temperature, irradiance, and humidity partially explain distributional differences between Fucus serratus and more desiccation-tolerant congeners (Fernández et al., 2015).

Adult piddocks and the algae that characterize this biotope have no mobility and cannot migrate up or down shore to adapt to changes in emergence. Within the chalk substratum, adult piddocks will be afforded some protection from desiccation and temperature increases, following increased emergence, by their burrows, which will retain some moisture. However, the shells of piddocks do not completely enclose the animals and cannot be closed to prevent water loss. The tolerance of piddocks to increased and decreased emergence varies between species. Pholas dactylus inhabits the shallow sub-tidal and lower shore, and Barnea candida and Petricolaria pholadiformis live slightly higher up the shore (Duval, 1977). During extended periods of exposure, Pholas dactylus squirts some water from their inhalant siphon and extend their gaping siphons into the air (Knight, 1984). This may result in increased detection and predation by birds. Hiatella arctica occurs within the intertidal and subtidal, and the presence of suitable substratum rather than the emergence regime is a more significant factor determining its distribution.

A decrease in emergence will reduce exposure to desiccation and extremes of temperature and allow Pholas dactylus, Barnea candida and Petricolaria pholadiformis to feed for longer periods and hence grow faster. No information was found on factors controlling the lower limit of piddock populations, and it is possible, for example, that predation (predominantly siphon nipping by gobies, and other species, Micu, 2007) may increase at the lower edge of the biotope. Competition for space with species better adapted to the changed conditions may also alter habitat suitability for this biotope.

The red algae within the biotope are likely to be sensitive to increased emergence. Corallina officinalis is sensitive to desiccation (Dommasnes, 1969) and are generally not found on open rock unless protected by algal canopies or where the surfaces are damp or wet. At Hinkley Point (Somerset, England), for example, seawater run-off from deep pools high in the intertidal supports dense turfs of Corallina spp. lower on the shore (Bamber & Irving, 1993). Fronds are highly intolerant of desiccation and do not recover from a 15% water loss, which might occur within 40 to 45 minutes during a spring tide in summer (Wiedemann, 1994). Bleached corallines were observed 15 months after the 1964 Alaska earthquake that elevated areas in Prince William Sound by 10 m. Similarly, increased exposure to air caused by upward movement of 15 cm due to nuclear tests at Amchitka Island, Alaska adversely affected Corallina pilulifera (Johansen, 1974). During an unusually hot summer, Hawkins & Hartnoll (1985) observed damaged Corallina officinalis and other red algae. Littler & Kauker (1984) suggest that the basal crustose stage is adaptive, allowing individuals to survive periods of physical stress as well as physiological stress such as desiccation and heating. The basal crust stage may persist for extended periods, with frond regrowth occurring when conditions are favourable.

Osmundea pinnatifida turfs growing on the upper extent of its usual zone in the Isle of Man experience greater desiccation and are shorter and less dense than those lower on the shore (Prathep, 2001), suggesting that habitat quality (measured through growth) decreases with increasing shore height. In laboratory experiments short-term photosynthesis of Osmundea pinnatifida was inhibited when fronds had lost more than 50% of their water content (Prathep, 2001). Following resubmergence, fronds that had lost 50 % of water content had fully recovered (measured as photosynthesis reaching a maximal value) after 1 hour, while fronds exposed to 70% water loss took 5 hours to recover (Prathep, 2001). Repeated exposure to high levels of desiccation would clearly impact growth.

Experimental grazer removal has allowed algae, including Palmaria palmata, Ceramium sp. and Osmundea (as Laurencia) pinnatifida to grow higher on the shore (during winter and damp summers) than usual suggesting that grazing also limits the upper shore extent of this biotope (Hawkins & Hartnoll, 1985). Palmaria palmata grew more abundantly higher up the shore following the massive mortality of molluscan grazers after the Torrey Canyon oil spill (Hawkins & Hartnoll, 1983). Palmaria palmata also grew more abundantly higher up the shore following the massive mortality of molluscan grazers after the Torrey Canyon oil spill (Hawkins & Hartnoll, 1983). These observations and further grazer removal experiments by Boaventura et al. (2003), indicate that grazing, in combination with physiological tolerances, limits the upper shore extent of biotopes characterized by red algal turfs on moderately and more exposed shores, where grazing is greater than on sheltered shores (Hawkins & Hartnoll, 1983, Boaventura et al., 2003). These results agree with Underwood (1980) who showed grazing and emersion stress limit the height to which red algal turfs can extend.

The occurrence of encrusting coralline algae seems to be critically determined by exposure to air and sunlight. Colonies survive in damp conditions under algal canopies or in pools, but not on open rock, where desiccation effects are important. Increased emergence leading to drying out of shallow pools would reduce habitat suitability for this group. Spore release by the crustose coralline Lithophyllum incrustans is triggered by small changes in salinity and temperature, and therefore changes in emergence may alter patterns in reproduction and recruitment (see relevant pressures for further information). However, this species does occur both high and low in the intertidal (Edyvean & Ford, 1986) and presumably such impacts are limited.

Mobile epifauna are likely to relocate to more suitable habitats. Species such as Patella vulgata and Littorina littorea that are found throughout the intertidal zone are adapted to tolerate desiccation to some extent. For example, littorinids can seal the shell using the operculum, while limpets clamped tightly to the rock will reduce water loss.

Sensitivity assessment. The biotope occurs in the eulittoral zone, where it experiences regular immersion and emersion. Species present are therefore tolerant of periods of emergence to some extent. However, changes in the emergence regime may alter habitat suitability and increase levels of predation and competition. Other species better able to tolerate desiccation are likely to competitively displace Fucus serratus following increased emergence. A significant, long-term increase in emergence is therefore considered likely to lead to replacement of this biotope with one that is similar but more typical of the changed conditions with less red algae. Following a decrease in emergence, Fucus serratus may be replaced by Laminaria digitata leading to biotope reclassification. An increase in emergence will lead to desiccation and osmotic stress, which may increase mortality. Based on these considerations, resistance to changes in emergence is assessed as ‘Low’ as changes may alter the upper or lower margins of the biotope. Resilience is assessed as ‘Medium’, and sensitivity is assessed as ‘Medium’.

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

Fucus serratus is highly flexible but not physically robust, and wave action leads to frequent mechanical damage, breaking fronds or even dislodging algae from the substratum (Knight & Parke, 1950). Fucoids are permanently attached to the substratum and are not able to re-attach if removed. An increase or reduction in wave action at the pressure benchmark is considered to have little effect, as Fucus serratus is naturally found in both sheltered conditions and on moderately exposed and exposed shores. Wave action plays a role in determining the shore level at which Fucus serratus is found, with stands tending to occur lower on the shore where it is exposed to greater wave action (Southward & Orton, 1954). Modelled predictions of the change in percentage cover of Fucus serratus in response to climate change induced increases in wind speed (and hence wave height) suggest that Fucus serratus would be resilient to 20% and 40% increases in wind speed (Hawkins et al., 2009). Although this is not directly related to the pressure benchmark, it suggests that Fucus serratus occurs across a range of wave exposures and wave heights. The Fucus serratus canopy will provide some protection from wave action to the under-canopy species assemblage. Wave exposure has been shown to limit the size of fucoids (Blanchette, 1997) as smaller individuals create less resistance to wave action. Mature plants are therefore more sensitive to this pressure. As exposure increases, the fucoid population would become dominated by small juvenile algae. In addition to physical stress, water movement also strongly influences reproduction, with calm conditions generally required for successful gamete release in fucoids (Brawley & Johnson, 1991; Serrão et al., 1996; Berndt et al., 2002), indicating that changes in wave exposure could also affect recruitment even if adult thalli remain intact.

The piddocks and Hiatella arctica are unlikely to be directly affected by changes in wave exposure, owing to their environmental position within the substratum, which protects them. Trudgill & Crabtree (1987) found Hiatella arctica at both sheltered and wave exposed sites, suggesting that substratum, rather than wave action, is a more significant factor determining distribution. On chalk and clay substrata, it is possible, however, that wave action actively erodes the substratum at a faster rate than the piddocks can burrow, leaving them exposed to predators or displaced. At higher densities, bioerosion by piddocks may destabilise the substratum, increasing vulnerability to erosion.

The most damaging effect of increased wave heights would be the erosion of the substratum, as this could eventually lead to loss of the habitat. Increased erosion would lead to the loss of habitat and removal of piddocks. No evidence was found to link significant wave height to erosion. Some erosion will occur naturally, and storm events may be more significant in loss and damage of clays than changes in wave height at the pressure benchmark. For example, Micu (2007) observed numerous Pholas dactylus that had been washed out of the clay substratum or exposed due to storm damage to the clay in the Romanian Black Sea. Erosion rates at the Cretaceous chalk cliffs in East Sussex on the south coast of the UK have accelerated by 22 to 32 cm/year due to natural and anthropogenic modification of the coast (Hurst et al., 2016).

As water velocity increases, algae can flex and reconfigure to reduce the size of the algae when aligned with the direction of flow, which minimises drag and the risk of dislodgement (Boller & Carrington, 2007). Within a canopy, the friction between the fronds and water slows flow, reducing drag. On exposed shores, larger, dense patches of Osmundea pinnatifida were more able to withstand increased wave action in winter than small patches, which were severely damaged, presumably due to the number and density of stolons (Prathep, 2001). These characteristics allow these species to persist on shores that experience a range of wave action levels. Flat growth forms also minimise drag, and crustose bases and encrusting corallines are able to withstand high levels of water movement. Colonies of Lithophyllum incrustans appear to thrive in conditions exposed to strong water movement, and Irvine & Chamberlain (1994) observe that the species is best developed on wave exposed shores.

A decrease in wave exposure may ultimately reduce Patella vulgata abundance because the species does not favour thick algal cover that is often present on more sheltered shores. Alternatively, an increase in significant wave height, linked to increased exposure, may result in population changes with fewer macroalgae and with more Chthamalus sp. present than Semibalanus balanoides (Ballantine, 1961) and the limpet Patella ulyssiponensis present, or present in greater numbers, rather than Patella vulgata (Thompson, 1980). These changes are not considered to lead to a significant change in biotope character as species replacements are functionally similar.

Sensitivity assessment. No direct evidence was found to assess this pressure at the benchmark level. Based on the distribution of Fucus serratus on shores of varying exposure and the position of the characterising bivalve species, the biotope is considered to be ‘Not sensitive’ to this pressure at the benchmark level.

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.

Fucus serratus has been shown to accumulate transition metals such as copper (Cu) and cadmium (Cd) at higher rates than other algal species in the White Sea, Russia (Andreev & Plakhotskaya, 2019). In Anglesey, Wales, macroalgae, including Fucus serratus, collected directly below an abandoned copper mine outflow showed extremely elevated metal concentrations (>250 mg Fe/g, >6 mg Cu/g, >2 mg Zn/g, >190 μg As/g) and evidence of toxicity, although pollution effects were largely confined within 200 m of the source (Chalkley et al., 2019). These studies demonstrate the species’ capacity for metal bioaccumulation and its potential use as a bioindicator of metal contamination.

Bryan (1984) suggested that the general order for heavy metal toxicity in seaweeds is: Organic Hg > inorganic Hg > Cu > Ag > Zn > Cd > Pb. Cole et al. (1999) reported that Hg was very toxic to macrophytes. The sub-lethal effects of Hg (organic and inorganic) on the sporelings of another intertidal red algae, Plumaria elegans, were reported by Boney (1971), where 100% growth inhibition was caused by 1 ppm Hg in his study. Heavy metals have the potential to accumulate in plant tissue; therefore, it may take some time for tissue levels to fall before recovery can begin.

No evidence was found for the sensitivity of piddocks.

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.

However, exposure to contaminants at levels greater than the benchmark may lead to impacts, although no evidence was found for the sensitivity of piddocks. O’Brien & Dixon (1976) suggested that red algae were the most sensitive group of algae to oil or dispersant contamination, possibly due to the susceptibility of phycoerythrins to destruction, but that the filamentous forms were the most sensitive. Laboratory studies of the effects of oil and dispersants on several red algae species, including Palmaria palmata (Grandy, 1984, cited in Holt et al. 1995), concluded that they were all sensitive to oil/ dispersant mixtures, with little differences between adults, sporelings, diploid or haploid life stages.

In areas of moderate oil deposit, up to about 1/2cm thick, on rocks after the Torrey Canyon oil spill, limpets had survived unscathed for over a month after the event, and feeding continued even though a coating of oil smothered their food source of algae and diatoms (Smith, 1968). Limpets can ingest thick oil and pass it through their gut. However, thick layers of oil smothering individuals will interfere with respiration and spoil normal food supplies for Patella vulgata. After the Braer oil spill, in common with many other oil spills, the major impact in the intertidal zone was on the population of limpets and other grazers. In West Angle Bay, where fresh oil from the Sea Empress tanker reached rocky shores within one day of the spill, limpet mortality was 90% (Glegg et al., 1999). Patella vulgata has higher intolerance to fresh oil, which has a high component of volatile hydrocarbons remaining. A significant reduction in the density of juvenile limpets was also observed at all sites known to have been oiled by the Sea Empress spill (Moore, 1997). In longer-term studies into the environmental effects of oil refinery effluent discharged into Littlewick Bay, Milford Haven, the number of limpets, usually found in substantial numbers on this type of shore, was considerably reduced in abundance in areas close to the discharge (Petpiroon & Dicks, 1982). In particular, only large individuals were found close to the outfall point, and juveniles were completely absent, suggesting that observed changes in abundance resulted from effluent effects on larval stages rather than upon adults directly.

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.

However, exposure to synthetic chemicals at levels greater than the benchmark may lead to impacts, although no evidence was found for the sensitivity of piddocks. O’Brien & Dixon (1976) suggested that red algae were the most sensitive group of algae to oil contamination, although the filamentous forms were the most sensitive. Laboratory studies of the effects of oil and dispersants on several red algae species (Grandy, 1984 cited in Holt et al., 1995) concluded that they were all intolerant of oil/ dispersant mixtures, with little differences between adults, sporelings, diploid or haploid life stages. Cole et al. (1999) suggested that herbicides, such as simazina and atrazine, were very toxic to macrophytes. Hoare & Hiscock (1974) noted that all red algae were excluded from Amlwch Bay, Anglesey by acidified halogenated effluent discharge.

Limpets are extremely intolerant of aromatic solvent-based dispersants used in oil spill clean-up. During the clean-up response to the Torrey Canyon oil spill, nearly all the limpets were killed in areas close to dispersant spraying. Viscous oil will not be readily drawn in under the edge of the shell by ciliary currents in the mantle cavity, whereas detergent, alone or diluted in seawater, would creep in much more readily and be liable to kill the limpet (Smith, 1968). A concentration of 5 ppm killed half the limpets tested in 24 hours (Southward & Southward, 1978; Hawkins & Southward, 1992). Acidified seawater affects the motility of Patella vulgata. At a pH of 5.5, motility was reduced whilst submerged, but individuals recovered when returned to normal seawater. At a pH of 2.5, total inhibition of movement occurred, and when returned to normal seawater, half had died (Bonner et al., 1993). Reduced motility reduces time for foraging and may result in decreased survival of individuals. Acidified seawater can also change the shell composition, which will lead to a decrease in its protective nature and hence survival (Bonner et al., 1993). Short periods (48 hours) are unlikely to have much effect on a population but long periods (1 year) may cause reduced grazing and an increase in algal growth. However, seawater is unlikely to reach pH 2.5 therefore, intolerance to slight changes in pH will be low. Hoare & Hiscock (1974) reported that in Amlwch Bay, Patella vulgata was excluded from sites within 100 to 150 m of the discharge of acidified, halogenated effluent.

Radionuclide contamination

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

Evidence

No evidence was found to assess this pressure at the benchmark. Algae bioaccumulate radionuclides (with extent depending on the radionuclide and the algae species). Adverse effects have not been reported at low levels.

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

No specific information on oxygen consumption and reduced oxygen tolerances for important characterizing species within the biotope was found. Cole et al. (1999) suggested possible adverse effects on marine species below 4 mg O₂/L and probable adverse effects below 2 mg O₂/l.

Reduced oxygen concentrations have been shown to inhibit both photosynthesis and respiration in macroalgae (Kinne, 1977). Despite this, macroalgae are thought to buffer the environmental conditions of low oxygen, thereby acting as a refuge for organisms in oxygen-depleted regions, especially if the oxygen depletion is short-term (Frieder et al., 2012). Duval (1963a) observed that conditions within the borings of Petricolaria pholadiformis were anaerobic and lined with a loose blue/black sludge, suggesting that the species may be relatively tolerant to conditions of reduced oxygen.

The associated invertebrate species also show high tolerances for reduced oxygen at levels that exceed the pressure benchmark. Semibalanus balanoides can respire anaerobically, so they can tolerate some reduction in oxygen concentration (Newell, 1979). When placed in wet nitrogen, where oxygen stress is maximal and desiccation stress is low, Semibalanus balanoides have a mean survival time of 5 days (Barnes et al., 1963). Limpets can also survive for a short time in anoxic seawater; Grenon & Walker, (1981) found that in oxygen free water limpets could survive up to 36 hours, although Marshall & McQuaid (1989) found a lower tolerance for Patella granularis, which survived up to 11 hours in anoxic water. Patella vulgata and littorinids can respire in air, mitigating the effects of this pressure during the tidal cycle.

This biotope would only be exposed to low oxygen in the water column intermittently during periods of tidal immersion. In addition, in areas of wave exposure and/or moderately strong current flow, low oxygen levels in the water are unlikely to persist for very long, as oxygen levels will be recharged by the incorporation of oxygen in the air into the water column or flushing with oxygenated waters.

Sensitivity assessment. As the biotope will only be exposed to this pressure when emersed and respiration of characterizing and associated species will occur in air, biotope resistance is assessed as ‘High’ and resilience as ‘High’ (no effect to recover from), resulting in a sensitivity of ‘Not sensitive’. 

Nutrient enrichment

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

Evidence

This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations. No evidence was found to assess the sensitivity of piddocks to this pressure, and it is unlikely that they, and other animal species present in the biotope, would be directly affected by this pressure.

Marine algae are often nutrient limited, by nitrogen in particular, so an increase in nutrient levels usually results in increased growth and fecundity. In the Bay of Fundy, for example, where there is a tidal flux of nutrients from the marshes, there is luxurious growth of Palmaria palmata. However, very high levels of nutrients can be toxic to macroalgae (Morgan et al., 1980). In general, the great majority of reports refer to an increase in the number of green algae associated with eutrophicated waters, usually at the expense of red and brown algae. Nutrient enrichment generally stimulates ephemeral macroalgae growth (Duarte et al., 1995). This stimulation of annual ephemerals may accentuate the competition for light and space and hinder perennial species development or harm their recruitment (Kraufvelin, 2007). Kraufvelin et al. (2006) found only minor effects on the fucoid community structure as a response to high nutrient levels (32 μM N and 2 μM P) during the first three years of their experiment. However, during the fourth year of exposure, Fucus serratus started to decline, and the population consequently crashed in the fifth year. The study observed full recovery of the algal canopy and animal community in less than two years after conditions returned to normal. The results indicated that established rocky shore communities of perennial algae with associated fauna were able to persist for several years, even at very high nutrient levels, but that community shifts may suddenly occur if eutrophication continues. They also indicated that rocky shore communities have the ability to return rapidly to natural, undisturbed conditions after the termination of nutrient enhancement.

Knoop et al. (2022) investigated nutrient enrichment associated with finfish aquaculture wastewater in Wales, UK. Fucus serratus, grown in mono- and polycultures with other macroalgal species, showed increased biomass production under nutrient-rich conditions (871 ± 22 μM TON (total organic nitrogen), 32 ± 0.4 μM NH₄⁺, and 141 ± 2 μM PO₄³). Ammonium removal increased by 25% and total oxidised nitrogen by nearly 10% in polycultures, indicating that Fucus serratus can persist and contribute to nutrient processing under elevated nitrogen inputs. Viana & Bode (2015) assessed δ¹⁵N (a stable nitrogen isotope commonly used to trace the source of nitrogen) in Fucus serratus along a salinity gradient in NW Spain. Tissue δ¹⁵N values reflected cumulative nitrogen inputs over long periods, rather than short-term fluctuations, demonstrating the species’ ability to tolerate and integrate chronic nutrient enrichment.

Atalah & Crowe (2010) added nutrients to rockpools occupied by a range of algae, including encrusting corallines, turfs of Mastocarpus stellatus, Chondrus crispus and Corallina officinalis and green and red filamentous algae. The invertebrates present were mostly Patella ulyssiponensis, the winkle Littorina littorea and the flat top shell Steromphala umbilicalis. Nitrogen and phosphorus enhancement were achieved via the addition of fertilisers at either 40 g/l or 20 g/l. The treatments were applied for seven months, and experimental conditions were maintained every two weeks. The experimental treatments do not directly relate to the pressure benchmark but indicate some general trends in sensitivity. The cover of green filamentous algae was significantly increased by both reduced grazing and increased nutrients. However, the effect size was synergistically magnified by the combined effect of grazer removal and nutrients. Nutrient enrichment caused an absolute increase in the average cover of green filamentous algae of 19% (±3.9 S.E.) with respect to the control treatments, while the cover of red turfing algae was not affected by nutrient addition (Atalah & Crowe, 2010).

Nutrient enrichment that enhances phytoplankton productivity may indirectly benefit the characteristic bivalve species of this biotope by increasing food supply.

Sensitivity assessment. The above evidence suggests that Fucus serratus beds are resistant to nutrient enriched conditions, except prolonged exposure to very high nutrient conditions. Therefore, both resistance and resilience are assessed as ‘High’. The biotope group is therefore assessed as ‘Not Sensitive’ to this pressure, but with 'Low' confidence due to a lack of direct evidence on piddocks.

Organic enrichment

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

Evidence

Organic enrichment and nutrient enrichment commonly co-occur; for example, sewage deposits or outputs from fish farms may enhance nitrogen, phosphorus, and organic matter. Nutrient enrichment alters the selective environment by favouring fast growing, ephemeral species such as Ulva lactuca and Ulva intestinalis (Berger et al., 2004; Kraufvelin, 2007). Rohde et al. (2008) found that both free growing filamentous algae and epiphytic microalgae can increase in abundance with nutrient enrichment. This stimulation of annual ephemerals may accentuate the competition for light and space and hinder perennial species development or harm their recruitment (Berger et al., 2003; Kraufvelin, 2007). Nutrient enrichment can also enhance fouling of Fucus fronds by biofilms (Olsenz, 2011). Nutrient enriched environments not only increase algae abundance, but also the abundance of grazing species (Kraufvelin, 2007). High nutrient levels may directly inhibit spore settlement and hinder the initial development of Fucus vesiculosus (Bergström et al., 2003). Bellgrove et al. (2010) found that coralline turfs outcompeted fucoids at a site associated with organic enrichment caused by an ocean sewage outfall. 

No evidence was found to assess the sensitivity of piddocks to this pressure.

Sensitivity assessment. The algae within the biotope are not considered likely to be directly affected by an increase in organic matter at the pressure benchmark. The fronds of algae may intercept particles and may remove these from the chalk surfaces when emersed, by the movement of fronds. Suspension feeders, including barnacles and piddocks, may be able to utilise particles as food. Piddocks are likely to be able to withstand a small level of deposition of organic matter (at the pressure benchmark). Resistance is assessed as ‘High’ and resilience as ‘High’, so the biotope is considered to be ‘Not sensitive’, but with 'Low confidence due to the lack of direct evidence.

Physical loss (to land or freshwater habitat)

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

Evidence

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

Physical change (to another seabed type)

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

Evidence

Wangkulangkul et al. (2016) demonstrated that substratum stability influences adult persistence of Fucus serratus on rocky shores in North Wales. Although settlement could occur within Mytilus edulis beds, mature thalli attached to mussel shells required less force to detach than those attached directly to rock, with failure typically occurring between mussels and the underlying rock. Thalli larger than 60 cm were recorded only on rock and not on mussel shells. The authors concluded that mussel-dominated habitat provides a less stable attachment surface for adult individuals and may reduce reproductive output. This evidence supports the importance of stable hard substratum for the maintenance of mature canopy-forming populations. This biotope is characterized by the soft chalk substratum which supports populations of burrowing piddocks (JNCC, 2022). A change to a sedimentary or hard impenetrable substratum would result in the loss of piddocks significantly altering the character of the biotope. Resistance to this pressure is therefore ‘None’, resilience is assessed as ‘Very low’ and sensitivity is assessed as ‘High’.

Physical change (to another sediment type)

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

Evidence

Not relevant. This pressure is relevant only to sedimentary biotopes.

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 removal of substratum to a depth of 30 cm will remove the entire biological assemblage in the impact footprint. Resistance in that case would therefore be assessed as ‘None’, and resilience would be ‘Medium’ (2 to 10 years), but see caveats in the resilience section above. However, the biotope is dependent on the presence of chalk or other soft rock. So, in cases where the substratum is entirely removed, resilience would be ‘Very low’. Sensitivity is therefore assessed as ‘High’, based on the lack of recovery on the substratum.

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

This biotope is found in the intertidal, an area easily accessible by humans, especially at low tide. Most macroalgae are very flexible but not physically robust. The trampling of shores by humans will result in increased breakage of algal thalli, decreased thallus height, and a net reduction in biomass (Tyler-Walters, 2005; Tyler-Walters & Arnold, 2008).

In the UK, Boalch et al. (1974) and Boalch & Jephson (1981) noted a reduction in the cover of fucoids at Wembury, South Devon, when compared to surveys conducted by Colman (1933). The size ranges of Ascophyllum nodosum, Fucus vesiculosus and Fucus serratus were skewed to a smaller length, and the abundance of Ascophyllum nodosum, in particular, was reduced (Boalch & Jephson, 1981). It was suggested that visitor pressure, especially after the construction of a car park, was responsible for the reduced cover of fucoids (Boalch et al., 1974). They suggested that the raised edges of the slatey rock severed fronds when the rocks were walked over. However, no quantitative data were provided.

Pinn & Rodgers (2005) compared a heavily visited ledge with a less visited ledge at Kimmeridge Bay, Dorset. Although the mean species richness was similar at both sites, the total number of species was greater at the less utilised site. Comparatively, the heavily utilised ledge displayed a reduction in larger, branching algal species (e.g. Fucus serratus) and increased abundances of ephemeral and crustose species (e.g. Ulva linza and Lithothamnia spp., respectively). Fletcher & Frid (1996a; 1996b) examined the effects of persistent trampling on two sites on the northeast coast of England. The trampling treatments used were 0, 20, 80, and 160 steps per m² per spring tide for eight months between March and November. Using multivariate analysis, they noted that changes in the community dominated by fucoids (Fucus vesiculosus, Fucus spiralis and Fucus serratus) could be detected within 1 to 4 months of trampling, depending on intensity. Intensive trampling (160 steps/m² /spring tide) resulted in a decrease in species richness at one site. The area of bare substratum also increased within the first two months of trampling but declined afterwards, although bare space was consistently most abundant in plots subject to the greatest trampling (Fletcher & Frid, 1996a, 1996b). The abundance of fucoids was consistently lower in trampled plots than in untrampled plots. Fletcher and Frid (1996a) noted that the species composition of the algal community was changed by as little as 20 steps per m² per spring tide of continuous trampling since recolonization could not occur. A trampling intensity of 20 steps per m² per spring tide could be exceeded by only five visitors taking the same route out and back again across the rocky shore in each spring tide. Both of the sites studied receive hundreds of visitors per year, and damage is generally visible as existing pathways, which are sustained by continuous use (Fletcher & Frid, 1996a, 1996b). However, the impact was greatest at the site with the lower original abundance of fucoids.

Brosnan & Crumrine (1994) noted that trampling significantly reduced algal cover within 1 month of trampling. Foliose algae were particularly affected and decreased in cover from 75% to 9.1% in trampled plots. Mastocarpus papillatus decreased in abundance from 9% to 1% in trampled plots but increased in control plots. Fucus distichus decreased in the summer months only to recover in winter, but in trampled plots, it remained in low abundance (between 1 and 3% cover). Trampling resulted in a decrease in the cover of Pelvetiopsis limitata from 16% to 1.5%. Iridaea cornucopiae decreased from 38 to 14% cover within a month and continued to decline to 4 to 8% cover. However, after trampling ceased, recovery of algal cover, including Iridaea cornucopiae and Mastocarpus papillatus, was rapid (approx. 12 months) (Brosnan & Crumrine, 1994). Fletcher & Frid (1996a; 1996b) reported a decrease in the understorey algal community of encrusting coralline algae and red algae, which was probably an indirect effect due to increased desiccation after removal of the normally protective fucoid canopy (see Hawkins & Harkin, 1985) by trampling. They also noted that opportunistic algae (e.g. Ulva sp.) increased in abundance. Schiel & Taylor (1999) also observed a decrease in understorey algae (erect and encrusting corallines) after 25 or more tramples, probably due to an indirect effect of increased desiccation as above. However, Schiel & Taylor (1999) did not detect any variation in other algal species due to trampling effects. Similarly, Keough & Quinn (1998) did not detect any effect of trampling on algal turf species.

Algal turfs seem to be relatively tolerant of the direct effects of trampling (based on the available evidence), and some species may benefit from the removal of canopy-forming algae (Tyler-Walters, 2005). Their tolerance may result from their growth form, as has been shown for vascular plants and corals (Liddle, 1997). Brosnan (1993) suggested that algal turf dominated areas (on shores usually dominated by fucoids) were indicative of trampling on the rocky shores of Oregon. However, tolerance is likely to vary with species, and their growth form and little species-specific data was found. Furthermore, algal turfs may suffer negative indirect effects when they form an understorey below canopy forming species.

Conversely, fucoid algae are particularly intolerant of trampling, depending on intensity. Fucoid algae demonstrate a rapid (days to months) detrimental response to the effects of trampling, depending on species, which has been attributed to either the breakage of their fronds across rock surfaces (Boalch et al., 1974) or their possession of small discoid holdfasts that offer little resistance to repeated impacts (Brosnan & Crumrine, 1992; Fletcher & Frid, 1996b). Foliose species such as Mastocarpus papillatus, Pelvetiopsis limitata and Iridaea cornucopiae are also likely to be intolerant of trampling (Brosnan & Crumrine, 1994). Brosnan (1993) suggested that the presence or absence of foliose algae (e.g. fucoids) could be used to indicate the level of trampling on the rocky shores of Oregon.

Once Fucus serratus has been removed, understorey algae will become exposed. Macroalgae canopies buffer the effects of high temperatures and water loss on organisms below their fronds, in particular when exposed to air. For instance, Bertness et al. (1999) determined that the substratum temperature was on average 8 to 10°C lower under the canopy than on bare rock. Desiccation of understorey algae will create bare patches (see ‘changes in emergence regime’ pressure). These bare patches can lead to invasions by grazing limpets, which in turn can promote even greater changes in community composition (Little et al., 2009). The removal of the macroalgal canopy due to abrasion will therefore have a direct impact on the entire community. However, cracks and crevices are ideal places for germlings to develop and sessile species to settle, as these sites may be protected from abrasion. Stagnol et al. (2013) found that opportunistic ephemeral green algae such as Ulva sp. responded positively to disturbance. These green ephemeral algae are major competitors of Fucus serratus for space colonization and nutrient uptake. Blooms of ephemeral algae facilitated by disturbance may then slow the development of longer-lived perennial algae, especially fucoids. Disturbance is a structuring factor in intertidal habitats. Perturbation events often remove organisms, increase mortality, and release resources such as space, nutrients and light that may enhance the appearance of new colonists (Connell et al., 1997). As a result of these contrasting effects, post-disturbance communities are frequently different from initial communities in terms of composition and dominance of species. Overall, disturbance causes a shift towards a disturbance tolerant seaweed community (Little et al., 2009).

Epifaunal species are particularly adversely affected by physical disturbance, either due to direct damage or modification of the habitat (Jennings & Kaiser, 1998). Similarly, Dayton (1971) observed a greatly reduced abundance of species living on, under, and among fucoids following large disturbance events. Hydroids, bryozoans, and encrusting fauna are easily ripped from the substratum and are unlikely to re-attach and will die. The shells of limpets, tubeworms and periwinkles may be crushed by the weight and force of the abrasion. However, some epifaunal species have been reported to exhibit increased abundances in high fishing effort areas, probably due to their ability to colonize and grow rapidly (Bradshaw et al., 2000). For instance, Ascidiella species had increased in abundance in an area subject to scallop dredging (Bradshaw et al., 2002). The breadcrumb sponge Halichondria panicea is attached to the substratum and will not survive abrasion and physical disturbance. Hiscock (1983) noted that a community, under conditions of scour and abrasion from stones and boulders moved by storms, developed into a community consisting of fast-growing species such as Spirobranchus triqueter due to decreased competition. A shift in community composition is therefore expected immediately after the disturbance event.

The effects of trampling are dependent on intensity, expressed as frequency and force per unit area of the impacting ‘footprint’ (see Liddle, 1997; Tyler-Walters & Arnold, 2008). Clearly, mechanical abrasion due to vehicles, jack-up-barges, or grounding vessels will exceed the abrasive ‘intensity’ of trampling by humans or livestock.

Although the piddocks are afforded some protection from surface abrasion by living in their burrows, damage to the chalk may leave individuals, especially those near the surface, vulnerable to damage and death through exposure, sediment damage, and compaction. For example, Micu (2007) observed that after storms in the Romanian Black Sea, the round goby, Neogobius melanostomus, removed clay from damaged or exposed burrows to eat piddocks.

The most significant impact may be on the chalk substratum by removing or damaging surface layers, resulting in the chalk being more vulnerable to erosion. Natural erosion processes are, however, likely to be ongoing within this habitat type. Where abundant, the boring activities of piddocks contribute significantly to bioerosion, which can make the substratum habitat more unstable and can result in increased rates of coastal erosion (Evans 1968a; Trudgill 1983; Trudgill & Crabtree, 1987). Pinn et al. (2005) estimated that over the lifespan of a piddock (12 years), up to 41% of the shore could be eroded to a depth of 8.5 mm. Therefore, surface erosion is a natural part of the environmental processes, although rates could be enhanced by surface abrasion and disturbance. Erosion rates at the Cretaceous chalk cliffs in East Sussex on the south coast of the UK have accelerated by 22 to 32 cm/year due to natural and anthropogenic modification of the coast (Hurst et al., 2016).

Sensitivity assessment. Surface abrasion may remove the algae and epifauna, resulting in the loss of some piddocks and damage to the habitat. Resistance is therefore assessed as ‘Low’. As the substratum cannot recover, resilience is assessed as ‘Very Low’, and the sensitivity of the overall biotope is considered to be ‘High’.

Penetration or disturbance of the substratum subsurface

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

Evidence

Penetration and disturbance below the surface of the substratum will damage and remove the algae and surface fauna, and could damage and expose piddocks. Piddocks in damaged burrows or those that are removed from the substratum are unlikely to be able to rebury and will be predated by fish and other mobile species (Micu, 2007). The most significant impact may be the damage and removal of the chalk substratum. Where abundant, the boring activities of piddocks can make the substratum habitat more unstable and can exacerbate erosion (Evans 1968a; Trudgill 1983; Trudgill & Crabtree, 1987). Pinn et al. (2005) estimated that over the lifespan of a piddock (12 years), up to 41% of the shore could be eroded to a depth of 8.5 mm. This would make the substratum more vulnerable to damage and removal.

Sensitivity assessment. Sub-surface penetration and disturbance will result in the loss of piddocks and damage to the habitat. Resistance is therefore assessed as ‘Low’ for the macroalgae, the piddocks, and the substratum. Resilience for the macroalgae and piddocks is assessed as ‘Medium’ where suitable substratum remains, so their sensitivity is ‘Medium’. Clay and chalk habitats are restricted in distribution and have been identified as irreplaceable habitats (Tillin et al., 2022). When removed, there is no mechanism by which the substratum can be replaced. Therefore, when removed in part or entirely, no recovery of habitat is possible, so resilience is assessed as ‘Very low’ (>25 years) and sensitivity as ‘High’.

Changes in suspended solids (water clarity)

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

Evidence

Intertidal biotopes will only be exposed to this pressure when submerged during the tidal cycle and therefore have limited exposure. Siltation, which may be associated with increased suspended solids and the subsequent deposition of these, is assessed separately (see siltation pressures). In general, increased suspended particles reduce light penetration and increase scour and deposition. Changes in suspended solids may enhance food supply to filter or deposit feeders (where the particles are organic in origin) or decrease feeding efficiency (where the particles are inorganic and require greater filtration effort). This biotope is sand or silt affected (Connor et al., 2004) and may therefore experience episodes with relatively high levels of turbidity from resuspension, and an increase at the pressure benchmark for extended periods may exceed the tolerances of sensitive species.

Light is an essential resource for all photoautotrophic organisms. Changes in suspended solids affecting water clarity will have a direct impact on photosynthesis in Fucus serratus. Irradiance below the light compensation point of photosynthetic species can compromise carbon accumulation (Middelboe et al., 2006). However, turbidity is only relevant when the biotope is covered with water, as seaweed photosynthesis declines on emersion and recommences when it is covered with water. Increased siltation may cover the frond surface of Fucus serratus with a layer of sediment, further reducing photosynthesis and growth rate. Sediment deposition can also interfere with the attachment of microscopic stages of seaweeds, reducing recruitment (see ‘siltation’ pressures). In extreme turbidity, such as found in the Bristol Channel, Fucus serratus is excluded from the bottom of the intertidal (below 2 m above chart datum) due to the lack of light for sustained growth (Chapman, 1995). 

Hiatella arctica is a filter-feeding bivalve. Many other bivalves have efficient mechanisms to remove inorganic particles via pseudofaeces. For example, Petricolaria pholadiformis can cope in water laden with suspended material by binding the material in mucus and using the palps to reject it (Purchon, 1955). Hiatella arctica is protected from scour within burrows, and increased organic particles may provide a food subsidy. Increased suspended sediments may impose sub-lethal energetic costs on bivalves by reducing feeding efficiency and requiring the production of pseudofaeces, with impacts on growth and reproduction.

Regression models developed by Bourget et al. (2003) found that temperature and water transparency (measured in metres and indicating the level of inorganic suspended solids) explained only 40% of the variation in biomass of Hiatella arctica fouling navigation buoys in the Gulf of St Lawrence system (Canada). These findings suggest that other variables play a more significant role in determining settlement, survival, and growth over a year in this system. However, the models indicated that biomass was higher where water transparency was greater (around 15 m) and declined at higher levels of suspended solids (transparency 5 m), although a causal link was not identified (Bourget et al., 2003).

Increases in sediment trapping, turf-forming red algae at the expense of canopy-forming species have been observed worldwide in temperate systems and have been linked to increased suspended solids linked to human activities worldwide (Airoldi, 2003). As turfs of Osmundea pinnatifida trap sediments (Prathep et al., 2003), it is clear that this species has some resistance to abrasion and scour from sediment particles. Corallina species accumulate more sediment than any other alga (Hicks, 1985). Hence, an increase in suspended sediment is likely to accumulate in the patches of Corallina officinalis. A significant increase may result in smothering (see siltation pressures). An accumulation of sediment within the turf may attract more sediment-dwelling interstitial invertebrates such as nematodes, harpacticoids, and polychaetes, although in more wave exposed locations, accumulation of sediment is likely to be minimal. Increased suspended sediment may also result in increased scour, which may adversely affect Fucus vesiculosus and foliose red algae, and interfere with settling spores and recruitment if sedimentation coincides with their major reproductive period. However, coralline algae, especially the crustose forms, are thought to be resistant to sediment scour (Littler & Kauker, 1984) and will probably not be adversely affected at the benchmark level.

Red algae and coralline algae, especially, are known to be shade tolerant and are common components of the understorey on seaweed dominated shores. Therefore, limited shading from suspended sediments is not considered to negatively affect this genus. Palmaria palmata is often found under partially shaded conditions as an epiphyte on the stems of Laminaria spp. (Morgan et al. 1980) in the sublittoral zone (Lüning 1990). In the Bay of Fundy, where the tidal flux of nutrients from the marshes includes a high level of suspended sediment, Palmaria palmata grows well despite high turbidity. Irvine (1983) observed morphological adaptation of the plant in fairly sheltered, silty conditions; sometimes the blade divisions are wedge-shaped and finely dissected above, or the blade has numerous linear divisions throughout. This form likely reduces possible smothering from increased siltation due to increased suspended sediment. In the absence of nutrients, a short-term increase in turbidity may affect growth and reproduction; however, as a perennial, the adults will probably survive. Other red algal species tolerate high levels of suspended solids. Chondrus crispus occurs in areas of sand covered rock in the subtidal biotope IR.HIR.KSed.ProtAhn suggesting it is very resistant to high levels of turbidity and scour associated with high levels of resuspended particles.

Hyslop et al. (1997) found that Palmaria palmata and Ulva spp. were reduced or absent on sites affected by high levels of resuspended colliery waste particles; however, the tougher fucoids were less affected. It is not clear how the levels of suspended solids experienced by these sites relate to the pressure benchmark.

No direct evidence was found to assess this pressure for piddocks and other invertebrate species. The piddocks are protected from scour within burrows, and increased organic particles could provide a food subsidy. Pholas dactylus occurs in habitats such as soft chalks, where turbidity may be high and is therefore unlikely to be affected by an increase in suspended sediments at the pressure benchmark. Piddocks, in common with other suspension feeding bivalves, have efficient mechanisms to remove inorganic particles via pseudofaeces. Experimental work on Pholas dactylus showed that large particles can either be rejected immediately in the pseudofaeces or passed very quickly through the gut (Knight, 1984). Similarly, Petricolaria pholadiformis can tolerate high levels of suspended solids through the production of pseudofaeces (Purchon, 1955). Increased suspended sediments may impose sub-lethal energetic costs on piddocks by reducing feeding efficiency and requiring the production of pseudofaeces, with impacts on growth and reproduction.

A significant decrease in suspended organic particles may reduce food input to the biotope resulting in reduced growth and fecundity of piddocks. However, local primary productivity may be enhanced where suspended sediments decrease, increasing food supply. Decreased suspended sediment may increase macroalgal competition, enhancing diversity, but is unlikely to significantly change the character of the biotope.

Sensitivity assessment. While the characteristic fauna are resistant to this pressure, changes in suspended solids, reducing water clarity, could have adverse effects on Fucus serratus photosynthesis. Resistance is therefore assessed as ‘Medium’ at the benchmark level. Once conditions return to ‘normal’, Fucus serratus is likely to rapidly regain photosynthesis and growth rate. Resilience is assessed as ‘High’, and sensitivity as ‘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

Intertidal biotopes will only be exposed to this pressure when submerged during the tidal cycle and therefore have limited exposure. Siltation, which may be associated with increased suspended solids and the subsequent deposition of these, is assessed separately (see siltation pressures). In general, increased suspended particles reduce light penetration and increase scour and deposition. Changes in suspended solids may enhance food supply to filter or deposit feeders (where the particles are organic in origin) or decrease feeding efficiency (where the particles are inorganic and require greater filtration effort). This biotope is sand or silt affected (Connor et al., 2004) and may therefore experience episodes with relatively high levels of turbidity from resuspension, and an increase at the pressure benchmark for extended periods may exceed the tolerances of sensitive species.

Light is an essential resource for all photoautotrophic organisms. Changes in suspended solids affecting water clarity will have a direct impact on photosynthesis in Fucus serratus. Irradiance below the light compensation point of photosynthetic species can compromise carbon accumulation (Middelboe et al., 2006). However, turbidity is only relevant when the biotope is covered with water, as seaweed photosynthesis declines on emersion and recommences when it is covered with water. Increased siltation may cover the frond surface of Fucus serratus with a layer of sediment, further reducing photosynthesis and growth rate. Sediment deposition can also interfere with the attachment of microscopic stages of seaweeds, reducing recruitment (see ‘siltation’ pressures). In extreme turbidity, such as found in the Bristol Channel, Fucus serratus is excluded from the bottom of the intertidal (below 2 m above chart datum) due to the lack of light for sustained growth (Chapman, 1995). 

Hiatella arctica is a filter-feeding bivalve. Many other bivalves have efficient mechanisms to remove inorganic particles via pseudofaeces. For example, Petricolaria pholadiformis can cope in water laden with suspended material by binding the material in mucus and using the palps to reject it (Purchon, 1955). Hiatella arctica is protected from scour within burrows, and increased organic particles may provide a food subsidy. Increased suspended sediments may impose sub-lethal energetic costs on bivalves by reducing feeding efficiency and requiring the production of pseudofaeces, with impacts on growth and reproduction.

Regression models developed by Bourget et al. (2003) found that temperature and water transparency (measured in metres and indicating the level of inorganic suspended solids) explained only 40% of the variation in biomass of Hiatella arctica fouling navigation buoys in the Gulf of St Lawrence system (Canada). These findings suggest that other variables play a more significant role in determining settlement, survival, and growth over a year in this system. However, the models indicated that biomass was higher where water transparency was greater (around 15 m) and declined at higher levels of suspended solids (transparency 5 m), although a causal link was not identified (Bourget et al., 2003).

Increases in sediment trapping, turf-forming red algae at the expense of canopy-forming species have been observed worldwide in temperate systems and have been linked to increased suspended solids linked to human activities worldwide (Airoldi, 2003). As turfs of Osmundea pinnatifida trap sediments (Prathep et al., 2003), it is clear that this species has some resistance to abrasion and scour from sediment particles. Corallina species accumulate more sediment than any other alga (Hicks, 1985). Hence, an increase in suspended sediment is likely to accumulate in the patches of Corallina officinalis. A significant increase may result in smothering (see siltation pressures). An accumulation of sediment within the turf may attract more sediment-dwelling interstitial invertebrates such as nematodes, harpacticoids, and polychaetes, although in more wave exposed locations, accumulation of sediment is likely to be minimal. Increased suspended sediment may also result in increased scour, which may adversely affect Fucus vesiculosus and foliose red algae, and interfere with settling spores and recruitment if sedimentation coincides with their major reproductive period. However, coralline algae, especially the crustose forms, are thought to be resistant to sediment scour (Littler & Kauker, 1984) and will probably not be adversely affected at the benchmark level.

Red algae and coralline algae, especially, are known to be shade tolerant and are common components of the understorey on seaweed dominated shores. Therefore, limited shading from suspended sediments is not considered to negatively affect this genus. Palmaria palmata is often found under partially shaded conditions as an epiphyte on the stems of Laminaria spp. (Morgan et al. 1980) in the sublittoral zone (Lüning 1990). In the Bay of Fundy, where the tidal flux of nutrients from the marshes includes a high level of suspended sediment, Palmaria palmata grows well despite high turbidity. Irvine (1983) observed morphological adaptation of the plant in fairly sheltered, silty conditions; sometimes the blade divisions are wedge-shaped and finely dissected above, or the blade has numerous linear divisions throughout. This form likely reduces possible smothering from increased siltation due to increased suspended sediment. In the absence of nutrients, a short-term increase in turbidity may affect growth and reproduction; however, as a perennial, the adults will probably survive. Other red algal species tolerate high levels of suspended solids. Chondrus crispus occurs in areas of sand covered rock in the subtidal biotope IR.HIR.KSed.ProtAhn suggesting it is very resistant to high levels of turbidity and scour associated with high levels of resuspended particles.

Hyslop et al. (1997) found that Palmaria palmata and Ulva spp. were reduced or absent on sites affected by high levels of resuspended colliery waste particles; however, the tougher fucoids were less affected. It is not clear how the levels of suspended solids experienced by these sites relate to the pressure benchmark.

No direct evidence was found to assess this pressure for piddocks and other invertebrate species. The piddocks are protected from scour within burrows, and increased organic particles could provide a food subsidy. Pholas dactylus occurs in habitats such as soft chalks, where turbidity may be high and is therefore unlikely to be affected by an increase in suspended sediments at the pressure benchmark. Piddocks, in common with other suspension feeding bivalves, have efficient mechanisms to remove inorganic particles via pseudofaeces. Experimental work on Pholas dactylus showed that large particles can either be rejected immediately in the pseudofaeces or passed very quickly through the gut (Knight, 1984). Similarly, Petricolaria pholadiformis can tolerate high levels of suspended solids through the production of pseudofaeces (Purchon, 1955). Increased suspended sediments may impose sub-lethal energetic costs on piddocks by reducing feeding efficiency and requiring the production of pseudofaeces, with impacts on growth and reproduction.

A significant decrease in suspended organic particles may reduce food input to the biotope resulting in reduced growth and fecundity of piddocks. However, local primary productivity may be enhanced where suspended sediments decrease, increasing food supply. Decreased suspended sediment may increase macroalgal competition, enhancing diversity, but is unlikely to significantly change the character of the biotope.

Sensitivity assessment. While the characteristic fauna are resistant to this pressure, changes in suspended solids, reducing water clarity, could have adverse effects on Fucus serratus photosynthesis. Resistance is therefore assessed as ‘Medium’ at the benchmark level. Once conditions return to ‘normal’, Fucus serratus is likely to rapidly regain photosynthesis and growth rate. Resilience is assessed as ‘High’, and sensitivity as ‘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

Sedimentation can directly affect assemblages inhabiting rocky shores by burial/smothering and scour/abrasion of organisms. Fucus serratus is attached to the substratum permanently by a holdfast and cannot relocate in response to increased sedimentation. Smothering will prevent photosynthesis, resulting in reduced growth and eventually death. Sedimentation of bedrock can impede attachment of Fucus embryos and decrease the survival and growth of juveniles through both scour and burial (Schiel et al., 2006). An increase in the vertical sediment overburden can also reduce growth whilst hindering the regeneration abilities of adults (Umar et al., 1998). 

The state of the tide will mediate the extent of the impact. If smothering occurs at low tide when the algae are lying flat on the substratum, then most of the organisms, as well as the associated community, will be covered by the deposit of fine material at the level of the benchmark. However, if smothering occurs whilst the alga is submerged, standing upright, then the photosynthetic surfaces of adult plants will be left uncovered. The resistance of this biotope group to this pressure may therefore vary with the time of day. Germlings, however, are likely to be smothered and killed in both scenarios and are inherently most susceptible to this pressure. Smothering will cause direct mortalities in the associated community, particularly in sessile organisms unable to relocate. Lower densities of herbivores have also been attributed to increased sedimentation, as silt will reduce their feeding activity and limit their movements (Airoldi & Hawkins, 2007; Schiel et al., 2006). The biotope group occurs in sheltered to moderately exposed conditions. In areas with greater water flow, excess sediments can be readily removed, reducing the time of exposure to this pressure.

The burrowing mechanisms of the piddocks Pholas dactylus and Barnea candida, and other Pholads, mean that the burrows have a narrow entrance excavated by the juvenile. As the individual grows and excavates deeper, the burrow widens, resulting in a conical burrow from which the adult cannot emerge. Petricolaria pholadiformis excavates a cylindrical burrow (Ansell, 1970) and hence may be able to relocate in sandy sediments. Although burrowing mechanisms have been studied, no evidence was found to suggest this species can re-emerge through sediments and re-bury. Therefore, piddocks cannot emerge from layers of deposited silt as other, more mobile bivalves can.

Sometimes the substratum in which piddocks reside is covered by a thin layer of loose sandy material, through which the piddocks maintain contact with the surface via their siphons. It is likely that the piddocks would be able to extend their siphons through loose material, particularly where tidal movements shift the sand around. Pholas dactylus have been found living under layers of sand in Aberystwyth, Wales (Knight, 1984) and in Eastbourne, with their siphons protruding at the surface (Pinn et al., 2008). Barnea candida has also been found to survive being covered by shallow layers of sand in Merseyside (Wallace & Wallace, 1983). Wallace & Wallace (1983) were unsure how long the Barnea candida could survive smothering, but noted that, on the coast of the Wirral, the piddocks have survived smothering after periods of rough weather. Where smothering is constant, survival can be more difficult. The redistribution of loose material following storms off Whitstable Street, in the Thames Estuary, is thought to be responsible for the suffocation of many Petricolaria pholadiformis, and this species may be the most intolerant of the three piddock species associated with this biotope. However, it was not known how deep the layer of ‘loose material’ was, nor how long it lasted, or what type of material it was made up of.

Indirect indications for the impacts of siltation are provided by studies of Witt et al. (2004) on the impacts of harbour dredge disposal. Petricolaria (syn. Petricola) pholadiformis was absent from the disposal area, and Witt et al. (2004) cite reports by Essink (1996, not seen) that smothering of Petricolaria (syn. Petricola) pholadiformis from siltation could lead to mortality within a few hours. Hebda (2011) also identified that sedimentation may be one of the key threats to Barnea truncata populations. At Agigea (Micu, 2007) reported that smothering of clay beds by sand and finer sediments had removed populations of Pholas dactylus. In this area, sand banks up to one metre thick frequently shift position, driven by storm events and currents (Micu, 2007). Similar smothering was described in the case of Barnea candida populations boring into clay beds (Gomoiu & Muller 1962, cited from Micu, 2007). As Hiatella arctica are essentially sedentary with relatively short siphons, siltation from fine sediments rather than sands, even at low levels for short periods, could be lethal.

Sensitivity assessment. As the characteristic bivalves are essentially sedentary and as siphons are relatively short, smothering by fine sediments rather than sands, even at low levels for short periods, could be lethal. Resistance to siltation is assessed as ‘None’, resilience as ‘Medium’ (2 to 10 years) and sensitivity as ‘Medium’. 

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 effects of electromagnetic fields (EMFs) on benthic organisms is still severely lacking. No studies examining the effect of EMFs on macroalgae were found. 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. No studies investigating the effect of EMFs at the population or community level for benthic organisms were found.

Sensitivity assessment. Given the lack of data at the level of individual biotopes, resistance and resilience to EMFs cannot be robustly assessed. Sensitivity is therefore recorded as ‘Insufficient evidence’.

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

Increased levels of diffuse irradiation correlate with increased growth in macroalgae (Aguilera et al., 1999). Macroalgae require light to photosynthesise, so changes in light intensity are likely to affect photosynthesis, growth, competition, and survival. Chapman (1995) noted that too little or too much light is likely to be stressful. There is considerable literature on the light compensation point of marine algae (see Lüning, 1990), but it is difficult to correlate such evidence with ‘shading’, as light saturation and compensation points depend on light availability, light quality, season, and turbidity. As fucoids are outcompeted in sublittoral conditions, permanent shading would likely affect their growth and allow them to be outcompeted by other, more shade tolerant species, within the affected area.

Fucoid macroalgae are strongly regulated by light intensity, light quality, and photoperiod. Light influences photosynthesis, growth, reproductive development, and the timing of gamete release. Consequently, both shading and artificial light at night (ALAN) have the potential to alter physiological performance and reproductive phenology.

Experimental manipulation of irradiance indicates that intertidal fucoids are generally tolerant of variation in daytime light levels within natural ranges. Schmid et al. (2021) reported that increased irradiance (from 30 to 90 μmol photons/m²/s) enhanced growth rates in both Ascophyllum nodosum and Fucus serratus, although pigment and polyunsaturated fatty acid concentrations declined with increasing light. Olabarria et al. (2018) found relatively few significant effects of altered light intensity or ultraviolet radiation on physiological traits in Fucus serratus, suggesting an ability to acclimate to variation in light regime. Similarly, Ferreira et al. (2015) found that experimental shading did not enhance recruitment success of Fucus spp. in either northern or southern European populations, indicating that reduced isolation alone does not necessarily limit establishment.

Field observations also demonstrated substantial tolerance to high daytime irradiance. Migné et al. (2021) showed that Fucus serratus maintained high aerial photosynthetic rates during emersion, even under high light and temperature. Although photoinhibition occurred in thalli at the top of the canopy under harsh conditions, self-shading within the canopy maintained effective photosynthesis in lower layers. This structural buffering suggests that intact stands possess intrinsic resistance to short-term increases in light intensity.

In contrast to changes in daytime irradiance, artificial light at night (ALAN) primarily alters the natural dark phase and therefore perceived photoperiod. Reproductive development in fucoids is strongly influenced by environmental cues, particularly light regime. Light availability supports receptacle growth through photosynthetic activity and is also involved in regulating the onset of gametogenesis and the timing of gamete release (Pearson & Brawley, 1996; Serrão et al., 1996). Photoperiod is widely regarded as a reliable seasonal signal for macroalgal gametogenesis (Brawley & Johnson, 1991), and experimental exposure to continuous light has been shown to interfere with normal patterns of gamete release in fucoids (Pearson et al., 1998). In natural populations, spawning rhythms may also reflect interactions between circadian timing mechanisms and tidal immersion cycles, in some cases producing semilunar reproductive peaks (Andersson et al., 1994; Pearson & Brawley, 1996; Ladah et al., 2003; Monteiro et al., 2012). Because ALAN extends or modifies night-time illumination, it has the potential to disrupt these regulatory processes. Factors that alter perceived day length, including turbidity and canopy shading, have previously been associated with shifts in reproductive timing (Breeman et al., 1984). Both short-day and long-day conditions can act as cues for gametogenesis depending on species and context (Dring, 1984; Kain & Norton, 1990), indicating that responses to altered photoperiod are likely to be complex and species-specific.

Field evidence has demonstrated that elevated night-time illumination can modify fucoid reproductive phenology. Moyse et al. (2025) quantified receptacle ripeness in fucoids, including Fucus serratus, along an ALAN gradient in Plymouth Sound, UK. At the most strongly illuminated site (16.15 mag/arcsec², 3.75 × 10⁻² cd/m²), Fucus serratus showed elevated winter ripening relative to lower illumination sites. Differences among moderate and low illumination sites were not significant for Fucus serratus, suggesting a possible threshold response. Disruption to photoperiod cues has the potential to affect reproductive success even where adult thalli appear physiologically tolerant, because fucoid recruitment depends on synchronised gamete release and suitable settlement conditions. Early developmental stages are also light-sensitive, with embryonic polarity (cellular organization within the developing alga) known to respond to external light gradients (Siméon & Hervé, 2017), suggesting additional pathways through which altered light regimes could influence recruitment dynamics. Where artificial illumination is sustained and of sufficient intensity, effects on population dynamics cannot be discounted.

Red algae, in general, are shade-tolerant, often occurring under a macroalgal canopy that reduces light penetration. In areas of higher light levels, the fronds may be lighter in colour due to bleaching (Colhart & Johansen, 1973). Other red algae in the biotope are flexible with regard to light levels. Canopy removal experiments in a rocky subtidal habitat in Nova Scotia, Canada, by Schmidt & Scheibling (2007) did not find a shift in understorey macroalgal turfs (dominated by Corallina officinalis, Chondrus crispus and Mastocarpus stellatus) to more light-adapted species over 18 months.

Pholas dactylus can perceive and react to light (Hecht, 1928). However, there is no evidence that this pressure would impact the piddocks within the biotope.

Sensitivity assessment. Experimental and field evidence indicate that Fucus serratus and associated fucoids show substantial tolerance to variation in daytime irradiance within natural ranges. Increased irradiance enhances growth under experimental conditions, while shading does not necessarily limit recruitment. Intact canopies also provide structural buffering against high light exposure during emersion. There is no evidence that realistic changes in daytime light intensity alone would result in mortality or loss of the biotope. Artificial light at night alters photoperiod rather than daytime irradiance. Field evidence indicates that high levels of night-time illumination can modify the timing of receptacle ripening in Fucus serratus. However, these effects are sublethal and do not demonstrate direct mortality, canopy loss, or immediate changes in biotope structure. Effects appear intensity-dependent, with detectable responses only at the highest recorded illumination levels and limited evidence of change at moderate levels. Most plausible scenarios are unlikely to result in significant mortality of adult thalli or loss of canopy cover. Therefore, resistance is assessed as ’High’. Where any localised reduction in reproductive output occurred, fucoids are capable of relatively rapid recruitment where propagule supply remains available. Resilience is therefore assessed as ’High’. Sensitivity is consequently assessed as ’Not sensitive’. Confidence in this assessment is ‘Medium’, reflecting strong evidence for physiological tolerance to irradiance change and emerging but limited field evidence for sublethal effects of artificial night-time illumination.

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.

Death or injury by collision

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

Evidence

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

Visual disturbance

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

Evidence

This pressure is not relevant to the algae within the biotope. Most animals will be able to sense changes in light, but are unlikely to be affected by visual disturbance. Pholas dactylus reacts quickly to changes in light intensity, after a couple of seconds, by withdrawing its siphon (Knight, 1984). This reaction is ultimately an adaptation to reduce the risk of predation by, for example, approaching birds (Knight, 1984). However, its visual acuity is probably very limited, and it is unlikely to be sensitive to visual disturbance. Birds are highly intolerant of visual presence and are likely to be scared away by increased human activity, therefore reducing the predation pressure on piddocks. Therefore, visual disturbance may be of indirect benefit to piddock populations.

Sensitivity assessment. The biotope is considered to be ‘Not sensitive’. Resistance and resilience are therefore assessed as ‘High’ by default.

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

The species characterizing this biotope are not farmed or translocated and therefore this pressure is ‘Not relevant’ to this biotope.

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

Very little is known about infections in Fucus (Wahl et al., 2012). Coles (1958) identified parasitic nematodes that caused galls on Fucus serratus in southwest Britain. More recently, Zuccaro et al. (2008) detected a number of fungal species associated with Fucus serratus. So far, no mortalities have been associated with the introduction of microbial pathogens. However, the potential for increased biotic interactions involving parasites or pathogens is on the rise in many marine systems (Torchin et al., 2002).

Sensitivity assessment. Based on the lack of evidence for major pathogens or significant mortalities of the key characterizing species, this biotope is considered to have ‘High’ resistance and hence ‘High’ resilience and is classed as ‘Not sensitive’ at the pressure benchmark. 

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

Fucus serratus is one of several harvested and exploited algal species. Seaweeds were collected from the middle of the 16th century for the iodine industry. Nowadays, seaweeds are harvested for their alginates, which are used in the cosmetic and pharmaceutical industries, for agricultural supply, water treatment, and for human food and health supplements (Bixler & Porse, 2010).

The commercial harvest removes seaweed canopies, which will have important direct and indirect effects on the wider ecosystem. Stagnol et al. (2013) investigated the effects of commercial harvesting of intertidal Fucus serratus on ecosystem biodiversity and functioning. The study found that the removal of macroalgae affected the metabolic flux of the area. Flows from primary production and community respiration were lower in the impacted area as the removal of the canopy caused changes in temperature and humidity conditions. Suspension feeders were the most affected by the canopy removal, as canopy-forming algae are crucial habitats for these species, and most are sessile organisms.

Migné et al. (2015) experimentally removed the canopy of Fucus serratus (cut with a knife and regrowth repeatedly removed) over 18 months on a mid-low rocky shore in the Southwest English Channel. Gross primary productivity (GPP) and community respiration were consistently and markedly reduced in canopy removal plots compared to controls throughout the study. Net primary productivity (NPP) was initially maintained due to the proliferation of opportunistic green algae, but declined substantially after nine months. While overall species richness and trophic structure were not significantly altered, the abundance and biomass of mobile invertebrates were greatly reduced in the absence of the canopy, indicating strong effects on higher trophic levels. The authors concluded that at this shore level, the canopy primarily supported ecosystem functioning by providing food resources and habitat rather than by ameliorating physical stress.

Other studies confirm that loss of canopy had both short and long-term consequences for benthic communities in terms of diversity, resulting in shifts in community composition and a loss of ecosystem functioning such as primary productivity (Lilley & Schiel, 2006; Gollety et al., 2008). Removal of the canopy caused bleaching and death of the understorey red turfing algae. Stagnol et al. (2013) observed Patella vulgata recruiting in bare patches of disturbed plots. Experimental studies have shown that limpets control the development of macroalgae by consuming microscopic phases (Jenkins et al., 2005) or the adult stages (Davies et al., 2007). The increase in Patella vulgata abundance could therefore limit the recruitment and growth of Fucus serratus in the impact zone. Due to the high intolerance of macroalgae communities to human exploitation, the European Union put in place a framework to regulate the exploitation of algae, establishing an organic label that implies that ‘harvest shall not cause any impact on ecosystems’ (no. 710/2009 and 834/2007).

Meichssner et al. (2021) investigated cultivation scenarios for Fucus serratus in the western Baltic Sea. Modelled annual yields for Fucus serratus ranged from 4.23 to 6.99 kg/m², with optimal yields (6.88 to 6.99 kg/m²) achieved at relatively high initial stocking densities (2.5 to 4 kg/m²) and harvest densities of 5 kg/m², requiring between two and six harvests per year. Growth showed strong seasonality, with reduced rates in winter and evidence of summer growth limitation. Nitrogen availability was identified as the likely limiting nutrient under field conditions. These results demonstrated the capacity for repeated biomass removal under controlled cultivation, although the study did not examine associated community responses or recovery dynamics in natural intertidal assemblages.

Piddocks may be removed as bait, and across Europe, they have traditionally been harvested for food. However, high levels of habitat damage are associated with the removal of boring molluscs (Fanelli et al., 1994), and this practice has largely been banned. The most sensitive component of this biotope to targeted harvesting is the chalk substratum, which may be damaged and removed if piddocks are excavated from their burrows.  This effect is considered through the physical damage pressures, abrasion, penetration and sub-surface damage.

The red crustose coralline alga Phymatolithon lenormandii is found in rock pools and shaded areas. Removal of the Fucus serratus and red algal turf canopy could result in desiccation, leading to bleaching and mortality, depending on the temperatures experienced and the degree of insolation.

Sensitivity assessment. Resistance to harvesting of any of the characterising species is deemed to be ‘Low’. Resilience is assessed as ‘Medium’. Therefore, the biotope’s sensitivity is assessed as ‘Medium’.

Removal of non-target species

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

Evidence

Direct physical impacts from harvesting are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of non-target species on this biotope. The loss of characterizing and associated species due to incidental removal as bycatch would alter the character of the habitat from the biotope description. The ecological services, such as primary production and habitat structure, would also be lost. The red crustose coralline alga Phymatolithon lenormandii is found in rock pools and shaded areas. Removal of the Fucus serratus and red algal turf canopy could result in desiccation, leading to bleaching and mortality, depending on the temperatures experienced and the degree of insolation.

Sensitivity assessment. Removal of a large percentage of the characterizing species, resulting in bare rock, would alter the character of the biotope, species richness, and ecosystem function. Resistance is therefore assessed as ‘Low’ and resilience as ‘Medium’ (based on the loss of fronds and holdfasts and piddocks, but see resilience section for caveats), so sensitivity is assessed as ‘Medium’. If a high proportion of holdfasts remained, recovery would be assessed as ‘High’, and sensitivity would be assessed as ‘Low’.

The American slipper limpet, Crepidula fornicata

Evidence

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and on the Essex coast in 1887-1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 1999, 2018; Hinz et al., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). It ranges from the Baltic Sea, the Kattegat and Skagerrak, the North Sea coasts of the UK, Germany, and Belgium, through the English Channels and into the Irish sea coasts of Ireland and south Wales with records in east and west Scotland, Northern Ireland, northwest France, Spain and south into the Mediterranean (NBN, 2024; OBIS, 2025).

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

Its density and ability to spread within and between sites (e.g., Bays) depend on the availability of suitable habitat, competition with other species, larval retention within the site, human activity (e.g., dredging) and summer and winter temperatures (especially in the intertidal). For example, the Crepidula fornicata population in the Bay of Mont-Saint-Michel grew by 50% between 1996 and 2004 and covered 25% at a high density (51 to 100% cover), aided by local oyster farming and shellfish dredging (Blanchard, 2009). However, in Arcachon Bay, France, Crepidula fornicata was limited to only 155 tonnes in 1999 and 312 tonnes in 2011 (De Montaudouin et al., 2001, 2018). Crepidula was limited to muddy sediments that were only approx. 8% of the bay and were colonized by Zostera beds and represented only 0.4% of suspension feeder biomass of the oyster Magallana gigas in the bay, and did not show signs of increasing biomass at a 12-year scale. In addition, benthic trawling was prohibited in the bay (De Montaudouin et al., 2001, 2018). As a result, De Montaudouin et al. (2018) concluded that Crepidula was not invasive in the Bay of Arcachon.

Crepidula fornicata is recorded from shallow, sheltered bays, lagoons, estuaries or the sheltered sides of islands, in variable salinity (from 18 to 40), although it prefers approx. 30 PSU (Tillin et al., 2020). It is recorded from the lower intertidal to approx. 160 m in depth, but it is most common in the shallow subtidal and low water springs (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 2025; Tillin et al., 2020).

Larvae require hard substrata for settlement. It prefers muddy, gravelly, shell-rich substrata that include gravel, shells of other Crepidula, or other species, e.g., oysters and mussels. It is highly gregarious and seeks out adult shells for settlement, forming characteristic ‘stacks’ of adults, but it is also recorded from rock, artificial substrata, and Sabellaria alveolata reefs (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015; Tillin et al., 2020). For example, 75% to 98% of Crepidula larvae settled on dead Crepidula shells in the eastern Solent harbours of Portsmouth, Langstone, and Chichester, while approx. 4% settled on stone, 2.5% on live Crepidula, 0.3% oyster shell, 0.6% cockle shell, 0.3% winkle shell and 0.1% periwinkle shell (Preston et al., 2020). However, in the Milford Harbour Waterway, the highest densities of Crepidula were found in areas of sediment with hard substrata (e.g., mixed fine sediment with shell, gravel, or both). While Crepidula density increased with increasing gravel cover in the subtidal zone, the opposite pattern was observed in the intertidal zone (Bohn et al., 2015). Gravel formed the base of most stacks of Crepidula in the intertidal, which suggested that initial colonization occurred on available hard substrata (i.e., gravel) in the absence of adult shells of Crepidula. The availability of hard substrata (e.g., gravel) may only restrict initial colonization as higher densities of Crepidula function as substrata for subsequent colonization (Thieltges et al., 2004; Blanchard, 2009). 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.

Bohn et al. (2015) also noted that Crepidula density was low in areas of homogenous fine sediment and absent in areas dominated by boulders. Bohn et al. (2015) suggested that wave action (exposure) probably prevented the establishment of large numbers of Crepidula in high-energy areas. However, Hinz et al. (2011) recorded Crepidula off the Isle of Wight in the English Channel, at approx. 60 m on rough ground in areas of high tidal flow. Tillin et al. (2020) suggested that the effect of oscillatory wave mediated flow might have a greater effect on Crepidula than tidal flow, presumably due to mobilisation of the substratum. Similarly, Crepidula was absent from sandy substrata in Swansea Bay but was most abundant in the shelter of the breakwater at Swansea East site (Powell-Jennings & Calloway, 2018).

The availability of hard substrata (e.g., gravel) may only restrict initial colonization as higher densities of Crepidula function as substrata for subsequent colonization (Thieltges et al., 2004; Blanchard, 2009). Bohn et al. (2015) suggested that wave action (exposure) probably prevented the establishment of large numbers of Crepidula in high-energy areas. Blanchard (2009) noted that sandy areas in the Bay of Saint-Mont Michel were not colonized by Crepidula because of surface sand mobility. Thieltges et al. (2003) also noted that storm events removed some clumps of mussels and presumably Crepidula onto tidal flats where they disappeared, which caused their abundance to fluctuate. Powell-Jennings & Calloway (2018) noted that Crepidula is killed by sudden burial and, possibly, burial due to deposition, which could mitigate Crepidula density. 

The density of Crepidula populations in northern Europe (Germany, Denmark, and Norway) is significantly lower (<100 /m²) than in southern waters. Thieltges et al. (2004) reported that the population of Crepidula was affected strongly by cold winters in the Wadden Sea. The winters of 2001 and 2003 resulted in approx. 56 to 64% mortality of intertidal Crepidula and up to 97% on one mussel bed, compared to only 11 to 14% in southern areas without frost. Crepidula almost vanished from the Wadden Sea after the 1978/79 winter and took ten years to recover due to moderate winters, which regularly affected the population. Similarly, 25% mortality was observed in Crepidula populations on the south coast of the UK after the extreme 1962/63 winter (Crisp, 1964; Bohn et al., 2012). Thieltges et al. (2003) suggested that global warming may allow Crepidula populations to become more abundant in northern Europe. Valdizan et al. (2011) noted that higher water temperatures between 2000 and 2001 and 2006 to 2007, together with elevated chlorophyll-a corresponded to an increase in gametogenesis and the duration of broods in the Crepidula population in Bournerf Bay, France. They suggested that rising temperatures in northern Europe could increase its reproductive success due to favourable breeding temperatures and increased phytoplankton (Valdizan et al., 2011). Nehls et al. (2006) noted that the decline in mussel (Mytilus edulis) beds in the Wadden Sea was due to mild winters that favoured non-native oysters (Magallana gigas) and slipper limpets, which co-existed with the mussels.

Crepidula fornicata has one or two reproductive periods per year (depending on location), is highly fecund, and has long-lived pelagic larvae. Hence, dispersal is potentially high. However, Bohn et al. (2012, 2013a, 2013b, 2015) suggested that the lack of suitable habitat, rather than larval supply, together with local hydrography, may limit the northward spread of Crepidula from Milford Harbour Waterway, and that post-settlement mortality is particularly important in the intertidal. Dupont et al. (2007) reported genetic isolation with distance along the English Channel but a high degree of genetic connectivity between the bays of northern France, which were consistent with hydrographic models of larval transport. They noted marked genetic isolation of the population in the semi-enclosed Bay of Brest. Dupont et al. (2007) suggested that Crepidula populations were isolated by hydrographic barriers over distances of approx. 100 km. Bohn et al. (2012) suggested that homogenous sediments and boulders at the entrance to the Milford Harbour Waterway formed a barrier to dispersal and, together with high larval export, probably explained the slow northward expansion of Crepidula along the Welsh coast. Nevertheless, the initial spread of Crepidula was facilitated by human activities such as shipping, shellfish culture (e.g. oysters and mussels), ballast water (Blanchard, 1997) and fisheries (e.g., dredging) (Blanchard, 1997, 2009; De Montaudouin et al., 2018; Kostecki et al., 2011; McNeill et al., 2010; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015).

Sensitivity Assessment. No evidence of Crepidula fornicata presence in clay or chalk habitats was found. According to Tillin et al. (2020), clay exposures are unsuitable for Crepidula fornicata settlement, although this is stated with low confidence. Therefore, resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’.

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 worldwide (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).

A lack of published descriptions and an incomplete historical record has led to the widespread misidentification of Didemnum vexillum, and it is often recorded as Didemnum spp. Hence, the native range of the species is not known conclusively (Lambert, 2009; Stefaniak et al., 2012; McKenzie et al., 2017; Holt, 2024). However, molecular data and limited historical evidence have suggested that the species may be native to Japan, with its native range possibly extending into continental Asia and north-western Pacific (Stefaniak et al., 2012; Tillin et al., 2020; Holt, 2024). Previously unrecorded populations of a colonial ascidian have been recently identified as Didemnum vexillum (Tillin et al., 2020).

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 limited natural dispersal because pelagic larvae remain in the water column for only a short time (up to 36 hours). Therefore, it has a short dispersal phase that can allow the species to build localised 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 from 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; Griffiths 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 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 winter 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 winter (Valentine et al., 2007a). The seasonal growth cycle is also likely influenced by location. For example, the Didemnum sp. growth cycle for colonies in the 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 and 20°C and slow or cease below 9 to 11°C as summer ends (Griffith et al., 2009; McKenzie et al., 2017). In New Zealand, recruitment occurs from November to July, where the highest average temperatures were recorded in February (18 to 22°C) and the lowest in July (9 to 10°C) (Fletcher et al., 2013a). In this New Zealand study, higher water temperatures were associated with higher 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). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems by its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock, gravel, pebble, cobble, or boulders (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).

In addition, Didemnum vexillum is commonly found associated with artificial hard substrata, being mostly found in harbours and marinas where it covers a variety of maritime structures such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ships hulls and at aquaculture facilities (Valentine et al., 2007a & b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012). Didemnum vexillum was abundant in the marinas at Terschelling, Texel, and Vlieland, in the Wadden Sea (Gittenberger et al., 2015). In the UK, Didemnum vexillum was initially recorded in marinas and adjacent shallow man-made structures (Tillin et al., 2020). In Wales, it was first recorded in Holyhead Marina, then subsequently reported in Plymouth Marina and other marinas around the UK (Griffith et al., 2009; Minchin & Nunn, 2013; Bishop et al., 2015).

Didemnum sp. can colonize both horizontal and vertical surfaces of fouling and benthic communities, commonly occurring on upper horizontal surfaces in benthic habitats (Dijkstra et al., 2007; Tillin et al., 2020). It has been recorded on overhangs or the underside of boulders (Hitchin, 2012) or on the underside of docks, boat hulls, pontoons (Griffiths et al., 2009; Minchin & Nunn, 2013). In sheltered areas, colonies are lobed and beard-like, forming long tendrils that drop down from the underside of docks or other artificial substrata to establish new colonies if there are suitable substrata available (Valentine et al., 2007a). In areas of stronger current, colonies are low undulating mats (Valentine et al., 2007 a&b).

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 & Grosholz, 2015; Vercaemer et al., 2015).

Some species have been shown to tolerate overgrowth by Didemnum vexillum. Such as anemones (did not specify species name), which were observed in high densities of 10 to 339 individuals in transects with high percentage cover of Didemnum vexillum (Lengyel et al., 2009). In the Netherlands, the sea anemone Sagartia elegans and Sabella pavonia tubes were not overgrown by Didemnum sp. (Gittenberger, 2007). Botrylloides violaceus can overgrow Didemnum sp. (Gittenberger, 2007), although it was noted to be overgrown in other studies (Valentine et al., 2007a). In addition, Styela clava and Ascidiella aspera survived overgrowth by Didemnum vexillum as long as their siphons remained free (Gittenberger, 2007). However, Gittenberger (2007) stated that the boring sponge Clione celata, the sea anemone Diadumene cincta, Mytilus edulis, Magallana (syn. Crassostrea) gigas, Ostrea edulis, a variety of hydroids, the colonial ascidians Aplidium (Fig. 4) and Diplosoma listerianum and the solitary ascidians Ciona intestinalis start to die on contact with Didemnum sp.

There are few observations of Didemnum vexillum on soft bottom habitats as evidence suggests it is unable to establish or grow easily on mud, mobile sand or other unstable substrata, and it is vulnerable to smothering by fine sediment (Bullard et al., 2007; Valentine et al., 2007a; Griffith et al., 2009). The species is usually found in areas where the colony is protected from sedimentation and wave action (Valentine et al., 2007b; McKenzie et al., 2017; Tillin et al., 2020).

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 been recorded from less than 1 m to at least 81 m deep (Bullard et al., 2007; Tagliapietra et al., 2012; Tillin et al., 2020). It is abundant across various shore heights, thriving in both nearshore and offshore sites, particularly in subtidal areas. For example, colonies of Didemnum vexillum were dominant at depths between 45 to 60 m, occupying 50 to 90% of available space in two gravelly areas (more than 230 km²) composed of immobile pebble and cobble pavement on Georges Bank fishing ground, USA (Bullard et al., 2007; Valentine et al., 2007b; Lengyel et al., 2009). In addition, patchy mats have been observed covering approximately 1 to 1.5 km² of the pebble cobble seabed, which is interspersed with large boulders and 30 m deep in Long Island Sound, USA (Mercer et al., 2009). In an offshore scallop dredge survey, Didemnum sp. was found attached to cobbles and boulders at 10 to 34 m (Vercaemer et al., 2015).

In northern Kent, Didemnum vexillum has been recorded covering London clay boulders on Whitstable Flats, West Beach; tabulate sandstone boulders (0.5 to 2 m across) on the mid shore; and sediment mounds on the low shore, characterized by larger areas of sand, mud, and low-lying sediment at Reculver and Bishopstone (Hitchin, 2012). It was also recorded in muddy substrata at that site. Hitchin (2012) noted that the site was exposed to enough waves and currents to cause sedimentation. However, Didemnum vexillum grew hanging from the underside of sandstone boulders nestled on sediment, on consolidated sediment mounds and firm clays, hence burial may prevent colonization and its survival rather than sedimentation alone.

The Sandwich tide pools were subject to air exposure at low tide, daily changes in water depth and temperature (Valentine et al., 2007a). Didemnum vexillum colonies can survive exposure to air at low tides during rapid colony growth in the summer months, July to September (Valentine et al., 2007a). However, parts of the large established colonies, which were artificially exposed to air for two to three hours in October, were observed desiccated or predated on by grazing periwinkles 30 days later, in the winter month of November (Valentine et al., 2007a). They suggested that the invasive tunicates’ ability to tolerate exposure to air varies with the seasonal growth cycle. Didemnum vexillum also tolerated emersion in Kent, as colonies on the mid-shore at Reculver flourish and survive in air exposure for up to three hours per cycle during spring tides (Hitchin, 2012). Hitchin (2012) suggested that the porous nature of the sandstone boulders, which the species colonized retained water. The Kent shore was sheltered but held water due to its shallow slope and flats, which may allow Didemnum sp. to survive in the low to mid-shore. There is evidence that Didemnum vexillum died when exposed to air for more than 6 hours (Laing et al., 2010).

Didemnum vexillum tolerates a wide range of environmental conditions, including temperature and salinity (Herborg et al., 2009; Tillin et al., 2020). Didemnum vexillum can withstand a wide range of salinities from 20 to 44 PSU, is commonly found in marine waters around 33 PSU, but is unable to survive in salinities below 20 PSU (Bullard & Whitlatch, 2009; Groner et al., 2011; Tillin et al., 2020). It has been recorded in estuarine conditions and tidal lagoons (Dijkstra et al., 2007; Tillin et al., 2020). In the Lagoon of Venice, Didemnum vexillum is found in waters at 30 PSU. It was absent in low salinity, such as the estuary and around the salt marshes, but well established in the euhaline and tidally well-flushed zones of the Lagoon of Venice (Tagliapietra et al., 2012). Similar results were found in Connecticut and Rhode Island, where Didemnum vexillum was not found in environments with salinity less than 20 PSU (Bullard & Whitlatch, 2009). However, in the Wadden Sea, colonies of Didemnum vexillum were abundant in salinities between 17.91 and 25.97 PSU (Gittenberger, 2007; Gittenberger et al., 2015).

Salinity can influence the growth rates of Didemnum vexillum. For example, in an experiment in the Thames River estuary, Connecticut, Bullard & Whitlatch (2009) found growth rates were significantly higher in high salinity areas (26 to 30 PSU) and although survival at different salinities was not significantly different, the Didemnum vexillum colonies in low (10 to 26 PSU) and medium (15 to 28 PSU) salinities were bloated, discoloured and appeared to be dying. In unpublished data from Bullard & Whitlatch (2009), similar results were found in the laboratory, as most colonies appeared to be dying after one week in 20 PSU and healthy in 30 PSU.

A study on Didemnum vexillum colonies from Holyhead Marina, Isle of Anglesey, found colony growth within a week was significantly impaired and reduced by two-thirds at lower salinities (27 PSU and 20 PSU), while in ambient Holyhead Marina salinity (34 PSU), the growth increased and surface area doubled (Groner et al., 2011). Mortality was described as negligible in colonies of Didemnum vexillum in ambient salinity (34 PSU) after two weeks. However, mortality increased as salinity decreased. At the end of the two-week experiment, 72% of invasive colonies survived in 27 PSU, and 55% of colonies survived in 20 PSU (Groner et al., 2011). When exposed to severe low salinity of 10 PSU for two hours, Didemnum vexillum showed no mortality, which suggested the duration of exposure influences mortality, not the stress intensity (Groner et al., 2011). Colonies of Didemnum vexillum collected from Anglesey, Wales, experienced more mortality under severe hypo-salinity (20 PSU, 38% colonies survived) compared to moderate hypo-salinity (27 PSU, 82% colonies survived) after two weeks, showing severe hypo-salinity creates more stressful conditions for Didemnum vexillum (Lenz et al., 2011). Therefore, Didemnum vexillum can tolerate a short-term severe decline in salinity, but prolonged exposure over two weeks caused chronic stress and increases in mortality.

Didemnum vexillum is a temperate species that can survive a broad temperature range of -2 to 24°C, with an upper survival limit suggested to be 25°C (Bullard et al., 2007; Valentine et al., 2007a; Herborg et al., 2009; Kleeman, 2009; McKenzie et al., 2017; Holt, 2024). It thrives best at 14 to 20°C, with optimal growth temperature between 14 to 18°C during summer months (May, June, September, October) (Gittenberger, 2007; Kleeman, 2009; McKenzie et al., 2017). Didemnum vexillum has been recorded surviving in 4 to 15°C in Georges Bank and 5 to 22°C in Holyhead (Bullard et al., 2007; Valentine et al., 2007b; Griffith et al., 2009).

In New England, colonies tolerate temperatures as low as -2°C (Bullard et al., 2007), but reports from the Netherlands show colonies “die-off” when temperatures drop below 5°C during winter months from November to April (Gittenberger, 2007; Herborg et al., 2009). Cold winter 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). Temperature changes are an important factor influencing the seasonal growth cycle and reproduction of Didemnum vexillum (Valentine et al., 2007a).

Didemnum vexillum requires suitable hard substrata for successful settlement and establishment of invasive populations. It grows quickly and can 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). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems due to its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata, gravel, pebble, cobble, or boulders (Tillin et al., 2020). The extensive mats formed by the invasive species over cobble-pebble substrata can bind or ‘glue’ small pebbles and cobbles together by filling spaces between the sediment particles, which alters the habitat complexity of the seafloor turning it into a more homogenous two-dimensional habitat rather than heterogeneous three-dimensional one (Griffith et al.,2009; Mercer et al., 2009; Lengyel et al., 2009).

Once established, Didemnum vexillum can expand rapidly, taking over most available hard substrata. Studies have hypothesized that this may alter species diversity and community composition and may decrease species abundance and biodiversity. Gittenberger (2007) stated that at this site, Didemnum sp. could cover around 95% of hard substrata, leaving little space for recruitment and growth of other species.

On Georges Bank, USA, Didemnum vexillum has altered the benthic community (Lengyel et al., 2009; Tillin et al., 2020). In Georges Bank, Lengyel et al. (2009) analysed photographs of the seabed and suggested that Didemnum vexillum outcompeted other epifaunal and macrofaunal species. Changes were seen in hydroids, the second most abundant epifaunal species at the location, which were overgrown by the invasive tunicate and negatively correlated with the percentage cover of Didemnum vexillum (Lengyel et al., 2009). The number of non-colonial macrofauna was also negatively related to the percentage cover of Didemnum vexillum (Lengyel et al., 2009). Dredge samples revealed clear differences in benthic species composition and revealed a significant difference in the species abundance before and after the colonization of Didemnum vexillum (Lengyel et al., 2009).

In contrast, some studies have suggested that potentially the overgrowth of Didemnum vexillum has little impact on benthic communities. In Long Island Sound, USA, Mercer et al. (2009) found the total abundance and richness of native epifaunal and infaunal species were either not different or significantly higher in samples taken inside Didemnum vexillum mats compared with samples collected outside the mats. While the mats did lead to subtle changes in community structure and shifts in species dominance, the authors suggested that benthic species may use Didemnum vexillum mats as a novel habitat and species living beneath the mats may use it for shelter and protection from epibenthic predators (Mercer et al., 2009).

Didemnum vexillum mats may alter the flux of materials by creating a barrier from the water column to the sediment column, influencing the biogeochemical cycling of many nutrients. This barrier can prevent light and food from reaching the sessile community underneath it, prevent predators from feeding on the bottom and hinder larvae settlement (Mercer et al., 2009; Dijkstra, 2009, cited in Tillin et al., 2020). This has been seen in Zostera marina (Carman & Grunden, 2010; Long & Grosholz, 2015). The barrier may also influence the dissolved oxygen exchange between sediments and overlaying water, creating hypoxic conditions (Mercer et al., 2009).

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

Sensitivity Assessment. There is no evidence of Didemnum vexillum colonization on chalk. However, it has been recorded on clay boulders (Hitchin, 2012). According to Tillin et al. (2020), clay exposures are potentially suitable substrata for Didemnum vexillum colonization, although this is stated with low confidence. Resistance is therefore ‘Low’, resilience is ‘Very Low’ as Didemnum vexillum would need to be physically removed to allow recovery, and sensitivity is assessed as ‘High’, albeit with low confidence due to a lack of direct evidence.

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, 2011b, 2012a). 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). 

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, 2011b, 2012a; Humphreys et al., 2014, cited in Alves et al., 2021; Hansen et al., 2023). It was also introduced to the northeast Adriatic Sea (Ezgeta-Balic et al., 2019) and southwest England from France, possibly via fouling on ships (GBNNSIP, 2011b, 2012a; Padilla, 2010; Ezgeta-Balic et al., 2019).

Magallana gigas has a high fecundity, a long-lived pelagic larval phase (2 to 4 weeks) and can produce up to 200 million eggs during spawning (Herbert et al., 2012, 2016; Alves et al., 2021; Wood et al., 2021; Hansen et al., 2023). Hence, as a broadcast spawner, it has a high dispersal potential of more than 1000 km (Padilla, 2010; Wood et al., 2021). Although larval mortality can be as large as 99% due to sensitivity to environmental conditions (Alves et al., 2021), adults are long-lived so that populations can survive with infrequent recruitment (Padilla, 2010).

Larval dispersal has facilitated the establishment of populations in various regions, such as the Oosterschelde estuary in the Netherlands and the Scandinavian coastlines, where northward drift on tidal and wind-driven currents has been suggested (Hansen et al., 2023). Offshore structures and aquaculture operations can enhance spread (Wood et al., 2021).

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

Settlement requires hard substrata, including rock, bedrock, chalk, bare boulders, cobbles and pebbles and shells (Kochmann, 2012; Kochmann et al., 2013; McKinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). Magallana gigas also attaches to available hard materials in mixed sediment environments such as shingle and sand within otherwise unsuitable mudflats (Spencer et al., 1994; McKinstry & Jensen, 2013; Tillin et al., 2020).

Populations of Magallana gigas have been found wave exposed rocky shores to wave-sheltered soft sediment environments and it has been described as a habitat generalist (Troost, 2010; Kochmann, 2012; Kochmann et al., 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). Patches of Pacific oyster reefs have been recorded on littoral rock in Kent, southern England and on littoral sediments in southern England, the North Sea, and the English Channel (Herbert et al., 2012, 2016; Morgan et al., 2021).

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

Although shorelines comprised mainly of mud were suggested to be unsuitable for spat settlement (Spencer et al., 1994), the presence of smaller hard substrata, such as shells or pebbles, can enable larvae to settle (Tillin et al., 2020). For example, in the River Teign estuary, Pacific oyster settlement was observed on shell-covered ground mainly attached to mussel shells, and occasionally attached to cockles, stones and common periwinkle (Littorina littorea) shells on a mud flat in the estuarine intertidal zone, otherwise mainly comprised of sand and mud (Spencer et al., 1994). In addition, the Blue Lagoon on the north shore of Poole Harbour had the highest abundance of oysters on mud mixed with shingle and shell (McKinstry & Jensen, 2013). Outside of the Blue Lagoon, oysters were also recorded on mixed substrata composed of mud, gravel, and shell (McKinstry & Jensen, 2013). Tillin et al. (2020) concluded that while successful invasions occurred on mudflats, Magallana gigas prefers mixed substrata. Fine mud sediments without hard substrata (such as small stones, gravel, and shell) are unlikely to be suitable (Tillin et al., 2020).

The speed of Magallana gigas reef formation on soft substrata seems to be dependent on the amount of hard substrata present (Troost, 2010). Bergstrom et al. (2021) reported that the presence of Magallana gigas was partially dependent on increasing gravel content up to 15% but remained stable with increasing percentages (measured up to 80%).

While often described as an intertidal and shallow subtidal species, Magallana gigas has been observed across a broader depth range. Although rocky habitats deeper than 10 m are generally considered unsuitable, it has been recorded down to 42 m in the Oosterschelde, Netherlands (Herbert et al., 2012, 2016; Tillin et al., 2020; Smaal et al., 2009).

It frequently occurs between Mean High Water and Mean Low Water in intertidal zones but has also been recorded at 1 to 10 m depth in regions like Sweden, Ireland, and the UK (Kochmann et al., 2013; Herbert et al., 2016; Bergstrom et al., 2021). In Lough Swilly and Lough Foyle, Ireland, oysters were found on shallow subtidal mussel beds and mixed mud and sand habitats (Kochmann, 2012). In the Thames Estuary and parts of Essex and Kent, oysters have also been found subtidally, 2 to 3 m below chart datum (Tillin et al., 2020).

Bergstrom et al. (2021) suggested the optimal depth in the Skagerrak is around 0.5 m, although presence is documented down to 5 m. In Lim Bay (Adriatic Sea), Magallana gigas occurs in the intertidal and shallow subtidal (down to 1 m), but not beyond 3 m depth (Stagličić et al., 2020). The species has not been recorded below extreme low water on rocky habitats, although it has been found subtidally on soft sediments in some areas (Herbert et al., 2012).

The Pacific oyster prefers wide intertidal areas with shallow gradients; it is generally absent from steep shores (McKinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). In Ireland and the Solway Firth, it is more commonly found on intertidal shores over 40 to 50 m wide (Kochmann et al., 2013; Cook et al., 2014).

It has been suggested that recruitment is enhanced, and abundances are higher in wave-sheltered conditions (Robinson et al., 2005; Ruesink, 2007, cited in Teschke et al., 2020; Tillin et al., 2020). Teschke et al. (2020) found the abundance of Magallana gigas was significantly higher at wave-protected sites within the artificial harbours of Helgoland, North Sea, compared to wave exposed sites outside the harbours. The authors suggested that the successful colonization in wave-protected sites could be due to the relative retention of water masses in the harbours that reduces larval drift and the whiplash effect on newly settled larvae. In addition, better growth and higher survival rates were observed at wave-protected sites, whereas mortality rates increased at wave exposed sites, due to the wave exposure causing dislodgement or detachment from the settlement substratum (Teschke et al., 2020; Tillin et al., 2020). Similarly, Bergstrom et al. (2021) noted that the occurrence of high densities of both Ostrea edulis and Magallana gigas decreased with increasing wave exposure.

Magallana gigas can withstand a wide range of salinities (from 11 to 34 PSU), but no oysters were observed in areas on the west Swedish coast which had salinities less than 20 PSU (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, native and Pacific oyster densities increased with rising salinity above 15 to 27 PSU. Larvae can survive salinities between 19 and 35 PSU (Troost, 2010; Tillin et al., 2020). Growth of Pacific oysters can occur between 10 to 30 PSU (Troost, 2010).

Carrasco & Baron (2010) suggested that Magallana gigas has successfully adapted to colonize a range of thermal niches. Temperature is important for the life cycle of the Pacific oyster and influences the establishment of feral and wild populations (Alves et al., 2021). Within its native range, Magallana gigas occurs in areas where the sea surface temperatures range from 14.0°C to 28.6°C in the warmest month of the year, and between -1.9°C and 19.8°C in the coldest month (Carrasco & Baron, 2010).

Magallana gigas has a seasonal reproductive cycle (Alves et al., 2021). Spawning occurs in the summer months, when temperatures are 16 to 34°C and larvae require a water temperature of 18°C or above for successful development (Mann 1979; Troost, 2010; Kochmann, 2012; Ezgeta-Balic et al., 2020; Alves & Tidbury, 2022). In Poole, UK, spawning temperatures were estimated at 19.7°C (Alves & Tidbury, 2022). Ezgeta-Balic et al.’s (2020) study indicated that temperatures in the Mediterranean and the Adriatic were favourable for Pacific oyster larval development, with gametogenesis initiated at temperatures from around 10 to 15°C and spawning initiated at around 24°C. However, the lower thermal limit for spawning was recognised as 16°C (Carrasco & Baron, 2010) and once settled, larvae are unable to survive in temperatures below 3°C (Alves & Tidbury, 2022).

Adults can survive in water temperatures up to 40°C and at low tide, freezing air temperatures as low as -17°C, depending on the salinity of the water in their shells (Troost, 2010; Tillin et al., 2020; Hansen et al., 2023). Growth of Pacific oysters occurs between 3 to 40°C (Troost, 2010; Kochmann, 2012).

Dense macroalgal cover is unsuitable for the Magallana gigas (Herbert et al., 2012, 2016; Tillin et al., 2020), being rarely found under macroalgal cover in Northern Ireland, absent from exposed bedrock or large boulders with macroalgae cover in the Solway Firth, Scotland, and absent in Poole Harbour, where there was competition with macroalgae (Kochmann, 2012; Kochmann et al., 2013; McKinstry & Jensen, 2013; Cook et al., 2014; Tillin et al., 2020). Fucus cover significantly reduced larval recruitment of the Pacific oyster in the Wadden Sea (Diederich, 2005). Hence, the Pacific oyster is more likely to colonize bare rock, boulders, or mussel beds without macroalgae (Diederich, 2005; Cook et al., 2014). Kochmann et al. (2013) suggested that macrophyte canopies prevent larvae from settling on the rock underneath, and macroalgae fronds inhibit settlement and recruitment by exuding metabolites.

Magallana gigas is a trophic competitor of other bivalves and other filter feeders (Decottignies et al., 2007, cited in Tillin et al., 2020), likely to compete with native species including native oyster and filter feeders such as Sabellaria alveolata (Cognie et al., 2006; Tillin et al., 2020). However, evidence has suggested Magallana gigas and some native species coexist, often forming more diverse reefs and habitats (e.g. Mytilus edulis and Ostrea edulis). For example, all sites studied in the Skagerrak area, Sweden colonized by Magallana gigas contained thriving populations of native oyster Ostrea edulis (Bergstrom et al., 2021) and there is no spatial competition identified between native Ostrea edulis and the Pacific oyster in the Northern Adriatic Sea, although densities of the Pacific oyster were significantly higher (Stagličić et al., 2020). In Balgzand, Wadden Sea, the impact on the food web and the biomass of Magallana gigas remained low (Jung et al., 2020).

The global spread of the Pacific oyster has facilitated the introduction of macrofauna and microparasites associated with oysters, including harmful algae and disease agents (Padilla, 2010). It is recognised that copepod parasites of Magallana gigas, Mytilicola orientalis and Myicola ostreae were introduced with imports of the oyster from France to Ireland (Tillin et al., 2020). Mytilicola orientalis was introduced into the Wadden Sea by Magallana gigas and infected blue mussels (Goedknegt et al., 2020). Predator avoidance by blue mussels in biogenic oyster reefs can indirectly affect parasite-host interactions. For example, in the Wadden Sea, one mixed mussel and oyster reef had significantly higher abundance of parasitic Mytilicola spp. in mussels at the top of the reef compared to at the bottom (Goedknegt et al., 2020). In contrast, with increasing oyster density, an increase in the presence of the trematode Renicola roscovita was seen in mussels (Goedknegt et al., 2019). Magallana gigas is also the predominant host of the shell-boring parasites Polydora ciliata and Polydora websteri in the Wadden Sea, with relatively higher densities of Polydora ciliata found in the Pacific oyster compared to the blue mussels (Waser et al., 2021).

Sensitivity assessment. While most of the evidence suggests the environmental conditions within this biotope are suitable for Magallana gigas, it is unlikely that they would be able to colonize this biotope without the removal of the macroalgal canopy. In addition, populations may be limited to low densities due to very wave exposed to wave exposed conditions. Although Herbert et al. (2016) found that Magallana gigas has colonized chalk habitats, its settlement may be mitigated by the macroalgal canopy and the level of wave exposure characteristic of this biotope. Therefore, resistance to Magallana gigas invasion is assessed as ‘High’, resilience as ‘High’ (no impact to recover from), and sensitivity as ‘Not sensitive’, albeit with ‘Low’ confidence due to the lack of direct evidence.

Wireweed, Sargassum muticum

Evidence

Sargassum muticum is a circumglobal invasive species (Engelen et al., 2015). It is recorded 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, can tolerate temperatures from -1°C to 30°C, and survives salinities below 10 PSU. Although fertilization does not occur below 15 PSU and the 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). 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). 

Sargassum muticum was shown to replace and outcompete leathery, canopy-forming macroalgae such as Saccharina latissima, Halidrys siliquosa, and Fucus spp. and, to a lesser degree, understorey species such as Codium fragile, Chondrus crispus and Dictyota dichotoma in Limfjorden, Denmark, between 1984 and 1997 (Staehr et al., 2000; Engelen et al., 2015; de Bettignies et al., 2021). The invasion in Limfjorden had stabilised by 2005, although many of the native macroalgal species continued to decline (Engelen et al., 2015). In Limfjorden, the distribution of Sargassum muticum was limited to areas with hard substratum, in particular stones >10 cm in diameter, while smaller stones, gravel and sand were unsuitable. It was most abundant between 1 and 4 m in depth but had low cover at 0 to 0.5 m and 4 to 6 m, in the turbid waters of the Limfjorden. Limfjorden is wave sheltered, but wave exposure has been reported to restrict the growth and survival of Sargassum muticum (Staehr et al., 2000). Viejo et al. (1995) reported that Sargassum muticum transplanted to wave exposed shores in Spain experienced >80% breakages within a month and that the growth of undamaged plants was significantly lower than that of plants on sheltered shores. Similarly, Andrew & Viejo (1998) noted that Sargassum muticum was restricted to intertidal rockpools in wave exposed sites in the Bay of Biscay. 

Sargassum muticum can outcompete some native fucoids. The cover of Fucus vesiculosus was inversely correlated with the cover of the invasive Sargassum muticum, indicating competitive interaction between the two species (Stæhr et al., 2000). Stæhr et al. (2000) determined that the invasion of Sargassum muticum could affect local algal communities through competition mainly for light and space.

Armitage et al. (2017) examined competitive interactions between Fucus serratus, Saccharina latissima and Sargassum muticum in a field experiment in south-west Norway across two summer seasons differing in temperature (mean water temperature at 1 m depth: 15.1°C in a warm year versus 12.4°C in a cooler year). Under warmer summer conditions, Fucus serratus exhibited greater weight gain than both competitors. The effect of Sargassum muticum on native species was no greater than the effect of intraspecific competition within those species. At the end of both summers, Sargassum muticum was in poor condition, potentially due to low seawater nutrient concentrations and low internal nitrogen status. These findings indicate that competitive outcomes between Fucus serratus and Sargassum muticum may be context-dependent and influenced by temperature and nutrient availability.

However, Atkinson et al. (2020) investigated the response of an intertidal rock pool assemblage invaded by Sargassum muticum to simulated marine heatwaves (+1.5, +2.0, and +3.5°C above 14°C for 14 days, followed by a 14-day recovery period) in both summer and winter. Marine heatwave treatments negatively affected the growth and photosynthetic performance of native seaweeds, including Fucus serratus, while enhancing the performance of Sargassum muticum. This pattern was consistent across seasons, indicating that Sargassum muticum may gain a physiological advantage over Fucus serratus under elevated temperature events.

Sensitivity assessment. Sargassum muticum is capable of competing with native fucoids for light and space (Stæhr et al., 2000). However, evidence specific to Fucus serratus indicates that competitive outcomes are environmentally contingent. However, Sargassum muticum has not been recorded in chalk or clay habitats. In addition, the level of wave exposure experienced by this biotope is generally unfavourable for Sargassum muticum, which thrives in sheltered sites. Based on the evidence, resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity is assessed as ‘Not Sensitive’, albeit with low confidence.

Wakame, Undaria pinnatifida

Evidence

Undaria pinnatifida, also known as Wakame or Asian Kelp, is a large brown seaweed and an Invasive Non-Indigenous Species (INIS) that could outcompete native UK kelp species (see Farrell & Fletcher, 2006; Thompson & Schiel, 2012; Brodie et al., 2014; Heiser et al., 2014; Arnold et al., 2016; Epstein & Smale, 2017; 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, the Netherlands, 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 all coastlines. Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound.

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). James (2017) suggested that Undaria pinnatifida could outcompete native species on artificial substrata (such as marinas and wharf structures). In Plymouth, UK, De Leij et al. (2017) found that natural habitats with dense native macroalgal canopies, such as Laminaria hyperborea, Laminaria ochroleuca, Laminaria digitata and Saccharina latissima, had more resistance to Undaria pinnatifida invasion than disturbed or sparse canopies, due to limited space and light availability for Undaria pinnatifida recruits. However, the dense canopies did not always prevent the invasion of Undaria pinnatifida as sporophytes were still recorded within dense Laminaria canopies, so canopy disturbance was not always required (De Leij et al., 2017; Epstein & Smale, 2018).

Undaria behaves as a winter annual, and recruitment occurs in winter followed by rapid growth through spring, maturity and then senescence through summer, with only the microscopic life stages persisting through autumn. It exhibits multiple dispersal strategies, such as short-range spore dispersal and long-range dispersal as whole drift plants or fragments. Undaria pinnatifida has spread rapidly across the UK and Europe, resulting in community-wide responses and impacts (Vaz-Pinto et al., 2014; Epstein & Smale, 2017). Its impacts are complex and context-specific, depending on space, time, and taxa present in the introduced location (Epstein & Smale, 2017; Teagle et al., 2017; Tidbury, 2020). 

Undaria pinnatifida has a wide physiological niche, meaning it can occur in both coastal and estuarine environments, showing 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, but increased in abundance and cover in very sheltered sites (Epstein & Smale, 2018). 

Thompson & Schiel (2012) found that native fucoids on New Zealand shores showed high resistance to invasions by Undaria pinnatifida. Undaria was able to recruit to artificial experimental gaps in the cover of the locally dominant fucoid Carpophyllum maschalocarpum. However, subsequent high local recruitment by the fucoid, coupled with annual die back of Undaria, reduced subsequent recruitment by Undaria, and the native fucoid had re-established its dominance within a year (Thompson & Schiel, 2012).

Sensitivity assessment. Undaria prefers sheltered conditions, usually in artificial habitats or in sheltered embayments, usually at the bottom of the shore and shallow subtidal. In addition, the evidence from New Zealand suggests that dense fucoid canopies can exclude Undaria. Furthermore, Undaria pinnatifida has not been recorded in chalk or clay habitats. It is therefore unlikely that Undaria pinnatifida poses a threat to this biotope. Resistance is assessed as ‘High’, resilience as ‘High’ by default, and sensitivity as ‘Not Sensitive’.

Other INIS

Evidence

The surface of the chalk is friable and subject to ongoing erosion, and therefore, the suitability of this habitat for large, long-lived, attached species is limited. The red algal mat characterizing this biotope consists of small, ephemeral species. Replacement of these species by invasive non-indigenous algal species could occur and alter the character of the biotope. However, no evidence was found in the literature to suggest that invasive non-indigenous species are present in UK marine and coastal peat habitats.

The American piddock, Petricolaria pholadiformis, is a non-native, boring piddock that was unintentionally introduced from America with the American oyster, Crassostrea virginica, not later than 1890 (Naylor, 1957). Rosenthal (1980) suggested that from the British Isles, the species has colonized several northern European countries by means of its pelagic larva and may also spread via driftwood, although it usually bores into clay, peat, or soft rock shores. In Belgium and the Netherlands, Petricolaria pholadiformis almost completely displaced the native piddock, Barnea candida (ICES, 1972). However, this has not been observed elsewhere. Later studies have found that Barnea candida is now more common than Petricolaria pholadiformis in Belgium (Wouters, 1993). There is no documentary evidence to suggest that Barnea candida has been displaced in the British Isles (J. Light & I. Kileen pers. comm. to Eno et al., 1997). Petricolaria pholadiformis is considered unlikely to exclude Pholas dactylus, which is more likely to occur subtidally. Should Petricolaria pholadiformis be present in this biotope, it would probably not alter the character or ecological function of the biotope.

Although not currently established in UK waters, the whelk Rapana venosa, may spread to habitats. This species has been observed predating on Pholas dactylus in the Romanian Black Sea by Micu (2007).

Sensitivity assessment. Based on the lack of records of invasive non-indigenous species in this biotope, and the unsuitability of the habitat for algae and other attached epifauna, this biotope is considered to have ‘High’ resistance to this pressure and, by default, 'High’ resilience. Therefore, the biotope is considered to be ‘Not sensitive’. This assessment may need revising in light of future invasions, e.g. the introduction of the whelk Rapana venosa.