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

Populations of Fucus ceranoides are restricted to estuaries, but examples this biotope are found throughout estuaries in the British Isles.  The distribution of Fucus ceranoides in the north east Atlantic extends from northern Norway (Lein, 1984) to the River Mondego in Portugal (Neiva et al., 2012).  Although species ranges may not accurately describe their ability to withstand localized changes in temperature.  They may to some extent display the limits of the species genetic ability to acclimatize to temperatures.  Growth rates in Fucus ceranoides have not been recorded, and there is no empirical evidence on the upper and lower temperature tolerances.  Fucus ceranoides occur from high to low shore (depending on locality) (Connor et al., 2004; physical comparative tables).  This range in location on the shore suggests that Fucus ceranoides can tolerate a range of temperatures.  Fertilization at extremely high (25 °C) temperatures can only occur in full salinity (Burrows, 1964).

Ulva sp. are distributed globally (Guiry & Guiry, 2015 and references therein) and occur in warmer waters than those surrounding the UK suggesting that they can withstand increases in temperature at the pressure benchmark.  Ulva sp. are characteristic of upper shore rock pools, where water and air temperatures are greatly elevated on hot days.  Empirical evidence for thermal tolerance to anthropogenic increases in temperature is provided by the effects of heated effluents on rocky shore communities in Maine, USA.  Ascophyllum and Fucus were eliminated from a rocky shore heated to 27-30 °C by a power station whilst Ulva intestinalis (as Enteromorpha intestinalis) increased significantly near the outfall (Vadas et al., 1976).

Although adults may be able to withstand acute and chronic increases in temperature at the pressure benchmark, increased temperatures may have sub-lethal effects on the population through impacts on reproduction.  The distribution of Semibalanus balanoides is ‘northern’ with their range extending to the Arctic circle. Populations in the southern part of England are relatively close to the southern edge of their geographic range.  Long-term time series show that successful recruitment of Semibalanus balanoides is correlated to sea temperatures (Mieszkowska, et al., 2014) and that due to recent warming its range has been contracting northwards. Increased temperatures are likely to favour chthamalid barnacles or Austrominius modestus in the sheltered variable salinity biotopes rather than Semibalanus balanoides (Southward et al. 1995).

Sensitivity assessment.  An increase in temperature at the pressure benchmark may cause some Fucus ceranoides to die off at the top of the shore due to increased levels of desiccation caused by higher temperatures, especially during the summer months.  However, it will not cause large losses of the characterizing species.  Ulva sp. are resistant to high temperatures and are very quick to recolonize.  Examples of the biotope in a more southerly location may see an increase in the number of Austrominius modestus and a decrease in Semibalanus.  This is unlikely to cause any significant changes in the biotope as the functional niche will still be filled.  Both resistance and resilience have been assessed as ‘High’, meaning the biotope is ‘Not sensitive’ to an increase in temperature at the pressure benchmark.

Temperature decrease (local)

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

Evidence

Populations of Fucus ceranoides are restricted to estuaries, but examples this biotope are found throughout estuaries in the British Isles.  The distribution of Fucus ceranoides in the north east Atlantic extends from northern Norway (Lein, 1984) to the River Mondego in Portugal (Neiva et al., 2012).  Although species ranges may not accurately describe their ability to withstand localized changes in temperature.  They may to some extent display the limits of the species genetic ability to acclimatize to temperatures.  Growth rates in Fucus ceranoides have not been recorded, and there is no empirical evidence on the upper and lower temperature tolerances.  This biotope is occasionally found on the low shore (17.65%) but is almost as common on the upper shore (52.94%) as it is on the mid shore (47.06%) (Connor et al., 2004; physical comparative tables).  This range in location on the shore suggests that Fucus ceranoides can tolerate a range of temperatures.  Fertilization at extremely low (2 °C) temperatures can only occur in full salinity (Burrows, 1964).  At a water temperature of 10 °C fertilization is entirely inhibited below 8 ppt (Burrows, 1964). 

Ulva intestinalis and Ulva lactuca, occur in Arctic regions and Alaska and are therefore found in colder waters than those around the UK (Guiry & Guiry, 2015 and references therein), Ulva sp. (as Enteromorpha) were reported to be tolerant of a temperature of -20 °C (Kylin, 1917).  Vermaat & Sand-Jensen (1987) found that rapid deep freezing of Ulva lactuca collected in Roskilde Fjord, Denmark killed the plants.  However, individuals from the same area when collected from frozen ice, survived and resumed growth, the plants are able to survive more gradual natural freezing (Vermaat & Sand-Jensen, 1987).

The tolerance of Semibalanus balanoides collected in the winter (and thus acclimated to lower temperatures) to low temperatures was tested in the laboratory.  The median lower lethal temperature tolerance was -14.6 ^oC (Davenport & Davenport, 2005).  A decrease in temperature at the pressure benchmark is, therefore, unlikely to negatively affect this species.

Sensitivity assessment.  A decrease in temperature at the pressure benchmark is not likely to have a significant effect on Fucus ceranoides.  There may be some decrease in growth rate, and a greater effect of winter temperatures.  However, it will not cause large die backs of the characterizing species.  Ulva sp. are resistant to low temperatures and are very quick to recolonize.  Examples of the biotope in a more southerly location may see a decrease in the number of Austrominius modestus and an increase in Semibalanus.  This is unlikely to cause any significant changes in the biotope as the functional niche will still be filled.  Both resistance and resilience have been assessed as ‘High’, meaning the biotope is ‘Not sensitive’ to an increase in temperature at the pressure benchmark.

Salinity increase (local)

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

Evidence

Species found in the intertidal are likely to have some form of behavioural or physiological adaptations to changes in salinity.  Fucus ceranoides can grow and become fertile between salinities of 7 – 34 ppt (Burrows, 1964).  Its optimal salinity for growth rate is 16 ppt, however, successful fertilization occurs most commonly at full salinities (Burrows, 1964).  The eggs of Fucus ceranoides can tolerate immersion in freshwater within 2 hours of release (Burrows, 1964).  Germlings and excised tips of Fucus ceranoides develop tissue damage and decay in full salinity water (Khfaji & Norton, 1979).  If Fucus ceranoides is subject to fully marine conditions it decays within a few months (Khfaji & Norton, 1979; Suryono & Hardy, 1997).  The evidence suggests that the fluctuation in salinity is important to the survival of Fucus ceranoides.  And even if Fucus ceranoides was able to tolerate fully marine conditions, it is unlikely to survive as it grows slower than other fucoids such as Fucus vesiculosus (Khfaji & Norton, 1979) and would be out-competed for space.  An increase in the salinity regime would also mean that the habitat would become suitable for littorinids.  This grazing species has a significant negative impact on Fucus ceranoides through grazing (Norton, 1986), and can completely remove Fucus ceranoides as the macroalga doesn’t have any mechanism to deter it.

Ulva sp. can survive hypersaline conditions in supralittoral rockpools subjected to evaporation and is considered to be a very euryhaline species, tolerant of extreme salinities ranging from 0 psu to 136 psu (Reed & Russell, 1979).  Some variations in salinity tolerance between populations of Ulva intestinalis have been found, however, suggesting that plants have some adaptation to the local salinity regime.

Semibalanus balanoides is tolerant of a wide range of salinities and has the ability to isolate themselves from water by closing their opercula valves (Foster, 1971b).  They can also withstand large changes in salinity over moderately long periods of time by falling into a "salt sleep" and can be found on shores (example from Sweden) with large fluctuations in salinity around a mean of 24 (Jenkins et al., 2001).  Barnes & Barnes (1974) found that larvae from Semibalanus (as Balanus) balanoides, complete their development to nauplii larvae at salinities between 20-40%.

Sensitivity assessment.  An increase in salinity at the benchmark within this biotope would have a negative impact on the health and reproductive capacity of Fucus ceranoides.  The change would lead to other intertidal fucoids, such as Fucus serratus and Fucus vesiculosus, recruiting into the habitat.  These species will out-compete Fucus ceranoides consequently forcing them from the shore and removing the biotope.  The increase in salinity will also mean that littorinid grazers would move into the habitat and begin grazing.  It is likely that species richness will rise as the substratum would probably to colonized by marine species which were previously excluded by an intolerance to reduced salinity.  Resistance has been assessed as ‘Low’ as the change in salinity at the benchmark would lead to Fucus ceranoides being both grazed and out-competed by species with a greater tolerance to salinity. The resilience is assessed as ‘Medium’ resulting in ‘Medium’ sensitivity.

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 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.  Fucus ceranoides can grow and become fertile between salinities of 7 and 34 ppt (Burrows, 1964).  Its optimal salinity for growth rate is 16 ppt. However, successful fertilization occurs most commonly at high salinities (Burrows, 1964).  The eggs of Fucus ceranoides can tolerate immersion in freshwater within two hours of release (Burrows, 1964).  Germlings and excised tips of Fucus ceranoides develop tissue damage and decay in full salinity water (Khfaji & Norton, 1979). Cotas et al., (2019) noted that Fucus ceranoides was well-adapted to low salinity (approx. 25 ppt) conditions in the Mondego River (Portugal), and that its main obstacle to successfully colonizing the higher salinity (approx. 35 ppt) regions of this habitat was inter-species competition with Fucus vesiculosus, which is better adapted to full salinity conditions. A decrease in salinity would lead to a low salinity (<18 ppt) (Connor et al., 2004).  Low salinity would decrease the success rates of fertilization but would still allow Fucus ceranoides to survive.

Ulva sp. can survive hypersaline conditions in supralittoral rockpools subjected to evaporation and is considered to be a very euryhaline species, tolerant of extreme salinities ranging from 0 psu to 136 psu (Reed & Russell, 1979).  Variations in salinity tolerance in populations of Ulva intestinalis have been found suggesting that plants have some adaptation to the local salinity regime.  Alströem-Rapaport et al. (2010) found that in the brackish Baltic Sea, Ulva intestinalis uses a variety of reproductive modes which was considered to partly explain the high rates of colonisation and adaptability of the species.  Reed & Russell (1979) found that the ability to regenerate from cut thalli varied according to the salinity conditions of the original habitat, and that the pattern of euryhalinity in parental material and offspring was in broad agreement (Reed & Russell, 1979).  For example, eulittoral zone material showed decreased percentage regeneration in all salinities (dilute: 0, 4.25, 8.5, 17 & 25.5 psu, and concentrated seawater: 51, 68, 95, 102 & 136 psu) except 34 psu, when compared to littoral fringe populations of Ulva intestinalis (studied as Enteromorpha intestinalis).  None of the eulittoral zone material was able to regenerate in freshwater or concentrated seawater, whilst littoral fringe and rock pool material was able to do so.

Semibalanus balanoides is tolerant of a wide range of salinities and has the ability to isolate themselves from water by closing their opercula valves (Foster, 1971b).  They can also withstand large changes in salinity over moderately long periods of time by falling into a "salt sleep" and can be found on shores (example from Sweden) with large fluctuations in salinity around a mean of 24 (Jenkins et al., 2001).  Barnes & Barnes (1974) found that larvae from Semibalanus (as Balanus) balanoides, complete their development to nauplii larvae at salinities between 20-40%.

Sensitivity assessment.  A decrease in salinity at the benchmark within this biotope would reduce the fertilization success of Fucus ceranoides, but would still allow growth and reproduction.  Other species in the biotope would still be able to survive for some time at low salinities, but their abundance would reduce over the year, and they may die back altogether.  When the salinity regime returned to normal (reduced / variable), species such as Semibalanus balanoides and Ulva sp. would return quickly.  Resistance has been assessed as ‘Medium’, as has the resilience, resulting in the biotope having a ‘Medium’ sensitivity to this pressure 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

This biotope is found in moderately strong – very weak (1.5 m/sec. - negligible) conditions.  However, the characterizing species Fucus ceranoides is limited to the salinity regimes found within estuaries, environments which also happen to have these water flow conditions.  Therefore, the level of water flow experienced within the biotope may only be a factor of the physical environment and not a limiting factor for the species.  No empirical evidence is available on water flow tolerances in Fucus ceranoides.  An investigation into Fucus spiralis (a species whose fronds reach a maximum length comparable to that of Fucus ceranoides) found that plants were torn from their anchorage at7 – 8 m/sec (Jonsson et al., 2006).  Even if water flows an increase in water flow didn’t directly remove adults from the substrata, it could have an impact on dispersal, fertilization, settlement and recruitment (Pearson & Brawley, 1996).

Both Ulva intestinalis and Ulva lactuca are flexible and conform to the direction of the flow reducing drag and breakage.  However, experimental studies show that exposure to currents results in sloughing of tissue and higher current velocities result in breakage of the thallus.  Kennison & Fong (2013) found that Ulva intestinalis, settled on ceramic tiles and deployed in the field were subject to greater losses at mean flow speeds of 0.2 m/s  (approximately 16 % of biomass) than the 8 % loss from individuals subject to lower flows (0.15 m/s).  Increased water velocities can enhance recruitment through increased larval supply (Kennison & Fong, 2013).  Houghton et al. (1973) observed that swarmers of Ulva were able to settle onto surfaces subjected to water speeds of up to 10.7 knots, suggesting that changes may not inhibit settlement.

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).  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.

Sensitivity assessment:  At the level of the benchmark a change in water flow will not have an impact on the biological communities within these biotopes.  Consequently, both resistance and resilience have been assessed as ‘High’, resulting in an overall sensitivity of ‘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.  Within this biotope Fucus ceranoides is found from the upper intertidal, down to the lower shore, where it is most sparse (Connor et al., 2004; Physical comparative tables).  A change in emergence would lead to a shift in the area covered by Fucus ceranoides and this biotope.  However, the presence of this biotope at all shore heights will allow the biotope to persist throughout.

Ulva sp. lose their water content very fast, but overcome the problem by growing in dense populations where they can cover and shade each other to some extent when exposed.  As Ulva intestinalis is able to tolerate dessication stress it is often very abundant on the high shore where desiccation stress is the primary factor controlling seaweed distribution, and may even be found above the tidal limits of the shore.  Ulva intestinalis (studied as Enteromorpha intestinalis) can survive several weeks of living in completely dried out rock pools, while becoming completely bleached on the uppermost layers, but remaining moist underneath the bleached fronds.  However, dessication stress of germlimgs may be lower than adults Hruby & Norton (1979) found that 7-14 day old germlings of Ulva (studied as Enteromorpha) were more tolerant of desiccation than earlier stages, so an increase in desiccation stress resulting from increased emergence may impact more adversely on newly settled germlings than more mature plants.

A change in emergence would alter the feeding time and increase desiccation for barnacles within the biotope.  Semibalanus balanoides is less tolerant of desiccation stress than other barnacle species.  If emergence were to change, there may be a shift in barnacle species within the biotope.  However, if the barnacle species were to change the functional niche would still be filled leading to the biotope remaining.

Sensitivity assessment.  This biotope extends from the high shore through to the low shore (Connor et al., 2004).  If there was a decrease in emergence then the lower limit of the biotope would shift up the shore.  However, if there was an increase in emergence the top part of the biotope would be lost and the biotope would shift down the shore.  The physiological tolerances of Fucus ceranoides would not be able to tolerate the changes in physical factors associated with the change in emergence.  Therefore, resistance is assessed as ‘Medium’ as some of the biotope would be lost.  However, the remaining Fucus ceranoides would provide a mature stand from which the shore could recolonise quickly.  This means the resistance is also ‘Medium’, giving the biotope a sensitivity of ‘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

This biotope is found in very sheltered to extremely sheltered conditions.  However, the characterizing species Fucus ceranoides is limited to the salinity regimes found within estuaries, environments which also happen to have these wave exposures.  Therefore, wave exposures experienced within the biotope may only be a factor of the physical environment and not a limiting factor for the biological despecies.  No empirical evidence is available on wave exposure tolerances in Fucus ceranoides.  However, the water flows experienced within this biotope (very weak to moderately strong (Connor et al., 2004) suggest that an increase in wave exposure would not have a negative impact on the characterizing species. The remaining species within this biotope are found within other biotopes where wave exposure exceeds that found within this biotope. 

Sensitivity assessment.  At the benchmark this biotope is ‘Not Sensitive’ to a change in wave exposure.  This means that both resistance and resilience are ‘High’.

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 ceranoides  bioaccumulates Al, Cr, Fe, Ni, Cu, Zn, As, Cd, and Hg. Hence,  Pacín et al. (2025) was able to use archived samples of this species to analyse trends in the environmental prevalence of these elements in northwest Spain over 30 years. Hassan et al. (2021) found that simultaneous exposure to high concentrations (36 μg/L) of copper and low salinity (8.5 ppt) significantly reduced photosynthetic efficiency in Fucus ceranoides. Increasing concentrations (50-150 μg/L) of silver inhibited growth in Fucus ceranoides from the Looe estuary in southwest England, with toxicity exacerbated by reduced salinity (10 ppt, compared to 28 ppt).

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.

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.

Radionuclide contamination

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

Evidence

No evidence.

Introduction of other substances

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

Evidence

This pressure is Not assessed.

De-oxygenation

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

Evidence

The biotope in question occurs in the eulittoral and consequently, a proportion of time will be spent in the air where oxygen is not limited so the metabolic processes of photosynthesis and respiration can take place.  This biotope is found in very sheltered to extremely sheltered and the moderately strong to negligible tidal flows, consequently water mixing is not very strong.  Therefore, water movement within this area will not reverse any oxygen depletion quickly, possibly exacerbating any negative effects. 

Sensitivity assessment. The characterizing species Fucus ceranoides and other species present within this biotope require oxygen for respiration.  The intertidal position of this biotope means that all species spend periods of time emersed, where oxygen is not limited.  This is advantageous for algae and can allow them to continue to photosynthesis and respire.  However, this is not the case for barnacles as they need to be immersed to obtain oxygen.  Barnacles appear are able to respire anaerobically for some time and consequently some individuals may be able to survive the pressure at this benchmark.  Resistance is assessed as ‘High’.  Resilience is assessed as ‘High’, giving the biotope a sensitivity score 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.  The nutrient enrichment of a marine environment leads to organisms no longer being limited by the availability of certain nutrients.  The consequent changes in ecosystem functions can lead to the progression of eutrophic symptoms (Bricker et al., 2008), changes in species diversity and evenness (Johnston & Roberts, 2009) decreases in dissolved oxygen and uncharacteristic microalgae blooms (Bricker et al., 1999, 2008).

No direct evidence was found to assess the effect of this pressure on this biotope.  A slight increase in nutrient levels could be beneficial for barnacles by promoting the growth of phytoplankton levels and therefore increasing zooplankton levels.  However, Holt et al. (1995) predict that smothering of barnacles by ephemeral green algae is a possibility under eutrophic conditions.

Opportunistic algae, including Ulva sp.,a contaminants can’t store nutrients in the thallus (unlike larger, long-lived species) and are adapted to efficiently capture and utilise available nutrients in the water column (Pedersen et al., 2009).  A large body of field observations and experiments, surveys and laboratory experiments confirm that Ulva sp., can utilise high levels of nutrients for growth (Martínez et al., 2012) and that enhanced recruitment (Kraufvelin et al., 2007) and growth of this genus can occur in enriched areas (Kennison & Fong, 2013; Vaudrey et al., 2010).  In areas where nutrient availability is lower either naturally or through management to reduce anthropogenic inputs, Ulva sp. may be negatively affected through reduced growth rate and species replacement (Martínez et al., 2012; Vaudrey et al., 2010).

Johnston & Roberts (2009) undertook a review and meta-analysis of the effect of contaminants on species richness and evenness in the marine environment.  Of the 47 papers reviewed relating to nutrients as contaminants, over 75% found that it had a negative impact on species diversity, <5% found increased diversity, and the remaining papers finding no detectable effect.  Not all of the 47 papers considered the impact of nutrients on intertidal rocky shores.  Yet this finding is still relevant as the meta-analysis revealed that the effect of marine pollutants on species diversity was ‘remarkably consistent’ between habitats (Johnston & Roberts, 2009).  It was found that any single pollutant reduced species richness by 30-50% within any of the marine habitats considered (Johnston & Roberts, 2009).  Throughout their investigation, there were only a few examples where species richness was increased due to the anthropogenic introduction of a contaminant.  These examples were almost entirely from the introduction of nutrients, either from aquaculture or sewage outfalls.  However, research into the impacts of nutrient enrichment from aquaculture or sewage on intertidal rocky shores often lead to shores lacking species diversity and the domination by algae with fast growth rates (Abou-Aisha et al., 1995, Archambault et al., 2001, Arévalo et al., 2007, Diez et al., 2003, Littler & Murray, 1975).

Nutrient enrichment alters the selective environment by favouring fast growing, ephemeral species such as Ulva lactuca and Ulva intestinalis (Berger et al., 2004, Kraufvelin et al., 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 et al., 2007).  Nutrient enrichment can also enhance fouling of Fucus fronds by biofilms (Olsenz, 2011).  Nutrient enriched environments can not only increase algae abundance but the abundance of grazing species (Kraufvelin et al, 2007).

Changes in community composition on intertidal rocky shores can happen rapidly, and fast growing ephemeral species can become established quickly in the presence of higher concentrations of nutrients.  The establishment and growth of these species are not controlled by wave exposure (Kraufvelin et al., 2007).  However, even though these fast growing ephemeral species can become well established quickly, healthy communities on intertidal rocky shores can survive long periods of time, and maintain ecological function after these species have become established (Bokn et al., 2002, 2003, Karez et al.,2004, Kraufvelin, 2007, Kraufvelin et al., 2006).

Sensitivity assessment. A slight increase in nutrients may enhance growth rates but high nutrient concentrations could lead to the overgrowth of the algae by ephemeral green algae and an increase in the number of grazers.  However, if the biotope is well established and in a healthy state the biotope could have the potential to persist.  Therefore, the resistance has been assessed as ‘High’.  As there will be nothing for the biotope to recover from therefore the resilience is also ‘High’. These two rankings give an overall sensitivity of ‘Not Sensitive’.

Organic enrichment

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

Evidence

The organic enrichment of a marine environment at this pressure benchmark leads to organisms no longer being limited by the availability of organic carbon.  The consequent changes in ecosystem functions can lead to the progression of eutrophic symptoms (Bricker et al., 2008), changes in species diversity and evenness (Johnston & Roberts, 2009) and decreases in dissolved oxygen and uncharacteristic microalgae blooms (Bricker et al., 1999, 2008).

Johnston & Roberts (2009) undertook a review and meta-analysis of the effect of contaminants on species richness and evenness in the marine environment.  Of the 49 papers reviewed relating to sewage as a contaminant, over 70% found that it had a negative impact on species diversity, <5% found increased diversity, and the remaining papers finding no detectable effect.  Not all of the 49 papers considered the impact of sewage on intertidal rocky shores.  Yet this finding is still relevant as the meta-analysis revealed that the effect of marine pollutants on species diversity was ‘remarkably consistent’ between habitats (Johnston & Roberts, 2009).  It was found that any single pollutant reduced species richness by 30-50% within any of the marine habitats considered (Johnston & Roberts, 2009).  Throughout their investigation, there were only a few examples where species richness was increased due to the anthropogenic introduction of a contaminant.  These examples were almost entirely from the introduction of nutrients, either from aquaculture or sewage outfalls.  However, research into the impacts of organic enrichment from aquaculture and sewage on intertidal rocky shores often led to shores lacking species diversity and the domination by algae with fast growth rates (Abou-Aisha et al., 1995, Archambault et al., 2001, Arévalo et al., 2007, Diez et al., 2003, Littler & Murray, 1975).

Organic enrichment alters the selective environment by favouring fast growing, ephemeral species such as Ulva lactuca and Ulva intestinalis (Berger et al., 2004, Kraufvelin et al., 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 et al., 2007).  Nutrient enrichment can also enhance fouling of fucoid fronds by biofilms (Olsenz, 2011).  Nutrient enriched environments can not only increase algae abundance but the abundance of grazing species (Kraufvelin et al., 2007).  Bellgrove et al. (2010) found that coralline turfs out-competed fucoids at a site associated with organic enrichment caused by an ocean sewage outfall.

Changes in community composition on intertidal rocky shores can happen rapidly, and fast growing ephemeral species can become established quickly in the presence of higher concentrations of nutrients. The establishment and growth of these species are not controlled by wave exposure (Kraufvelin, 2007).  However, even though these fast growing ephemeral species can become well established quickly, healthy communities on intertidal rocky shores can survive long periods of time, and maintain ecological function after these species have become established (Bokn et al., 2002, 2003; Karez et al.,2004, Kraufvelin et al., 2006b; Kraufvelin et al., 2007).

Sensitivity assessment.  Little empirical evidence was found to support an assessment of this biotope at this benchmark.  The contribution of Ulva sp. to the biological community within this biotope means that if conditions were to become more favourable for this species it could easily increase in abundance.  An increase in this species could over grow the slower growing characterizing species Fucus ceranoides.  Due to the potential negative impacts that have been reported to result from the introduction of excess organic carbon, resistance has been assessed as ‘Medium’ and resilience has been assessed as ‘Medium’.  This gives an overall sensitivity score of ‘Medium’.

Physical loss (to land or freshwater habitat)

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

Evidence

All marine habitats and benthic species are considered to have a resistance of ‘None’ to this pressure and to be unable to recover from a permanent loss of habitat (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

This biotope occurs on hard rock bedrock and boulders on which the associated biological community attach.  A change in substratum would significantly alter the character of the biotope and would lead to the development of a biological assemblage more typical of the changed conditions.  A change to an artificial substratum could also impact this biotope as species may have settlement preferences for particular surface textures.  Therefore, the replacement of natural surfaces with artificial ones may lead to changes in the biotope through changes in species composition, richness and diversity (Green et al., 2012; Firth et al., 2013).

Sensitivity assessment.  A change in substratum would result in the loss of the characterizing species Fucus ceranoides along with other species associated with this biotope.  Resistance is assessed as ‘None’.  As this pressure represents a permanent change, recovery is impossible as a suitable substratum for the biological community of this biotope is lacking. Consequently, resilience is assessed as ‘Very Low’.  The habitat scores a ‘High’ sensitivity. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.

Physical change (to another sediment type)

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

Evidence

Not relevant to this biotope which occurs predominantly on bedrock (Connor et al., 2004).

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 species characterizing this biotope occur on rock and would be sensitive to the removal of the habitat.  However, extraction of rock substratum is considered unlikely and this pressure is considered to be ‘Not relevant’ to hard substratum habitats.

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

Trampling on the rocky shore has been observed to reduce fucoid cover which decreased the microhabitat available for epiphytic species, increased bare space and increased cover of opportunistic species such as Ulva sp. (Fletcher & Frid, 1996a).  This biotope is found on the intertidal shore, an area easily accessible by humans especially at low tide.  Individual macroalgae are flexible but not physically robust.  Fucoids are intolerant of abrasion from human trampling, which has been shown to reduce the cover of seaweeds on a shore (Holt et al., 1997).  Brosnan (1993) investigated the effect of trampling on an intertidal rocky shore area dominated by algae and barnacle assemblages in Oregon.  The effects of 250 tramples per plot, once a month for a year were recorded.  Abundances of algae in each plot were reduced from 80% to 35% within a month of the introduction of the pressure and remained low for the remainder of the experiment.  As few as 20 steps / m² on stations on an intertidal rocky shore in the north east of England were sufficient to reduce the abundance of fucoids (Fletcher & Frid, 1996a).  This reduction in the complexity of the algae community, in turn, reduced the microhabitat available for epiphytic species.  Trampling pressure can thus result in an increase in the area of bare rock on the shore (Hill et al., 1998). Chronic trampling can affect community structure with shores becoming dominated by algal turf or crusts (Tyler-Walters, 2005).  Pinn & Rodgers (2005) compared the biological communities found on two intertidal rocky shore ledges in Dorset. They found that the ledge that had a higher number of visitors had few branching algal species, including fucoids, but had greater abundances of crustose and ephemeral species (Pinn & Rodgers, 2005).  The densities of fucoids were recorded from the intertidal rocky shore at Wembury, Devon in 1930 (Colman, 1933) and 1973 (Boalch et al., 1974).  Boalch et al. (1974) found a reduction in fucoids on the shore at Wembury compared to the abundances recorded by Coleman (1933). 

No direct evidence was found to assess how Ulva spp. respond to surface abrasion.  The fronds are very thin and could be torn and damaged and individuals may be removed from the substratum, altering the biotope through changes in abundance and biomass.  Ulva spp. cannot repair damage or reattach but torn fronds could still photosynthesise and produce gametes.  Tearing and cutting of the fronds have been shown to stimulate gamete production and damaged plants would still be able to grow and reproduce.  Therefore, damage and removal may not lead to mortality of impacted individuals.  Ulva spp. can also form unattached mats, particularly in response to nutrient enrichment.

The barnacles which contribute to this biotope typically occur on the rock surfaces where they will be exposed to abrasion.  Although barnacles are protected by hard shells or plates, abrasion may damage and kill individuals or detach these.  All removed barnacles would be expected to die as there is no mechanism for these to reattach.  Evidence for the effects of abrasion are provided by a number of experimental studies on trampling (a source of abrasion) and on abrasion by wave thrown rocks and pebbles.  The effects of trampling on barnacles appear to be variable with some studies not detecting significant differences between trampled and controlled areas (Tyler-Walters & Arnold, 2008).  However, this variability may be related to differences in trampling intensities and abundance of populations studied.  The worst case incidence was reported by Brosnan & Crumrine (1994) who reported that a trampling pressure of 250 steps in a 20 x 20 cm plot one day a month for a period of a year significantly reduced barnacle cover at two study sites.  Barnacle cover reduced from 66% to 7% cover in 4 months at one site and from 21% to 5% within 6 months at the second site.  Overall barnacles were crushed and removed by trampling.  Barnacle cover remained low until recruitment the following spring.  Long et al. (2011) also found that heavy trampling (70 humans / km^-1 shoreline h^-1) led to reductions in barnacle cover.  Single step experiments provide a clearer, quantitative indication of sensitivity to direct abrasion.  Povey & Keough (1991) in experiments on shores in Mornington peninsula, Victoria, Australia, found that in single step experiments 10 out of 67 barnacles, (Chthamalus antennatus about 3 mm long), were crushed.  However, on the same shore, the authors found that limpets may be relatively more resistant to abrasion from trampling.

Sensitivity assessment.  Fucoids are not tolerant of abrasion, and the effect of the pressure at this benchmark is very likely to cause a decrease in abundance of Fucus ceranoides.  Barnacles are also negatively affected by trampling.  Therefore, the resistance is ‘Low’; the resilience is assessed as ‘Medium’.  Consequently, the sensitivity of this biotope to abrasion has been assessed as ‘Medium’.

Penetration or disturbance of the substratum subsurface

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

Evidence

The characterizing species of this biotope occurs on rock which is resistant to subsurface penetration.  The assessment for abrasion at the surface only is therefore considered to equally represent sensitivity to this pressure.

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

In general, increased suspended particles may enhance food supply (where these are organic in origin) or decrease feeding efficiency (where the particles are inorganic and require greater filtration efforts).  Very high levels of silt may clog respiratory and feeding organs of the suspension feeding barnacle species.  In addition, increased turbidity will decrease light penetration reducing photosynthesis by macroalgae within this biotope.  Fucus ceranoides is found within all zones of intertidal rocky shores.  Water depths within this part of the intertidal will never be great.  So although an increase in turbidity at this pressure benchmark will increase light attenuation it is likely that light will still penetrate to the depth at which the algae are found.  Those examples of Fucus ceranoides found on the lower shore may be mildly affected by this and may experience a slight decrease in growth rate, however, there is no empirical evidence for this.  A decrease in SPM could is unlikely to have a negative impact on the macroalgae within this biotope.  Increased levels of particles may increase scour and deposition in the biotope depending on local hydrodynamic conditions, although changes in substratum are assessed through the physical change (to another seabed type) pressure.

A significant decrease in suspended organic particles may reduce food input to the biotope resulting in reduced growth and fecundity of barnacles.  However, local primary productivity may be enhanced where suspended sediments decrease, increasing food supply.

Sensitivity assessment.  At this pressure benchmark resistance and resilience of this biotope has been assessed as ‘High’.  The sensitivity of this biotope to this pressure at the benchmark is ‘Not Sensitive’.

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

In general, increased suspended particles may enhance food supply (where these are organic in origin) or decrease feeding efficiency (where the particles are inorganic and require greater filtration efforts).  Very high levels of silt may clog respiratory and feeding organs of the suspension feeding barnacle species.  In addition, increased turbidity will decrease light penetration reducing photosynthesis by macroalgae within this biotope.  Fucus ceranoides is found within all zones of intertidal rocky shores.  Water depths within this part of the intertidal will never be great.  So although an increase in turbidity at this pressure benchmark will increase light attenuation it is likely that light will still penetrate to the depth at which the algae are found.  Those examples of Fucus ceranoides found on the lower shore may be mildly affected by this and may experience a slight decrease in growth rate, however, there is no empirical evidence for this.  A decrease in SPM could is unlikely to have a negative impact on the macroalgae within this biotope.  Increased levels of particles may increase scour and deposition in the biotope depending on local hydrodynamic conditions, although changes in substratum are assessed through the physical change (to another seabed type) pressure.

A significant decrease in suspended organic particles may reduce food input to the biotope resulting in reduced growth and fecundity of barnacles.  However, local primary productivity may be enhanced where suspended sediments decrease, increasing food supply.

Sensitivity assessment.  At this pressure benchmark resistance and resilience of this biotope has been assessed as ‘High’.  The sensitivity of this biotope to this pressure at the benchmark is ‘Not Sensitive’.

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

A discrete event where sediment inundates this biotope to 30 cm would cover all species with a thick layer of fine materials.  During high tide, the characterizing species will be vertical in the water column.  Any deposition at this state of the tide will mean that only a proportion of the stipe and holdfast will be covered.  Fronds of Fucus ceranoides are between 30 – 60 cm in length (Bunker et al., 2012).  Consequently, sediment deposition at high tide could leave a proportion of Fucus ceranoides fronds sediment free.  In contrast, if the tide is out then fronds of the characterizing fucoid canopy will be flat on the substratum and will be smothered by the sediment deposit.  There is no empirical evidence to suggest how smothering effects Fucus ceranoides.  However, it may be assumed that inundation with sediment will reduce gaseous exchange limiting the macroalga's ability to photosynthesise and respire.  Without the ability to carry out these two functions fronds will become damaged and consequently die.  Due to the lack of information on this, timescales for this process are not available.  Moderately strong water flows caused by tidal movements within this biotope could remove from the shore within a few tidal cycles.  Germlings are likely to be smothered and killed regardless of the state of the tide and are inherently most susceptible to this pressure.  Sediment deposition can reduce macroalgal recruitment by (1) reducing the amount of substratum available for attachment of propagules; (2) scour, removing attached juveniles and (3) burial, altering the light and/or the chemical microenvironment (Devinny & Volse, 1978, Eriksson & Johansson, 2003).

Observations and experiments indicate that Ulva spp. have relatively high tolerances for the stresses induced by burial (darkness, hypoxia and exposure to sulphides).  Vermaat & Sand-Jensen , (1987) exposed thallus discs (113 mm²) of Ulva lactuca to darkness and anoxia and sulphides at winter temperatures.  It was found that these conditions did not affect survival over two months, although exposure to anoxia increased respiration and reduced growth (Vermaat & Sand-Jensen, 1987).   Although Ulva sp. present in sedimentary habitats may be able to survive the chemical stress of burial and re-grow from surviving fragments, evidence for attached individuals from rocky shores suggests that resistance to this pressure may be lower. 

Semibalanus balanoides and Austrominius modestus are likely to be negatively affected by siltation.  The lower limits of Semibalanus balanoides (as Balanus balanoides) appear to be set by levels of sand inundation on sand-affected rocky shores in New Hampshire (Daly & Mathieson, 1977), suggesting that this species is sensitive to the deposition of relatively coarse sediments, although whether this is due to repeated scour events removing juveniles rather than siltation effects (i.e. smothering, prevention of feeding) is not clear. Experiments have shown that the addition of even thin layers of sediment (approximately 4 mm) inhibit grazing and result in loss of attachment and death after a few days (Airoldi & Hawkins, 2007).

Sensitivity assessment.  Within this biotope, there is some water flow from tides and currents which can reach moderately strong (1.5 m/sec.) (Connor et al., 2004).  This water movement will allow the deposited sediment to be washed away relatively quickly.  However, sediment may still be present long enough to cause damage to the fronds of Fucus ceranoides and the other species within this biotope.  For this reason, resistance has been assessed as ‘Medium’, as has the resilience.  The sensitivity of this biotope is, therefore, also ‘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 effect 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

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

Introduction of light or shading

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

Evidence

Increased levels of diffuse irradiation correlate with increased growth in macroalgae (Aguilaria et al., 1999).  Levels of diffuse irradiation increase in summer, and with a decrease in latitude.  As Fucus ceranoides is found in the middle its natural range in the British Isles an increase or decrease in the level of diffuse irradiation will not cause a negative impact on the species or the biotope.

A number of experiments have demonstrated that Ulva lactuca, has high tolerance for shading and can survive periods of darkness.  Vermaat & Sand-Jensen (1987) found that Ulva lactuca, collected from Roskilde Fjord in Denmark in late autumn had extremely high shade tolerances.

Semibalanus balanoides sheltered from the sun grew bigger than un-shaded individuals (Hatton, 1938; cited in Wethey, 1984), although the effect may be due to indirect cooling effects rather than shading.  Barnacles are also frequently found under algal canopies suggesting that they are tolerant of shading. 

Since 2016, research on artificial light at night (ALAN) has expanded considerably in the marine and coastal environment. Light was previously assumed to be of low ecological significance in subtidal and intertidal habitats, but there is now evidence that ALAN is widespread in the marine environment, with biologically relevant levels of light penetrating to depths of up to 50m (Davies et al., 2020; Smyth et al., 2021). ALAN can alter biological processes across taxa and at multiple levels of organisation. Documented responses include disruption of diel and circalunar rhythms, changes in activity and foraging, altered predator–prey interactions, shifts in community composition, and impacts on algal growth and phenology (Davies et al., 2014, 2015; Gaston et al., 2017; Tidau et al., 2021; Lynn et al., 2022; Marangoni et al., 2022; Miller & Rice, 2023; Ferretti et al., 2025). Evidence for benthic habitats and assemblages specifically is beginning to emerge (e.g. Trethewy et al., 2023; Schaefer et al., 2025), but remains limited and fragmented, often focusing on single taxa or short-term experiments. Mortality thresholds, long-term consequences, and responses at the biotope scale are rarely addressed, and there are major gaps around indirect effects such as trophic cascades or habitat modification.

Sensitivity assessment. Given the rapid expansion of the evidence base but the continuing lack of data at the level of individual biotopes, resistance and resilience cannot be robustly assessed. Sensitivity is therefore recorded as Insufficient evidence.

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion (Barrier to species movement pressure definition).

Evidence

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

Death or injury by collision

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

Evidence

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

Visual disturbance

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

Evidence

Not relevant.

Genetic modification & translocation of indigenous species

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

Evidence

Key characterizing species within this biotope are not cultivated or translocated.  This pressure is, therefore, considered 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

There is insufficient evidence to support a review of the pressure at this benchmark.  No evidence is available on the microbes that are associated with the characterizing species Fucus ceranoides or Ulva sp.  Barnacles are considered subject to persistent, low levels of infection by pathogens and parasites.  But little specific information on the effects they have on their hosts is available.

Sensitivity assessment.  Due to the lack of information on the effect of microbial pathogens on the species found within  this biotope an assessment of 'No evidence' has been given. 

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

There is no evidence of target harvesting of the characterizing species, Fucus ceranoides.  Therefore, this pressure is assessed as ‘Not relevant’.

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 characterizing species Fucus ceranoides is a dominant species within this biotope.  The dominance of this characterizing species means it could easily be incidentally removed from this biotope as by-catch when other species are being targeted.  The loss of this species and other associated species would decrease species richness and negatively impact on the ecosystem function.

Sensitivity assessment. Incidental removal of a large percentage of the characterizing species would alter the character of the biotope. The resistance to removal is ‘Low’ due to the easy accessibility of the biotopes location and the inability of these species to evade collection. The resilience is ‘Medium’, with recovery only being able to begin when the harvesting pressure is removed altogether. This gives an overall sensitivity score of ‘Medium’.

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 Channel and into the Irish Sea coasts of Ireland and south Wales, with records in east and west Scotland, Northern Ireland, northwest France, Spain and south into the Mediterranean (NBN, 2023; OBIS, 2023).

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

Sensitivity assessment.

Biotopes characterised by dense macroalgae, such as this one, are unlikely to be suitable for Crepidula fornicata (Tillin et al., 2020). Sensitivity is assessed as 'Not sensitive' based on the existing evidence.

The carpet sea squirt, Didemnum vexillum

Evidence

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

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

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

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

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

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

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

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

In contrast to Didemnum vexillum’s preference for sheltered conditions, established colonies observed in Georges Bank and Long Island Sound were exposed to moderately strong tidal currents (1 to 2 knots; ca 0.5 to 1 m/s recorded at both sites) that may mobilise sediment (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020). However, Valentine et al. (2007b) describe the substratum as immobile, presumably consolidated gravel, cobbles and pebbles.

Kleeman (2009), stated that the presence of a consistent mild wave action or ‘swash zone’ appears to favour Didemnum sp. establishment in the intertidal. Although some evidence suggests that waves and currents can facilitate the fragmentation and spread of Didemnum vexillum (Mckenzie et al., 2017), the tidal current velocities at some sites where Didemnum vexillum has been reported (for example, New England, where current velocities reach up to around 3 m/s) is lower than the current velocity required for the dislodgement of Didemnum vexillum fragments (around 7.6 m/s) (Reinhardt et al., 2012). This suggests that not all tidal currents are likely to dislodge Didemnum vexillum fragments. When on boat hulls the species can experience higher current velocities which is enough to cause dislodgement (Reinhardt et al., 2012).  

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

Sensitivity assessment. Didemnum vexillum requires hard substrata for successful colonization, therefore, it could colonize the rock and mixed sediment substratum typical of this biotope. Also, Didemnum vexillum has a preference for wave sheltered conditions. Didemnum vexillum has been recorded in the lower intertidal but in the mid-shore examples of the biotope, the abundance and extent of colonies may be limited due to emersion. Didemnum vexillum colonies can survive exposure to air at low tides for a short time (not exceeding two hours) (Valentine et al., 2007a). However, wave splash may mitigate Didemnum vexillum decline in the mid-shore by providing moisture. There is evidence that Didemnum vexillum can overgrow and displace sessile organisms, including brown algae species Ascophyllum nodosum, Sargassum, Fucus evanescens and Fucus serratus. There is no evidence on how Didemnum vexillum affects fucoids or if it causes fucoid mortality. However, Didemnum vexillum competing for light and space with fucoids and epifauna could lead to mortality and a reduction in biodiversity and may interfere with the recruitment of characteristic species. Therefore, a resistance of 'Medium' (some mortality, <25%) is suggested as a precaution. Resilience is likely to be 'Very low' as Didemnum vexillum would need to be physically removed to allow recovery. Hence, sensitivity to invasion by Didemnum is assessed as 'Medium'. 

The Pacific oyster, Magallana gigas

Evidence

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

Since its introduction, the species has invaded and established self-sustaining natural populations throughout Europe from the North Sea, Wadden Sea and Scandinavian coastlines to the Atlantic coastlines of Spain and Portugal, as well as the Mediterranean and Adriatic Sea (Wrange et al., 2010; GBNNSIP, 2011, 2012; Ezgeta-Balic et al., 2019; Spagnolo et al., 2019; Bergstrom et al., 2021; Hansen et al., 2023). In the UK, the species predominantly occurs around the southern and western coastlines (OBIS, 2024; NBN, 2024). Shipping activity has also been associated with the introduction of Magallana gigas in the northeastern Adriatic Sea, where it was not introduced for aquaculture (Ezgeta-Balic et al., 2019) and possibly in southwest England from France possibly via fouling on ships (GBNNSIP, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).

Its distribution and environmental tolerances are considered to overlap with this biotope and, as a reef forming species, it can alter habitat structure. This species may also affect the grazers present in the biotope. No evidence was found for effects on rock pools, although on the Mediterranean coast Magallana gigas is cultivated in micro-tidal lagoons and has established wild populations (Miossec et al., 2009, cited from Herbert et al., 2012).  In the Wadden Sea and the North Sea, Magallana gigas overgrows mussel beds in the intertidal zone (Diederich, 2005, 2006; Kochmann et al., 2008), although larvae did show preference for settling on conspecifics before the mussels and struggled to settle on mussels with a fucoid covering.  It has been observed that mussel beds in the Wadden Sea that are adjacent to oyster farms were quickly converted to oyster beds (Kochmann et al., 2008).  Dense aggregations of Magallana gigas on a former mussel bed showed increased abundance and biomass of Littorina littorea in the Wadden Sea (Markert et al., 2010).  However, Eschweiler & Buschbaum (2011) found that juvenile Littorina littorea could carry Magallana gigas and Crepidula fornicata as epibionts. Body dry weight of snails without oyster overgrowth was twice as high compared to winkles covered with oysters. Crawling speed of snails with oyster epigrowth was significantly slowed down and about ten times lower than in unfouled periwinkles. Additionally, oyster epibionts caused a strong decrease in reproductive output. In laboratory experiments, egg production of fouled Littorina littorea was about 100-fold lower than in affected individuals. Field surveys in different years and habitats demonstrated that up to 10% of individuals occurring on epibenthic bivalve beds and up to 25% of snails living on sand flats may be fouled by Magallana gigas. Sargassum muticum is likely to have been imported to Europe in association with Magallana gigas imports (Engelen et al., 2015).  Sargassum muticum can grow on shells and pebbles and Magallana gigas shells would provide a suitable habitat. .

Sensitivity assessment. Magallana gigas populations may be limited to low densities due to very wave exposed to wave exposed conditions (Teschke et al., 2020). However, 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. Therefore, this biotope is probably unsuitable for colonization by the Pacific oyster, and sensitivity is assessed as 'Not sensitive' based on existing evidence.

Wireweed, Sargassum muticum

Evidence

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°C, +2.0°C, +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 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 from Fucus serratus studies indicates that competitive outcomes are environmentally contingent. Under warm summer conditions, Fucus serratus achieved greater biomass gain than Sargassum muticum, and competitive effects were comparable to intraspecific competition (Armitage et al., 2017). In contrast, simulated marine heatwave conditions enhanced the performance of Sargassum muticum while negatively affecting Fucus serratus (Atkinson et al., 2020). While these findings suggest the potential for altered competitive balance under elevated temperature events, there is no direct evidence of decline, displacement, or altered recovery of Fucus ceranoides within this biotope. Therefore, resistance and resilience cannot be confidently assessed and sensitivity is recorded as ‘Insufficient evidence’.

Wakame, Undaria pinnatifida

Evidence

Thompson & Schiel (2012) found that native fucoids show high resistance to invasions by Undaria pinnatifida. In the absence of evidence for significant displacement or decline of Fucus serratus within this biotope, resistance is likely to be ‘High’. As resistance is high, resilience is not considered further, and the biotope is assessed as ‘Not sensitive’ to this pressure.

Other INIS

Evidence

There are no known invasive non-native species in the British Isles that invade reduced salinity intertidal rocky shores where this biotope is found.  The Australasian barnacle Austrominius (syn. Elminius) modestus was introduced to British waters on ships during the Second World War.  Austrominius modestus is now described as a species that naturally occurs within this biotope.  Its overall effect on the dynamics of rocky shores has been small, as Austrominius modestus has replaced some individuals of a group of co-occurring barnacles (Raffaelli & Hawkins, 1999).

Sensitivity assessment.  Fucoid species have been negatively affected by both the direct and indirect consequences of INIS being present.  However, no evidence can be found on the impacts of INIS on Fucus ceranoides within this biotope.  Therefore, the effect of this pressure has been given as ‘Insufficient evidence’. Literature for this pressure should be revisited.