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

Saccharina latissimahas a latitudinal range of 41.3 degrees South to 79.8 degrees North, depth limits of 2.5 to 30 m, a thermal limit of -1.8 to 21.5°C, and there is little coastline left for poleward range expansion in the northwest Atlantic (Khan et al., 2018). The temperature isotherm of 19 to 20°C has been reported as limiting Saccharina latissima geographic distribution (Müller et al., 2009). The southernmost limit of this species in Europe is northern Portugal (Kerrison et al., 2015; Azevedo et al., 2016), and it grows well between 5 and 17°C (Druehl, 1967, Fortes & Luning, 1980 and Machalek, Davison & Falkowski, 1996 cited in Kerrison et al., 2015). At the high end of this temperature range, net photosynthesis declines and acclimation effort increases, involving the upregulation of many temperature-responsive genes (Davison, 1991 and Heinrich et al., 2012 cited in Kerrison et al., 2015). Tissue loss or death is commonly reported for this species above 17 to 20°C (Gerard & Du Bois, 1988; Gerard, Dubois & Greene, 1987 cited in Kerrison et al., 2015).

Temperature ecotypes may exist which have adapted to high seasonal temperature exposure. For example, populations from Helgoland, Germany, can tolerate temperatures of 18 to 20°C (Davison, 1987 cited in Kerrison et al., 2015), while populations in New York, USA, can survive at >20°C, albeit with substantially reduced growth (Gerard & Du Bois, 1988). In addition, Azevedo et al. (2016) cultured Saccharina latissima in tanks in northwest Portugal throughout the summer, withstanding average temperatures around 20°C from May onwards (temperature varied between 11.7°C in April and 24.9°C in August), well above published optimum temperatures for this species (10 to 15°C). Biomass increased until the third week of May, and afterwards it remained constant until the beginning of July, reaching a density of 13 kg/m ³ (Azevedo et al., 2016). This observation may be explained by their origin in populations located near the southern distribution boundary, which may have acquired adaptations that increased tolerance to high temperatures (Azevedo et al., 2016). In addition, colder water populations of Saccharina latissima may benefit from higher temperatures, with individuals during a warming experiment from Kongsfjorden, Svalbard, having a significantly higher growth rate in the warmer treatments (9°C) compared to colder treatments (4°C) (Iñiguez et al., 2016b).

The distribution of kelp is strongly influenced by climatic conditions; therefore, kelp species are extremely sensitive to the ongoing ocean warming (Kain, 1979; Van Den Hoek, 1982; Breeman, 1990; Lüning, 1990; Assis et al., 2016; Smale, 2020). Northern distribution boundaries are set by winter temperatures that are lethal, or summer temperatures too low for growth and/or reproduction, while southern limits are set by high lethal summer temperatures or winter temperatures too high for induction of a crucial step in the life cycle (Breeman, 1990). Kelps have a high dependence on ocean temperatures, which makes them highly vulnerable to ocean warming (Assis et al., 2014). As temperatures increase, populations found towards the upper limit of their temperature range may be adversely affected by warming as physiological thresholds are exceeded (Wiens, 2016). Thermal stress can lead to mortality and consequent population-level effects, such as decreased abundance, altered size structure, local extinction and range contractions (Smale, 2020).

Climate change is projected to increase the average sea surface temperature by between 1 and 3°C over the 21^st century and is predicted to cause the northward retreat of kelps (Solomon et al., 2007, Méléder et al., 2010 and Raybaud et al., 2013 cited in Kerrison et al., 2015).

Yesson et al. (2015b) examined the change in abundance of large brown seaweeds in the British Isles between 1974 and 2010. They found that for all sites, Saccharina latissima showed a negative trend in abundance. Regional Sea Surface Temperatures showed annual fluctuations between 1974 and 2010, and the general trend has been a 1 to 2°C increase during this time-period, with the East coast (North Sea) experiencing the greatest increases (Yesson et al., 2015b). In addition, only the abundance of Saccharina latissima responded negatively to both summer and winter temperatures (Yesson et al., 2015b). Saccharina latissima has an optimal growth temperature between 10 and 15°C (Li et al., 2020), with growth reducing by 50 to 70% at 20°C, and all experimental specimens disintegrating after seven days at 23°C (Bolton & Lüning, 1982). The temperature isotherm of 19 to 20°C has been reported as limiting Saccharina latissima growth (Müller et al., 2009). Armitage et al. (2017) noted that Saccharina latissima was the most successful species in the cool summer (approx. 12 to 15°C), but it was strongly negatively affected by the hot summer (≥18°C), during a field study in southwestern Norway to observe competition between a non-native and two native habitat-building seaweeds.

Simonson, Scheibling & Metaxas (2015a) investigated the impacts of four temperature treatments (11, 14, 18 and 21°C) on growth, net length change and mortality of Saccharina latissima in Nova Scotia. Histological analysis showed temperature-mediated tissue damage, including holes, splitting of the medulla, damage to the meristoderm and loss of differentiation between tissue layers at temperatures between 14 and 21°C. Exposure to 21°C for one week reduced blade tissue strength (breaking stress) and extensibility (breaking strain) by 40 to 70% and exhibited reduced strength after three-week exposure to 18°C (Simonson, Scheibling & Metaxas, 2015a). Since the middle of the 20^th century, kelp species in Nova Scotia have experienced large population declines, up to 85 to 99%, of which temperature could have been a contributing factor (Filbee-Dexter, Feehan & Scheibling, 2016). However, Krumhansl et al. (2023) analysed the changes in Nova Scotia kelp abundance over the past 40 years (1982 to 2022) and found that there has been a loss in cold-tolerant kelps (such as Alaria esculenta, Saccorhiza dermatodea, and Agarum clathratum) and an increase in favour of the more warm-tolerant kelps like Saccharina latissima and Laminaria digitata. Kelp abundance increased since 2000, with Saccharina latissima widely abundant in the region by 2022 (Krumhansl et al., 2023). The highest kelp cover occurred on wave-exposed shores and at sites where temperatures have remained below thresholds for growth (21°C) and mortality (23 °C) (Krumhansl et al., 2023). Moreover, kelp has recovered from turf dominance following losses at some sites during a warm period from 2010 to 2012 (Krumhansl et al., 2023). Krumhansl et al. (2023) concluded that the dramatic change seen in kelp community composition in Nova Scotia over the past 40 years was in part driven by the loss of sea urchin herbivory, but a broad-scale shift to turf-dominance had not occurred, and that resilience and persistence were still a feature of kelp forests in the region despite rapid warming over the past several decades.

Elevated temperatures can increase erosion of Saccharina latissima blades and the subsequent release of total organic carbon and total nitrogen. Ding, Brussaard & Timmermans (2025) collected Saccharina latissima samples from the coastal waters south of Texel, The Netherlands, and subjected samples to naturally increased temperatures (from 16.1°C to 22.5°C) and further elevated temperatures (from 16.1°C to 27.1°C). A significant increase in the erosion rate of the distal parts of blades was observed in both temperature treatments, and substantial amounts (4.24 ± 0.31 mg/cm of carbon and 0.32 ± 0.13 mg/cm of nitrogen) of nutrients were released from Saccharina latissima, especially under sub-lethal temperature conditions. Under further elevated temperatures, with a prolonged period of higher temperature and a maximum temperature of 27.1°C, the effects were stronger, and erosion occurred along the edges of the whole blade. Ding, Brussaard & Timmermans (2025) concluded that rising temperatures accelerate the erosion of Saccharina latissima blades, highlighting a reason for the decline of kelp forests under climate change, as well as the potential impacts on nutrient cycling in the oceans.

Müller, Wiencke & Bischof (2008) found that elevated temperatures can exacerbate stress from ultraviolet radiation from sunlight. They investigated the combined effects of temperature and light quality on early life stages of Laminaria digitata and Saccharina latissima from Arctic (Spitsbergen) and temperate (Helgoland) populations. Temperature treatments ranged from 2°C to 18°C, representing Arctic summer conditions and North Sea summer extremes. For Laminaria digitata, Arctic populations germinated well at 2 to 12°C but failed at 18°C, while Helgoland populations showed optimal germination at 7 to 18°C. Saccharina latissima exhibited very low germination in Arctic populations (8 to 35%) and complete inhibition at 18°C, whereas temperate populations maintained high germination (85 to 92%) across all temperatures. UV-B radiation was the most damaging factor, reducing germination by up to 99% in Arctic Laminaria digitata and 74 to 90% in Arctic Saccharina latissima, and strongly inhibiting egg release (from 19 to 34 eggs mm² under normal light to 1.5 to 4 eggs mm² under UV-B). UV-A occasionally enhanced gametogenesis at moderate temperatures but did not offset UV-B damage. Overall, increased light (UV exposure) combined with higher temperatures produced the greatest negative effects, while low light and moderate temperatures favoured Arctic populations, and these findings indicate that warming exacerbates UV-B stress and severely limits recruitment (Müller, Wiencke & Bischof, 2008).

In a warming experiment studying Arctic populations of Saccharina latissima, no gametophytes survived at 20°C, but most growth parameters were greater at 10 to 15°C than at 5°C (Park et al., 2017). Another warming experiment involving Saccharina latissima from Kongsfjorden (Svalbard, Norway) highlighted an increase in physiological performance and growth in samples at 15°C (compared to 0°C), and that at least Arctic populations of Saccharina latissima can adjust and might even benefit from increased temperatures (Li et al., 2020). However, Gordillo, Carmona & Jimenez (2022) observed how Arctic individuals of Saccharina latissima lost more biomass in the dark at higher temperatures than lower ones, with a warmer polar night posing a limit on multi-year seaweeds to occupy new ice-free illuminated areas of the Arctic coasts.

Temperature is an environmental factor controlling the development of the microscopic stages of Saccharina latissima, with crucial changes in survival, growth, and gametogenesis occurring within a few degrees of its upper thermal limits (Redmond, 2013). The optimal germination temperature for Saccharina latissima is between 2°C and 12°C, with gametophyte survival between 23 to 25°C (Müller et al., 2009). Germination rates drop at 22°C, with surviving gametophytes smaller than those grown at lower temperatures (Redmond, 2013). Park et al. (2017) observed reductions in the percentage of sporophytes produced at 15°C when compared to values produced at 5°C and 10°C. Fales et al. (2023) compared the physiological responses of Saccharina latissima sporophytes to high temperature stress (low: 9 and 13°C, moderate: 15 and 16°C, and warm: 21°C) and nitrogen limitation (low: 1 to 3 μM vs. high: >10 μM) over 8 to 9 days. Saccharina latissima responded negatively to elevated temperatures, but not to low nitrogen levels. Blades of Saccharina latissima showed signs of metabolic stress and reduced growth in the warmest temperature treatment (21°C), at both high and low nitrogen levels, suggesting that Saccharina latissima is susceptible to thermal stress over short time periods, and that nutrient additions may actually reduce kelp performance at supra-optimal temperatures (Fales et al., 2023).

Niedzwiedz et al. (2022) also studied the response of Saccharina latissima sporophytes (sampled from Helgoland, German Bight, in June 2018, August 2018 and August 2019) to warming (at treatment temperatures of 18, 20, 22 and 24°C) and found that survival decreased with increasing environmental and experimental temperatures. Growth also revealed seasonal patterns, being higher in June than in August (Niedzwiedz et al., 2022). Niedzwiedz et al. (2022) concluded that the thermal tolerance of Saccharina latissima towards heatwaves in summer is significantly affected by the environmental history it previously experienced. This result has been seen in other experiments involving Saccharina latissima as well, whereby its sporophytes are pre-exposed to moderate stress to improve the performance and tolerance of plants when exposed to harsher conditions. This is known as thermal priming, and this may happen naturally as kelp are continually exposed to a warming climate. Gauci et al. (2024) observed how gametophytes primed at 20°C for four and six weeks exhibited an 11-day longer tolerance at 22°C, a seven-day longer tolerance at 23°C, and a 1°C higher thermal tolerance over seven days compared to two-week priming.

In the field, Saccharina latissima has shown significant regional variation in its acclimation response to changing environmental conditions. For example, Gerard & Dubois (1988) observed sporophytes of Saccharina latissima that were regularly exposed to ≥20°C tolerated these high temperatures, whereas sporophytes from other populations, which rarely experience ≥17°C, showed 100% mortality after three weeks of exposure to 20°C. At higher temperatures (11, 18 and 21°C), the nutritional content (C/N) of Saccharina latissima seems unaffected (Simonson, Scheibling & Metaxas, 2015b). However, the sea snail Lacuna vincta was observed grazing more kelp at higher temperatures (21°C) and suggests that the effects of grazing will act additively with the direct effects of temperature and cause increased biomass loss from kelp beds (Simonson, Scheibling & Metaxas, 2015b).

Saccharina latissima has suffered a dramatic decline in the Skagerrak region, Norway, where community structure has shifted from Saccharina latissima forests to communities dominated by filamentous macroalgae (Moy & Christie, 2012). In 2006, Andersen et al. (2011) transplanted Saccharina latissima into areas from where this species had been lost previously to determine whether the kelp could grow and mature. High mortality occurred from August to November each year. In 2008, only six of the seventeen original transplanted Saccharina latissima sporophytes survived (approx. 65% mortality rate). All surviving sporophytes were heavily fouled by epiphytic organisms (estimated cover of 80 & 100%). Between 1960 and 2009, sea surface temperatures in the region had regularly exceeded 20°C, and so had the duration at which temperatures remained above 20°C. High sea temperatures have been linked to the slow growth of Saccharina latissima, which is likely due to a decrease in the photosynthetic ability of Saccharina latissima, and an increase in vulnerability to epiphytic loading, bacterial and viral attacks (Anderson et al., 2011).

Kelp forests, including populations of Saccharina latissima, across the coastline of New England, USA, have experienced population shifts since the start of the 21^st century. Suskiewicz et al. (2024) surveyed between 31 and 67 forests spanning >350 km of coastline in Maine between 2001 and 2018 and then modelled how temperature change and sea urchin density influenced kelp abundance. Notably, the time-period studied was marked by rapid regional warming and several marine heatwaves, and the length of coastline examined experiences a more than 6°C difference in summer seawater temperatures from north to south (Suskiewicz et al., 2024). The maximum summer Near-Surface Seawater Temperatures in southern Maine commonly exceeded 20°C and were, on average, approx. 5.6°C warmer than those observed in northeast Maine (Suskiewicz et al., 2024). Consequently, southwestern subregions now regularly experience temperatures (15°C) at which nitrate saturation reaches zero (García-Reyes et al., 2022 and Zimmerman & Kremer, 1984 cited in Suskiewicz et al., 2024) as well as temperatures (20°C) at which sugar kelp erodes faster than it grows (Lee & Brinkhuis, 1986, cited in Suskiewicz et al., 2024). Also, high seawater temperatures reduce nutrient availability to kelp, causing nutrient depletion at 15°C (García-Reyes et al., 2022 and Zimmerman & Kremer, 1984 cited in Suskiewicz et al., 2024); and reduced nutrients during periods of maximum growth (spring) or thermal stress (summer) can accelerate kelp loss over time, as seen across all subregions by the end of the study by Suskiewicz et al. (2024). Although forests (Saccharina latissima and Laminaria digitata) had broadly returned to Maine in the late 20^th century, forests in northeast Maine have since experienced slow but significant declines in kelp, and forest persistence in the northeast was juxtaposed by a rapid, widespread collapse in the southwest (Suskiewicz et al., 2024). Forests collapsed in the southwest likely because ocean warming has directly and indirectly made this area inhospitable to kelp (Suskiewicz et al., 2024).

Hill et al. (2025) used species distribution models to evaluate the potential of enhanced thermal tolerance to buffer the effects of climate change (an increase of 1 to 5°C in maximum sea surface temperature) on cold-adapted kelp species. The models demonstrated that an increase of 1 to 2°C in thermal tolerance could recover over 50% of predicted losses of suitable habitat for cold-adapted kelps, with Saccharina latissima peaking at 17°C (Hill et al., 2025). For example, in the East Atlantic, Saccharina latissima recovery was concentrated in the southeast UK, but all species had projected patches of recovery on the Iberian coastline (Hill et al., 2025). In the North Sea and Skagerrak regions, a tolerance increase of 4 to 5°C was required for complete recovery (Hill et al., 2025). In the Baltic Sea, Saccharina latissima recovered with a tolerance increase of 1 to 2°C except for the mouth of the Baltic, where some areas remained unrecovered, even with a 5°C increase in tolerance (Hill et al., 2025). Overall, Saccharina latissima had the highest recovery potential with 99% of its projected lost suitable habitat area recovered under all climate change scenarios explored using the species distribution models (Hill et al., 2025). However, relying on mitigation or adaptation alone will likely be insufficient to maintain their historic range under projected climate change (Hill et al., 2025).

Similarly, Goldsmit et al. (2021) used a Random Forest model to predict future habitat suitability and cover for the dominant kelp species under climate change scenarios in the Eastern Canadian Arctic. Saccharina latissima is projected to have the largest gain in suitable habitat in both 2050 and 2100, with declines projected for some areas (e.g., north of Baffin Bay, Foxe Basin and Hudson Bay) by 2100 (Goldsmit et al., 2021). In general, suitable habitat is projected to occur in the northernmost reaches of the Eastern Canadian Arctic and is expected to persist into the future (Goldsmit et al., 2021). As the ocean warms and ice recedes, the model by Goldsmit et al. (2021) projects that Saccharina latissima will gain suitable habitat along much of the west coast of Greenland and the northern arm of the Northwest Passage.

Assis et al. (2018) predicted that, under the highest emission scenario (RCP 8.5), the range of Saccharina latissima would move northwards, retreating from the coast of Portugal, France and the southwest coast of the UK. The authors projected that, under RCP 2.6, 13% suitable Laminaria hyperborea habitat would be lost from the Western English Channel, while under the RCP 8.5 emission, 87% of suitable habitat was expected to be lost.

The growth and uptake of nitrate (NO₃) and phosphate (PO₄³) of juvenile Saccharina latissima sporophytes vary with temperature. Ding, Soetaert & Timmermans (2025) examined this effect under five temperature treatments ranging from 7.6°C to 24.5°C and found that NO₃ uptake significantly decreased when temperature was at or above 15.7°C, while high temperatures had no effect on PO₄³ uptake rates, and nitrate uptake significantly correlated with growth only at lower temperatures of 7.6°C and 12.6°C. In contrast, PO₄³ uptake was significantly correlated with growth across all temperature treatments except the highest (24.5°C). Also, at high temperatures (20.9°C and 24.5°C), NO₃ release was observed, while PO₄³ uptake consistently showed positive values, suggesting distinct regulatory mechanisms for nitrogen and phosphorus in Saccharina latissima (Ding, Soetaert & Timmermans, 2025).

Jung et al. (2025) studied the effects of temperature on early sporophyte development of Saccharina latissima under different temperatures (5, 10, 15, and 20°C) for 20 days. The development of sporophytes was observed earlier at 10°C than all other temperatures, with no sporophytes observed at 20°C during the experiment. Ebbing et al. (2021) also observed optimal reproduction of Saccharina latissima at lower temperatures (10.2°C), but at high light intensities (≥29 µmol photons/m²/s), and at higher temperatures (≥12.6°C) at lower light intensities (≤15 µmol photons/m²/s), highlighting both spring and autumn as the optimal seasons for Saccharina latissima reproduction.

Gametophytes can develop in ≤23°C (Lüning, 1990). However, the optimal temperature range for sporophyte growth is 10 to 15 °C (Bolton & Lüning, 1982). Bolton & Lüning (1982) experimentally observed that sporophyte growth was inhibited by 50 to 70% at 20°C, and following seven days at 23°C, all specimens completely disintegrated. In the field, Saccharina latissima has shown significant regional variation in its acclimation to temperature changes. For example, Gerard & Dubois (1988) observed sporophytes of Saccharina latissima which were regularly exposed to ≥20°C could tolerate these temperatures, whereas sporophytes from other populations, which rarely experience ≥17°C, showed 100% mortality after 3 weeks of exposure to 20°C. Therefore, the response of Saccharina latissima to a change in temperature is likely to be locally variable.

Saccharina latissima has disappeared almost completely from the Danish estuary Limfjorden, where maximum surface temperatures in summer have increased by 0.7°C per decade over the last 40 years, while the number of days with temperatures above 20°C has increased dramatically from 1 to 2 days per year to >25 days per year (Pedersen, 2015). Similarly, Saccharina latissima has been lost from the Skagerrak coast of Norway, which is thought to be due to an increase in summer temperatures, coupled with eutrophication (Moy & Christie, 2012).

Davey et al. (2025) studied the effect of short-term sublethal heat shock (20°C vs. ambient 10°C) on the health (growth and productivity), physiological performance (photosynthetic variables) and potential for compensatory mechanisms (phenolic content) of Saccharina latissima. The effect of heat shock was tested over five time points (0, 6, 24, 48, 72 hours). Growth of Saccharina latissima increased by 56% under heat shock, and gross primary productivity was initially greater in heat shock treatments (after six hours) but declined after 48 hours (Davey et al. 2025). Davey et al. (2025) concluded that Saccharina latissima exhibits the potential for short-term acclimation to sublethal heat shock, which may provide resistance to extreme temperature events, but that responses are species-specific.

Ding, Derksen & Timmermans (2025) investigated physiological and biochemical responses of juvenile Saccharina latissima sporophytes to acute (1-day to 10-day) and chronic (20-day to 40-day) warming from 11°C to 21°C, followed by exposure to 25°C. Acute warming (mimicking marine heatwaves) impaired physiological performance and reduced survival of juvenile Saccharina latissima sporophytes, whereas chronic warming led to elevated carbon and nitrogen reserves, increased fucoidan and protein levels, and enhanced photosynthetic performance. Improved heat tolerance of juvenile Saccharina latissima sporophytes was observed only in sporophytes previously exposed to 25°C, only after prior chronic warming treatments (Ding, Derksen & Timmermans, 2025). Ding, Derksen & Timmermans (2025) concluded that while exposure to chronic (gradual) temperature increases may allow Saccharina latissima to acclimate, events can exceed their physiological limits, leading to low survival, especially acute warming, which ultimately determines the presence or distribution of Saccharina latissima.

Under experimental conditions, Nepper-Davidsen, Andersen & Pedersen (2019) exposed a northern (Denmark) population of Saccharina latissima to a simulated three-week heatwave of three different intensities, 18, 21 and 24°C. When exposed to heatwaves of 18 and 21°C, there was a decrease in photosynthesis and growth. When 24°C was simulated, 91% of sporophytes were dead within a week, and the fronds of the few survivors were disintegrating, so the experiment was terminated (Nepper-Davidsen, Andersen & Pedersen et al., 2019). The results show that exposure to high, but sub-lethal, temperatures can have significant long-term effects, which may cause loss of biomass and leave Saccharina latissima susceptible to other stressors (Nepper-Davidsen, Andersen & Pedersen, 2019). This suggests that Saccharina latissima is unlikely to survive heatwaves of the length and magnitude predicted by the end of this century for both the middle and high emission scenarios.

Simonson, Scheibling & Metaxas (2015a) investigated the impacts of four temperature treatments (11, 14, 18 and 21°C) on growth, net length change and mortality of Saccharina latissima in Nova Scotia. Histological analysis showed temperature-mediated tissue damage, including holes, splitting of the medulla, damage to the meristoderm and loss of differentiation between tissue layers at temperatures between 14 and 21°C. Exposure to 21°C for one week reduced blade tissue strength (breaking stress) and extensibility (breaking strain) by 40 to 70% in and exhibited reduced strength after three-week exposure to 18°C (Simonson, Scheibling & Metaxas, 2015a). At the time, kelp species in Nova Scotia were experiencing large population declines, of which temperature could have been a contributing factor. However, Krumhansl et al. (2023) analysed the changes in Nova Scotia kelp abundance over the past 40 years (1982 to 2022) and found that there has been a loss in cold-tolerant kelps (such as Alaria esculenta, Saccorhiza dermatodea, and Agarum clathratum) and an increase in favour of the more warm-tolerant kelps like Saccharina latissima and Laminaria digitata. Kelp abundance increased since 2000, with Saccharina latissima widely abundant in the region by 2022 (Krumhansl et al., 2023). The highest kelp cover occurred on wave-exposed shores and at sites where temperatures have remained below thresholds for growth (21°C) and mortality (23°C) (Krumhansl et al., 2023). Moreover, kelp has recovered from turf dominance following losses at some sites during a warm period from 2010 to 2012 (Krumhansl et al., 2023). Krumhansl et al. (2023) concludes that that the dramatic change seen in kelp community composition in Nova Scotia over the past 40 years is in part driven by the loss of sea urchin herbivory, but a broad-scale shift to turf-dominance has not occurred, and that resilience and persistence are still a feature of kelp forests in the region despite rapid warming over the past several decades.

Miller et al. (2024a) conducted a 23-day mesocosm experiment exposing mixed kelp communities to warming and heatwave scenarios projected for the year 2100 to assess their impact. Three treatments were considered: a constant warming (+1.8°C from the control), a medium magnitude and long duration heatwave event (+2.8°C from the control for 13 days), and two short-term, more intense, heatwaves (5-day long scenarios with temperature peaks at +3.9°C from the control). The results showed that both marine heatwave treatments reduced net community production, whereas the constant warm temperature treatment displayed no difference from the control (Miller et al., 2024a). The long marine heatwave scenario resulted in reduced accumulated net community production, indicating that prolonged exposure had a greater severity than two high-magnitude, short-term heatwave events (Miller et al., 2024a). Miller et al. (2024a) estimated an 11°C temperature threshold at which negative effects to primary production appeared present, and that marine heatwaves can induce sublethal effects on kelp communities by depressing net community production.

In southern Norway, the frequency and intensity of marine heatwaves have been correlated to decreasing kelp biomass resulting from an increasing trend in the duration of temperature anomalies at a rate of 0.17 days/year over the past 60 years (Filbee-Dexter et al., 2020 cited in Miller et al., 2024a). This leads to temperatures surpassing the mortality threshold of 19.7°C for populations of Saccharina latissima (Miller et al., 2024a). During the mesocosm experiment by Miller et al. (2024a), temperatures reached a maximum of approx. 14°C, which should be below the mortality threshold, although the populations studied are from an Arctic fjord in northern Norway. This is particularly relevant given that ecotypes may demonstrate a difference in temperature tolerance (King et al., 2019 cited in Miller et al., 2024a) and the acclimatization potential to marine heatwaves by specific ecotypes (Miller et al., 2024a). It is important to note that the negative effects of marine heatwaves presented here do not indicate mass mortality or significant senescence, but more of a sublethal effect on community production (Miller et al., 2024a).

Interestingly, Saccharina latissima has been shown to potentially carry a thermal history where exposure to a previously high temperature anomaly, or its accumulated exposure duration reduces its tolerance to future anomalies (Niedzwiedz et al., 2022). This result has been seen in other experiments involving Saccharina latissima as well, whereby its sporophytes are pre-exposed to moderate stress to improve the performance and tolerance of plants when exposed to harsher conditions. This is known as thermal priming, and this may happen naturally as kelp are continually exposed to a warming climate. Gauci et al. (2024) observed how gametophytes primed at 20°C for four and six weeks exhibited an 11-day longer tolerance at 22°C, a seven-day longer tolerance at 23°C, and a 1°C higher thermal tolerance over seven days compared to two-week priming.

Kelp forests, including populations of Saccharina latissima, across the coastline of New England, USA, have experienced population shifts since the start of the 21^st century. Suskiewicz et al. (2024) surveyed between 31 and 67 forests spanning >350 km of coastline in Maine between 2001 and 2018 and then modelled how temperature change and sea urchin density influenced kelp abundance. Notably, the time-period studied was marked by rapid regional warming and several marine heatwaves, and the length of coastline examined experiences a more than 6°C difference in summer seawater temperatures from north to south (Suskiewicz et al., 2024). Maximum summer Near-Surface Seawater Temperatures in southern Maine commonly exceeded 20°C and were, on average, ~5.6°C warmer than those observed in northeast Maine (Suskiewicz et al., 2024). Consequently, southwestern subregions now regularly experience temperatures (15°C) at which nitrate saturation reaches zero (García-Reyes et al., 2022 and Zimmerman & Kremer, 1984 cited in Suskiewicz et al., 2024) as well as temperatures (20°C) at which sugar kelp erodes faster than it grows (Lee & Brinkhuis, 1986 cited in Suskiewicz et al., 2024). Also, high seawater temperatures reduce nutrient availability to kelp, causing nutrient depletion at 15°C (García-Reyes et al., 2022 and Zimmerman & Kremer, 1984 cited in Suskiewicz et al., 2024); and reduced nutrients during periods of maximum growth (spring) or thermal stress (summer) can accelerate kelp loss over time, as seen across all subregions by the end of the study by Suskiewicz et al. (2024). Although forests (Saccharina latissima and Laminaria digitata) had broadly returned to Maine in the late 20^th century, forests in northeast Maine have since experienced slow but significant declines in kelp, and forest persistence in the northeast was juxtaposed by a rapid, widespread collapse in the southwest (Suskiewicz et al., 2024). Forests collapsed in the southwest likely because ocean warming has directly and indirectly made this area inhospitable to kelp (Suskiewicz et al., 2024).

Phyllophora crispa is widely distributed on the coasts of the British Isles, except in the east of England. It is widely distributed in the east Atlantic with a northern limit in Iceland and a southern limit in North Africa but is also present in the Mediterranean and Black Sea (Newroth, 1971; Dixon & Irvine, 1977; Guiry & Guiry, 2015; Bunker et al., 2012). Phyllophora crispa has been recorded in temperatures ranging from 5 to 20°C, with the vast majority of observations recorded in the 10 to 15°C range (OBIS, 2025). Kooistra et al. (1989) noted that it was limited to lower shore tide pools, and that oxygen levels and competition were more limiting factors for Phyllophora crispa survival than temperature and salinity. However, Gallon et al. (2014) reported that changes in red seaweed assemblages across Brittany were correlated with a 0.7°C increase in coastal water temperature over the prior twenty years. Species varied in their response but the occurrence of several species of red algae, including Phyllophora crispa, increased.

Phyllophora crispa and Phyllophora pseudoceranoides are sensitive to large changes in temperature. Through culture experiments, 30°C was found lethal to Phyllophora pseudoceranoides within 4 to 2 weeks (Molenaar & Breeman, 1994). At 27°C plants were severely damaged after 3 months but were able to recover when returned to lower temperatures. Furthermore, temperature was found to control the time at which Phyllophora pseudoceranoides begins sporulation. For example, ≥15°C sporulation occurred at 30 months were as 10 °C sporulation occurred at 8 months (Molenaar & Breeman, 1994).

Phyllophora crispa showed resistance to elevated temperatures combined with elevated light levels, in an experiment on Mediterranean macroalgae (Hesse et al., 2025). Temperature (21, 26, and 30°C) was fully-crossed with light levels (180, 320, and 760 µmol photons/m²/s), and photosynthesis and respiration were measured via oxygen flux. Net photosynthesis (µmol O₂ /m²/s) was reduced by roughly half in the medium light/medium temperature treatment but remained similar to controls in the high light/high temperature treatment. Photosynthesis to respiration ratio (P:R) ranged from 1.9 in the low light/medium heat treatment to 3.69 in the high light/low heat treatment. P:R values above 1 indicate good health, as the alga is producing more energy through photosynthesis than it is consuming. Hesse et al. (2025) concluded that Phyllophora crispa in the Mediterranean would be relatively resistant to increasing temperature and light regimes, especially compared to Cystoceira spp. which is also a dominant species in Mediterranean macroalgal assemblages.

Sensitivity assessment. The distribution of Phyllophora crispa suggests that it would be tolerant of an increase in temperature at the pressure benchmark level. However, ecotypes of Saccharina latissima have been shown to have different temperature optimums (Dubois, 1988; Kerrison et al., 2015; Azevedo et al., 2016). Both a 2 and 5°C increase in temperature, when combined with high UK summer temperatures in the south of the UK, could cause large-scale mortality of Saccharina latissima. Therefore, resistance has been assessed as ‘None’. Hence, resilience is assessed as ‘High’ and sensitivity as ‘Medium’.

Temperature decrease (local)

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

Evidence

Saccharina latissima is widespread throughout the Arctic. The species has a latitudinal range of 41.3 degrees South to 79.8 degrees North, depth limits of 2.5 to 30 m, a thermal limit of -1.8 to 21.5°C, and there is little coastline left for poleward range expansion in the northwest Atlantic (Khan et al., 2018). Saccharina latissima is known to grow well between 5 and 17°C (Druehl, 1967, Fortes & Luning, 1980 and Machalek, Davison & Falkowski, 1996 cited in Kerrison et al., 2015), and it has a lower temperature threshold for sporophyte growth at 0°C (Lüning, 1990).

Despite a low temperature tolerance, growth at low temperatures does affect the demography of natural kelp beds through a reduction in growth rate combined with increased longevity (Rinde & Sjøtun, 2005). Novaczek et al. (1986) observed that 99% of newly settled zoospores died at 0°C, but sporophytes transferred from 5°C to 0°C remained healthy and continued to grow for a period of two months. Novaczek et al. (1986) therefore demonstrated that sporophytes could tolerate exposure to low (≥0°C) temperatures, but that exposure could have negative effects on larval survival and recruitment processes. However, a reduction in growth due to cold temperatures may just be representative of that specific population of kelp. For example, Saccharina latissima grows in Danish waters below the low optimal temperature for sporophyte growth (10 to 15°C), but those temperatures reflect the natural autumn and winter temperatures of the region (Boderskov et al., 2016).

Jung et al. (2025) studied the effects of temperature on early sporophyte development of Saccharina latissima under different temperatures (5, 10, 15, and 20°C) for 20 days, and the development of sporophytes was observed earlier at 10°C than all other temperatures, with no sporophytes observed at 20°C during the experiment. Ebbing et al. (2021) also observed optimal reproduction of Saccharina latissima at lower temperatures (10.2°C), but at high light intensities (≥29 µmol photons/m²/s), and at higher temperatures (≥12.6°C) at lower light intensities (≤15 µmol photons/m²/s), highlighting both spring and autumn as the optimal seasons for Saccharina latissima reproduction.

The growth and uptake of nitrate (NO₃) and phosphate (PO₄³) of juvenile Saccharina latissima sporophytes vary with temperature. Ding, Soetaert & Timmermans (2025) examined this effect under five temperature treatments ranging from 7.6°C to 24.5°C and found that NO₃ uptake significantly decreased when temperature was at or above 15.7°C, while high temperatures had no effect on PO₄³ uptake rates, and nitrate uptake significantly correlated with growth only at lower temperatures of 7.6°C and 12.6°C. In contrast, PO₄³ uptake was significantly correlated with growth across all temperature treatments except the highest (24.5°C). Also, at high temperatures (20.9°C and 24.5°C), NO₃ release was observed, while PO₄³ uptake consistently showed positive values, suggesting distinct regulatory mechanisms for nitrogen and phosphorus in Saccharina latissima (Ding, Soetaert & Timmermans, 2025).

Monteiro et al. (2021) studied the acclimation mechanisms of Saccharina latissima towards temperature and salinity. Samples of Saccharina latissima sporophytes were collected from Brittany, France, and were exposed to a combination of three temperatures (0, 8 and 15°C) and two salinity levels (20 and 30 PSU). The Saccharina latissima samples experienced a fivefold increase in the osmolyte mannitol in response to low temperature (0°C) compared to 8 and 15°C, which may have ecological and economic implications (Monteiro et al., 2021). Low temperatures significantly affected all parameters, mostly in a negative way; chlorophyll-a, the accessory pigment pool, growth and the maximal quantum yield of photosystem II (Fv/Fm) were significantly lower at 0°C, while the de-epoxidation state (the light-harvesting state, aka how plants dissipate excess light energy as heat) of the xanthophyll cycle (a mechanism protecting plants against oxidative stress) was increased at both 0 and 8°C compared to 15°C (Monteiro et al., 2021).

In the UK, the northern to southern Sea Surface Temperature ranges from 8 to 16°C in summer and 6 to 13°C in winter (Beszczynska-Möller & Dye, 2013). The effect of temperature change is likely to be regionally variable.

Phyllophora crispa is widely distributed on the coasts of the British Isles, except in the east of England. It is widely distributed in the east Atlantic with a northern limit in Iceland and a southern limit in North Africa but is also present in the Mediterranean and Black Sea (Newroth, 1971; Dixon & Irvine, 1977; Guiry & Guiry, 2015; Bunker et al., 2012). Phyllophora crispa has been recorded in temperatures ranging from 5 to 20°C, with the vast majority of observations recorded in the 10 to 15°C range (OBIS, 2025) . Kooistra et al. (1989) noted that it was limited to lower shore tide pools and that oxygen levels and competition were more limiting factors for Phyllophora crispa survival than temperature and salinity.

Sensitivity assessment. Resistance has been assessed as ‘High’, resilience as ‘High’. Sensitivity has been assessed as ‘Not Sensitive’.

Sensitivity assessment. The evidence shows that Saccharina latissima is probably tolerant of a long-term 2°C change in temperature for a year. It is also likely to tolerate a 5°C change in the short-term. The distribution of Phyllophora spp. suggests they are also able to tolerate temperature decreases at the benchmark level. Therefore, a resistance of ‘High’ is suggested so that resilience is ‘High’ (by default) and the biotope is assessed as ‘Not sensitive’ at the benchmark level. However, confidence in the assessment is ‘Low’ as it is based on expert judgment and proxies for evidence.

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

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

Haniask (1979) reported that at 24°C Codium fragile thalli growth could occur from 12-42‰, with an optimum from 24-30‰. 100% mortality occurred at 6‰ and at 12‰ growth was reduced. At the extremes of Codium fragile temperature tolerance (6 & 30°C) salinity tolerances were restricted, thalli grown at 6°C had a tolerance of 18-36‰, and thalli grown at 30°C had a salinity tolerance of 18-48 ‰ (Haniask, 1979). Codium fragile sporelings had narrower salinity and thresholds than mature thalli; Spores did not germinate at <18‰.

Lindahl and Runnström (1929) showed (experimentally) that Psammechinus miliaris from the littoral (Z form) and sub-littoral (S form) had different salinity optima. Gezelius (1962) reported the littoral growth form had an optimal salinity range of 20-32 ppt, whereas the sub-littoral growth form 26-38ppt. Mature examples of the littoral growth form tolerated 15 ppt for a period of 27 days, however, were not able to produce gametes at this salinity.

Karsten (2007) tested the photosynthetic ability of Saccharina latissima under acute 2 and 5 day exposure to salinity treatments ranging from 5-60 psu. A control experiment was also carried at 34 psu. Saccharina latissima showed high photosynthetic ability at >80% of the control levels between 25-55 PSU. Hyposaline treatment of 10-20 PSU led to a gradual decline of photosynthetic ability. After 2 days at 5 PSU, Saccharina latissima showed a significant decline in photosynthetic ability at approx. 30% of control. After 5 days at 5 PSU, Saccharina latissima specimens became bleached and showed signs of severe damage. The affect of long-term salinity changes (>5 days) or salinity >60 psu on Saccharina latissima’ photosynthetic ability was not tested. The experiment was conducted on Saccharina latissima from the Arctic, and at extremely low water temperatures (1-5 °C) macroalgae acclimation to rapid salinity changes could be slower than at temperate latitudes. It is, therefore, possible that resident Saccharina latissima of the UK maybe be able to acclimate to salinity changes more effectively.

Sensitivity assessment. IR.LIR.KVS.Cod is recorded in full salinity but probably exposed to reduced (18-30 ppt) conditions (Connor et al., 2004). A salinity decrease to “Low” (<18 ppt) may cause declines in Codium spp. growth and detriment the biotope. As a result, the Codium abundance could fall resulting in the SlatPhyVS biotope. IR.LIR.KVS.SlatPsaVS and IR.LIR.KVS.SlatPhyVS are recorded in ‘reduced’ and ‘low’ salinity, A further reduction in salinity would result in close to freshwater conditions and, however unlikely, would result in loss of the biotopes.  Resistance has been assessed as ‘Low’ and resilience as ‘High’. Therefore, the sensitivity of this biotope to a decrease in salinity has been assessed as ‘High’.

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

For macroalgae, a salinity of 33 to 35 PSU commonly results in optimal growth, while areas of reduced salinity, such as at river mouths, are amenable to the survival of fewer, tolerant species (Kerrison et al., 2015). Saccharina latissima can be classed as semi-euryhaline but appears more sensitive to salinity than other kelp, such as Laminaria digitata (Kerrison et al., 2015). No reduction in growth rates is observed between 24 and 35 PSU, and there may even be an increase (Druehl, 1967 and Gerard, Dubois & Greene, 1987 cited in Kerrison et al., 2015). Between 25 and 55 PSU (under acute 2- and 5-day exposure), Saccharina latissima shows a high photosynthetic ability at >80% (Karsten, 2007). Below 24 PSU, a stress response can be observed: a 20 to 25% reduction in growth rate at 21 PSU (Dubois & Greene, 1987 cited in Kerrison et al., 2015) and a 20 to 30% reduction in photosynthetic performance at 15 to 20 PSU (Karsten, 2007). After two days at 5 PSU, Saccharina latissima showed a significant decline in photosynthetic ability at approx. 30% of control, and after five days at 5 PSU, Saccharina latissima specimens became bleached and showed signs of severe damage (Karsten, 2007). The experiment by Karsten (2007) was conducted on Saccharina latissima from the Arctic, and they suggest that at extremely low water temperatures (1 to 5°C) macroalgae acclimation to rapid salinity changes could be slower than at temperate latitudes. It is therefore possible that the resident Saccharina latissima of the UK may be able to acclimate to salinity changes more effectively.

Natural populations do occur at salinity tipping points, such as those in the White Sea, where salinity is 24 to 26 PSU (Drobyshev, 1971 cited in Kerrison et al., 2015) and in Danish fjords, where salinity is 22 to 24 PSU (Middelboe and Sand-Jensen, 2000 cited in Kerrison et al., 2015). However, these may represent locally adapted ecotypes. Nielsen et al. (2016) studied two populations of Saccharina latissima in Danish waters, one brackish and one marine, and noted how gene flow was reduced both between clusters and between populations within clusters. Thus, highlighting the high likelihood of locally adapted ecotypes, with both populations vulnerable to differing changes in salinity, such as an increase in salinity in the brackish ecotype, or a decrease in salinity in the marine ecotype (Nielsen et al., 2016).

Young Saccharina latissima sporophytes can survive a four-day exposure to 11 PSU, although significant stress is observed (Peteiro & Sánchez, 2012 cited in Kerrison et al., 2015), while exposure of only a few days to 5 or 6 PSU results in either a 95% reduction in photosynthetic performance and significant pigment loss, or death (Karsten, 2007; Peteiro & Sánchez, 2012 cited in Kerrison et al., 2015). Monteiro et al. (2021) studied the acclimation mechanisms of Saccharina latissima towards temperature and salinity. Samples of Saccharina latissima sporophytes were collected from Brittany, France, and were exposed to a combination of three temperatures (0, 8 and 15°C) and two salinity levels (20 and 30 PSU). Mannitol content and growth decreased with decreasing salinity; in contrast, pigment content and maximal quantum yield of photosystem II were to a large extent unresponsive to salinity (Monteiro et al., 2021).

Vettori, Nikora & Biggs (2020) studied the implications of hyposaline (freshwater) stress on the morphological and mechanical properties of Saccharina latissima. They noted how under hyposaline stress blades bleach, develop blisters underneath the cortex, change dimensions (increased volume and thickness, decreased width), and how blade material becomes more flexible and more difficult to break (i.e. tougher). However, it is important to note that the response to hyposaline stress reported may be specific to seaweeds living in waters with high salinity (salinity at the sample collection site is around 30 PSU and samples were held in tanks of 34 PSU) (Vettori, Nikora & Biggs, 2020).

Phyllophora crispa is recorded from shady places in the lower littoral, lower littoral pools and subtidally to approx. 30m (Dixon & Irvine, 1977; Bunker et al., 20102). Kooistra et al. (1989) noted that Phyllophora crispa was limited to lower shore tide pools and that oxygen levels and competition were more limiting factors for Phyllophora crispa survival than temperature and salinity. Maximova (2013; summary only) reported that ‘morphological and biological changes’ in Phyllophora crispa from the Black Sea changed in experiments where the ‘normal’ salinity was raised from 18 PSU to 25, 32 and 39, but no further details were available. Phyllophora crispa is found in seawater with salinities ranging from 15 to 40 PSU, with most records in the 30 to 35 PSU range (OBIS, 2025) .

A comparative study of salinity tolerances of macroalgae collected from North Zealand and the South Kattegat (Denmark) where salinity is 16 PSU showed that species generally had a high tolerance (maintained more than half of photosynthetic capacity in short-term exposures of four days) to salinities lower than 3.7 PSU. However, tolerances varied between species with Phyllophora pseudoceranoides exhibiting greater tolerance than Phycodrys rubens, which was the least resistant species tested (Larsen & Sand-Jensen, 2006).

Sensitivity assessment. The presence of Phyllophora crispa in the lower intertidal suggests that it might be exposed to changes in salinity due to evaporation or rainfall but only for very short periods. The observations from the Black Sea, the South Kattegat and OBIS (2025) suggest that Phyllophora crispa could survive reduced salinity conditions. However, a decrease in one MNCR salinity scale from “Full” (30 to 40 PSU) to “Reduced” (18 to 30 PSU) could inhibit Saccharina latissima photosynthesis and hence growth. However, a shift to reduced salinity conditions is likely to result in a change in the infauna community and an overall reduction in species diversity. Therefore, resistance has been assessed as ‘Medium’ and resilience as ‘High’. Sensitivity of this biotope to a decrease in salinity has been assessed as ‘Low’.

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

IR.LIR.KVS.Cod , IR.LIR.KVS.Cod, IR.LIR.KVS.SlatPhyVS & IR.LIR.KVS.SlatPsaVS are predominantly shallow biotopes recorded from 0-10 m BCD An increase in emergence will result in an increased risk of desiccation and mortality of the macro-algae of the biotope.  Removal of canopy-forming kelps, through desiccation, has also been shown to increase desiccation and mortality of the understorey macroalgae (Hawkins & Harkin, 1985). Thomsen & McGlathery (2007) demonstrated that Codium fragile biomass declined if artificially placed at higher tidal elevations, and would therefore likely be sensitive to changes in emergence regime. Many of the dominant species an also occur in the lower intertidal, however, the biotope would probably be replaced by a lower shore equivalent.  Providing that suitable substrata are present, the biotope is likely to re-establish further down the shore within a similar emergence regime to that which existed previously.

Sensitivity assessment. Resistance has been assessed as ‘Low’ and resilience  as ‘Medium’. The sensitivity of this biotope to a change in emergence is considered to be ‘Medium’.

Wave exposure changes (local)

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

Evidence

Saccharina latissima typically dominates sheltered shorelines (Gilson et al., 2023) and is rarely present in areas of wave exposure, where it is out-competed by Laminaria hyperborea (Birkett et al., 1998). However, off the coast of northern Portugal, Saccharina latissima grew in offshore exposed conditions, with growth rates of 3.3% to 4.5%/day between January and May, while withstanding high wave heights (ranging from 0.5 to 12.6 m during the study period of January to September) (Azevedo et al., 2019).

Mols-Mortensen et al. (2017) cultivated Saccharina latissima with different wave and current exposures (sheltered, current-exposed and wave-exposed) in the Faroe Islands (from March to August 2015) to understand their variation in growth, yield, and protein concentration. Location 1 was defined as the sheltered location with a current speed of <5 cm/s (0.05 m/s) approx. half of the time and an overall current speed of <10 cm/s (0.1 m/s). The maximum observed current speed on this location was 20 to 30 cm/s (0.2 to 0.3 m/s), but this was only observed for a short period of time (Mortensen et al. 2014b cited in Mols-Mortensen et al., 2017). The highest average wave heights on location 1 were 0.9 m, and therefore, the location was considered to be sheltered, both with regard to current speed and wave heights. Location 2 was defined as the current-exposed location with an overall current speed of >20 cm/s (0.2 m/s) and occasional current speeds of >40 cm/s (0.4 m/s) not maintained for long periods of time (Larsen 1999 cited in Mols-Mortensen et al., 2017). The highest average wave heights on location 2 were 0.9 m, and therefore, current speed was considered to be the most important exposure factor on this location. Location 3 was defined as the wave-exposed location with a current speed of <10 cm/s (0.1 m/s) approx. half of the time and an overall current speed of <20 cm/s (0.2 m/s). Single observations on current speeds of 40 to 60 cm/s (0.4 to 0.6 m/s) were reported (Mortensen et al. 2014a cited in Mols-Mortensen et al., 2017). The highest average wave heights on location 3 were 2.2 m, and therefore, wave height was considered to be the most important exposure factor on this location. Overall, Saccharina latissima individuals cultivated at the current exposed location were heavier compared to the individuals cultivated at the other locations; however, the total biomass yield was significantly lower at the current exposed location (Mols-Mortensen et al., 2017).

Zhu et al. (2021) studied the morphological and physiological plasticity of Saccharina latissima in response to different hydrodynamic conditions and nutrient availability (56 days under fully controlled conditions of waves or no waves, and high or low nutrients). They observed how waves primarily increased frond biomass, elongation rate, and carbon to nitrogen ratio (C:N ratio), and induced both a greater variety in and rougher frond surface shapes; the highest C:N ratio was observed in the low nutrient-wave treatment. Together, these results seem to suggest that the thready and spring-like shapes found in the central frond (i.e., rougher frond surface) in wave-exposed conditions can at least partly compensate for low nutrient availability by enhancing nutrient and photon acquisition, particularly in low nutrient conditions (Zhu et al., 2021). Zhu et al. (2021) concluded that frond surface shapes in the newly formed central frond of Saccharina latissima can be regarded as possessing high morphological and physiological plasticity that enables kelp to cope with contrasting environments.

Visch, Nylund & Pavia (2020) studied the effect of wave exposure (defined as 500,000 to 800,000 m²/s for exposed, 100,000 to 200,000 m²/s for moderately exposed, and 10,000 to 30,000 m²/s for sheltered) on growth and biofouling of Saccharina latissima along the Swedish west coast. Growth, measured as blade surface area, generally increased with decreased wave exposure, with approximately 40% less growth at exposed locations compared to sheltered or moderately exposed locations (Visch, Nylund & Pavia, 2020). Biofouling of kelp decreased with increased wave exposure, from 10 and 6% coverage at sheltered and moderately exposed locations, respectively, to 3% at exposed locations (Visch, Nylund & Pavia, 2020). In addition, exposure level affected the tissue composition, with a high carbon, but low nitrogen and water content at exposed locations compared to moderate and sheltered sites; isotope signatures (i.e. δ¹³C and δ¹⁵N) also differed between exposure levels (Visch, Nylund & Pavia, 2020).

Gilson et al. (2023) studied the seasonal and spatial variability in rates of primary production and detritus release by intertidal stands of Saccharina latissima on wave-exposed shores in the northeast Atlantic. On moderately exposed shores, productivity and erosion of Saccharina latissima remained low and showed no clear seasonal pattern (Gilson et al., 2023). Peak erosion rates of Saccharina latissima at both wave exposures were approx. 0.6 g dry weight/day (Gilson et al., 2023), which is higher than previous rates recorded for populations of other kelp species, such as L. hyperborea and L. ochroleuca, along the UK coastline (Pessarrodona, Moore, et al., 2018 cited in Gilson et al., 2023). The ruffled margins of Saccharina latissima create considerably more drag than the flat lamina of Laminaria digitata, accounting for their greater rates of dislodgment even at more sheltered sites (Buck & Buchholz, 2005 cited in Gilson et al., 2023). In addition, Saccharina latissima also routinely settles on semi-stable rocks and cobbles instead of emergent bedrock, particularly in sheltered conditions, increasing its susceptibility to dislodgement (Scheibling et al., 2009 and Smale & Vance, 2016 cited in Gilson et al., 2023).

Storm-induced increases in wave action can be detrimental to kelp biotopes. During the Northeast Atlantic storm season of 2013 to 2014, the south coast of the UK was subjected to some of the most intense storms in recent history, being classed as a '1-in-30 year' event, where inshore significant wave heights and periods exceeded 7 m and 13 seconds (Smale & Vance, 2015). Overall, kelp canopies were highly resistant to storm disturbance, however, at one study site, a mixed canopy comprising Laminaria ochroleuca, Saccharina latissima, and Laminaria hyperborea was significantly altered by the storms, due to a decreased abundance of the former two species (Smale & Vance, 2015). On the Atlantic coast of Nova Scotia, Hurricane Earl generated extreme wave heights of up to 25 m and strong bottom currents, which caused a large-scale defoliation of kelp beds in shallow subtidal zones (Filbee-Dexter & Scheibling, 2012). Saccharina latissima and Laminaria digitata were stripped of blades, leaving only stipes and fragments, resulting in a 46% average loss of kelp canopy cover across the surveyed sites; the strong bottom currents also caused the displacement of urchins (Strongylocentrotus droebachiensis) (Filbee-Dexter & Scheibling, 2012). In addition, coralline and filamentous red algae cover increased after the storm due to the loss of kelp (Filbee-Dexter & Scheibling, 2012).

Sensitivity assessment. Saccharina latissima has been reported to grow in high wave exposure environments (Mols-Mortensen et al., 2017; Azevedo et al., 2019; Visch, Nylund & Pavia, 2020). However, Phyllophora crispa are typically found in sheltered to extremely sheltered conditions (JNCC, 2022). A large increase in near-shore wave height is likely to significantly influence biotope structure, especially where wave action mobilises the sediment and removes hard substratum embedded in the sediment, on which the macroalgae depend for attachment, potentially increase dislodgment or relocation of the characterizing species (South & Burrows, 1967; Birkett et al., 1998b; Smale & Vance, 2015), or increase erosion (Gilson et al., 2023) and affect growth (Mols-Mortensen et al., 2017). Storm damage (as above) can even remove kelps from hard substrata. However, an increase in nearshore significant wave height of 3 to 5% is not likely to result in damage to or loss of this biotope but may result in subtle changes to the associated community depending on the interplay of wave action and tidal streams. Therefore, resistance has been assessed as ‘High’, resilience as ‘High’, and sensitivity as ‘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

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.

Saccharina latissima fronds, being predominantly subtidal, would not come into contact with freshly released oil but only to sinking emulsified oil and oil adsorbed onto particles (Birkett et al., 1998). The mucilaginous slime layer coating of laminariales may protect them from smothering by oil. Hydrocarbons in solution reduce photosynthesis and may be algicidal. However, Holt et al. (1995) reported that oil spills in the USA and from the 'Torrey Canyon' had little effect on kelp forests. Similarly, surveys of subtidal communities at a number sites between 1-22.5 m below chart datum showed no noticeable impacts of the Sea Empress oil spill and clean up (Rostron & Bunker, 1997). An assessment of holdfast fauna in Laminaria showed that although species richness and diversity decreased with increasing proximity to the Sea Empress oil spill, overall the holdfasts contained a reasonably rich and diverse fauna, even though oil was present in most samples (Somerfield & Warwick, 1999).

Echinoderms seem especially sensitive to the toxic effects of oil, likely because of the large amount of exposed epidermis (Suchanek, 1993). Schäfer & Köhler (2009) found 20 day exposure to sub-lethal concentrations of phenanthrene resulted in severe ovarian lesions of Psammechinus miliaris limiting the production of gametes.

Following the Torrey Canyon incident, large numbers of dead Psammechinus miliaris were found in the vicinity of Sennen, UK possibly due to exposure to the oil spill and the heavy spraying of hydrocarbon based dispersants in that area (Smith, 1968). Other significant effects have been observed in other species of urchins. For example, mass mortality of the echinoderm Echinocardium cordatum was observed shortly after the Amoco Cadiz oil spill (Cabioch et al., 1978) and reduced abundance of the species was detectable up to >1000 m away one year after the discharge of oil-contaminated drill cuttings in the North Sea (Daan & Mulder, 1996). In the Mediterranean around Naples, urchins were absent from areas which had visible signs of massive pollution of both sewage and oil. Echinus esculentus populations in the vicinity of an oil terminal in A Coruna Bay, Spain, showed developmental abnormalities in the skeleton. The tissues contained high levels of aliphatic hydrocarbons, naphthalenes, pesticides and heavy metals (Zn, Hg, Cd, Pb, and Cu) (Gommez & Miguez-Rodriguez, 1999). But the observed effects may have been due to a single contaminant or synergistic effects of all present.

Cullinane et al. (1975) found large quantities of Codium fragile washed up on Relane, Bantry Bay, USA shortly after a large oil spill. No other evidence could be located for the effect of hydrocarbon & PAH contamination on Codium spp.

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.

Johansson (2009) exposed samples of Saccharina latissima to several antifouling compounds, observing chlorothalonil, DCOIT, dichlofluanid and tolylfluanid inhibited photosynthesis. Exposure to Chlorothalonil and tolylfluanid, was also found to continue inhibiting oxygen evolution after exposure had finished, and may cause irreversible damage. Smith (1968) noted that epiphytic and benthic red algae were intolerant of dispersant or oil contamination due to the Torrey Canyon oil spill; only the epiphytes Crytopleura ramosa and Spermothamnion repens and some tufts of Jania rubens survived together with Osmundea pinnatifida, Gigartina pistillata and Phyllophora crispa from the sublittoral fringe. Considerable observations and work, mainly on Echinus esculentus but also on Psammechinus miliaris (Smith, 1968; Gommez & Miguez-Rodriguez, 1999; Dinnel et al., 1988) indicate high intolerance to synthetic contaminants. Newton & McKenzie (1995) state that echinoderms tend to be very intolerant of various types of marine pollution, but there is little more detailed information than this. Following the Torrey Canyon incident, large numbers of dead Psammechinus miliaris in the vicinity of Sennen, UK presumably due to the heavy spraying of dispersants in that area and exposure to the oil spill (Smith, 1968).

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

Reduced oxygen concentrations can inhibit both photosynthesis and respiration in macroalgae (Kinne, 1977). Despite this, macroalgae are thought to buffer the environmental conditions of low oxygen, thereby acting as a refuge for organisms in oxygen depleted regions especially if the oxygen depletion is short-term (Frieder et al., 2012). A rapid recovery from a state of low oxygen is expected if the environmental conditions are transient. If levels do drop below 4 mg/L negative effects on these organisms can be expected, with adverse effects occurring below 2 mg/L (Cole et al., 1999).

Kooistra et al. (1989) noted that Phyllophora crispa was limited to lower shore tide pools but concluded that temperature and salinity were not the limiting factors but that oxygenation and competition were possible limiting factors. However, no direct evidence was found.

Hypoxia caused by a severe eutrophication event in the north-western Black Sea was reported as a constraint on the recovery of ‘Zernov’s Phyllophora field’ (Stevens et al., 2019). This area had undergone significant degradation between 1964 and 2004 due to eutrophication, resultant algal blooms and increased turbidity (Black Sea Commission, 2008; Kostylev et al., 2010; Stevens et al., 2019; Alexandrov & Milchakova, 2022). The field was composed of several species of Phyllophora including Phyllophora crispa. The Phyllophora field has remained but the abundance of the Phyllophora, the range of Phyllophora species, their age structure, the extent of the field, and the ecosystem of fish and other algae declined. Despite the establishment of six MPAs along the southwestern coast of Crimea (the Black Sea), the species’ biomass, density and thallus weight continued to decrease 2.7-fold, 1.5-fold, and 2-fold respectively between 1964 and 2017 (Alexandrov & Milchakova, 2022). However, an increase in species richness and extent of the field was reported from 2005 to 2007, so that regeneration had begun (Kostylev et al., 2010). BSC (Black Sea Commission, 2008) suggest that eutrophication and its effects stabilised in the 1990s and decreased in the 2000s. From 2006 to 2008, isolated patches of Phyllophora spp. were found in its former range in the Black Sea (Stevens et al., 2019). Maximum Phyllophora spp. cover ranged between 9 to 13% compared to extensive beds of 100% cover reported in the 1960s. It was suggested that recovery is constrained by residual nutrient flux from sediments, persistent hypoxia and competition from opportunistic algae, which are direct consequences of the eutrophication that took place decades ago.

Sensitivity assessment. Based on the case of ‘Zernov’s Phyllophora field’, it is likely that deoxygenation could have severe consequences for this biotope. Due to the extent of Phyllophora loss and the lack of recovery observed at ‘Zernov’s Phyllophora field’, resistance is assessed as ‘Low’, resilience as ‘Very low’, and sensitivity as ‘High’.

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

Johnston & Roberts (2009) conducted a meta-analysis, which reviewed 216 papers to assess how a variety of contaminants (including sewage and nutrient loading) affected 6 marine habitats (including subtidal reefs). A 30 to 50% reduction in species diversity and richness was identified from all habitats exposed to the contaminant types. Johnston & Roberts (2009) also highlighted that macroalgal communities are relatively tolerant to contamination, but that contaminated communities can have low diversity assemblages which are dominated by opportunistic and fast growing species (Johnston & Roberts, 2009 and references therein).

Areas with high nutrient loading will sustain rapid macroalgal growth during the summer (Davison, Andrews & Stewart, 1984 cited in Kerrison et al., 2015), and where nutrient loading is lower, high water flow increases the nutrient uptake rate of macroalgae by refreshing the boundary layer (Wheeler & North (date) cited in Kerrison et al., 2015), so maximal growth rates can be sustained. It has been shown in Saccharina latissima that 10 μmol/L of nitrate is required to maximise growth rate and leads to internal storage for later use (Chapman, Markham & Lüning, 1978 cited in Kerrison et al., 2015). Boderskov et al. (2016) noted how Saccharina latissima grown under high nutrient availability in Denmark fulfils a higher degree of nutrient bioremediation and has an improved biomass quality in regard to increased concentrations of pigments and nitrogen-rich compounds.

Conolly & Drew (1985) found Saccharina latissima sporophytes had relatively higher growth rates when in close proximity to a sewage outlet in St Andrews, UK, compared to other sites along the east coast of Scotland. At St Andrews, nitrate levels were 20.22 µM, which represents an approx. 25% increase compared to other sites (approx. 15.87 µM). Read et al. (1983) reported that after the installation of a new sewage treatment works, which reduced the suspended solid content of liquid effluent by 60% in the Firth of Forth, Saccharina latissima became abundant where previously it had been absent.

The association of fish and shellfish mariculture can also lead to an increased growth rate of macroalgae, while removing excess nutrients from the environment (Sanderson et al, 2008, Sanderson et al., 2012 and Wang et al., 2014 cited in Kerrison et al., 2015). Handå et al. (2013) reported Saccharina latissima sporophytes grew approx. 1% faster per day when in close proximity to Norwegian salmon farms, where elevated ammonium could be readily absorbed by sporophytes. However, experimentation in Denmark did not show any benefit in terms of growth, nitrogen, phosphorus, or amino acid content of Saccharina latissima cultured in proximity to fish and mussel aquaculture (Marinho, Holdt & Angelidaki, 2015 and Marinho et al., 2015 cited in Kerrison et al., 2015).

Rugiu et al. (2021) exposed Saccharina latissima to natural seawater, water enriched to levels of ammonium and nitrate simulating finfish cage waste (test IMTA1), and a combination of such enrichment with natural effluents coming from mussels (test IMTA2). The Saccharina latissima biomass was higher and produced elevated total organic content when exposed to both IMTA1 and IMTA2 nutrient scenarios, including a significant enhancement in pigment content only when algae were exposed to the strongest enrichment (IMTA2) (Rugiu et al., 2021). In addition, the photosynthetic responses in terms of relative electron transfer rate, PSII (photosystem II) saturation irradiance, total nitrogen content, and the content of chlorophyll-a and fucoxanthin were also positively affected by both IMTA1 and IMTA2 (Rugiu et al., 2021). Rugiu et al. (2021) concluded that Saccharina latissima showed a significant physiological response to nutrient enrichment mimicking aquaculture settings, as well as the benefit of added nutrients through a boost in photosynthetic activity that leads to higher kelp biomass and pigment production.

Fales et al. (2023) compared the physiological responses of Saccharina latissima sporophytes to high temperature stress (low: 9 and 13°C, moderate: 15 and 16°C, and warm: 21°C) and nitrogen limitation (low: 1 to 3 μM vs. high: >10 μM) over eight to nine days. Saccharina latissima responded negatively to elevated temperatures, but not to low nitrogen levels. Blades of Saccharina latissima showed signs of metabolic stress and reduced growth in the warmest temperature treatment (21°C), at both high and low nitrogen levels, suggesting that Saccharina latissima is susceptible to thermal stress over short time periods, and that nutrient additions may actually reduce kelp performance at supra-optimal temperatures (Fales et al., 2023).

Jevne, Forbord & Olsen (2020) examined how differences in light conditions and nutrient availability affect the growth and intracellular nitrogen of Saccharina latissima through cultivating sporophytes in land-based tanks with four different combinations of high/low light and high/low nutrient supply over an experimental period of 20 days. The results revealed that the mean growth rate and the intracellular nitrogen component of the sporophytes were positively related to the external nitrate concentration during the experimental period, indicating that Saccharina latissima requires high nutrient concentration to maintain a rapid growth (Jevne, Forbord & Olsen, 2020).

Bokn et al. (2003) conducted a nutrient loading experiment on intertidal fucoids. Within three years of the experiment, no significant effect was observed in the communities. However, four to five years into the experiment, a shift occurred from perennials to ephemeral algae. Although Bokn et al. (2003) focused on fucoids, the results could indicate that long-term (>4 years) nutrient loading can result in a community shift to ephemeral algae species. Disparities between the findings of the aforementioned studies are likely to be related to the level of organic enrichment.

Johnston & Roberts (2009) conducted a meta-analysis, which reviewed 216 papers to assess how a variety of contaminants (including sewage and nutrient loading) affected six marine habitats (including subtidal reefs). A 30 to 50% reduction in species diversity and richness was identified from all habitats exposed to the contaminant types. Johnston & Roberts (2009), however, also highlighted that macroalgal communities were relatively tolerant to contamination, but that contaminated communities could have low diversity assemblages dominated by opportunistic and fast-growing species (Johnston & Roberts, 2009).

‘Zernov’s Phyllophora field’ in the north-western Black Sea has undergone significant degradation between 1964 and 2004 due to eutrophication, resultant algal blooms and increased turbidity (Black Sea Commission, 2008; Kostylev et al., 2010; Stevens et al., 2019; Alexandrov & Milchakova, 2022). The ‘field’ was composed of several species of Phyllophora including Phyllophora crispa. The Phyllophora field has remained but the abundance of the Phyllophora, the range of Phyllophora species, their age structure, the extent of the ‘field’, and the ecosystem of fish and other algae decreased or declined. Despite the establishment of six MPAs along the southwestern coast of Crimea (the Black Sea), the species’ biomass, density and thallus weight continued to decrease 2.7-fold, 1.5-fold, and 2-fold respectively between 1964 and 2017 (Alexandrov & Milchakova, 2022). However, an increase in species richness and extent of the field was reported from 2005 to 2007, so that regeneration had begun (Kostylev et al., 2010). BSC (Black Sea Commission, 2008) suggest that eutrophication and its effects stabilised in the 1990s and decreased in the 2000s. From 2006 to 2008, isolated patches of Phyllophora spp. were found in its former range in the Black Sea (Stevens et al., 2019). Maximum Phyllophora spp. cover ranged between 9 to 13% compared to extensive beds of 100% cover reported in the 1960s. It was suggested that recovery is constrained by residual nutrient flux from sediments, persistent hypoxia and competition from opportunistic algae, which are direct consequences of the eutrophication that took place decades ago.

Sensitivity assessment. Based on the case of ‘Zernov’s Phyllophora field’, it is likely that nutrient enrichment and resultant algal blooms could have severe consequences for this biotope. Due to the extent of Phyllophora loss and the lack of recovery observed at ‘Zernov’s Phyllophora field’, resistance is assessed as ‘Low’, resilience as ‘Very low’, and sensitivity as ‘High’.

Organic enrichment

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

Evidence

Areas with high nutrient loading will sustain rapid macroalgal growth during the summer (Davison, Andrews & Stewart, 1984 cited in Kerrison et al., 2015), and where nutrient loading is lower, high water flow increases the nutrient uptake rate of macroalgae by refreshing the boundary layer (Wheeler & North (date) cited in Kerrison et al., 2015), so maximal growth rates can be sustained. It has been shown in Saccharina latissima that 10 μmol/l of nitrate is required to maximise growth rate and leads to internal storage for later use (Chapman, Markham & Lüning, 1978 cited in Kerrison et al., 2015). Boderskov et al. (2016) noted how Saccharina latissima grown under high nutrient availability in Denmark fulfils a higher degree of nutrient bioremediation and has an improved biomass quality in regard to increased concentrations of pigments and nitrogen-rich compounds.

Conolly & Drew (1985) found Saccharina latissima sporophytes had relatively higher growth rates when in close proximity to a sewage outlet in St Andrews, UK, compared to other sites along the east coast of Scotland. At St Andrews, nitrate levels were 20.22 µM, which represents an approx. 25% increase compared to other sites (approx. 15.87 µM). Read et al. (1983) reported that after the installation of a new sewage treatment works, which reduced the suspended solid content of liquid effluent by 60% in the Firth of Forth, Saccharina latissima became abundant where previously it had been absent.

The association of fish and shellfish mariculture can also lead to an increased growth rate of macroalgae, while removing excess nutrients from the environment (Sanderson et al, 2008, Sanderson et al., 2012 and Wang et al., 2014 cited in Kerrison et al., 2015). Handå et al. (2013) reported Saccharina latissima sporophytes grew approx. 1% faster per day when in close proximity to Norwegian salmon farms, where elevated ammonium could be readily absorbed by sporophytes. However, experimentation in Denmark did not show any benefit in terms of growth, nitrogen, phosphorus, or amino acid content of Saccharina latissima cultured in proximity to fish and mussel aquaculture (Marinho, Holdt & Angelidaki, 2015 and Marinho et al., 2015 cited in Kerrison et al., 2015).

Rugiu et al. (2021) exposed Saccharina latissima to natural seawater, water enriched to levels of ammonium and nitrate simulating finfish cage waste (test IMTA1), and a combination of such enrichment with natural effluents coming from mussels (test IMTA2). The Saccharina latissima biomass was higher and produced elevated total organic content when exposed to both IMTA1 and IMTA2 nutrient scenarios, including a significant enhancement in pigment content only when algae were exposed to the strongest enrichment (IMTA2) (Rugiu et al., 2021). In addition, the photosynthetic responses in terms of relative electron transfer rate, PSII (photosystem II) saturation irradiance, total nitrogen content, and the content of chlorophyll-a and fucoxanthin were also positively affected by both IMTA1 and IMTA2 (Rugiu et al., 2021). Rugiu et al. (2021) concluded that Saccharina latissima showed a significant physiological response to nutrient enrichment mimicking aquaculture settings, as well as the benefit of added nutrients through a boost in photosynthetic activity that leads to higher kelp biomass and pigment production.

Fales et al. (2023) compared the physiological responses of Saccharina latissima sporophytes to high temperature stress (low: 9 and 13°C, moderate: 15 and 16°C, and warm: 21°C) and nitrogen limitation (low: 1 to 3 μM vs. high: >10 μM) over eight to nine days. Saccharina latissima responded negatively to elevated temperatures, but not to low nitrogen levels. Blades of Saccharina latissima showed signs of metabolic stress and reduced growth in the warmest temperature treatment (21°C), at both high and low nitrogen levels, suggesting that Saccharina latissima is susceptible to thermal stress over short time periods, and that nutrient additions may actually reduce kelp performance at supra-optimal temperatures (Fales et al., 2023).

Jevne, Forbord & Olsen (2020) examined how differences in light conditions and nutrient availability affect the growth and intracellular nitrogen of Saccharina latissima through cultivating sporophytes in land-based tanks with four different combinations of high/low light and high/low nutrient supply over an experimental period of 20 days. The results revealed that the mean growth rate and the intracellular nitrogen component of the sporophytes were positively related to the external nitrate concentration during the experimental period, indicating that Saccharina latissima requires high nutrient concentration to maintain a rapid growth (Jevne, Forbord & Olsen, 2020).

Bokn et al. (2003) conducted a nutrient loading experiment on intertidal fucoids. Within three years of the experiment, no significant effect was observed in the communities. However, four to five years into the experiment, a shift occurred from perennials to ephemeral algae. Although Bokn et al. (2003) focused on fucoids, the results could indicate that long-term (>4 years) nutrient loading can result in a community shift to ephemeral algae species. Disparities between the findings of the aforementioned studies are likely to be related to the level of organic enrichment.

Johnston & Roberts (2009) conducted a meta-analysis, which reviewed 216 papers to assess how a variety of contaminants (including sewage and nutrient loading) affected six marine habitats (including subtidal reefs). A 30 to 50% reduction in species diversity and richness was identified from all habitats exposed to the contaminant types. Johnston & Roberts (2009), however, also highlighted that macroalgal communities were relatively tolerant to contamination, but that contaminated communities could have low diversity assemblages dominated by opportunistic and fast-growing species (Johnston & Roberts, 2009).

‘Zernov’s Phyllophora field’ in the north-western Black Sea has undergone significant degradation between 1964 and 2004 due to eutrophication, resultant algal blooms and increased turbidity (Black Sea Commission, 2008; Kostylev et al., 2010; Stevens et al., 2019; Alexandrov & Milchakova, 2022). The ‘field’ was composed of several species of Phyllophora including Phyllophora crispa. The Phyllophora field has remained but the abundance of the Phyllophora, the range of Phyllophora species, their age structure, the extent of the ‘field’, and the ecosystem of fish and other algae decreased or declined. Despite the establishment of six MPAs along the southwestern coast of Crimea (the Black Sea), the species’ biomass, density and thallus weight continued to decrease 2.7-fold, 1.5-fold, and 2-fold respectively between 1964 and 2017 (Alexandrov & Milchakova, 2022). However, an increase in species richness and extent of the field was reported from 2005 to 2007, so that regeneration had begun (Kostylev et al., 2010). BSC (2008) suggest that eutrophication and its effects stabilised in the 1990s and decreased in the 2000s. From 2006 to 2008, isolated patches of Phyllophora spp. were found in its former range in the Black Sea (Stevens et al., 2019). Maximum Phyllophora spp. cover ranged between 9 to 13% compared to extensive beds of 100% cover reported in the 1960s. It was suggested that recovery is constrained by residual nutrient flux from sediments, persistent hypoxia and competition from opportunistic algae, which are direct consequences of the eutrophication that took place decades ago.

Sensitivity assessment. Based on the case of ‘Zernov’s Phyllophora field’, it is likely that organic enrichment due to algal blooms or organic wastes, together with turbidity could have severe consequences for this biotope. Due to the extent of Phyllophora loss and the lack of recovery observed at ‘Zernov’s Phyllophora field’, resistance is assessed as ‘Low’, resilience as ‘Very low’, and sensitivity as ‘High’.

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 forms on hard rock substrata, i.e. bedrock, boulders and cobbles.  A change from hard rock to sedimentary substrata would result in permanent loss of the biotope. Therefore, resistance is assessed as None, resilience as Very low and sensitivity as High. confidence is assessed as 'High' due to the incontrovertible nature of the pressure.

Physical change (to another sediment type)

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

Evidence

Not relevant on hard rock substrata.

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

Not relevant on hard rock substrata.

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

Dixon & Irvine (1977) noted that Phyllophora crispa regenerates after erosion and animal grazing. Bunker et al. (2012) noted that it tolerated sediment cover and thrived in areas subject to shell gravel. Both observations suggest that it can either resist or regrow from damage due to sediment scour or animal grazing.

In 2012, the Costa Concordia cruise ship crashed into a rocky reef near Giglio Island in the Mediterranean (Piazzi et al., 2020). It took until July 2017 to fully remove the wreckage. Surveys that took place immediately after removal showed that the impacted site and two reference sites had macroalgal assemblages that were dominated by Phyllophora crispa. Percentage cover was initially 98.25 and 94.58% at the two reference sites, compared to 70.67% at the impacted site. Four months later, surveys showed that percentage cover only changed marginally at the reference sites but had declined by a further 56.5% at the impacted site. This decline coincided with an increase in native turf-forming macroalgae, which could colonize faster than Phyllophora crispa.

Abrasion of the substratum e.g. from bottom or pot fishing gear, cable laying etc. may cause localised mobility of the substrata and mortality of the resident community. The effect would be situation dependent however if bottom fishing gear were towed over a site, it may mobilise a high proportion of the rock substrata and cause high mortality in the resident community. 

Sensitivity assessment. Resistance has been assessed as ‘None’, resilience as ‘High’, and sensitivity as ‘Medium’.

Penetration or disturbance of the substratum subsurface

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

Evidence

Penetration is unlikely to be relevant to hard rock substrata. Therefore, the pressure is Not relevant.  However, physical disturbance of the surface is assessed under 'abrasion' above.

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

Suspended Particle Matter (SPM) concentration has a positive linear relationship with sub-surface light attenuation (Kd) (Devlin et al., 2008). Also, although short-term exposure (<4 years) to organic enrichment may not affect seaweeds directly, indirect effects such as turbidity may significantly affect photosynthesis (Read 1983), and result in reduced growth and reproduction and increased competition form fast growing but ephemeral species.

Light availability and water turbidity are principal factors in determining the depth range at which macro-algae can be found (Birkett et al., 1998b). Light penetration influences the maximum depth at which laminarians can grow, and it has been reported that laminarians grow at depths at which the light levels are reduced to 1% of incident light at the surface. Maximal depth distribution of laminarians therefore varies from 100 m in the Mediterranean to only 6 to 7 m in the silt-laden German Bight. In Atlantic European waters, the depth limit is typically 35 m. In very turbid waters, the depth at which kelp is found may be reduced, or in some cases excluded completely (e.g. Severn Estuary), because of the alteration in light attenuation by suspended sediment (Lüning, 1990; Birkett et al. 1998b). Laminarians show a decrease of 50% photosynthetic activity when turbidity increases by 0.1/m (light attenuation coefficient =0.1-0.2/m; Staehr & Wernberg, 2009).

Red algae are shade tolerant macroalgae. Phyllophora crispa is particularly shade tolerant and is recorded at greater depths than many other red algae. For example, Smith & Jones (1970) reported that Phyllophora crispa grew at a greater depth (25 m) than other red algae examined on the west coast of Anglesey and Norton (1968) reported it at 33 m at St Mary’s Isles of Scilly. Norton et al. (1971) noted that Phyllophora crispa penetrated up to 15 m in a sea cave (Bullock Island Cave, near Lough Ine), although its growth was stunted at its limit within the cave. The degradation of ‘Zernov’s Phyllophora field’ in the north-western Black Sea was attributed to eutrophication, algal blooms and turbidity. All three species of Phyllophora present survived but their biomass was reduced by an order to magnitude (Kostylev et al., 2010).

Phyllophora crispa-dominated communities between Cape Kosa Severnaya and Cape Tolsty in the Black Sea have experienced significant changes since 1964 (Pankeeva & Mironova, 2022; Parkhomenko et al., 2024). Field data from surveys conducted in 1964, 1997, 2006, and 2017, were combined with historic hydrological and hydrochemical data from 1998 to 2021 and were analysed to identify what was driving these changes. Light limitation from increased total suspended matter (TSM) was most strongly linked to the changes in Phyllophora crispa biomass (Parkhomenko et al., 2024). Increased turbidity led to a decline in the biomass of the deeper (≥10 m) Phyllophora crispa community to near zero by 1997. This trend continued until Phyllophora crispa were no longer recorded at these depths by 2006. Only by 2017 did small, isolated patches re-establish, but with biomass levels ten times lower than in 1964, indicating limited recovery and persistent sensitivity to increased turbidity (Pankeeva & Mironova, 2022; Parkhomenko et al., 2024). Model outputs further showed that when TSM levels exceeded approx. 1.5 mg/L, Phyllophora crispa could grow only in shallow water (≤4 m), highlighting its high vulnerability to turbidity driven light limitation (Parkhomenko et al., 2024). Overall, the study concluded that declines in water transparency, rather than nutrient trends, were the dominant driver of long-term biomass loss and structural change in the Phyllophora crispa community.

Sensitivity Assessment. A decrease in turbidity is likely to support enhanced growth (and possible habitat expansion) and is therefore not considered in this assessment. An increase in water turbidity is likely to primarily affect photosynthesis, therefore growth and density of the characteristic species. Resistance to this pressure is assessed as ‘Low’ and resilience to as ‘High’ at the benchmark level due to the scale of the impact. Hence, this biotope is regarded as having a sensitivity of ‘Low‘.

Smothering and siltation rate changes (light)

Benchmark. ‘Light’ deposition of up to 5 cm of fine material added to the seabed in a single discrete event (Smothering pressure definition).

Evidence

Suspended Particle Matter (SPM) concentration has a positive linear relationship with sub-surface light attenuation (Kd) (Devlin et al., 2008). Also, although short-term exposure (<4 years) to organic enrichment may not affect seaweeds directly, indirect effects such as turbidity may significantly affect photosynthesis (Read 1983), and result in reduced growth and reproduction and increased competition form fast growing but ephemeral species.

Light availability and water turbidity are principal factors in determining the depth range at which macro-algae can be found (Birkett et al., 1998b). Light penetration influences the maximum depth at which laminarians can grow, and it has been reported that laminarians grow at depths at which the light levels are reduced to 1% of incident light at the surface. Maximal depth distribution of laminarians therefore varies from 100 m in the Mediterranean to only 6 to 7 m in the silt-laden German Bight. In Atlantic European waters, the depth limit is typically 35 m. In very turbid waters, the depth at which kelp is found may be reduced, or in some cases excluded completely (e.g. Severn Estuary), because of the alteration in light attenuation by suspended sediment (Lüning, 1990; Birkett et al. 1998b). Laminarians show a decrease of 50% photosynthetic activity when turbidity increases by 0.1/m (light attenuation coefficient =0.1-0.2/m; Staehr & Wernberg, 2009).

Red algae are shade tolerant macroalgae. Phyllophora crispa is particularly shade tolerant and is recorded at greater depths than many other red algae. For example, Smith & Jones (1970) reported that Phyllophora crispa grew at a greater depth (25 m) than other red algae examined on the west coast of Anglesey and Norton (1968) reported it at 33 m at St Mary’s Isles of Scilly. Norton et al. (1971) noted that Phyllophora crispa penetrated up to 15 m in a sea cave (Bullock Island Cave, near Lough Ine), although its growth was stunted at its limit within the cave. The degradation of ‘Zernov’s Phyllophora field’ in the north-western Black Sea was attributed to eutrophication, algal blooms and turbidity. All three species of Phyllophora present survived but their biomass was reduced by an order to magnitude (Kostylev et al., 2010).

Phyllophora crispa-dominated communities between Cape Kosa Severnaya and Cape Tolsty in the Black Sea have experienced significant changes since 1964 (Pankeeva & Mironova, 2022; Parkhomenko et al., 2024). Field data from surveys conducted in 1964, 1997, 2006, and 2017, were combined with historic hydrological and hydrochemical data from 1998 to 2021 and were analysed to identify what was driving these changes. Light limitation from increased total suspended matter (TSM) was most strongly linked to the changes in Phyllophora crispa biomass (Parkhomenko et al., 2024). Increased turbidity led to a decline in the biomass of the deeper (≥10 m) Phyllophora crispa community to near zero by 1997. This trend continued until Phyllophora crispa were no longer recorded at these depths by 2006. Only by 2017 did small, isolated patches re-establish, but with biomass levels ten times lower than in 1964, indicating limited recovery and persistent sensitivity to increased turbidity (Pankeeva & Mironova, 2022; Parkhomenko et al., 2024). Model outputs further showed that when TSM levels exceeded approx. 1.5 mg/L, Phyllophora crispa could grow only in shallow water (≤4 m), highlighting its high vulnerability to turbidity driven light limitation (Parkhomenko et al., 2024). Overall, the study concluded that declines in water transparency, rather than nutrient trends, were the dominant driver of long-term biomass loss and structural change in the Phyllophora crispa community.

Sensitivity Assessment. A decrease in turbidity is likely to support enhanced growth (and possible habitat expansion) and is therefore not considered in this assessment. An increase in water turbidity is likely to primarily affect photosynthesis, therefore growth and density of the characteristic species. Resistance to this pressure is assessed as ‘Low’ and resilience to as ‘High’ at the benchmark level due to the scale of the impact. Hence, this biotope is regarded as having a sensitivity of ‘Low‘.

Smothering and siltation rate changes (heavy)

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

Evidence

Smothering by sediment e.g. 30 cm material during a discrete event, is unlikely to damage Saccharina latissima sporophytes but may affect holdfast fauna, gametophyte survival, interfere with zoospore settlement and, therefore, recruitment processes (Moy & Christie, 2012). Given the short life expectancy of Saccharina latissima (2-4 years (Parke, 1948)), IR.LIR.KVS.SlatPhyVS is likely to be dependent on annual Saccharina latissima recruitment (Moy & Christie, 2012). Given the microscopic size of the gametophyte, 30cm of sediment could be expected to significantly inhibit growth. However, laboratory studies showed that kelp gametophytes can survive in darkness for between 6-16 months at 8°C and would probably survive smothering by a discrete event. Once returned to normal conditions the gametophytes resumed growth or maturation within 1 month (Dieck, 1993). Intolerance to this factor is likely to be higher during the peak periods of sporulation and/or spore settlement.

Mature Codium tomentosum thalli can grow up to 30 cm long (Marlin, 2015). 30cm of deposited sediment is likely to inundate mature thalli. During winter, thalli fragment and individuals are reduced to a holdfast. It is unknown whether retained sediment would inhibit growth if the holdfast was inundated the following spring. 

Psammechinus miliaris is quite small (typically up to 40 mm) and is likely to be inundated by 30 cm of sediment (Jackson, 2008). If unable to 'dig out' of the sediment, deposited sediment may cause mortality.

Phyllophora crispa fronds can grow to a length of 15 cm and Phyllophora pseudoceranoides can grow to a length of 10cm (Bunker et al., 2012). Deposition of 30 cm sediment is likely to completely inundate Phyllophora crispa and Phyllophora pseudoceranoides.

IR.LIR.KVS.Cod, IR.LIR.KVS.SlatPhyVS and IR.LIR.KVS.SlatPsaVS are classed as low energy habitats, and are therefore unlikely to experience >moderate tidal streams (>0.5 m/sec) or wave action.

Sediment could, therefore, remain within the host habitat and recovery rate would be related to sediment retention but will probably be dissipated within a year. Deposited sediments could affect macroalgae recruitment (Birkett et al., 1998) and the survival of Psammechinus miliaris.

Sensitivity assessment Deposition of 30 cm of sediment is likely to inundate all but large macroalgae, e.g. mature Saccharina lattisima, and cause mortality in Codium spp. and understorey red seaweeds. As the deposit may remain for some time (depending on local conditions) and mortality is likely. Resistance has been assessed as of ‘None’; resilience has been assessed as ‘Medium’. Sensitivity has been 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. 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; MacKenzie, 2024), depending on the study species and duration and intensity of exposure. There have been no studies investigating the effect of EMFs at the population or community level for benthic organisms. 

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

No evidence.

Introduction of light or shading

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

Evidence

Barrier to species movement

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

Evidence

Not relevant. This pressure is considered applicable to mobile species, e.g. fish and marine mammals rather than seabed habitats. Physical and hydrographic barriers may limit the dispersal of spores.  But spore 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. Collision from grounding vessels is addressed under abrasion above.

Visual disturbance

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

Evidence

Not relevant.

Genetic modification & translocation of indigenous species

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

Evidence

Saccharina latissima has shown significant regional acclimation to environmental conditions. Gerard & Dubois (1988) found Saccharina latissima sporophytes which were regularly exposed to ≥20°C could tolerate these high temperatures whereas sporophytes from other populations which rarely experience ≥17°C showed 100% mortality after 3 weeks of exposure to 20°C. It is, therefore, possible that transplanted eco-types of Saccharina latissima may react differently to environmental conditions that differ from those of their origin.

However, there was little evidence for translocation of any other characteristic species over significant geographic distances. Nor was there any evidence regarding the genetic modification or effects of translocation.

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

Little is currently known about diseases in kelp, or seaweeds in general, although various causative agents have been implicated (Gachon et al., 2010 cited in Kerrison et al., 2015). The bacteria Pseudoalterom spp. and Alteromonas spp. are known to be responsible for some diseases (Egan et al., 2014 cited in Kerrison et al., 2015), but in numerous cases, the agent has not been identified. The prevalence of endophytic infection is known to be high in wild kelp populations (Ellertsdóttir & Peters, 1997 cited in Kerrison et al., 2015), and so there are concerns that pathogens may be transplanted with seaweed stocks, infecting nearby natural seaweed beds, and as physicochemical stress is often a trigger for outbreaks in cultivated kelp (FAO, 2015 cited in Kerrison et al., 2015), climate change impacts such as rising seawater temperatures may in the future lead to more severe disease impacts.

Saccharina latissima may be infected by the microscopic brown alga Streblonema aecidioides. Infected algae show symptoms of Streblonema disease, i.e. alterations of the blade and stipe ranging from dark spots to heavy deformations and completely crippled thalli. Infection can reduce growth rates of host algae (Peters & Scaffelke, 1996). The marine fungi Eurychasma spp can also infect early life stages of Laminarians, however, the effects of infection are unknown (Müller et al., 1999).

Sensitivity assessment. The evidence of diseases in laminarians suggests that growth and, possibly, survival may be affected. Hence, resistance is assessed as ‘Medium’ to represent the possible loss of a proportion of the population, but with ‘Low’ confidence due to the lack of direct evidence of mortality in the dominant characteristic species. Hence, resilience is assessed as ‘High’ and sensitivity to the introduction of microbial pathogens is assessed as ‘Low’.

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

Removal of non-target species

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

Evidence

Direct, physical impacts from harvesting are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of non-target species on this biotope. Incidental removal of the key characterizing species and associated species would alter the character of the biotope. IR.LIR.KVS.Cod, IR.LIR.KVS.SlatPhyVS and IR.LIR.KVS.SlatPsaVS are characterized by a canopy of Saccharina lattisima. Saccharina lattisima provides a canopy under which a variety of red seaweeds grow, including, Phyllophora sp. (as in IR.LIR.KVS.SlatPhyVS). The loss of the canopy due to incidental removal as by-catch would, therefore, alter the character of the habitat and result in the loss of species richness. The ecological services such as primary and secondary production provided by these species would also be lost.

The main grazers of natural kelp forests are benthic invertebrates such as sea urchins, snails, abalone and small crustaceans. Natural kelp beds can be decimated by outbreaks of these grazers, although these may be prevented by top-down pressure from fishing (Johnson et al., 2013) or predators such as carnivorous fish or otters (Estes & Palmisano, 1974 and Davenport & Anderson, 2007 cited in Kerrison et al., 2015). The removal of sea urchins could lead to increases in kelp abundances (Miller et al., 2024b).

Incidental non-targeted catch (e.g. by trawls or dredges) could mobilise sediment, remove large kelp species, overturn boulders and cobbles and bury smaller seaweeds and cause high mortality within the affected area.

Sensitivity assessment. Resistance has been assessed as ‘None’, resilience as ‘High’, and Sensitivity as ‘Medium’.

The American slipper limpet, Crepidula fornicata

Evidence

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

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

Its density and ability to spread within and between sites (e.g., bays) depend on the availability of suitable habitat, competition with other species, larval retention within the site, human activities (e.g., dredging), and seasonal temperatures, particularly in the intertidal zone. For example, the Crepidula fornicata population in the Bay of Mont-Saint-Michel grew by 50% between 1996 and 2004, covering 25% of the area at high density (51–100% cover), aided by local oyster farming and shellfish dredging (Blanchard, 2009). However, in Arcachon Bay, France, Crepidula fornicata was limited to only 155 tonnes in 1999 and 312 tonnes in 2011 (De Montaudouin et al., 2001, 2018). It was confined to muddy sediments, which accounted for approximately 8% of the bay and were colonized by Zostera beds. These areas represented just 0.4% of the suspension feeder biomass compared to the oyster Magallana gigas in the bay, and there was no indication of increasing biomass over a 12-year period. In addition, benthic trawling was prohibited in the bay (De Montaudouin et al., 2001, 2018). As a result, De Montaudouin et al. (2018) concluded that Crepidula fornicata was not invasive in the Bay of Arcachon.

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

In the eastern Solent harbours of Portsmouth, Langstone, and Chichester, 75% to 98% of Crepidula larvae settled on dead Crepidula shells, while ~4% settled on stone, 2.5% on live Crepidula, 0.3% oyster shell, 0.6% cockle shell, 0.3% winkle shell and 0.1% periwinkle shell (Preston et al., 2020). In the Milford Harbour Waterway, the highest densities of Crepidula were found in areas of sediment with hard substrata, e.g., mixed fine sediment with shell, or gravel or both, while Crepidula density increased as gravel cover increased in the subtidal, the reverse was found in the intertidal (Bohn et al., 2015). However, gravel formed the base of most stacks of Crepidula in the intertidal, which suggested that initial colonization occurred on available hard substrata (i.e., gravel) in the absence of adult shells of Crepidula. The availability of hard substrata (e.g., gravel) may only restrict initial colonization as higher densities of Crepidula function as substrata for subsequent colonization (Thieltges et al., 2004; Blanchard, 2009). Bohn et al. (2015) also noted that Crepidula density was low in areas of homogenous fine sediment and absent in areas dominated by boulders.

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

Crepidula fornicata has been recorded from the lower intertidal to ~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., 2011b; OBIS, 2025; 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 approx.0.7 m) and was absent from high tidal levels (approx.1.8 m) in the Milford Harbour Waterway (Bohn et al., 2015).

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

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

High densities of Crepidula fornicata cause ecological impacts on sedimentary habitats. The species can smother the seabed in shallow bays, changing and modifying the habitat structure (Blanchard, 1997, 2009; Helmer et al., 2019; Tillin et al., 2020). At high densities, the species physically smothers the sediment, and the resultant build-up of silt, pseudofaeces, and faeces is deposited and trapped within the bed (Tillin et al. 2020, Fitzgerald, 2007, Blanchard, 2009, Stiger-Pouvreau & Thouzeau, 2015). The biodeposition rates of Crepidula are extremely high and once deposited, form an anoxic mud, making the environment suitable for other species, including most infauna (Stiger-Pouvreau & Thouzeau, 2015, Blanchard, 2009). For example, in fine sands, the community is replaced by a reef of slipper limpets, that provide hard substrata for sessile suspension-feeders (e.g., sea squirts, tube worms and fixed shellfish), while mobile carnivorous microfauna occupy species between or within shells, resulting in a homogeneous Crepidula dominated habitat (Blanchard, 2009). Blanchard (2009) suggested the transition occurred and became irreversible at 50% cover of the limpet. De Montaudouin et al. (2018) suggested that homogenization occurred above a threshold of 20 to 50 Crepidula /m². However, Blanchard (2009) noted that sandy areas in the Bay of Mont Saint-Michel were not colonized by Crepidula due to sediment mobility, although adjacent areas were colonized. Thieltges et al. (2003) noted that storm events removed some clumps of mussels and presumably Crepidula onto tidal flats where they disappeared, which caused their abundance to fluctuate.

Impacts on the structure of benthic communities will depend on the type of habitat that Crepidula colonizes. De Montaudouin & Sauriau (1999) reported that in muddy sediment dominated by deposit-feeders, species richness, abundance and biomass increased in the presence of high densities of Crepidula (~562 to 4772 ind./m²), in the Bay of Marennes-Oléron, presumably because the Crepidula bed provided hard substrata in an otherwise sedimentary habitat. In medium sands, Crepidula density was moderate (330 to 1300 ind./m²) but there was no significant difference between communities in the presence of Crepidula. Intertidal coarse sediment was less suitable for Crepidula with only moderate or low abundances (11 ind./m²) and its presence did not affect the abundance or diversity of macrofauna. However, there was a higher abundance of suspension–feeders and mobile Crustacea in the absence of Crepidula (De Montaudouin & Sauriau, 1999). The presence of Crepidula as an ecosystem engineer has created a range of new niche habitats, reducing biodiversity as it modifies habitats (Fitzgerald, 2007). De Montaudouin et al. (1999) concluded that Crepidula did not influence macroinvertebrate diversity or density significantly under experimental conditions, on fine sands in Arcachon Bay, France. De Montaudouin et al. (2018) noted that the limpet reef increased the species diversity in the bed, but homogenised diversity compared to areas where the limpets were absent. In the Milford Haven Waterway, the highest densities of Crepidula were found in areas of sediment with hard substrata, e.g., mixed fine sediment with shell or gravel or both but, while Crepidula density increased as gravel cover increased in the subtidal, the reverse was found in the intertidal (Bohn et al., 2015). Bohn et al. (2015) suggested that high densities of Crepidula in high-energy environments were possible in the subtidal but not the intertidal. Hinz et al. (2011b) reported a substantial increase in the occurrence of Crepidula off the Isle of Wight, between 1958 and 2006, at a depth of ~60 m, on hard substrata (gravel, cobbles, and boulders), swept by strong tidal streams. Presumably, Crepidula is more tolerant of tidal flow than the oscillatory flow caused by wave action (Tillin et al., 2020).

Crepidula creates more muddy substrata, this impacts the larval settlement and survival of other species such as the King scallop (Pecten maximus) and Queen scallop (Aequipecten opercularis), causing a decrease in stocks (Stiger-Pouvreau & Thouzeau, 2015). This impact is more significant to the environment in more densely colonized areas (Blanchard, 2009).

Crepidula invasion on sediment also affects the hydrodynamics and transport properties of the benthic boundary layer. Results have suggested that seabed erosion and velocity measurements of flows over an artificial Crepidula shell bed decreases as roughness density increased, suggesting a sheltering effect by the shells (Stiger-Pouvreau & Thouzeau, 2015). Higher particle resuspension was observed in the study in muddy sand substrates with few stacks of Crepidula when compared with higher density areas (Stiger-Pouvreau & Thouzeau, 2015).

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

Sensitivity assessment. There is currently a lack of evidence of Crepidula fornicata colonization on bedrock in the infralittoral or circalittoral. Tillin et al. (2020) suggested that Crepidula could colonize circalittoral rock due to its presence on tide-swept rough grounds in the English Channel. At present, there is insufficient evidence to suggest that this biotope is sensitive to colonisation by Crepidula fornicata.

The carpet sea squirt, Didemnum vexillum

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions worldwide (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009).

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

Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024).

In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Minchin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

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

While natural fragmentation of tendrils is thought to allow Didemnum vexillum to invade longer distances and increase its dispersal potential, Stefaniak & Whitlatch (2014) found that only one tendril out of 80 reattached to the flat, bare substrata used in their study, because tendrils required an extensive (at least eight 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; Griffiths et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009). For example, in oyster farms in British Columbia, large fragments of Didemnum sp. come off oyster strings when they are pulled out of water, and other fragments can be pulled off oysters and mussels and thrown back into the water, which is likely to aid dispersal of the invasive species (Bullard et al., 2007). Dijkstra et al. (2007) hypothesised that Didemnum sp. was introduced to the Gulf of Maine with oyster aquaculture in the Damariscotta River and transported via Pacific oysters.

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

Didemnum vexillum has a seasonal growth cycle that is influenced by temperature (Valentine et al., 2007a). In warmer months (June and July), colonies may be large and well-developed encrusting mats. Populations experience more rapid growth from July to September, sometimes continuing into December. Colonies begin to decline in health and ‘die off’ when temperatures drop below 5°C during winter months from around October to April (Gittenberger, 2007; Valentine et al., 2007a; Herborg et al., 2009). Cold winter months cause colonies to regress and reduce in size, yet they often regenerate as temperatures warm (Griffith et al., 2009; Kleeman, 2009; Mercer et al., 2009), although some populations may not survive winter at all (Dijkstra et al., 2007). The early growth phase, from May to July, is initiated by smaller colonies developing from remnants of colonies that survived the cold winter (Valentine et al., 2007a). The seasonal growth cycle is also likely influenced by location. For example, the Didemnum sp. growth cycle for colonies in Sandwich tide pool (temperature range from -1°C to 24°C, with daily fluctuations), probably does not occur in deep offshore subtidal habitats in Georges Bank (annual temperature range from 4°C to 15°C, and daily fluctuations are minimal) (Valentine et al., 2007a). Larval release and recruitment typically occur between 14 and 20°C and slow or cease below 9 to 11°C as summer ends (Griffith et al., 2009; Mckenzie et al., 2017). In New Zealand, recruitment occurs from November to July, where the highest average temperatures were recorded in February (18 to 22°C) and the lowest in July (9 to 10°C) (Fletcher et al., 2013a). In this New Zealand study, higher water temperatures were associated with 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). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems by its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock, gravel, pebble, cobble, or boulders (Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020).

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).

Didemnum vexillum has been found colonizing the stipes of Laminaria spp. in the Gulf of Maine (Dijkstra et al., 2007) and in Norway (Legrand et al., 2025). However, it has not been recorded in sites exposed to wave action, that is, 'very wave exposed', 'wave exposed' and 'moderately wave exposed' (sensu MNCR, Hiscock, 1996), especially in the intertidal, where wave action is not ameliorated by depth (see Hiscock, 1983).

This species requires suitable hard substrata for successful settlement and the establishment of colonies. It can grow quickly and can establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems by its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock, gravel, pebble, cobble, or boulders (Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020). It is therefore likely that the substrate in this biotope is suitable for Didemnum vexillum colonisation. In addition, the depth range at which Laminaria hyperborea biotopes are found (0 to 30 m) overlaps with the depth range that is suitable for Didemnum vexillum colonization. Didemnum vexillum has been recorded from less than 1 m to at least 81 m deep (Bullard et al., 2007; Tagliapietra et al., 2012; Tillin et al., 2020).

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

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

Reinhart et al. (2012) examined the effects of water flow and hydrodynamics on the encrusting and tendril forms of Didemnum vexillum. They reported that a current speed of approx. 7.6 m/s was required to induce fragmentation of tendrils, but that natural tidal flow alone was insufficient to cause fragmentation of tendrils. They suggested that rare instances of wave action, such as storms that resulted in wave orbital velocities of ca 8 m/s or (more likely) human activity, could cause fragmentation of tendrils.

Reinhart et al. (2012) noted that the tensile strength of Didemnum vexillum was an order of magnitude higher than that of Botrylloides sp. and was similar to that of Alyconium digitatum. Alyconium digitatum is reported from sheltered to very wave-exposed conditions, but in the sublittoral. Reinhart et al. (2012) also suggested that seasonal changes in the condition of Didemnum vexillum reduced the tensile strength of colonies and were associated with the period of greater larval production, and implied that fragmentation aided dispersal.

Considering other epiphytes, such as bryozoans, mussels, and other seaweeds, their overgrowth of the cultivated thalli on Saccharina latissima has led to its degradation and tissue loss during late spring or early summer (Marinho, Holdt & Angelidaki, 2015, cited in Kerrison et al., 2015; Kerrison et al., 2015).

Sensitivity assessment. There is no evidence of Didemnum vexillum colonizing this biotope in the UK. However, it has been recorded in similar kelp habitats in Norway (Järnegren et al., 2023). Didemnum vexillum requires hard substrata for successful colonization, therefore, it could colonize the bedrock and boulders that characterize this biotope. Didemnum vexillum can overgrow sessile organisms, including kelp Laminaria sp. However, no direct evidence was found on how Didemnum vexillum affects kelp or if it contributes to Laminaria sp. mortality (Järnegren et al., 2023), although epifaunal growth by Membranipora membrancea was reported to reduce the physical strength of kelp fronds (inc. Laminaria digitata) and make them susceptible to removal by wave action (Krumhansl et al., 2011). In addition, overgrowth by epiphytes contributed to the decline of Saccharina latissima in Norway (Andersen et al., 2011). However, Didemnum vexillum may compete for light and space with kelp and epifauna and could interfere with recruitment, which could lead to the mortality of some epifauna, the loss of kelp, and a reduction in biodiversity. In addition, the sheltered conditions of this biotope are also suitable for Didemnum vexillum. 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'. However, confidence in the assessment is ‘Low’ due to the lack of direct evidence of damage to kelp beds.

The Pacific oyster, Magallana gigas

Evidence

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

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). It was also introduced to the northeast Adriatic Sea (Ezgeta-Balic et al., 2019) and southwest England from France, possibly via fouling on ships (GBNNSIP, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).

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, 2011b, 2012a; Ezgeta-Balic et al., 2019; Spagnolo et al., 2019; Bergstrom et al., 2021; Hansen et al., 2023). In the UK, the species occurs predominantly around the southern and western coastlines (OBIS, 2025; 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 by fouling on ships (GBNNSIP, 2011b, 2012a; Padilla, 2010; Ezgeta-Balic et al., 2019).

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

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

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

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

Populations of Magallana gigas have been found on wave-exposed rocky shores to wave-sheltered soft sediment environments, and it has been described as a habitat generalist (Troost, 2010; Kochmann et al., 2012, 2013). For example, in Scotland, wild Magallana gigas are mainly located in the lower intertidal on bedrock, bedrock encrusted with barnacles, within bedrock crevices, and large and small boulders (Cook et al., 2014). Patches of Pacific oyster reefs have been recorded on littoral rock in Kent, southern England and on littoral sediments in southern England, the North Sea, and the English Channel (Herbert et al., 2012, 2016; Morgan et al., 2021).

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

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

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

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

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

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

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

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

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

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

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

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

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

Sensitivity assessment. While most of the evidence suggests the environmental conditions within this biotope are suitable for Magallana gigas, it is unlikely that they would be able to colonize this biotope without the removal of the characterising macroalgal species. Therefore, this biotope is assessed as ‘Not Sensitive’ to this pressure.

Wireweed, Sargassum muticum

Evidence

Competition with invasive macroalgae may be a potential threat to this biotope (de Bettignies et al., 2021). Sargassum muticum is a circumglobal invasive species (Engelen et al., 2015). It is recorded (2015) from Norway to Morocco and into the Mediterranean in the eastern Atlantic and from Alaska to Baja California in the eastern Pacific and from southern Russia to southern China in the western Pacific (Engelen et al., 2015). It colonizes a variety of habitats and can tolerate -1°C to 30°C and survive salinities below 10 PSU. Although fertilization does not occur below 15 PSU and growth of germlings is limited below 10°C, it can complete its life cycle as long as temperatures are over 8°C for at least four months of the year (Engelen et al., 2015). However, its distribution is limited by the availability of hard substratum (e.g. stones >10 cm) and light (Staehr et al., 2000; Strong & Dring, 2011; Engelen et al., 2015). It is most abundant between 1 and 3 m below mean water, but it has been recorded at 18 m or 30 m in the clear waters of California. However, it is a poor competitor under low light and only develops dense canopies in shallow areas (Engelen et al., 2015).

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

Strong & Dring (2011) used canopy removal experiments to investigate inter- and intra-species competition between Sargassum muticum and Saccharina latissima in the Dorn, Strangford Lough, N. Ireland. The Dorn consists of tidal pools, very sheltered from wave action but with moderately strong tidal streams (1-2 knots). Sargassum muticum grew better in mixed stands with Saccharina latissima than in the highest-density monospecific stands examined. However, the growth of Saccharina was not affected by the proportion of Sargassum in mixed stands. They concluded that Saccharina was not impacted significantly by the alien species, while Sargassum benefited from growth in mixed stands. Experimental manipulation of subtidal algal canopies in San Juan Islands, Washington State, USA, showed that Sargassum muticum reduced the abundance of native macroalgae, including the kelp Laminaria bongardiana, due to shading. However, experimental removal of Sargassum resulted in the recovery of native species within about one year (Britton-Simmons, 2004; Engelen et al., 2015). The negative effects of Sargassum muticum on native macroalgae are mainly due to competition for light, rather than changes in nutrient availability, sedimentation or water flow (Britton-Simmons, 2004; Engelen et al., 2015).

Sensitivity assessment. The above evidence suggests that Sargassum muticum can both compete with and co-exist with Saccharina latissima, depending on local conditions. For example, Sargassum muticum out-competed Saccharina latissima in the Limfjorden but coexisted in the Dorn in Strangford Lough. The evidence above suggests that Sargassum also prefers wave-sheltered conditions and shallow water (approx. 1 to 4 m depth). Therefore, Sargassum muticum is likely to threaten the biotope. Sargassum muticum may either co-exist with or out-compete Saccharina latissima, resulting in a potentially significant (25 to 75%) reduction in the abundance or extent of the native kelp and a possible decrease in the diversity of other macroalgae. Therefore, resistance is assessed as ‘Low’. Recovery after invasion by Sargassum, would require direct intervention (removal), so resilience is assessed as ‘Very low’. Hence, the sensitivity of the biotope is assessed as ‘High’. Overall, confidence is assessed as ‘Low’ due to evidence of variation and the site-specific nature of competition between native kelps and Sargassum muticum.

Wakame, Undaria pinnatifida

Evidence

Competition with invasive macroalgae may be a potential threat to this biotope (de Bettignies et al., 2021). Undaria pinnatifida (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; Hieser et al., 2014; Arnold et al., 2016; Epstein & Smale, 2017; Epstein & Smale, 2018; Kraan, 2017; Epstein et al., 2019a,b; Tidbury, 2020). Undaria pinnatifida originates from Japan but is currently established on the coastlines of New Zealand, Australia, Northern France, Spain, Italy, the UK, Portugal, Belgium, Holland, Argentina, Mexico, and the USA (De Leij et al., 2017). Undaria pinnatifida was first recorded in the UK in the Hamble Estuary in 1994 (Macleod et al., 2016). It has since proliferated along UK coastlines. One year after its discovery at the Queen Anne Battery marina, Plymouth, it had become a major fouling plant on pontoons (Minchin & Nunn, 2014). Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound.

Undaria pinnatifida seems to settle better on artificial substrata (e.g. floats, marinas, or piers) than on natural rocky shores among local kelps (Vaz-Pinto et al., 2014). It is found predominantly in low intertidal to shallow subtidal habitats (Epstein et al., 2019b) and is significantly more abundant on artificial substrata compared to natural rocky substrata (Heiser et al., 2014; Epstein & Smale, 2018). James (2017) suggested that Undaria pinnatifida could out-compete native species on artificial substrata (such as marinas and wharf structures). In Plymouth, UK, De Leij et al. (2017) found that natural habitats with dense native macroalgal canopies, such as Laminaria hyperborea, Laminaria ochroleuca, Laminaria digitata, and Saccharina latissima, had more resistance to Undaria pinnatifida invasion than disturbed or sparse canopies, due to limited space and light availability for Undaria pinnatifida recruits. However, the dense canopies did not always prevent the invasion of Undaria pinnatifida as sporophytes were still recorded within dense Laminaria canopies, so canopy disturbance was not always required (De Leij et al., 2017; Epstein & Smale, 2018).

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

Undaria pinnatifida has a wide physiological niche, meaning it can occur in both coastal and estuarine environments, showing tolerance for varying salinities, turbidity, and siltation (Heiser et al., 2014; Epstein & Smale, 2018). Undaria pinnatifida can inhabit a broad range of habitats, including reefs; coastal brackish/saline lagoons; large shallow inlets and bays; estuaries; estuarine rocky habitats; natural or near-natural estuary; coastal lagoons; and tidal rivers, estuaries, mudflats, sandflats and lagoons (James 2017). Undaria pinnatifida prefers sites sheltered with low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018). In natural habitats, Undaria pinnatifida was not recorded if the wave fetch was greater than 642 km but increased in abundance and cover in very sheltered sites (Epstein & Smale, 2018).

In Plymouth Sound (UK), Epstein et al. (2019b) found that within its depth range (+1 to –4 m), Undaria pinnatifida co-existed with seven species of canopy-forming brown macroalgae, including Saccharina latissima. However, they reported that Undaria pinnatifida biomass was negatively related to Saccharina latissima in both intertidal and subtidal habitats. This was only statistically significant in subtidal habitats, which suggested that there was some competition between the two species (Epstein et al., 2019b). Heiser et al. (2014) surveyed 17 sites within Plymouth Sound, UK and found that Saccharina latissima was significantly more abundant at sites with Undaria pinnatifida, with ca 5 Saccharina latissima individuals present per m², compared to ca 0.5 Saccharina latissima individuals per m² present at sites without Undaria pinnatifida.

Undaria pinnatifida has been reported to both co-exist with and out-compete Saccharina latissima (Farrell & Fletcher, 2006; Heiser et al., 2014; Epstein et al., 2019b). For example, in Torquay Marina, UK, Farrell & Fletcher (2006) completed a canopy removal experiment between 1996 and 2002. They reported that Saccharina latissima decreased in both control and treatment plots from ca 3 plants per 0.45 m² in 1996 to ca 1 plant per 0.45 m² in 1997 and had disappeared completely from pontoons by 2002. This coincided with a significant increase in Undaria pinnatifida from zero plants per 0.45 m² in 1996 to ca 6 plants per 0.45 m² in 1997. However, there was a slight decrease in Undaria pinnatifida in both control and treatment plots between 1997 and 1998. By 2002, Undaria pinnatifida had recovered at control and treatment plots to ca 4-6 plants per 0.45 m², whereas Saccharina latissima had not.

Undaria pinnatifida was successfully eradicated on a sunken ship in Clatham Islands, New Zealand, by applying a heat treatment of 70°C (Wotton et al., 2004). However, numerous other eradication attempts have failed and as noted by Fletcher & Farrell (1998), once established, Undaria pinnatifida resists most attempts at long-term removal.

The proliferation of Undaria pinnatifida and competition with native species may cause a reduction in local biodiversity (Valentine & Johnson, 2003; Vaz-Pinto et al., 2014; Arnold et al., 2016; Teagle, 2017; Tidbury, 2020). A shift towards Undaria pinnatifida-dominated beds could result in diminished epibiotic assemblages and lower local biodiversity compared with assemblages associated with native perennial kelp species, such as Laminaria spp. and Saccharina latissima (Arnold et al., 2016; Teagle et al., 2017). In Plymouth, UK, Arnold et al. (2016) found that Undaria pinnatifida supported less than half the number of taxa and had no unique epibionts compared to Laminaria ochroleuca and Saccharina latissima (Arnold et al., 2016).

Sensitivity assessment. The above evidence suggests that Undaria pinnatifida can both compete with and co-exist with Saccharina latissima, depending on local conditions. For example, Undaria pinnatifida can out-compete Saccharina latissima in artificial habitats, such as in Torquay Marina, but within natural habitats, it can co-exist with native kelp species within its depth range (-1 to 4 m), as shown in Plymouth Sound, UK.

The evidence above suggests that Undaria prefers sheltered conditions, in the shallow subtidal and sublittoral fringe (approx. +1 to 4 m in depth). Undaria pinnatifida may therefore either co-exist with or out-compete Saccharina latissima, resulting in a potentially significant (25-75%) reduction in the abundance or extent of the native kelp and a possible decrease in the diversity of other macroalgae. Therefore, resistance is assessed as ‘Low’. Recovery after invasion by Undaria would require direct intervention (removal) so resilience is assessed as ‘Very low’. Therefore, the sensitivity of the biotope is assessed as ‘High’. Overall, confidence is assessed as ‘Low’ due to evidence of variation and the site-specific nature of competition between native kelps and Undaria pinnatifida.

Other INIS

Evidence

The golden kelp Laminaria ochroleuca is a warm-temperate Lusitanian kelp with a distribution ranging from Morocco to the south of the UK. It was first recorded in the southwest UK in 1946 (Parke, 1948) and is projected to expand further northwards under future climate change scenarios (Franco et al., 2018). A small population was recorded in northwest Ireland in 2018 (Schoenrock et al., 2019), further suggesting ongoing poleward expansion. In Plymouth Sound, southwest UK, estimates of Laminaria ochroleuca standing stock are now comparable to those of the native kelp Laminaria hyperborea (Taylor-Robinson et al., 2024; also see Smale et al., 2016 for the standing stock of Laminaria hyperborea). While not considered a traditional invasive species, its northward expansion into the UK has led to competition with native kelps such as Laminaria hyperborea, and since the environmental range of Saccharina latissima overlaps with Laminaria hyperborea, Laminaria ochroleuca can be considered as a potential invasive competitor. No evidence could be found on the interactions of Laminaria ochroleuca on Saccharina latissima, so evidence for Laminaria hyperborea is presented instead.

It is suggested that Laminaria ochroleuca may have a competitive advantage over Laminaria hyperborea due to its tolerance of warmer waters. Barrientos et al. (2025) investigated changes in kelp forests in northwest Spain between 1997 and 2023. They found that kelp forests had disappeared or severely declined in density at 29 of 50 sites, and the canopy was now dominated by Laminaria ochroleuca at the surviving sites, while Laminaria hyperborea is almost entirely absent, occurring at only two sites. These changes were linked to sea surface temperature (an average increase of 0.01 to 0.02°C per year over the 26-year study period), which suggested that Laminaria ochroleuca was more resistant to warming and could, therefore, outcompete Laminaria hyperborea under global warming scenarios.

Saccharina latissima grows well between 5 and 17°C (Druehl, 1967, Fortes & Luning, 1980 and Machalek, Davison & Falkowski, 1996 cited in Kerrison et al., 2015) and has a limiting temperature isotherm of 19 to 20°C (Müller et al., 2009). However, temperature ecotypes exist, having adapted to high seasonal temperature exposure: populations from Helgoland, Germany, can tolerate temperatures of 18 to 20°C (Davison, 1987 cited in Kerrison et al., 2015), while populations in New York, USA, can survive at >20°C, albeit with substantially reduced growth (Gerard & Du Bois, 1988). In addition, Azevedo et al. (2016) cultured Saccharina latissima in tanks in northwest Portugal throughout the summer, withstanding average temperatures around 20°C from May onwards (temperature varied between 11.7°C in April and 24.9°C in August), well above published optimum temperatures for this species (10 to 15°C). Therefore, Saccharina latissima may be able to compete alongside Laminaria ochroleuca better than Laminaria hyperborea, but both native species share a similar geographic range from Portugal to the Arctic (Birkett et al., 1998b; Azevedo et al., 2016).

There is contrasting evidence on the relative resilience of Laminaria ochroleuca, Laminaria hyperborea, and Saccharina latissima to storm damage. Pereira et al. (2017) reported no recovery of Laminaria hyperborea populations in the two years following a storm in northern Portugal, whereas Laminaria ochroleuca showed partial recovery. In contrast, during the Northeast Atlantic storm season of 2013 to 2014, the south coast of the UK was subjected to some of the most intense storms in recent history, being classed as a '1-in-30 year' event, where inshore significant wave heights and periods exceeded 7 m and 13 seconds (Smale & Vance, 2015). Overall, kelp canopies were highly resistant to storm disturbance, however, at one study site, a mixed canopy comprising Laminaria ochroleuca, Saccharina latissima, and Laminaria hyperborea was significantly altered by the storms, due to a decreased abundance of the former two species (Smale & Vance, 2015). In particular, the breakage of mature Laminaria hyperborea stipes ranged between 2.3 and 6.9%, while broken Laminaria ochroleuca stipes were on average 8.7 times more prevalent (Smale & Vance, 2015). Given this conflicting evidence, it remains unclear whether Laminaria ochroleuca biotopes could displace Laminaria hyperborea and Saccharina latissima biotopes following storm events.

Another potential advantage of Laminaria ochroleuca is its greater average stipe length compared to Laminaria hyperborea, potentially reducing light availability for Laminaria hyperborea recruits in mixed-population forests (Smale et al., 2014). This shading effect may exaggerate the impacts of marine heatwaves on Laminaria hyperborea, as elevated temperatures increase metabolic demands that cannot be met under light-limited conditions (Bass et al., 2023). However, in terms of Saccharina latissima, Laminaria ochroleuca, on average, is shorter (1.5 m compared to 4 m – see Laminaria ochroleuca and Saccharina latissima), and may be at a lower risk of shading.

The introduction of Laminaria ochroleuca into Laminaria hyperborea or Saccharina latissima forests can have negative impacts on biodiversity. Kelp stipe assemblages differ significantly between Laminaria ochroleuca and Laminaria hyperborea due to the texture of the stipe. Laminaria hyperborea stipes are rough and pitted and, therefore, have a larger surface area, while Laminaria ochroleuca stipes are uniformly smooth. Teagle & Smale (2018) found species from up to 15 different taxonomic groups on Laminaria hyperborea stipes in spring, compared to 2 taxa at most on Laminaria ochroleuca stipes all year round. In addition, the biomass of Laminaria hyperborea stipe assemblages was >3,600 more than that of Laminaria ochroleuca stipe assemblages. Therefore, the proliferation of Laminaria ochroleuca could reduce available habitat space for epibionts that are associated with Laminaria hyperborea biotopes, and potentially Saccharina latissima biotopes as well, as 111 taxa have previously been recorded on Arctic Saccharina latissima individuals (Shunatova et al., 2018 cited in Diehl et al., 2024).

Bonnemaisonia hamifera (and the Trailliella-phase) is a non-native red algae introduced to the British Isles from Japan and first recorded in 1890 (Dixon & Irvine, 1977; Maggs & Stegenga, 1998; Gollasch, 2006). It is thought to have been introduced by shipping or with shellfish and to have dispersed by drifting on water currents (Gollasch, 2006). Bonnemaisonia hamifera (and the Trailliella-phase) has spread around the British Isles and Europe, into the Mediterranean and the Canary Isles and north to the Orkneys and the Norwegian coast (Lüning, 1990, Maggs & Stegenga, 1998; Gollasch, 2006). It grows rapidly, reproduces vegetatively, and can spread by fragmentation and drifting (Maggs & Stegenga, 1998).

Bonnemaisonia hamifera (and the Trailliella-phase) occurs in biotopes with Phyllophora crispa but at a lower abundance than its characteristic biotope (SS.SMp.KSwSS.Tra). SS.SMp.KSwSS.Tra occurs in shallower waters but otherwise similar conditions. Therefore, if the abundance of Phyllophora crispa was reduced by an external factor then the Trailliella might be able to take over the available space, especially in the shallow examples of the biotope. However, there is No evidence that this has happened to date. Hence, there is insufficient evidence on which to base an assessment.

Sensitivity Assessment. Considering Laminaria hyperborea in the place of Saccharina latissima, and the evidence for the poleward range shift for Laminaria hyperborea and Saccharina latissima (Moy & Christie, 2012; Assis et al., 2016; Casado-Amezúa et al., 2019; Simkanin et al., 2005 cited in Veenhof et al., 2024), alongside the expansion of Laminaria ochroleuca into higher latitudes (Franco et al., 2018), Laminaria ochroleuca could displace existing kelp biotopes in the southern UK. In Plymouth Sound, Laminaria ochroleuca is already rivalling Laminaria hyperborea, which used to be the dominant kelp in the area (Saccharina latissima is also common in this region) (Smale et al., 2014; Taylor-Robinson et al., 2024). Its stipe length could reduce light availability for smaller and/or juvenile kelps, and when combined with elevated temperatures, could create unfavourable conditions for the persistence and recovery of native species. Laminaria ochroleuca, however, does form mixed forests with Laminaria hyperborea and Saccharina latissima in moderately sheltered to exposed shores, and has physiological and morphological advantages that could allow it to proliferate if Laminaria hyperborea and/or Saccharina latissima density was reduced. Resistance to Laminaria ochroleuca is assessed as ‘Low’ based on the evidence of Laminaria ochroleuca rivalling Laminaria hyperborea in Plymouth Sound, southwest UK, and potentially rivalling Saccharina latissima. Hence, resilience is assessed as ‘Very Low’, and sensitivity as ‘High’. While the quality and applicability of the evidence is high, there is contrasting evidence regarding both species’ resistance and resilience to storm damage. Therefore, confidence in this sensitivity assessment is ‘Medium’.

Regarding Bonnemaisonia hamifera, if the abundance of Phyllophora crispa was reduced by an external factor then the Trailliella-phase of this species may be able to take over the available space, especially in the shallow examples of the biotope. However, there is No evidence that this has happened to date. Hence, there is insufficient evidence on which to base an assessment for this INIS.