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 north-east 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°C, +2.0°C 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) did report that germlings were negatively affected by increased temperature, indicating that early life stages are more vulnerable than mature algae to this pressure.

Strong tidal currents, characteristic of this biotope, encourage communities of sponges and ascidians. Changes to the hydrological regime are therefore likely to directly influence the presence of these species. Once removed, however, these species are likely to rapidly recolonize via planktonic larvae, thereby facilitating recruitment. Most species associated with this biotope are poor long-distance dispersers. However, the moderately strong tidal currents of this biotope enable these species to disperse over greater distances than in slow-flowing environments.

Sensitivity assessment. In the British Isles, Fucus serratus occurs centrally within its geographic range, and temperature does not appear to limit its present distribution (Yesson et al., 2015). Experimental studies indicate that adult thalli tolerate temperatures at or above typical UK summer maxima, with no reduction in growth reported at 22°C and short-term survival at 25°C (Arrontes, 1993; Nielsen et al., 2014). Field measurements demonstrate substantial thermal buffering within intact canopies (Migné et al., 2021), and short-term heatwave simulations indicate interactions among fronds that reduce physiological stress under elevated temperatures (Harris et al., 2025).

The adverse effects documented in Iberia concern populations at the southern range edge that are already near upper thermal limits and frequently experience interacting stressors such as extreme emersion heat, reduced salinity or chronic warming. While laboratory and mesocosm experiments show reductions in growth, photosynthetic performance and carbon balance under elevated temperatures (e.g. Atkinson et al., 2020; Díaz-Acosta et al., 2021), these effects generally occur at temperatures exceeding typical UK baseline conditions or under multi-stressor scenarios. Germlings may be more sensitive than adults (Nielsen et al., 2014), but there is no evidence that the benchmark-level increase would result in widespread adult mortality or loss of canopy structure within UK populations.

Given the central geographic position of UK populations, evidence of adult tolerance above current summer maxima, and absence of documented canopy loss in response to temperature increases within the benchmark magnitude and duration, resistance is assessed as High. As no significant mortality or structural loss of the biotope is predicted at the benchmark level, resilience is assessed as High (no recovery required). The biotope is therefore assessed as Not Sensitive to this pressure at the 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

Lüning (1984) reported that Fucus serratus survived in the laboratory for a week a range temperature 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'.

Sensitivity assessment. A decrease in acute or chronic temperature above average British and Irish temperatures is not likely to have a detrimental effect of Fucus serratus and associated communities, based on global distribution. However, it should be noted that phenotypic plasticity will influence the tolerance of individual population. Resistance and resilience are therefore both assessed as ‘High’ (no impacts to recover from) and the biotope is assessed as ‘Not Sensitive’ to a change 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

This biotope group is found in the intertidal zone and is, therefore, likely to experience cyclical periods of hypo- and hyper-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).

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.

Sensitivity assessment. Fucus serratus occurs in fully marine conditions and is regularly exposed to short-term fluctuations in salinity in fjords and estuarine-influenced coasts. Experimental evidence indicates tolerance across a broad salinity range, with maximal growth at 20 PSU and recruitment possible above 7 PSU (Malm et al., 2001). Field observations suggest stable or increased biomass under relatively high salinity conditions (Kolbeeva et al., 2025), and species distribution modelling at the southern range edge indicates greater probability of occurrence at salinities above 33 PSU (Viejo et al., 2024).

However, no information on the effects of hypersaline effluents on fucoid communities was found. While intertidal communities are probably tolerant of short-term variation in salinity, especially at low tide, the effects of exposure to hypersaline effluent (>40) are unclear.  Hence, the current evidence is insufficient to form the basis of an assessment.

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

This biotope group is found in the intertidal zone and is therefore likely to experience cyclical periods of hypo- and hyper-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 distribution edge 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.

Sensitivity assessment. Fucus serratus commonly inhabits narrow fjords where salinity can vary widely along a spatial (km) and/or temporal (hours to daily) scale. A reduction in salinity at the level benchmark (e.g. from 'Full' to 'Reduced'  for one year) could have beneficial effects on Fucus serratus as growth rates are maximal below full saline conditions (Malm et al., 2001). UK populations normally experience salinities well above thresholds causing reproductive or germling failure (Ardehed et al., 2016; Kinnby et al., 2019), and most other biotope characterizing species are also likely to tolerate minor decreases. While extreme low salinity (<5 psu) or combined stressors (e.g., high temperature) can negatively affect survival and early life stages (Pedersen et al., 2025; Viejo et al., 2024), a decrease of one MNCR salinity category remains within tolerable limits for the biotope. Resistance and resilience are therefore both assessed as ‘High’ (no impacts to recover from). The biotope is assessed as ‘Not Sensitive’ to a decrease in salinity at the pressure benchmark. 

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s and 0.2 m/s for more than one year (Water flow pressure definition). 

Evidence

The rich community of suspension feeders in this biotope is, in part, due to the strong tidal streams with which it is associated. Strong currents provide suspension feeders with a continual supply of food and remove sediment that would otherwise interfere with their feeding apparatus. A decrease in water flow rate could lead to siltation, to the detriment of filter feeders. Furthermore, grazers unable to cope with the strong flow rates normally associated with this biotope may be able to graze more efficiently, increasing herbivory pressure.  High water flow rates increase mechanical stress on macroalgae by increasing drag. This can result in individuals being torn off the substratum. Once removed, the attachment cannot be reformed, causing the death of the algae. Any sessile organisms attached to the algae are also lost. Fucoids are, however, highly flexible and are able to reorient their position in the water column to become more streamlined. By going with the flow, fucoids can reduce the relative velocity between algae and the surrounding water, thereby reducing drag and lift (Denny et al., 1998). Propagule dispersal, fertilization, settlement, and recruitment are also influenced by water movement (Pearson & Brawley, 1996). In addition, increased water flow will cause scour through increased sediment movement, affecting in particular small life stages of macroalgae by removing new recruits from the substratum and hence reducing successful recruitment (Devinny & Volse, 1978) (see ‘siltation’ pressures).  Changes in water motion can thus strongly influence local distribution patterns of Fucus spp. (Ladah et al., 2008). Increases in drag can, however, be counterbalanced in the long-term by changes in morphology resulting in structurally more resistant thalli and holdfasts (Haring et al., 2002).

An increase in water flow rate could also adversely affect this biotope. The mixed nature of the substratum means that suspended sediment of various sizes will be re-suspended, and cobbles and pebbles are likely to move across the seabed. The movement of pebbles, cobbles and boulders across the sea floor may scour epilithic sponges, bryozoans and hydroids off the rocks and clumps of algae may become dislodged. Re-suspended sediments are likely to interfere with the suspension feeders' feeding apparatus and respiratory currents to their detriment (see ‘suspended sediment’ pressure). In wave sheltered areas where this biotope occurs, the breadcrumb sponge Halichondria panicea grows in massive forms. Poorly attached massive forms may be ripped off by an increase in water flow rate, leading to the death of large colonies.

Sensitivity assessment.  Moderately strong tidal flow, characteristic of this biotope, encourages communities of sponges and ascidians. Changes to the hydrological regime are therefore likely to directly influence the presence of these species. As the biotope occurs in moderately strong tidal flow (0.5 to 1.5 m/s), an increase in water flow may result in disturbance or mobilisation of the hard substrata (i.e. cobbles and pebbles), resulting in abrasion or loss of attached epilfora and epifauna. A reduction in water flow is likely to result in a loss of the suspension feeding species, an overall reduction in species richness, and result in loss of this biotope, as it is replaced by another Fucus serratus dominated biotope, e.g. LR.LLR.F.FserrX.  However,  a change of 0.1 to 0.2 m/s(the benchmark) is unlikely to adversely affect the biotope, although a reduction may decrease feeding and hence growth rates.  Therefore, a resistance of 'High' is recorded, with a resilience of 'High' and a sensitivity of 'Not sensitive' at the benchmark level. 

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

This biotope is found in the eulittoral zone from MTL (mean tide level) to MLWN (mean low water neap) and is, therefore, subjected to cyclical immersion and emersion over the tidal cycle. 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. in 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). Germlings are, however, 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).

Mathieson & Burns (1971) measured the photosynthetic rate of Chondrus crispus at varying degrees of desiccation and found that after loss of 65% of its water content, the rate of photosynthesis dropped to 55% of the control rate. In Palmaria palmata, 50% of the plant's water content can be lost in less than 4 hours in dry air at 25°C (Kain & Norton, 1990). This scenario can reasonably be expected at low tide in summer in Britain, although the Fucus canopy is likely to protect the underlying red algae to some extent.

A decrease in submergence is likely to adversely affect the suspension feeder population by reducing feeding opportunities, as immersion is a prerequisite of feeding. This can prove fatal for short-lived species such as bryozoans and ascidians. The tissue of Halichondria panicea holds some water and can tolerate a certain degree of desiccation. On the other hand, the soft-bodied sea squirt Ascidiella scabra has a greater vulnerability to this pressure. The sea squirt is commonly found in damp crevices or under the canopy of macroalgae, offering protection from desiccation, but individuals at the highest point on the shore may dry out and die at the benchmark level.

On the other hand, an increase in submergence is likely to benefit this biotope. Feeding opportunity for suspension feeders will increase; desiccation and temperature stresses for all flora and fauna will decrease, as will predation from birds. The biotope may extend further up the shore, but this extension is likely to be counteracted by a reduction in the lower shore extent of the biotope, likely to be taken over by seaweeds more characteristic of the sublittoral fringe. Furthermore, predation by the common shore crab Carcinus maenas is likely to increase.

Sensitivity assessment. Severe desiccation and associated osmotic stress can increase mortality (Perason et al., 2009). Other species better able to tolerate desiccation will competitively displace Fucus serratus following changes in the emergence regime. Juvenile stages are more susceptible but are largely protected from desiccation by the canopy of adults. Resistance is thus assessed as ‘Low’, as most of the Fucus serratus cover will probably be lost or replaced by other species. Resilience is thus assessed as Medium and sensitivity as ‘Medium’ at the level of the benchmark.

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year (Wave action pressure definition). 

Evidence

This biotope is limited to coasts with moderate exposure or less. Fucus serratus is highly flexible but not physically robust, and an increase in wave exposure above this level will cause mechanical damage, break fronds or even dislodge algae from the substratum. Fucoids are permanently attached to the substratum and would not be able to re-attach if removed. Organisms living on the fronds and holdfasts will be washed away with the algae, whereas free-living community components could find new habitat in the surrounding areas. 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, 1992; Serrão et al., 1996; Berndt et al., 2002), indicating that changes in wave exposure could also affect recruitment even if adult thalli remain intact.

An increase in wave action would lead to the dominance of the community by grazers and barnacles at the expense of fucoids. Increased wave action may also reduce light penetration, thereby lowering overall photosynthesis (see ‘changes in suspended solids’ pressure). A reduction in wave action would have little effect as the species is naturally found in sheltered conditions.

The biotope is found in wave sheltered to extremely sheltered habitats. In these locations, the breadcrumb sponge Halichondria panicea grows in massive forms. Poorly attached massive forms may be ripped off by an increase in water flow rate, leading to the death of large colonies. 

Sensitivity assessment. Fucus serratus and associated communities are sensitive to an increase in wave action, as increased exposure would result in important losses both in biomass and species richness. However, a 3-5% change in significant wave height is unlikely to be significant, and the biotope is considered to be 'Not sensitive' 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.

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

Sustained reduction of dissolved oxygen can lead to hypoxic (reduced dissolved oxygen) and anoxic (extremely low or no dissolved oxygen) conditions. Sustained or repeated episodes of reduced dissolved oxygen have the potential to severely degrade an ecosystem (Cole et al., 1999). Josefson & Widbom (1988) investigated the response of benthic macro and meiofauna to reduced dissolved oxygen levels in the bottom waters of a fjord in Sweden. At dissolved oxygen concentrations of 0.21 mg/l, the macrofaunal community was eradicated and was not fully re-established 18 months after the hypoxic event. Meiofauna seemed, however, unaffected by deoxygenation.

Sensitivity assessment. Macroalgae may be negatively impacted by reduced dissolved oxygen level at the level of the benchmark (2 mg/l for 1 week) resulting in direct mortalities.  However, the strong water movement in tide-swept tidal currents combined with turbulent flow over rocks would aerate the water column, and emersion at low tide would mean that any oxygen depletion was transient. Therefore, resistance is assessed as ‘High’. Hence, resilience is assessed as ‘High’, and the biotope as '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

Nutrient enrichment generally stimulates ephemeral macroalgae growth (Duarte, 1995). This stimulation of annual ephemerals may accentuate the competition for light and space and hinder perennial species development or harm their recruitment (Kraufvelin et al., 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 indicate that established rocky shore communities of perennial algae with associated fauna are able to persist for several years, even at very high nutrient levels, but that community shifts may suddenly occur if eutrophication continues. They also indicate 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, 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 sources 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.

Johansson et al. (1998) investigated the changes in the algal vegetation of the Swedish Skagerrak coast, an area heavily affected by eutrophication, between 1960 and 1997. Slow-growing species, including Chondrus crispus, declined in abundance, probably due to competition from faster-growing red algal species such as Phycodrys rubens and Delesseria sanguinea. However, this biotope occurs in areas with moderately strong to very strong tidal currents, rapidly renewing depleted oxygen levels (see de-oxygenation pressure).  

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, at the level of the benchmark, both resistance and resilience are assessed as ‘High’. The biotope group is therefore ‘Not Sensitive’ to this pressure at the pressure benchmark.

Organic enrichment

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

Evidence

Organic enrichment can stimulate the production of primary consumers and may lead to eutrophication (see ‘nutrient enrichment’ pressure). Husa et al. (2014) found that the macroalgal communities beyond the immediate proximity of fish farms in Hardangerfjord, Norway, seemed to be little affected by the deposition of organic matter from the salmon farming industry. Bellgrove et al. (2010) however determined that coralline turfs out-competed fucoids at a site associated with organic enrichment caused by an ocean sewage outfall.

Sensitivity assessment. At the level of the benchmark, resistance is assessed as ‘Medium’ as some mortalities are likely to occur. Recovery will be rapid resulting in ‘High’ resilience score. The biotope has thus a ‘Low’ sensitivity to organic enrichment at the level of the benchmark. 

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 rock substratum. A change towards a sedimentary substratum would lead to the direct loss of suitable attachment areas, resulting in the loss of Fucus serratus and associated communities. 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 habitats provide a less stable attachment surface for adult individuals and may reduce reproductive output. This evidence supports the importance of a stable hard substratum for the maintenance of mature canopy-forming populations. Resistance is assessed as ‘None’. As this pressure represents a permanent change, recovery is impossible as a suitable substratum for fucoids is lacking. Consequently, resilience is assessed as ‘Very low’.  Therefore, the habitat is assessed as 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

This biotope group occurs on hard substratum (boulders, pebbles and cobbles) on muddy sediment. A change towards a soft sediment and loss of associated mixed hard substrata (boulders, pebbles and cobbles)  would lead to the direct loss of suitable attachment areas, resulting in the loss of Fucus serratus, Chondrus crispus and other red seaweeds. The loss of macroalgae will result in the loss of habitat for associated sponge and ascidian communities. Resistance is assessed as ‘Low’. As this pressure represents a permanent change, recovery is impossible as a suitable substratum for fucoids is lacking. Consequently, resilience is assessed as ‘Very low’.  The habitat, therefore, scores a ‘High’ sensitivity. Although no specific evidence is described, confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.  

Habitat structure changes - removal of substratum (extraction)

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

Evidence

Extraction of sediment to 30 cm could remove the underlying sediment but also remove a proportion of the boulders, cobbles and pebbles on which the biotope develops. As a result, the biotope could be lost or severely damaged, depending on the scale of the activity (see abrasion). Therefore, a resistance of 'None' is suggested. Resilience is probably 'Low', therefore the biotope is likely to have a 'High' sensitivity to this pressure. 

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

The biotope group is found in the lower 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 (see Tyler-Walters & Arnold, 2008 for review).  

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 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 utilized site.  Comparatively, the heavily utilized 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 and Frid (1996a; 1996b) examined the effects of persistent trampling on two sites on the north east coast of England. The trampling treatments used were 0, 20, 80, and 160 steps per m² per spring tide for 8 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 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 4-8% cover. However, after trampling ceased, recovery of algal cover, including Iridaea cornucopiae and Mastocarpus papillatus was rapid (ca 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 where 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 substratum temperatures were on average 8-10°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 macroalgae canopy due to abrasion will thus 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 have been found to be 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 thus 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. In addition, any potential abrasive activity that drags or moves across the substratum could overturn the boulders, cobbles and pebbles on which this biotope occurs, or even remove them from the area, resulting in significant damage or loss of the biotope. 

Sensitivity assessment. Physical disturbance resulting from activities such as trampling (by humans and livestock) or abrasive activities (e.g. vehicles, jack-up-barges, or grounding vessels) could cause a significant loss of fucoid cover and an important reduction in species abundance and diversity. Resistance is thus assessed as ‘Low’. If some of the Fucus serratus population remains, recovery will be fairly rapid. However, recruitment mortality, grazing by limpets and the presence of turfs and encrusting algae can slow down and limit recovery. Resilience is thus assessed as ‘Medium’. The biotope, therefore, scores a ‘Medium’ sensitivity to abrasion pressure. If the entire population of Fucus serratus is removed, other species may come to dominate, and the recovery will take considerably longer. Re-establishment of the seaweed may depend on the ability to out-compete other species, and this may be dependent on suitable environmental conditions.

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

Penetrative activities (e.g. anchoring, scallop or suction dredging) are likely to remove or displace the cobbles, pebbles, or small boulders on which this biotope develops. As a result, the biotope could be lost or severely damaged, depending on the scale of the activity (see abrasion above). Therefore, a resistance of 'None' is suggested. Resilience is probably 'Low', therefore the biotope is likely to have a 'High' 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

Light is an essential resource for all photoautotrophic organisms. Changes in suspended solids affecting water clarity will have a direct impact on the photosynthesising capabilities of Fucus serratus. Irradiance below the light compensation point of photosynthetic species can compromise carbon accumulation (Middelboe et al., 2006). However, water clarity is only relevant when the biotope is covered with water. Seaweed photosynthesis declines on emersion and recommences when recovered with water. In addition, 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). Red algae can tolerate a wider range of light levels than any other group of photosynthetic plants (Kain & Norton, 1990) and will therefore be less affected by a reduction in water clarity. 

In turbid waters, the feeding apparatus of suspension feeders may become clogged with particles, interfering with their feeding and respiratory currents, resulting in net losses. For instance, the hydroid Dynamena pumila experienced a marked decline in areas with increased silt content in Strangford Lough, Northern Ireland (Seed et al., 1983). Some filter feeders have the ability to cope with siltation and excess suspended material. For example, the sea squirt  Ascidiella scabra can extend its siphons to a small extent and can maintain a passage through the silt to the siphons. However, Robbins (1985b) found that increased inorganic particulate concentrations reduced growth rates of Ascidiella scabra, with mortalities occurring at high levels of suspended sediments. The breadcrumb sponge Halichondria panicea has a cleaning mechanism sloughing off its complete outer tissue layer together with any debris (Barthel & Wolfrath, 1989). However, there is an energetic cost in cleaning, resulting in reduced growth. For short-lived species, such as the star ascidian Botryllus schlosseri, reduced growth could prove fatal.

Sensitivity assessment. Changes in suspended solids, reducing water clarity, will have adverse effects on the biotope group, reducing species richness. Resistance is thus assessed as ‘Medium’. Once conditions return to 'normal', Fucus serratus is likely to rapidly regain photosynthesising capabilities as well as growth rate. Associated communities will also rapidly recover as most of the intolerant species produce planktonic larvae and are therefore likely to be able to recolonize quickly from surrounding areas. Resilience is thus assessed as ‘High’. Overall, this biotope is assessed as ‘Low’ sensitivity to this pressure.

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

Light is an essential resource for all photoautotrophic organisms. Changes in suspended solids affecting water clarity will have a direct impact on the photosynthesising capabilities of Fucus serratus. Irradiance below the light compensation point of photosynthetic species can compromise carbon accumulation (Middelboe et al., 2006). However, water clarity is only relevant when the biotope is covered with water. Seaweed photosynthesis declines on emersion and recommences when recovered with water. In addition, 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). Red algae can tolerate a wider range of light levels than any other group of photosynthetic plants (Kain & Norton, 1990) and will therefore be less affected by a reduction in water clarity. 

In turbid waters, the feeding apparatus of suspension feeders may become clogged with particles, interfering with their feeding and respiratory currents, resulting in net losses. For instance, the hydroid Dynamena pumila experienced a marked decline in areas with increased silt content in Strangford Lough, Northern Ireland (Seed et al., 1983). Some filter feeders have the ability to cope with siltation and excess suspended material. For example, the sea squirt  Ascidiella scabra can extend its siphons to a small extent and can maintain a passage through the silt to the siphons. However, Robbins (1985b) found that increased inorganic particulate concentrations reduced growth rates of Ascidiella scabra, with mortalities occurring at high levels of suspended sediments. The breadcrumb sponge Halichondria panicea has a cleaning mechanism sloughing off its complete outer tissue layer together with any debris (Barthel & Wolfrath, 1989). There is, however, an energetic cost in cleaning, resulting in reduced growth. For short-lived species, such as the star ascidian Botryllus schlosseri, reduced growth could prove fatal.

Sensitivity assessment. Changes in suspended solids, reducing water clarity, will have adverse effects on the biotope group, reducing species richness. However, a deposit of 5 cm of fine sediment (the benchmark) is likely to be removed within a few tidal cycles in the tide-swept conditions experienced in this biotope. Hence, resistance is assessed as 'High', resilience as 'High', and sensitvity as 'Not sensitive' at the benchmark level. 

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

Several studies found that increasing the vertical sediment burden negatively impacts fucoids' survival and associated communities (see above). At the level of the benchmark (30 cm of fine material added to the seabed in a single event), smothering could result in significant mortalities, especially in sheltered examples of the biotope where the sediment burden could remain for many tidal cycles. However, the tide-swept conditions may remove the deposited sediment more rapidly. Resistance is assessed as ‘Low’ as some individuals exposed to siltation at the benchmark level may die, albeit with 'Low' confidence due to the lack of direct evidence. Hence, resilience is probably ‘Medium’ and sensitivity is assessed 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 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 biotope 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

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 insolation 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., 1999). 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. Because fucoid recruitment depends on synchronised gamete release and suitable settlement conditions, disruption to photoperiod cues has the potential to affect reproductive success even where adult thalli appear physiologically tolerant. 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.

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. 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 group.

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 the Southwest of Britain. More recently, Zuccaro et al. (2008) detected a number of fungal species associated with Fucus serratus. So far no mortalities have been associated to 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). Other characteristic species, for example Chondrus crispus and Mytilus edulis are known to be adversely affected by infestation by microbial pathogens (see relevant MarLIN reviews). However, even if microbial infestation resulted in the loss of these two species from the biotope, the recognizable biotope per se would not be affected.

Sensitivity assessment. Both resistance and resilience are assessed as ‘High’; the biotope is therefore ‘Not Sensitive’ to this pressure. However the assessment has a low confidence score as more research is needed into the effects of microbial pathogen on Fucus serratus and associated communities. 

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, most of them being sessile organisms.

Migné et al. (2015) experimentally removed the canopy of Fucus serratus (cut with a knife and regrowth repeatedly removed) over an 18-month period on a mid to low rocky shore in the south-west 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 thus 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.

Sensitivity assessment. Removal of the Fucus serratus canopy will have a negative impact on the diversity of the animal community and the metabolism of the area. The harvesting impact on the animal community was amplified by the settlement of an ephemeral canopy of Ulva spp., a seasonal opportunistic green alga (ref). Resistance is thus assessed as ‘Low’. If some Fucus serratus population remains, recovery will be fairly rapid. However, recruitment mortality, grazing by limpets and the presence of turfs and encrusting algae can slow down and limit recovery. A switch to a disturbance community will also slow the recovery of Fucus serratus and the associated community. Experimental cultivation studies demonstrate that Fucus serratus can exhibit substantial seasonal growth and tolerate repeated biomass removal under controlled conditions (Meichssner et al., 2021), indicating high intrinsic regrowth capacity; however, recovery in natural assemblages remains constrained by grazing, competition and recruitment dynamics. Resilience is thus assessed as ‘Medium’. Therefore, the biotope is assessed as a ‘Medium’ sensitivity to this pressure.

If the entire population of Fucus serratus is removed, other species may come to dominate, and the recovery will take considerably longer. Re-establishment of the seaweed may depend on the ability to out-compete other species, dependent on suitable environmental conditions.

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

The harvest of algae, crabs, snails, mussels, and some species of fish from the shore is a widespread practice. None of the components of this biotope have known obligate relationships, and the removal of non-target species will therefore not have a significant impact.  Resistance to this pressure is deemed ‘High’.  Resilience is also ‘High’ as there are no ecological impacts to recover from, resulting in a ‘Not Sensitive’ score. The assessment is based on expert opinion, resulting in a 'Low' confidence score.

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

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).

Abundances at its northern and southern extremes may be low, but densities in the UK and France are often over 1000/m², and it may carpet the seafloor in the Solent and Essex. In the UK, it was reported to reach abundances of >1000/m² (max. 2,748/ m²) in the Milford Harbour Waterway (MHW) (Bohn et al., 2012), 84 /m² in Portsmouth, 174/m2 in Langstone and 306/ m² in Chichester harbours in 2017 (Helmer et al., 2019). In France, it has been reported to reach >4,700/m2 in the Bay of Marennes-Oleron, 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).

It is recorded from the lower intertidal to ca 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, 2023; Tillin et al., 2020). Bohn et al. (2012, 2013a, 2013b, 2015) suggested that extreme conditions in the intertidal limited its upward distribution due to early post-settlement mortality. It reached its highest densities in the lower shore (below ca 0.7 m) and was absent from high tidal level (ca 1.8 m) in the MHW (Bohn et al., 2015). Bohn et al. (2013b) noted that Crepidula spat in their experimental intertidal panels suffered high mortality, 78 to 100%, during emersion by low water spring tides. Thieltges et al. (2003) noted that Crepidula abundance at the intertidal to subtidal transition zone (ca 21 / m²) was significantly higher than in the upper, mid, and lower intertidal ca <3 / m²). Similarly, Diederich & Pechenik (2013) noted that Crepidula densities were not significantly different in the low intertidal (+0.2 m) and shallow subtidal (-1 m) but became lower at +0.4 and were absent above +0.6 m in Bissel Cove, Rhode Island, where the mean high water was +1.38 m. Diederich et al. (2015) reported that most juvenile Crepidula died after aerial exposure under laboratory conditions (20°C, 75% relative humidity), while adults from the intertidal and subtidal survived (26°C, 75% relative humidity).  Bohn et al. (2015) suggested that wave action (exposure) probably prevented the establishment of large numbers of Crepidula in high-energy areas.

Sensitivity assessment. Crepidula prefers wave sheltered habitats, so the moderately exposed examples of this biotope are probably unsuitable. In addition, Tillin et al. (2020) suggested that the sweeping action of the fronds of macroalgae may prevent recruitment in Crepidula, based on its low abundance from Zostera beds in Arcachon Bay, France. Therefore, this biotope is probably unsuitable for colonization by Crepidula or limited to small numbers of individuals. The biotope is assessed as ‘Not sensitive’ (resistance and resilience are ‘High’) but with ‘Low’ confidence based on limited 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; Minchin & Nunn, 2013; Bishop et al., 2015; McKenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

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

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

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

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

Didemnum vexillum has a seasonal growth cycle 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 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 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., 2007a & b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012; Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent, and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020).

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

In contrast 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 are 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 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 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). It was also suggested that some Magallana gigas populations were established in southwest England from France, possibly via fouling on ships (GBNNSIP, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).

Magallana gigas 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). Larval mortality can be as large as 99%, as larvae are sensitive to environmental conditions (Alves et al., 2021). But adults are long-lived so that populations can survive with infrequent recruitment (Padilla, 2010). Larval dispersal and mass spawning events have facilitated the settlement and establishment of Pacific oysters, as seen in the Oosterschelde estuary, Netherlands (Hansen et al., 2023). It has been suggested that the spread of the Pacific oyster in Scandinavia is due to northward larval drift on tidal and wind-driven currents (Hansen et al., 2023). Wood et al. (2021) suggested that larval dispersal of the Pacific oyster from populations within and outside the UK was possible via unaided (passive) transport by currents, but that aquaculture and offshore structures (e.g. windfarms) increased the risk of the invasive species spreading and the geographical extent of spread.

Magallana gigas is considered an intertidal and shallow subtidal species rarely recorded below extreme low water (Herbert et al., 2012, 2016; Tillin et al., 2020). On littoral rock in Brittany, the Pacific oyster colonizes all intertidal levels from Mean High Water to Mean Low Water on sheltered (low energy), moderately exposed (moderate energy) and high energy rock shores (Herbert et al., 2012). However, in the northwest Pacific, Magallana gigas is commonly found on sheltered low energy littoral rock and has less than 10% cover on exposed high energy littoral rock shores (Herbert et al., 2012, 2016). Magallana gigas has not been found at extreme low water levels or subtidally beneath rocky habitats, as it has been in soft sediment areas (Herbert et al., 2012).

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 rocks underneath, and macroalgae fronds inhibit settlement and recruitment by exuding metabolites.

Sensitivity assessment. The evidence above suggests that competition with macroalgae can exclude Magallana from rocky shores (Kochmann et al., 2012, 2013; McKinstry & Jensen, 2013; Cook et al., 2014; Tillin et al., 2020). While a few individuals may occur in gaps, it is probably unable to form beds, and the habitat is unlikely to be adversely affected. Therefore, resistance is assessed as ‘High’, resilience as ‘High’ and sensitivity as ‘Not sensitive’.

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 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. 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 serratus within this biotope. Therefore, resistance and resilience cannot be confidently assessed, and sensitivity is recorded as ‘Insufficient evidence’.

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 out-compete 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, 2018; Kraan, 2017; Epstein et al., 2019a, b; Tidbury, 2020). Undaria pinnatifida originates from Japan but is currently established on the coastlines of New Zealand, Australia, Northern France, Spain, Italy, the UK, Portugal, Belgium, 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. One year after its discovery at the Queen Anne Battery marina, Plymouth, it had become a major fouling plant on pontoons (Minchin & Nunn, 2014). 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 predominantly found in low intertidal to shallow subtidal habitats (between -1 and 4 m depth) (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 can out-compete native species on artificial substrata (such as marina and wharf structures). De Leij et al. (2017) suggested that in natural substrata, Undaria pinnatifida can be inhibited by the presence of native competitors, such as large perennial species. The species behaves as a winter annual, with recruitment occurring 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).

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. Hence, only the lowest extent of the sheltered examples of these biotopes is likely to be suitable for colonization. In addition, the evidence from New Zealand suggests that dense fucoid canopies can exclude Undaria.  Therefore, in the absence of evidence for significant displacement or decline of Fucus serratus due to invasion by Undaria, resistance is assessed as ‘High’, resilience as ‘High and the biotope is assessed as ‘Not sensitive’ to this pressure. However, the confidence is ‘Low’ due to the limited available evidence.

Other INIS

Evidence

No evidence was found