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

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.

Laminaria hyperborea

Laminaria hyperborea is a cold-temperate species of kelp with an optimum temperature for growth of 15°C, and an upper temperature limit of 21°C (Bolton & Lüning, 1982). Germination success can decrease by almost two thirds at temperatures as low as 17°C. According to OBIS (2026), Laminaria hyperborea occurs in temperatures from 5 to 15°C, with most records coming from the 10 to 15°C range. According to OBIS (2026), Laminaria hyperborea occurs in temperatures from 5 to 15°C, with most records coming from the 10 to 15°C range.

Kain (1964) stated that Laminaria hyperborea sporophyte growth and reproduction could occur within a temperature range of 0 to 20°C. Upper and lower lethal temperatures were estimated at between 1 to 2°C above or below the extremes of this range (Birkett et al., 1988). Gamete survival is reduced above 17°C (Kain, 1964 and 1971) and gametogenesis is inhibited at 21°C (Dieck, 1992). It is, therefore, likely that Laminaria hyperborea recruitment would be impaired at a sustained temperature increase of above 17°C. Sporophytes, however, can tolerate slightly higher temperatures of 20°C. Temperature tolerances for Laminaria hyperborea are seasonally variable, with more sensitivity to temperature change in winter months than summer months (Birkett et al., 1998b).

Laminaria hyperborea is a boreal northern species with a geographic range from mid-Portugal to Northern Norway (Birkett et al., 1998b), and a mid-range within southern Norway (60° to 65° North) (Kain, 1971). The average seawater temperature for southern Norway in October is 12 to 13°C (Miller et al., 2009), and average annual sea temperature, from 1970 to 2014, is 8°C (Beszczynska-Möller & Dye, 2013). In Portugal and the southwest UK, Laminaria hyperborea is near the southern limit of its range where sea surface temperatures are closer to the upper thermal limit for this species. These populations are known as ‘trailing edge’ populations, where the species’ geographic range is contracting due to ocean warming. Trailing edge populations are known to be more sensitive to temperature increases than populations in the centre of their geographic range because they are already living close to or at the limit of their thermal tolerance (Smale, 2020; Hereward et al., 2020; Leathers et al., 2024).

Trailing edge Laminaria hyperborea populations assimilate less carbon than populations in colder waters (Pessarrodona et al., 2018) and therefore accumulate less biomass (Smale et al., 2016). In colder parts of the UK, lamina extension, regrowth, and carbon standing stock were 1.5, 2 and 3 times higher, respectively, than in warmer areas (Smale et al., 2020). Wernberg et al. (2025) observed differences in morphological features between populations at opposing ends of the species’ range, such as stipe height, lamina width, stipe diameter, lamina thickness, and the number of digits. They also found that stipe epiphyte load was far greater in the colder region than in the warmer region (Wernberg et al., 2025), most likely due to greater stipe surface area (Teagle & Smale, 2018).

Temperature increases beyond the thermal optimum for kelp can negatively affect photosynthesis in kelps. Photosynthetic efficiency (measured as F_V/F_M) is widely used for measuring physiological stress in photosynthetic organisms (Trautmann et al., 2024). Photosynthetic efficiency responses to heat spikes can vary by season, light availability, and by the degree of warming. Bass et al. (2023) observed a decline in average photosynthetic efficiency of 0.33 in high light conditions and 0.11 in low light conditions, with both values falling below 0.7. The biggest decline was observed in the 22°C treatment, while the control (18°C) and 20°C treatments showed no significant change in photosynthetic efficiency. This interactive effect was also observed by Diehl et al. (2024), where photosynthetic efficiency was reduced significantly only in the coldest (0°C) treatment combined with a long photoperiod (24:0 hours light:dark) treatment. Cold and long light conditions significantly decreased chlorophyll-a, accessory pigments and VAZ pigments, which indicates a photoprotective stress response. In the 10°C treatment, these pigments either decreased or showed no change, suggesting that the relatively higher temperature mitigated light stress. Dry weight increased significantly, despite no measurable change in surface area, when the highest temperature (10°C) treatment was combined with moderate (16:8 h) and long (24:0 h) photoperiods. This increase in dry weight was not detrimental to the kelp and was likely due to the accumulation of storage carbohydrates rather than growth. No significant responses were observed in phlorotannin (compounds that protect against light stress) levels. Mannitol (a storage carbohydrate) decreased under the long night treatment, but this effect is expected and not detrimental to the kelp. Laminarin (the other storage carbohydrate that was measured) increased significantly under both light treatments and the two warmer treatments (5°C and 10°C), which is a positive metabolic response.

The loss of Laminaria hyperborea in some parts of the UK has been attributed to increasing sea surface temperatures in the last several decades (Yesson et al., 2015b). A reduction in abundance was observed in 187 out of 496 sites between 1974 and 2010. Declines were recorded in the English Channel and West Channel and Celtic Sea, the southernmost regions in the study. The English Channel decline was strongly correlated to the sea surface temperature increase of 1 to 2°C in this period. Northernmost sites (around Scotland) were overall unchanged, while populations on the west coast of Ireland increased in abundance (Yesson et al., 2015b).

Between 1977 and 2007, many cool-water macroalgae including Laminaria hyperborea and many other species found along its distribution have almost disappeared and been replaced by warm-water macroalgae on the west coast of Asturias, northern Spain (Fernández, 2016). Moreover, only 21 out of 50 (42%) locations that were surveyed between 1997 and 2023 on the northwest coast of Spain still had dense kelp forests, of which nine were completely dominated by Laminaria hyperborea (Barrientos et al., 2025). Sea surface temperatures overall increased by around 0.01°C to 0.02°C per year in this region and across this period. In 2023, only eight of the 21 remaining dense forests still had the same canopy-forming species as they did in 1997. Laminaria hyperborea was no longer the dominant kelp in any of these sites, and it only persisted in two sites, which it shared with Laminaria ochroleuca. The persistence of these forests was strongly correlated with winter and summer sea surface temperatures as well as higher wave action (Barrientos et al., 2025). Globally, Laminaria hyperborea has experienced a range contraction of 14% between the 1980s and 2010s (Casado-Amezúa et al., 2019). This range loss is estimated to continue to up to 39.34% under the most extreme greenhouse gas emissions projections (Assis et al., 2016).

Saccharina latissima

Saccharina latissima has a latitudinal range of 41.3° South to 79.8° 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). According to OBIS (2026), Saccharina latissima occurs in temperatures from 5 to 15°C, with just over half of all records coming from the 10 to 15°C range. 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).

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.

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). These factors combined with establishment of annual filamentous algae in Skagerrak, Norway are likely to prevent the establishment of self-sustaining populations in the area (Anderson et al., 2011; Moy & Christie, 2012).

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/m³ (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).

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 sublethal 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.

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.

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 1-2 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).

Another feature of this biotope, subtidal red algae, are less tolerant of temperature extremes than intertidal red algae, surviving between -2°C and 18 to 23°C (Lüning 1990; Kain & Norton, 1990). Temperature increase may affect growth, recruitment or interfere with reproduction processes. For example, there is some evidence to suggest that blade growth in Delesseria sanguinea is delayed until ambient sea temperatures fall below 13°C. Blade growth is also likely to be intrinsically linked to gametangia development (Kain, 1987), and maintenance of sea temperatures above 13°C may affect recruitment success.

Warming can indirectly affect this biotope through cascading effects in the food chain. For example, it is suggested that the decline of cod (Gadus morhua) in the northeast Atlantic due to fishing, combined with increased sea surface temperatures in the past 50 years, has led to the spread of crabs (Cancer pagurus and Carcinus maenas) into this region. In addition, king crabs have spread into the northeast Atlantic since their introduction to Russia from the Pacific in the 1960s. These species are known predators of urchins, and this is believed to be a contributing factor in the recovery of kelp forests in this region (Christie et al., 2019).

Echinus esculentus

Warming can indirectly affect this biotope through cascading effects in the food chain. For example, it is suggested that the decline of cod (Gadus morhua) in the northeast Atlantic due to fishing, combined with increased sea surface temperatures in the past 50 years, has led to the spread of crabs (Cancer pagurus and Carcinus maenas) into this region. In addition, king crabs have spread into the northeast Atlantic since their introduction to Russia from the Pacific in the 1960s. These species are known predators of urchins, and this is believed to be a contributing factor in the recovery of kelp forests in this region (Christie et al., 2019).

Bishop (1985) suggested that Echinus esculentus cannot tolerate high temperatures for prolonged periods due to increased respiration rate and resultant metabolic stress. Ursin (1960) reported Echinus esculentus occurred at temperatures between 0 to 18°C in Limfjord, Denmark. Bishop (1985) noted that gametogenesis occurred at 11 to 19°C however, continued exposure to 19°C disrupted gametogenesis. Embryos and larvae developed abnormally after 24hr exposure to 15°C but normally at 4, 7 and 11°C (Tyler & Young 1998).

Marine Heatwaves

Marine heatwaves are extreme weather events defined as periods of extreme sea surface temperature that persists for days to months (Frölicher et al., 2018). Marine heatwaves are predicted to occur more frequently, last for longer and at increased intensity by the end of this century under both middle and high emission scenarios (Frölicher et al., 2018). Marine heatwaves are known to cause significant impacts to kelp forests, particularly if a population is found towards the edge of its southern limit (Smale et al., 2019). 

Therefore, it is expected that like other kelp species, Laminaria hyperborea will be highly sensitive to marine heatwaves.

Temperature tolerances for Laminaria hyperborea are seasonally variable, with more sensitivity to temperature change in winter months than summer months (Birkett et al., 1998). If marine heatwaves (MHWs) occurred during the in autumn and winter (Jacobs et al., 2024) during the reproductive phase, the success of recruitment could be reduced.

Trailing edge populations are known to be more sensitive to temperature increases than populations in the centre of their geographic range because they are already living close to or at the limit of their thermal tolerance (Smale, 2020; Hereward et al., 2020; Leathers et al., 2024). The effects of MHWs on this biotope could therefore differ between locations, i.e., in the southwest UK, this biotope would be more sensitive to MHWs than it would be if an MHW of the same duration and intensity occurred in the north of the UK.

Burdett et al. (2019) found that simulated heat spikes (+2°C and +4°C) for three days had no overall effect on Laminaria hyperborea oxygen flux or photosynthetic efficiency, with the latter remaining above 0.72 for all treatments (with 0.7 being the widely accepted value which indicates physiological stress – Bass et al., 2023).

Laminaria digitata and Laminaria ochroleuca photosynthetic efficiency is negatively affected by simulated MHWs, with more pronounced effects in the highest MHW intensity treatments (King et al., 2024; Leathers et al., 2024). This pattern was also observed by Leathers et al. (2024), who also found that the duration of MHWs had a significant effect on photosynthetic efficiency and bleaching of these two species.

MHWs have also been linked to mass mortalities of Saccharina latissima. In the spring of 2018, a severe category III MHW in southern Norway led to a reduction in kelp cover from above 50% to below 30% (Filbee-Dexter et al., 2020). In addition, four strong category II MHWs occurred on the east coast of the USA, which led to a reduction in kelp density from above 40 individuals per m² to less than 5 individuals per m². In southern Norway an increasing trend in the duration of temperature anomalies at a rate of 0.17 days/year over the past 60 years has led to a decrease in kelp biomass (Filbee-Dexter et al., 2020). This led to temperatures surpassing the mortality threshold of 19.7°C for populations of Saccharina latissima (Miller et al., 2024a).

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 sublethal, 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 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.

During a 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).

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.

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.

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, 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).

In Baja California, Mexico, an extreme heat even between 2014– 2016, led to both a decrease in density of Macrocystis pyrifera and a decrease in the number of fronds per individual in Baja California, Mexico (Arafeh-Dalmau et al., 2019). Additionally, there was a significant change to the understory algal composition, and half of the fish and invertebrates associated with this habitat disappeared. The same heatwave, coupled with a loss of starfish through disease and an increase in urchin grazing, led to the loss of > 90% of Macrocystis pyrifera from 350 km of coastline in northern California (Rogers-Bennett & Catton, 2019).

Sensitivity Assessment

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, Saccharina latissima resistance has been assessed as ‘None’. Although Laminaria hyperborea may be more resistant to this pressure, their recovery would be slower than that of Saccharina latissima. Echinus esculentus show signs of physiological stress (Bishop, 1985) and reduced gametogenesis with continued exposure to 19°C (Ursin, 1960). Therefore, the biotope’s resilience to this pressure is assessed as ‘Medium’ due to the lower resilience of Laminaria hyperborea, and sensitivity is ‘Medium’.

Please note: This pressure relates to increases in local water temperature over short time periods. For evidence of Laminaria hyperborea and Saccharina latissima sensitivity to long-term increases in sea surface temperature, see Global Warming evidence for section for IR.MIR.KT.XKT.

Temperature decrease (local)

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

Evidence

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.

According to OBIS (2026), Laminaria hyperborea occurs in temperatures from 5 to 15°C, with most records coming from the 10 to 15°C range. Kain (1964) stated that Laminaria hyperborea sporophyte growth and reproduction could occur within a temperature range of 0 to 20°C. Upper and lower lethal temperatures have been estimated at between 1 to 2°C above or below the extremes of these ranges (Birkett et al., 1988). 

Laminaria hyperborea is a boreal northern species with a geographic range from mid-Portugal to Northern Norway (Birkett et al., 1998b), and a mid-range within southern Norway (60° to 65° North) (Kain, 1971). The average seawater temperature for southern Norway in October is 12 to 13°C (Miller et al., 2009), and average annual sea temperature, from 1970 to 2014, is 8°C (Beszczynska-Möller & Dye, 2013). The available information suggests that Laminaria hyperborea and biotope structure would not be affected by a change in sea temperature at the benchmark level.

Laminaria hyperborea populations in the middle and northern regions of their latitudinal range assimilate more carbon than populations in warmer waters (Pessarrodona et al., 2018) and therefore accumulate more biomass. In colder parts of the UK, lamina extension, regrowth, and carbon standing stock were 1.5, 2 and 3 times higher, respectively, than in warmer areas (Smale et al., 2020). Wernberg et al. (2025) observed differences in morphological features between populations at opposing ends of the species’ range, such as stipe height, lamina width, stipe diameter, lamina thickness, and the number of digits. They also found that stipe epiphyte load was far greater in the colder region than in the warmer region (Wernberg et al., 2025), most likely due to greater stipe surface area (Teagle & Smale, 2018).

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). According to OBIS (2026), Saccharina latissima occurs in temperatures from 5 to 15°C, with just over half of all records coming from the 10 to 15°C range.

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 (F_V/F_M) 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).

Subtidal red algae can survive at temperatures between -2°C and 18 to 23°C (Lüning, 1990; Kain & Norton, 1990).

Sensitivity Assessment

Both Laminaria hyperborea and Saccharina latissima have northern distributions (Birkett et al., 1998). An acute or long-term decrease in temperature within the UK, at the benchmark level, is not likely to have any dramatic effect on biotope structure. Resistance has been assessed as ‘High’, resilience as ‘High’ and sensitivity as ‘Not sensitive’.

Salinity increase (local)

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

Evidence

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). Lüning (1990) suggest that kelps are stenohaline, their general tolerance to salinity as a phenotypic group covering 16 to 50 PSU over a 24-hour period. Optimal growth probably occurs between 30 to 35 PSU, and growth rates are likely to be affected by periodic salinity stress. Birkett et al. (1998) suggested that long-term increases in salinity may affect Laminaria hyperborea growth and may result in loss of affected kelp, and therefore loss of the biotope. According to OBIS (2026), Laminaria hyperborea occurs in salinities from 25 to 40 PSU, with most records coming from the 30 to 35 PSU range.

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 acclimation to rapid salinity changes could be slower at extremely low water temperatures (1 to 5°C) 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. According to OBIS (2026), Saccharina latissima occurs in salinities from 20 to 40 PSU, with most records coming from the 30 to 35 PSU range.

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).

Sensitivity Assessment

The evidence suggests that Saccharina latissima can tolerate exposure to hypersaline conditions of ≥40 PSU. However, optimal salinities for Laminaria hyperborea growth are assumed to be 30 to 35 PSU. Hence, increases in salinity to >40 PSU may cause mortality for Laminaria hyperborea. Resistance has been assessed as ‘Low’, resilience as ‘Medium’, and sensitivity as ‘Medium’.

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

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). Lüning (1990) suggest that kelps are stenohaline, their general tolerance to salinity as a phenotypic group covering 16 to 50 PSU over a 24-hour period. Optimal growth probably occurs between 30 to 35 PSU, and growth rates are likely to be affected by periodic salinity stress. Birkett et al. (1998) suggest that long-term changes in salinity may result in loss of affected kelp.

According to OBIS (2026), Laminaria hyperborea occurs in salinities from 25 to 40 PSU, with most records coming from the 30 to 35 PSU range. Hopkin & Kain (1978) tested Laminaria hyperborea sporophyte growth at various low salinity treatments. The results showed that sporophytes could grow normally at 19 PSU, while growth was reduced at 16 PSU and completely stopped at 7 PSU.

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. According to OBIS (2026), Saccharina latissima occurs in salinities from 20 to 40 PSU, with most records coming from the 30 to 35 PSU range.

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 largely 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. 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).

Echinoderms are generally unable to tolerate low salinity (stenohaline) and possess no osmoregulatory organ (Boolootian, 1966). At low salinity, urchins gain weight, and the epidermis loses its pigment as patches are destroyed; prolonged exposure is fatal.

Echinus esculentus has shown acute and chronic sensitivity to low salinity conditions (Barrett et al., 2024). In short-term (24 hour) exposures ranging from 11 to 31 PSU, oxygen consumption, activity levels, and coelomic fluid osmolality all declined in direct relation to falling salinity, with 100% survival at >21 PSU but 0% survival at <16 PSU. Over longer-term exposures (25 days), individuals maintained at 21 PSU showed elevated metabolic rates, reduced feeding, and reduced activity compared to controls at 31 PSU. Individuals held at 26 PSU displayed physiological metrics comparable to controls by the experiment’s end, indicating acclimation. Righting ability (when an upside-down urchin flips itself back to its correct orientation) and metabolic capacity also improved at 26 PSU, suggesting functional resistance to subsequent acute salinity stress. Despite these adjustments, coelomocyte composition and concentration remained unchanged, and overall results indicate that 21 PSU represents the lower acclimation limit for this species.

However, JNCC (2022) records show Echinus esculentus is found within a number of variable and reduced (18 to 30) salinity biotopes, e.g. IR.LIR.KVS.SlatPsaVS.

Sensitivity Assessment

IR.LIR.K.LhypSlat and its sub-biotopes are recorded in full salinity (30 to 40 PSU). A decrease in one MNCR salinity scale to ‘Variable Salinity (18 to 40 PSU) or ‘Reduced Salinity’ (18 to 30 PSU) for one year (as per the pressure benchmark) may result in a decrease in Laminaria hyperborea and Saccharina latissima growth and mortality of Echinus esculentus. Resistance has been assessed as ‘Low’ and resilience as ‘Medium’. Therefore, sensitivity of this biotope to a decrease in salinity has been assessed as ‘Medium’.

Water flow (tidal current) changes (local)

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

Evidence

The presence of water motion is important for kelp. The diffusive boundary layer becomes thinner as water motion increases and is further reduced by thallus pitching and flapping (Denny & Roberson, 2002 and Huang, Rominger & Nepf cited in Kerrison et al., 2015). The rates of photosynthesis and nutrient uptake increase in correlation to the current velocity until these metabolic processes become saturated (Wheeler, 1980, Hurd, 2000 and Hepburn et al., 2007 cited in Kerrison et al., 2015). This occurs with current velocities of approximately 0.1 m/s, with anything less defined as low water motion (Wheeler, 1980 cited in Kerrison et al., 2015). Anything above 0.25 m/s can be considered as high water motion (Stevens & Hurd, 1997 cited in Kerrison et al., 2015). Fast flow may also reduce the settlement and abundance of epiphytes, grazers, and sediment, which could smother or degrade the macroalgae and reduce growth rate (Méléder et al., 2010 cited in Kerrison et al., 2015). Very high-water motion may also dislodge fully-grown adults (Hurd, 2000 cited in Kerrison et al., 2015). On exposed coasts, kelps can be exposed to wind-wave induced orbital water velocities as high as 2 to 3 m/s (De Bettignies, Wernberg & Lavery, 2013 cited in Kerrison et al., 2015). Macroalgae grown in such high-water motion invest more energy into the development of large holdfasts to attach more firmly to the substratum (Kawamata, 2001, Duggins et al., 2003 cited in Kerrison et al., 2015), at the cost of reduced growth size and productivity (Gerard & Mann, 1979 cited in Kerrison et al., 2015).

Water flow influences the abundance, morphology and distribution of Laminaria hyperborea. This species commonly occurs in strong tidal streams and can persist in very strong currents (>3 m/s), such as those recorded in the Menai Strait, Wales, where tidal velocities may exceed 4 m/s (NBN, 2015), and in tidal rapids in Norway (J. Jones, pers. comm.). However, exceptionally strong currents can reduce kelp abundance by increasing detachment rates and limiting successful settlement (Bekkby et al., 2019).

The morphology of Laminaria hyperborea varies with hydrodynamic conditions. In areas exposed to strong water movement, plants typically develop longer and more flexible stipes, which are thought to represent an adaptation to increased flow and wave action (Sjøtun et al., 1998). In addition, the lamina becomes narrower and thinner in strong currents (Sjøtun & Fredriksen, 1995). However, Kregting et al. (2013) observed no significant differences in blade growth or stipe elongation between sites with different levels of wave height and maximum water velocity.

Norderhaug et al. (2014) investigated the effects of current speed on species richness and diversity on Laminaria hyperborea holdfasts. Species richness was significantly linked to current speed, with the highest richness observed at intermediate current speeds between 0.12 and 0.18 m/s, (defined as “intermediate” based on the range of values observed in their study). It has been suggested that higher current speeds increase nutrient flow in the area, thereby promoting the growth of epiphytic algae on kelp stipes (Bekkby et al., 2015). Moreover, current speed has an interactive effect with wave exposure, where areas with high tidal flow and high wave action have a higher stipe epiphyte density. In contrast, in areas with high tidal flow and low wave activity, the bidirectional flow of water from tidal forces may increase canopy shading due to drag, while orbital and stochastic wave action can allow more light penetration through the canopy and facilitate epiphyte growth (Bekkby et al., 2015).

Peteiro & Freire (2013) measured Saccharina latissima growth from two sites; the first had maximal water velocities of 0.3 m/sec, and the second 0.1 m/sec. At site one, Saccharina latissima had significantly larger biomass than at site two (16 kg/m to 12 kg/m, respectively). Peteiro & Freire (2013) suggested that faster water velocities were beneficial to Saccharina latissima growth. However, Gerard & Mann (1979) measured Saccharina latissima productivity at greater water velocities and found that Saccharina latissima productivity was reduced in moderately strong tidal streams (≤1 m/sec) when compared to weak tidal streams (<0.5 m/sec).

Flow rate also has a considerable effect on the blade and stipe morphology of Saccharina latissima. In sheltered conditions, the thallus becomes wide and thin, and in some species, corrugated (Kerrison et al., 2015). It is thought that the increased surface area enables maximal photon capture and gas/nutrient exchange, while thallus undulations may increase turbulence (Hurd, 2000, Fowler-Walker, Wernberg & Connell, 2006, Wing et al., 2007, Koehl, 2008, Hurd & Pilditch, 2009 cited in Kerrison et al., 2015). In exposed conditions, the thallus becomes narrow and thick with a robust stipe (Hurd, 2000 cited in Kerrison et al., 2015). This makes the thallus more hydrodynamic and reduces its drag, which is reasoned to prevent breakage or dislodgement of the adult sporophyte (Fowler-Walker, Wernberg & Connell, 2006 and Hurd & Pilditch, 2009 cited in Kerrison et al., 2015).

Saccharina latissima prefers low to moderate water motion areas and is usually absent in locations with high motion or surf (Southward & Orton, 1954, Druehl, 1967 and Burrows, 2012 and cited in Kerrison et al., 2015). However, Saccharina latissima has been successfully cultured in high motion areas with currents up to 1.53 m/s (Buck & Buchholz, 2005 cited in Kerrison et al., 2015). This suggests that their exclusion from high water motion locations may be due to other factors, such as competition for space by a species more adapted to this environment.

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 regarding 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 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 to60 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 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).

Echinus esculentus occurred in kelp beds on the west coast of Scotland in currents of about 0.5 m/sec. Outside the beds specimens were occasionally seen being rolled by the current (Comely & Ansell, 1988), which may have been up to 1.4 m/sec. Urchins are removed from the stipe of kelps by wave and current action. Echinus esculentus are also displaced by storm action. After disturbance Echinus esculentus migrates up the shore, an adaptation to being washed to deeper water by wave action (Lewis & Nichols, 1979). Therefore, increased water flow may remove the population from the affected area; probably to deeper water although individuals would probably not be killed in the process and could recolonize the area quickly.

Sensitivity Assessment

IR.LIR.K.LhypSlat and its sub-biotopes are recorded in weak (>1 kn) to very weak (negligible) tidal streams. Due to the range of tidal velocities that these biotopes are recorded within, a change in flow of between 0.1 to 0.2 m/sec would likely have no significant effect on Laminaria hyperborea or Saccharina latissima growth or productivity. Resistance has been assessed as ‘High’, resilience as ‘High’. Sensitivity has been assessed as ‘Not Sensitive’ at the benchmark level.

Larger increases in tidal streams (>3 m/s) may result in a change in biotope. An increase of this magnitude could decrease siltation which separates IR.LIR.K.LhypSlat and its associated sub-biotopes from the tide-swept IR.MIR.KT.XKT and IR.MIR.KT.XKTX biotopes. Moreover, IR.LIR.K.LhypSlat and its sub-biotopes have a macroalgae-dominated understory community as opposed to the rich filter-feeding fauna-dominated communities of the tide-swept IR.MIR.KT.XKT and IR.MIR.KT.XKTX biotopes. Therefore, it is possible that an increase in water flow could be unsuitable for the recruitment of understory algae. It may also increase urchin dislodgement and reduce their abundance so that IR.LIR.K.LhypSlat.Gz transitions into one of the other LhypSlat biotopes. Lastly, an extreme increase in water flow could dislodge Laminaria hyperborea and increase the abundance of ephemeral kelps such as Saccharina latissima or Alaria esculenta which are both fast-growing species and are tolerant of fast water movement (Birkett et al., 1998b). However, changes of this magnitude are outside of the scope of this habitat sensitivity assessment.

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.K.LhypSlat plus associated sub-biotopes are recorded predominantly in the sublittoral. An increase in emergence will result in an increased risk of desiccation and mortality of the dominant kelp species (Laminaria hyperborea & Saccharina latissima) in shallow examples of the biotope. Removal of canopy forming kelps has also been shown to increase desiccation and mortality of the understorey macro-algae (Hawkins & Harkin, 1985). Several mobile species such as sea urchins, brittle stars and feather stars are likely to move away. However, 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. Resilience has been assessed as ‘Low’. Resistance as ‘Medium’. The sensitivity of this biotope to a change in emergence is considered as ‘Medium’.

Wave exposure changes (local)

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

Evidence

The morphology of Laminaria hyperborea also can vary with wave exposure, with local adaptations likely aimed at reducing drag in highly wave exposed sites. In wave exposed areas, for example, Laminaria hyperborea develops a long and flexible stipe, and this is probably a functional adaptation to strong water movement (Sjøtun, 1998). In addition, the lamina becomes narrower and thinner in strong currents (Sjøtun & Fredriksen, 1995). Other studies have found that populations in areas with higher wave exposure tend to have greater levels of growth, biomass and density than their counterparts in sites with lower exposure (Pedersen et al., 2012; Smale et al., 2016; Gundersent et al., 2021). At low wave exposure, Laminaria hyperborea canopy-forming plants were smaller, had lower densities and had higher mortality rates than at exposed sites. This may be due to high epiphytic loading in low-exposure sites, which could reduce light and nutrient uptake, and increase the drag of the host Laminaria hyperborea during extreme storm events (Pedersen et al. (2012). However, Kregting et al. (2013) measured Laminaria hyperborea blade growth and stipe elongation from an exposed and a sheltered site in Strangford Lough, Northern Ireland, from March 2009-April 2010. Maximal significant wave height (H_m0) was 3.67 & 2 m at the exposed and sheltered sites, and maximal water velocity (Vel_rms) was 0.6 & 0.3 m/s at the exposed and sheltered sites respectively. Despite the differences in wave exposure and water velocity, there was no significant difference in Laminaria hyperborea growth between the exposed and sheltered sites.

Bekkby et al. (2019) observed a strong interactive effect of wave action, tidal flow and depth on Laminaria hyperborea abundance. They found that the probability of Laminaria hyperborea occurrence was 0.2 in areas with low wave exposure (0.002 km²/s), while it was almost 1.0 in areas with high wave exposure (up to 1.95 km²/s). These values were calculated using a simplified wave model which uses fetch, wind speed and frequency as input variables, is widely used in Scandinavian coastal studies and aligns with the EUNIS wave exposure classification system (described by Rinde et al., 2005).

Laminaria hyperborea responses to extreme levels of wave exposure vary in the literature. In the winter of 2013-14, the south of the UK experienced 6 of the 12 most intense storms that had been recorded in the prior five years (Smale & Vance, 2015). Despite wave heights exceeding 7 m and periods exceeding 13 s, Laminaria hyperborea showed much more resistance to the storm than the negatively impacted Laminaria ochroleuca, which had three times more density of severed stipes than Laminaria hyperborea. Earp et al. (2024) observed significant changes in canopy cover after Storm Arwen in the UK in November 2021. In the most wave exposed site (most northerly facing), canopy cover was reduced from above 75% to less than 10%. In the second most exposed site, cover was reduced from 100% to less than 60%. In addition to canopy loss, the understorey algae communities had also almost completely changed in the four most exposed sites. In northwest Portugal, Laminaria hyperborea abundance was reduced from 80 individuals/m² to 60 individuals/m² in the month following a storm (Pereira et al., 2017). The abundance was recorded at 10 individuals/m² ten months later, and the remaining individuals were almost entirely recruits.

Smaller changes in local wave height have the potential to cause changes to the understorey community. For example, species richness and diversity were found to be greater in Laminaria hyperborea forests with intermediate levels of wave exposure than those with higher levels of exposure (Norderhaug et al., 2014).

Saccharina latissima typically dominates sheltered shorelines (Gilson et al., 2023) and is rarely present in areas of wave exposure, where it is outcompeted 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). If present at wave exposed sites, it develops a short, thick stipe and a short, narrow and tightly wrinkled blade (Birkett et al., 1998).

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 sheltered, regarding both current speed and wave height. 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 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 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).

Echinus esculentus occurred in kelp beds on the west coast of Scotland in currents of about 0.5 m/sec. Outside the beds specimens were occasionally seen being rolled by the current (Comely & Ansell, 1988), which may have been up to 1.4 m/sec. Urchins are removed from the stipe of kelps by wave and current action. Echinus esculentus are also displaced by storm action. After disturbance Echinus esculentus migrates up the shore, an adaptation to being washed to deeper water by wave action (Lewis & Nichols, 1979). Keith Hiscock (pers. comm.) reported Echinus esculentus occurred in significant numbers as shallow as 15m below low water at the extremely wave exposed site of Rockall, Scotland. Therefore, localised increases in wave height may remove the population from the affected area; probably to deeper water although individuals would probably not be killed in the process and could recolonize the area quickly.

Sensitivity Assessment

IR.LIR.K.LhypSlat and its sub-biotopes are recorded from wave sheltered sites, so an increase in wave exposure (e.g. to moderate or higher) could lead to a change in abundance of the characterising kelp species. There is a wealth of evidence suggesting that Saccharina latissima persists in wave exposed sites, albeit with different morphology and levels of performance, but Birkett et al. (1998) suggest that Saccharina latissima is typically outcompeted by Laminaria hyperborea in exposed sites. Therefore, as a precaution, resistance has been assessed as ‘Medium’, resilience as ‘Medium’ and sensitivity as ‘Medium’ at the benchmark level.

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant Transitional metal or organometal (e.g. TBT) contaminants via uncontrolled releases or incidental spills (Transitional metals and organometals pressure definition).

Evidence

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

Very few studies have investigated whether these contaminants are harmful to kelp. Most studies on this topic instead focus on implications for human health and other trophic levels in kelp forest ecosystems.

While the effects of heavy metals on Laminaria hyperborea have not been extensively studied in recent literature, closely related species such as Laminaria digitata and Laminaria ochroleuca have. Given the close taxonomic relationship between these species, it is possible that physiological responses to metal contamination may be similar.

Maulvault et al. (2015) investigated toxic elements in seafood samples around Europe, including Laminaria digitata, bivalves, and fish. The highest levels of arsenic (As) in the study (41 mg/kg) were detected in Laminaria digitata. Cadmium (Cd) and copper (Cu) have been shown to significantly decrease photosynthetic efficiency in Laminaria digitata to 0.65 (Anacleto et al., 2017), which is below the commonly accepted indicative value for stress, 0.7 (Bass et al., 2023).

Laminaria ochroleuca significantly reduces Cu concentrations in the surrounding water (Cereja et al., 2026). This effect was strong enough to increase Mytilus edulis survival compared to the treatment without Laminaria ochroleuca. However, this also resulted in a decrease in chlorophyll-a concentration, an increase in phaeophytin, and tissue bleaching, indicating physiological stress. After 15 days of exposure, Cu accumulation led to kelp mortality (Cereja et al., 2026).

Sporophytes of Saccharina latissima have a low tolerance to heavy metals, especially early life stages. The effects of Cu, zinc (Zn), and mercury (Hg) on Saccharina latissima have been investigated by Thompson & Burrows (1984). They observed that the growth of sporophytes was significantly inhibited at 50 µg Cu /l, 1000 µg Zn/l and 50 µg Hg/l. Zoospores were found to be more intolerant and significant reductions in survival rates were observed at 25 µg Cu/l, 1000 µg Zn/l and 5 µg/l.

Bryan (1984) suggested that the general order for heavy metal toxicity in seaweeds is: Organic Hg > inorganic Hg > Cu > silver (Ag) > Zn > Cd > lead (Pb). Cole et al. (1999) reported that Hg was very toxic to macrophytes. Similarly, Hopkin & Kain (1978) demonstrated sublethal effects of heavy metals on Laminaria hyperborea gametophytes and sporophytes, including reduced growth and respiration. Although macroalgae species may not be killed, except by high levels of contamination, reduced growth rates may impair the ability of the biotope to recover from other environmental disturbances.

Sheppard et al. (1980) noted that increasing levels of heavy metal contamination along the west coast of Britain reduced species number and richness in holdfast fauna, except for suspension feeders which became increasingly dominant. Gastropods may be relatively tolerant of heavy metal pollution (Bryan, 1984). 

Little is known about the effects of heavy metals on echinoderms. Bryan (1984) reported that early work had shown that echinoderm larvae were intolerant of heavy metals, e.g. the intolerance of larvae of Paracentrotus lividus to Cu had been used to develop a water quality assessment. It is therefore possible that Echinus esculentus recruitment is likely to be impaired by heavy metal contamination due to the intolerance of echinoderm larvae to heavy metals. However, they are long-lived and have poor recruitment so they may not reduce grazing pressure in the short-term. Kinne (1984) reported developmental disturbances in Echinus esculentus exposed to waters containing 25 µg/l of Cu. Sea urchins, especially the eggs and larvae, are used for toxicity testing and environmental monitoring (reviewed by Dinnel et al. 1988). Migliaccio et al. (2014) reported exposure of Paracentrotus lividus larvae to increased levels of cadmium and manganese caused abnormal larval development and skeletal malformations.

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.

Laminaria hyperborea and 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 1-22.5m below chart datum, including Laminaria hyperbora communities, 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 (Sommerfield & Warwick, 1999). Laboratory studies of the effects of oil and dispersants on several red algae species, including Delesseria sanguinea (Grandy 1984; cited in Holt et al., 1995) concluded that they were all sensitive to oil/ dispersant mixtures, with little differences between adults, sporelings, diploid or haploid life stages. Holt et al. (1995) concluded that Delesseria sanguinea is probably generally sensitive of chemical contamination. Loss of red algae is likely to reduce the species richness and diversity of IR.LIR.K.LhypSlat.

Echinus esculentus is subtidal and unlikely to be directly exposed to oil spills. However, as with the ‘Prestige’ oil spill rough seas can cause mixing with the oil and the seawater, and therefore sub-tidal habitats can be affected by the oil spill. Castège et al., (2014) recorded the recovery of rocky shore communities following the Prestige oil spill which impacted the French Atlantic coast. Rough weather at the time of the spill increased mixing between the oil and seawater, causing sub-tidal communities/habitats to be affected. The urchin Echinus esculentus was reported absent after the oil spill however returned after 2-5 years. Large numbers of dead Echinus esculentus were found between 5.5 and 14.5 m in the vicinity of Sennen cove, presumably due to a combination of wave exposure and heavy spraying of dispersants following the ‘Torrey canyon’ oil spill (Smith 1968). Smith (1968) also demonstrated that 0.5 -1ppm of the detergent BP1002 resulted in developmental abnormalities in its echinopluteus larvae. 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) (Gomez & Miguez-Rodriguez 1999).

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.

Although Laminaria hyperborea sporelings and gametophytes are intolerant of atrazine (and probably other herbicides) overall they may be relatively tolerant of synthetic chemicals (Holt et al., 1995; Johansson, 2009). Laminaria hyperborea survived within >55m from the acidified halogenated effluent discharge polluting Amlwch Bay, Anglesey, albeit at low density. These specimens were greater than 5 years of age, suggesting that spores and/or early stages were more intolerant (Hoare & Hiscock, 1974). Patella pellucida was excluded from Amlwch Bay by the pollution and the species richness of the holdfast fauna decreased with proximity to the effluent discharge; amphipods were particularly intolerant although polychaetes were the least affected (Hoare & Hiscock, 1974). The richness of epifauna/flora decreased near the source of the effluent and epiphytes were absent from Laminaria hyperborea stipes within Amlwch Bay. The red alga Phyllophora membranifolia was also tolerant of the effluent in Amlwch Bay.

There is also currently insufficient evidence for harmful effects of synthetic compound contamination in the closely related kelp Laminaria digitata. Studies which investigate the presence of these compounds in seafood, including kelps, focus on implications for human health (see Alvarez-Munoz et al., 2015) rather than their effects on the species or ecosystems themselves. However, Anacleto et al. (2017) did investigate the effects of a range of pollutants on Laminaria digitata health but found no significant reduction in photosynthetic activity in response to pesticides (diflubenzuron and lindane).

Saccharina latissima has been found to be sensitive to antifouling compounds. 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 which may cause irreversible damage.

Large numbers of dead Echinus esculentus were found between 5.5 and 14.5 m depth in the vicinity of Sennen, presumably due to a combination of wave exposure and heavy spraying of dispersants in that area following the Torrey Canyon oil spill (Smith 1968). Smith (1968) also demonstrated that 0.5 to 1 ppm of the detergent BP1002 resulted in developmental abnormalities in echinopluteus larvae of Echinus esculentus. Echinus esculentus populations in the vicinity of an oil terminal in 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) (Gomez & Miguez-Rodriguez 1999).

The urchins Echinus esculentus, Paracentrotus lividus, and Psammechinus miliaris are sensitive to hydrogen peroxide, which is used to treat sea lice in Atlantic salmon aquaculture (Liu et al., 2023). In Echinus esculentus, coelomocytes (types of immune cells in invertebrates) show measurable oxidative DNA damage even at 0.5 mM (millimolar) H₂O₂, concentrations which are representative of diluted sea lice treatment effluent from aquaculture operations. Interspecific differences were clear, with Psammechinus miliaris exhibiting the highest DNA strand scission factors (a quantitative index of DNA strand breakage) at peak exposure, while Echinus esculentus and Paracentrotus lividus showed comparatively lower but still significant oxidative DNA damage, indicating variation in susceptibility and/or DNA repair capacity among species. In a later experiment, Liu et al. (2025) exposed Paracentrotus lividus to H₂O₂ concentrations of 50 and 500 µM and found that coelomocyte and red cell concentrations increased by approx. 50 and approx. 12%, respectively, after three hours of exposure. In addition, they observed DNA damage (approx. 0.05 strand scission factor) after one hour of exposure to 500 µM H₂O₂.

Evidence suggests that grazing gastropods, amphipods and red algae are more sensitive. Loss of red algae is likely to reduce the species richness and diversity of the biotope, potentially resulting in the understorey being dominated by encrusting corallines. However, red algae are likely to recover relatively quickly.

O'Brian & Dixon (1976) suggested that red algae were the most sensitive group of macrophytes to oil and dispersant contamination (see Smith, 1968). Smith (1968) also noted that epiphytic and benthic red algae were highly sensitive to dispersant or oil contamination following the Torrey Canyon oil spill. Only the epiphytes Crytopleura ramosa and Spermothamnion repens and some tufts of Jania rubens, and sublittoral fringe species such as Osmundea pinnatifida, Gigartina pistillata and Phyllophora crispa. Delesseria sanguinea was probably the most sensitive, showing damage at depths of 6 m (Smith, 1968). Holt et al. (1995) similarly suggested that Delesseria sanguinea is generally sensitive to chemical contamination.

Smith (1968) also 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. Delesseria sanguinea was probably to most intolerant since it was damaged at depths of 6m (Smith, 1968). Holt et al. (1995) suggested that Delesseria sanguinea is probably generally sensitive of chemical contamination. Although Laminaria hyperborea may be relatively tolerant to synthetic chemical pollution, evidence suggests that grazing gastropods, amphipods and red algae are sensitive. Loss of red algae is likely to reduce the species richness and diversity of the biotope, and the understorey may become dominated by encrusting corallines; however, red algae are likely to recover relatively quickly.

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 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).

In general, respiration in most marine invertebrates do not appear to be significantly affected until extremely low concentrations are reached. For many benthic invertebrates this concentration is about 2 ml l^-1, or even less (Herreid, 1980; Rosenberg et al., 1991; Diaz & Rosenberg, 1995).

Sensitivity Assessment. Reduced oxygen levels are likely to inhibit photosynthesis and respiration but not cause a loss of the macroalgae population directly. Resistance has been assessed as ‘High’, Resilience as ‘High’. Sensitivity has been assessed as ‘Not sensitive’ at the benchmark level.

Nutrient enrichment

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

Evidence

Laminaria hyperborea forests near high-effluent salmon farms show different stipe-associated community compositions to those near low-effluent farms and reference sites with no nearby aquaculture activity (Haugland et al., 2021). Bryozoan biomass was significantly higher at the high-effluent sites, whereas stipes at the low-effluent and reference sites were predominantly colonized by epiphytic macroalgae. At the high-effluent sites, the reduced epiphytic community was dominated by Ectocarpus spp., leading to lower heterogeneity within the stipe assemblage and reduced habitat heterogeneity. This suggests that changes in dissolved inorganic nitrogen could potentially shift this biotope from being fauna-dominated to an algae-dominated one (e.g. IR.HIR.KFaR.LhypR).Increased nutrients may result in phytoplankton blooms that increase turbidity (see above), and may favour sea urchins, e.g. Echinus esculentus, due their ability to absorb dissolved organics, potentially increasing grazing pressure leading to loss of understorey epiflora/fauna, reduced kelp recruitment, and possibly to the formation of urchin barrens. Therefore, although nutrients may not affect kelps directly, indirect effects such as turbidity, siltation and competition may significantly affect the structure of the biotope.

Holt et al. (1995) suggest that Laminaria hyperborea may be tolerant of organic enrichment since healthy populations are found at ends of sub littoral untreated sewage outfalls in the Isle of Man. Increased nutrient levels e.g. from sewage outfalls, has been associated with increases in abundance, primary biomass and Laminaria hyperborea stipe production but with concomitant decreases in species numbers and diversity (Fletcher, 1996). Increases in ephemeral and opportunistic algae are associated with reduced numbers of perennial macrophytes (Fletcher, 1996). Increased nutrients may also result in phytoplankton blooms that increase turbidity.

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 regarding to increased concentrations of pigments and nitrogen-rich compounds.

Conolly & Drew (1985) found Saccharina latissima sporophytes had relatively higher growth rates when near 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). Handå et al. (2013) also reported Saccharina latissima sporophytes grew approx. 1% faster per day when near Salmon farms, where elevated ammonium can be readily absorbed. 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 near 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 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).

Increased nutrients may also result in phytoplankton blooms that increase turbidity (see ‘changes in suspended solids’ section below), and may favour sea urchins, e.g. Echinus esculentus, due their ability to absorb dissolved organics, potentially increasing grazing pressure leading to loss of understorey epiflora/fauna, reduced kelp recruitment, and possibly to the formation of urchin barrens. Therefore, although nutrients may not affect kelps directly, indirect effects such as turbidity, siltation and competition may significantly affect the structure of the biotope.

It was suggested by Comely & Ansell (1988) that Echinus esculentus could absorb dissolved organic material as a nutritional source. Nutrient enrichment may encourage the growth of ephemeral and epiphytic algae and therefore increase sea-urchin food availability. Lawrence (1975) reported that sea urchins had persisted over 13 years on barren grounds near sewage outfalls, presumably feeding on dissolved organic material, detritus, plankton and microalgae, although individuals died at an early age.

Sensitivity Assessment

Although short-term exposure (<4 years) to nutrient enrichment may not affect seaweeds directly, indirect effects such as turbidity may significantly affect photosynthesis and result in reduced growth and reproduction and increased competition form fast-growing but ephemeral species. The above evidence suggests that increased nutrients may benefit Laminaria hyperborea kelp beds but alter the associated community, possibly resulting in changes in biotope classification. In extreme cases, turbidity and suspended sediment (see changes in suspended sediment) may attenuate light and be detrimental. Hence, resistance is assessed as ‘Medium’ as a precaution. Resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’.

Organic enrichment

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

Evidence

Laminaria hyperborea forests near high-effluent salmon farms show different stipe-associated community compositions to those near low-effluent farms and reference sites with no nearby aquaculture activity (Haugland et al., 2021). Bryozoan biomass was significantly higher at the high-effluent sites, whereas stipes at the low-effluent and reference sites were predominantly colonized by epiphytic macroalgae. At the high-effluent sites, the reduced epiphytic community was dominated by Ectocarpus spp., leading to lower heterogeneity within the stipe assemblage and reduced habitat heterogeneity. This suggests that changes in dissolved inorganic nitrogen could potentially shift this biotope from being fauna-dominated to an algae-dominated one (e.g. IR.HIR.KFaR.LhypR).Increased nutrients may result in phytoplankton blooms that increase turbidity (see above), and may favour sea urchins, e.g. Echinus esculentus, due their ability to absorb dissolved organics, potentially increasing grazing pressure leading to loss of understorey epiflora/fauna, reduced kelp recruitment, and possibly to the formation of urchin barrens. Therefore, although nutrients may not affect kelps directly, indirect effects such as turbidity, siltation and competition may significantly affect the structure of the biotope.

Holt et al. (1995) suggest that Laminaria hyperborea may be tolerant of organic enrichment since healthy populations are found at ends of sub littoral untreated sewage outfalls in the Isle of Man. Increased nutrient levels e.g. from sewage outfalls, has been associated with increases in abundance, primary biomass and Laminaria hyperborea stipe production but with concomitant decreases in species numbers and diversity (Fletcher, 1996). Increases in ephemeral and opportunistic algae are associated with reduced numbers of perennial macrophytes (Fletcher, 1996). Increased nutrients may also result in phytoplankton blooms that increase turbidity.

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 regarding increased concentrations of pigments and nitrogen-rich compounds.

Conolly & Drew (1985) found Saccharina latissima sporophytes had relatively higher growth rates when near 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). Handå et al. (2013) also reported Saccharina latissima sporophytes grew approx. 1% faster per day when near Salmon farms, where elevated ammonium can be readily absorbed. 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 near 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 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).

Sensitivity Assessment

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. Therefore, resistance has been assessed as ‘Medium’ as a precaution, resilience as ‘Medium’, and sensitivity as ‘Medium’.

Physical loss (to land or freshwater habitat)

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

Evidence

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

Physical change (to another seabed type)

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

Evidence

If rock substrata were replaced with sedimentary substrata this would represent a fundamental change in habitat type, which kelp species would not be able to tolerate (Birkett et al., 1998b). The biotope would be lost.

Sensitivity assessment. Resistance to the pressure is considered ‘None’, and resilience ‘Very Low’ or ‘None’. The sensitivity of this biotope to change from sedimentary or soft rock substrata to hard rock or artificial substrata or vice-versa is assessed as ‘High’.

Physical change (to another sediment type)

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

Evidence

Not relevant

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 to hard substratum (rock) biotopes.

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

Kelp harvesting can cause significant changes to this biotope through the removal of Laminaria hyperborea and the habitat space that it provides for its associated communities. Removing 26% of the canopy led to a 67% reduction in epiphytes and an 89% reduction of invertebrates (Norderhaug et al., 2020).

Christie et al. (1998) observed Laminaria hyperborea habitat regeneration following commercial Laminaria hyperborea trawling in south Norway. Trawling removed all large canopy-forming adult Laminaria hyperborea, however, sub-canopy recruits were largely unaffected. Within 2 to 6 years of harvesting, a new canopy had formed 1 m off the seabed. The associated holdfast communities recovered in six years. However, the epiphytic stipe community did not fully recover within the same period. Christie et al. (1998) suggested that kelp habitats were relatively resistant to direct disturbance/removal of Laminaria hyperborea canopy.

In Nord-Trøndelag, Norway, Laminaria hyperborea was harvested for the first time in 2010 (Steen et al., 2016). Video surveys and plant sampling conducted two days prior to the trawling and in each year for the following four years, showed that Laminaria hyperborea coverage had returned to pre-harvest levels (around 94%). However, the new canopy, was significantly lower in density, average plant age, length, weight, and epiphyte biomass. In addition, the density of understorey recruits had only recovered by one-third by the end of the study period. It was suggested that 80% of the new canopy consisted of understorey plants that had survived the harvesting, and that the resilience of this biotope was dependent on the frequency of harvesting (Steen et al., 2016).

Recurrent disturbance on a timescale shorter than the 2 to 6-year recovery period could prolong the recovery. Kain (1975) cleared sublittoral blocks of Laminaria hyperborea at different times of the year for several years. The first colonizers and succession communities differed between blocks and the time of year they were cleared. However, within two years of clearance, the blocks were dominated by Laminaria hyperborea. Leinaas & Christie (1996) also observed Laminaria hyperborea recolonizing urchin barrens following urchin removal. The substratum was initially colonized by filamentous macroalgae and Saccharina latissima. However, Laminaria hyperborea dominated the community after 2 to 4 years.

Laminaria hyperborea forests subjected to regular harvesting support different associated communities compared to unharvested, preserved forests (Leclerc et al., 2015). Macroalgal species richness was consistently higher at the harvested site across all parts of the kelp and on the surrounding rock. Sessile fauna richness was slightly higher on the stipes and surrounding rock at the harvested site, but lower on the holdfast compared to the preserved site. In contrast, mobile fauna richness and density were generally greater on all parts of the kelp in the preserved site, although both were higher on the surrounding rock at the harvested site. Following disturbance, or in areas experiencing frequent disturbance, Laminaria hyperborea recruitment may be affected by interspecific competition with Non-Indigenous Invasive Species (INIS) or ephemeral algae (Brodie et al., 2014; Smale et al., 2013) (see INIS below).

Low level disturbances (e.g. solitary anchors) are unlikely to cause harm to the biotope, due to the impact’s small footprint. Saccharina latissima is commercially cultivated but sporophytes are typically matured on ropes (Handå et al., 2013) and not directly extracted from the seabed. Evidence to assess the resistance of Saccharina latissima to in/direct harvesting or abrasion is limited.

Species with fragile tests, such as Echinus esculentus were reported to suffer badly because of scallop or queen scallop dredging (Bradshaw et al., 2000; Hall-Spencer & Moore, 2000a, b). Kaiser et al. (2000) reported that Echinus esculentus were less abundant in areas subject to high trawling disturbance in the Irish Sea. Jenkins et al. (2001) conducted experimental scallop trawling in the North Irish Sea and recorded the damage caused to several conspicuous megafauna species, both when caught as bycatch and when left on the seabed. The authors predicted 16.4% of Echinus esculentus were crushed/dead, 29.3% would have >50% spine loss/minor cracks, 1.1% would have <50% spine loss and the remaining 53.3% would be in good condition. Sea urchins can rapidly regenerate spines, e.g. Psammechinus miliaris were found to re-grow all spines within a period of 2 months (Hobson, 1930).

Sensitivity Assessment

Abrasion by passing trawls or harvesting of macroalgae is likely to remove all the large canopy-forming plants (Svendsen, 1972; Christie et al., 1998) and cause Echinus esculentus mortality (Bradshaw et al., 2000; Hall-Spencer & Moore, 2000a, b; Jenkins et al., 2001). Therefore, resistance has been assessed as ‘None’, resilience as ‘Medium’, and sensitivity as ‘High’.

Penetration or disturbance of the substratum subsurface

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

Evidence

Not Relevant to hard substratum (rock) biotopes

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 linear relationship with sub-surface light attenuation (Kd) (Devlin et al., 2008). An increase in SPM results in a decrease in sub-surface light attenuation. Light availability and water turbidity are principal factors in determining kelp depth range (Birkett et al., 1998). 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).

Kain (1964) suggested that early Laminaria hyperborea gametophyte development could occur in the absence of light. Furthermore, observations from south Norway found that a pool of Laminaria hyperborea recruits could persist growing beneath Laminaria hyperborea canopies for several years, indicating that sporophyte growth can occur in light-limited environments (Christe et al., 1998). However, in habitats exposed to high levels of suspended silts Laminaria hyperborea is outcompeted by Saccharina latissima, a silt tolerant species, and thus, a decrease in water clarity is likely to decrease the abundance of Laminaria hyperborea in the affected area (Norton, 1978).

Ecklonia radiata show a decrease of 50% photosynthetic activity when turbidity increases by 0.1/m (light attenuation coefficient =0.1 to 0.2/m; Staehr & Wernberg, 2009). An increase in water turbidity therefore will likely affect the photosynthetic ability of Laminaria hyperborea and Saccharina latissima and decrease their abundance and density.

Sensitivity Assessment

Changes in water clarity are likely to affect photosynthetic rates and enable Saccharina latissima to compete more successfully with Laminaria hyperborea. 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 clarity from clear to intermediate (10 to 100 mg/l) represents a change in light attenuation of approx. 0.67 to 6.7 Kd/m, and is likely to result in a greater than 50% reduction in photosynthesis of Laminaria spp. Therefore, Laminaria hyperborea will probably suffer a significant decline so resistance to this pressure is assessed as ‘Low’. Resilience to this pressure is ‘Medium’, so sensitivity is ‘Medium’.

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

Moore (1977) suggested that Echinus esculentus was unaffected by turbid conditions. Echinus esculentus is an important grazer of red macro-algae within CR.MCR.EcCr. Increased turbidity and resultant reduced light penetration is likely to negatively affect algal growth. However, Echinus esculentus can feed on alternative prey, detritus or dissolved organic material (Lawrence, 1975, Comely & Ansell, 1988). Suspended Particle Matter (SPM) concentration has a linear relationship with sub surface light attenuation (Kd) (Devlin et al., 2008). An increase in SPM results in a decrease in sub-surface light attenuation. Light availability and water turbidity are principal factors in determining depth range at which kelp can be found (Birkett et al., 1998). Light penetration influences the maximum depth at which kelp species can grow and it has been reported that laminarians grow at depths at which the light levels are reduced to 1 percent of incident light at the surface. Maximal depth distribution of laminarians therefore varies from 100 m in the Mediterranean to only 6-7m 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. 1998).

Laminaria spp. 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). An increase in water turbidity will likely affect the photosynthetic ability of Laminaria hyperborea and Saccharina latissima, decrease kelp abundance and density and increase the dominance of kelp park biotopes in shallow water (see sub biotope- IR.LIR.K.LhypSlat.Pk). Kain (1964) suggested that early Laminaria hyperborea gametophyte development could occur in the absence of light. Furthermore observations from south Norway found that a pool of Laminaria hyperborea recruits could persist growing beneath Laminaria hyperborea canopies for several years, indicating sporophytes growth can occur in light limited environments (Christie et al., 1998).

Sensitivity Assessment. An increase in water clarity from clear to intermediate (10-100mg/l) represent a change in light attenuation of ca 0.67-6.7 Kd/m, and is likely to result in a greater than 50% reduction in photosynthesis of Laminaria spp. Therefore the dominant kelp species will probably suffer a severe decline and resistance to this pressure is assessed as ‘None’.  Resilience is probably to this pressure is defined as ‘Medium’ at the benchmark. Hence, this biotope is regarded as having a sensitivity of ‘Medium ‘to this pressure.

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. 5 cm material during a discrete event, is unlikely to damage Laminaria hyperborea or 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.K.LhypSlat is likely to be dependent on annual 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 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.

Comely & Ansell (1988) recorded large Echinus esculentus from kelp beds on the west coast of Scotland in which the substratum was seasonally covered with "high levels" of silt. This suggests that Echinus esculentus is unlikely to be killed by smothering, however, smaller specimens and juveniles may be less resistant. A layer of sediment may interfere with larval settlement.  If retained within the host biotope for extended periods a layer of 5cm of the sediment may negatively affect successive recruitment events.

If inundation is long lasting then the understorey flora may be adversely affected. If clearance of deposited sediment occurs rapidly then understorey communities are expected to recover quickly. In moderately exposed examples of IR.LIR.K.LhypSlat, deposited sediment is unlikely to remain for more than a few tidal cycles (due to water flow or wave action). In wave sheltered examples of IR.LIR.K.LhypSlat sediment could remain and recovery rate would be related to sediment retention, which may take a few years to dissipate.

Sensitivity assessment. Resistance has been assessed as ‘Medium’, resilience 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), depending on the study species and duration and intensity of exposure.

Chapman et al. (2023) exposed Echinus esculentus to a 500 µT magnetic field for 24 hours, reflecting the expected EMF strength from subsea power cables used in offshore renewable energy infrastructure. They found no significant change in behaviour or physiology, although the authors noted that this may differ depending on exposure duration, life stage, and other stressors.

There is insufficient evidence to make a sensitivity assessment for this pressure due to the lack of evidence of this pressure affecting the characteristic species of this biotope.

Underwater noise changes

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

Evidence

Not Relevant

Introduction of light or shading

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

Evidence

The availability of light is highly spatiotemporally variable, and its oversupply can be a major threat to macroalgal survival (Airoldi & Beck, 2007 cited in Kerrison et al., 2015). If too much light is absorbed by kelp, the excess energy can inhibit photosynthesis (Dring, Wagner & Luning, 2001 cited in Kerrison et al., 2015) and may lead to cellular damage and death of the organism. This sets an upper depth limit for many species. Conversely, sufficient photosynthetically active radiation must be supplied to sustain growth, therefore setting a lower depth limit. In very clear waters, some kelps can grow down to 30 to 40 m (Smale et al., 2013; Khan et al., 2018), while in waters carrying suspended sediment, light penetration declines quickly, leading to a shallow limit of less than a metre. 

Light availability is a key environmental factor influencing the distribution, morphology, and productivity of Laminaria hyperborea. Several studies have demonstrated that biomass accumulation, canopy density, and morphological traits are positively correlated with light levels. For example, Smale et al. (2016) found that summer daytime light had a strong positive effect on canopy biomass and standing stock of carbon. Similarly, Smith et al. (2022) reported that percentage surface irradiance significantly predicted total and canopy density, canopy standing biomass, total fresh weight, blade fresh weight, and blade length in Laminaria hyperborea populations across the UK. In southern regions, blade width, total length, and age were also positively affected.

Stahl et al. (2024) conducted a study on the potential for Laminaria hyperborea afforestation in the German Bight, with a specific focus on light requirements and habitat suitability. Their study identified a minimum compensation irradiance of approximately 30 µmol photons /m²/s¹ under summer conditions. Monteiro et al. (2015) supported these findings, showing that over 75% of observed kelp species, including Laminaria hyperborea, occurred in areas where more than 3.65% of surface light reached the seafloor. These studies show that Laminaria hyperborea occurs in areas with moderate to high levels of light.

The effects of light on Laminaria hyperborea physiology can vary depending on temperature. For instance, at 10ºC, three different photoperiods: polar day (24:0 light:dark), long day (16:8 light:dark) and polar night (0:24 light:dark) had very little effect on photosynthetic efficiency, while 5ºC and 0ºC treatments had varied photosynthetic responses (Diehl et al., 2024). Photosynthetic efficiency declined significantly over 12 weeks in the polar day and long day photoperiods at 0ºC but increased slightly with the polar night treatment. In the same time frame, photosynthetic efficiency declined significantly at 5ºC with the polar day treatment but was relatively unchanged with the other two photoperiods. Other measured responses – such as dry weight, pigments, phlorotannins, and storage carbohydrates – all varied by light and temperature treatments. Cold and long light conditions significantly decreased chlorophyll-a, accessory pigments and VAZ pigments, which indicates a photoprotective stress response. In the 10°C treatment, these pigments either decreased or showed no change, suggesting that the relatively higher temperature mitigated light stress. Dry weight increased significantly, despite no measurable change in surface area, when the highest temperature (10°C) treatment was combined with moderate (16:8 h) and long (24:0 h) photoperiods. This increase in dry weight was not detrimental to the kelp and was likely due to the accumulation of storage carbohydrates rather than growth. No significant responses were observed in phlorotannin (compounds that protect against light stress) levels. Mannitol (a storage carbohydrate) decreased under the long night treatment, but this effect is expected and not detrimental to the kelp. Laminarin (the other storage carbohydrate that was measured) increased significantly under both light treatments and the two warmer treatments (5°C and 10°C), which is a positive metabolic response.

Shading of the biotope (e.g. by coastal development) could adversely affect the biotope in areas already low in water clarity. This may shift the balance toward shade-tolerant species, leading to loss of the biotope within shaded zones or a reduction in Laminaria hyperborea abundance, shifting from forest to park-type biotopes. Laminaria spp. show a decrease of 50% photosynthetic activity when turbidity increases by 0.1/m (light attenuation coefficient = 0.1 to 0.2/m; Staehr & Wernberg, 2009). Therefore, any activity that decreases incident light (e.g. shading) may be detrimental.

The optimum depth for growth of Saccharina latissima has been reported as 9 to 12 m in Maine, USA (Boden, 1979 cited in Kerrison et al., 2015), 5 m in mid-Norway (Handå et al., 2013 cited in Kerrison et al., 2015), or only 1.5 to 3 m in Scotland (Kerrison et al., 2015). Young Saccharina latissima sporophytes appear to have similar light requirements and tolerance as Laminaria digitata (Han & Kain, 1996 cited in Kerrison et al., 2015). While adults are light saturated at around 215 μmol m²/s and have their maximum photosynthetic rate at 200 μmol m²/s (Bartsch et al., 2008 cited in Kerrison et al., 2015). One or two hours of light at 500 to 700 μmol m²/s leads to significant dynamic photoinhibition and photodamage, with young sporophytes being more susceptible than adults (Bruhn & Gerard, 1996 and Hanelt, Wiencke & Karsten, 1997 cited in Kerrison et al., 2015). High light exposure can therefore lead to the death of thallus tissue and the loss of biomass (Kerrison et al., 2015).

Ebbing et al. (2020) studied how light and biomass density influence the reproduction of delayed Saccharina latissima gametophytes over 21 days. They reported that reproductive success decreased at high light intensities (≥80 µmol photons/m²/s) across all light qualities and that optimal reproduction occurred at light intensities between 14.2 µmol and 25.7 µmol photons/m²/s (Ebbing et al., 2020). In addition, white light led to the highest reproductive success under optimal Initial Gametophyte Density conditions, while blue light resulted in the lowest reproductive success, especially at higher intensities (Ebbing et al., 2020). Red light at low intensity (5 µmol photons/m²/s) significantly inhibited reproduction, but higher intensities of red light improve reproductive success (Ebbing et al., 2020). ​Finally, Photosynthetically Usable Radiation (PUR), which integrates light intensity and quality, is a strong abiotic factor regulating reproduction, and reproductive success decreases when PUR exceeds 26.8 µmol photons/m²/s, regardless of light quality (Ebbing et al., 2020). In a further study, 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. Furthermore, Ebbing et al. (2021) demonstrated that delayed gametophytes of Saccharina latissima could reliably reproduce sexually after more than a year of vegetative growth, depending on the effects of light intensity and temperature. These findings suggest that both the quantity and quality of light, along with temperature, play critical roles in regulating the reproduction of delayed Saccharina latissima gametophytes.

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. Although the results revealed that the mean growth rate and the intracellular nitrogen components of the sporophytes were positively related to the external nitrate concentration during the experimental period, the authors note how they saw no significant difference between the high light and the low light treatments (Jevne, Forbord & Olsen, 2020). Jevne, Forbord & Olsen (2020) did go on to highlight how Saccharina latissima grown between 10 and 15°C at a high light intensity of 250 μmol m²/s showed a 50% lower growth rate compared with kelps grown at 110 μmol m²/s, which was found to be the optimum for photon flux in this temperature interval (citing Fortes & Lüning, 1980). 

De Jong et al. (2021) also studied the effect of nutrient availability and light intensity on the sterol content of Saccharina latissima over a five-week period, subjected to a nutrient-replete and nutrient-depleted regime, then followed by the introduction of light-limited and light-saturated conditions in the sixth week. No significant inter-treatment differences were found in the sterol content in weeks one to five. However, significant intra-treatment differences were found in weeks three to five, regardless of nutrient treatment, wherein the fucosterol, 24-methylenecholesterol, and squalene contents of both treatment groups were found to correlate inversely with photosynthetic performance (De Jong et al., 2021). Concentrations of all other sterolic components increased with increasing irradiance and low nutrient conditions, while decreasing or remaining unchanged with increasing irradiance and high nutrient conditions (De Jong et al., 2021). From their data, De Jong et al. (2021) suggests, within their timeframe, the sterol content of Saccharina latissima is unaffected by nutrient availability alone but changes with combined alterations in irradiance and nutrient availability.

Niedzwiedz et al. (2024) studied the response of Saccharina latissima to realistic Arctic summer heatwave scenarios (4 to 10°C) under low- and high-light conditions (3 and 120 μmol photons/m²/s) for 12 days. They found that high-light caused physiological stress in Saccharina latissima (e.g., lower photosynthetic efficiency of photosystem II), which was enhanced by cold and mitigated by warm temperatures, and under low-light conditions, there was no temperature response, likely due to light limitation (Niedzwiedz et al., 2024). However, Saccharina latissima acclimated to light variations by adjusting its chlorophyll-a concentration, meeting cellular energy requirements (Niedzwiedz et al., 2024). Cobos et al. (2025) also studied the response of Arctic Saccharina latissima to light. They collected samples from Kongsfjorden, Svalbard, in early February and incubated them in dim light (6 μmol photons/m²/s) and dark (complete darkness) conditions for seven days. Saccharina latissima responded to light by decreasing its partial derivative carbon13 values, indicating some activation of its carbon concentrating mechanism, and increased its maximum photosynthetic electron transport rate; overall, showing that dim light had the potential to trigger photosynthetic metabolism and growth as early as February (Cobos et al., 2025).

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, more light (UV exposure) combined with higher temperatures produced the greatest negative effects, while low light and moderate temperatures favoured Arctic populations. These findings indicate that warming exacerbates UV-B stress and severely limits recruitment (Müller, Wiencke & Bischof, 2008).

While incident light has an overall positive effect on Laminaria hyperborea at optimal temperatures, the effects of artificial light on kelp are not yet fully understood. There is now a growing body of evidence to show that artificial light at night (ALAN) is widespread in the marine environment, with biologically relevant levels of light penetrating to depths of up to 50m (Davies et al., 2020; Smyth et al., 2021).

ALAN has been shown to change the timing of Ascophyllum nodosum and Fucus serratus reproduction, with receptacles (the reproductive tissues of fucoid macroalgae) continuing to ripen into the winter months instead of peaking in the summer (Moyse et al., 2025). This change in the timing of reproduction could result in gametes being released during suboptimal conditions, such as winter storms, and therefore reduce fertilisation success. Reduced recruitment may lead to shifts in macroalgal assemblages in favour of species which are less sensitive to ALAN, such as Fucus vesiculosus, which seems to be unaffected (Moyse et al., 2025). ALAN can also vary significantly on small spatial scales and therefore affect some macroalgal forests more than others even if they are close to one another. It is therefore possible that ALAN could cause changes in macroalgal assemblages over time.

Lastly, shading of the biotope (e.g. by the construction of a pontoon, pier, etc.) could adversely affect the biotope in areas where the water clarity is also low, and tip the balance to shade-tolerant species, resulting in the loss of the biotope directly within the shaded area, or a reduction in seaweed abundance.

Sensitivity Assessment

An increase in incident light is likely to increase plant productivity and increase the density of the kelps. Constant artificial light may affect the reproductive cues and recruitment in macroalgae, but no evidence was found specifically for Laminaria hyperborea or Saccharina latissima. However, shading, especially from permanent structures (e.g. pontoons, jetties) are likely to reduce incident light and will probably result in the reduction in kelp density, or even its exclusion from the affected area. Therefore, a resistance of ‘Low’ is suggested. Resilience is probably ‘High’ if the shading is temporary but ‘Very low’ if permanent. Therefore, a precautionary sensitivity of ‘High’ is suggested.

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 is little evidence for translocation of Saccharina latissima over significant geographic distances. Nor is there any evidence regarding the genetic modification or effects of translocation of native kelp populations.

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

Laminarians 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).

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.

Sensitivity Assessment

Resistance to the pressure is considered ‘Medium’, and resilience ‘Medium’. The sensitivity of this biotope to introduction of microbial pathogens is assessed as ‘Medium’.

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

Laminaria hyperborea is extracted on a commercial scale in southern Norway, primarily for alginate (Werner & Kraan, 2004). During harvesting, all large canopy-forming sporophytes are removed, although sub-canopy sporophytes and understorey community remain intact (Christie et al., 1998). There has also been recent commercial interest in Saccharina latissima as a consumable called ‘sea vegetables’ (Birkett et al., 1998). However, sporophytes are typically matured on ropes (Handå et al., 2013) and are not directly extracted from the seabed. Thus, evidence to assess the resistance of Saccharina latissima to direct harvesting is limited.

Kelp harvesting can cause significant changes to this biotope through the removal of Laminaria hyperborea and the habitat space that it provides for its associated communities. Removing 26% of the canopy led to a 67% reduction in epiphytes and an 89% reduction of invertebrates (Norderhaug et al., 2020).

Christie et al. (1998) observed Laminaria hyperborea habitat regeneration following commercial Laminaria hyperborea trawling in south Norway. Trawling removed all large canopy-forming adult Laminaria hyperborea, however, sub-canopy recruits were largely unaffected. Within 2 – 6 years of harvesting, a new canopy had formed 1 m off the seabed. The associated holdfast communities recovered in 6 years. However, the epiphytic stipe community did not fully recover in the same time. Christie et al. (1998) suggested that kelp habitats were relatively resistant to direct disturbance/removal of Laminaria hyperborea canopy.

In Nord-Trøndelag, Norway, Laminaria hyperborea was harvested for the first time in 2010 (Steen et al., 2016). Video surveys and plant sampling, conducted 2 days prior to the trawling and in each year for the following 4 years, showed that Laminaria hyperborea coverage had returned to pre-harvest levels (around 94%). However, the new canopy was significantly lower in density, average plant age, length, weight, and epiphyte biomass. In addition, the density of understorey recruits had only recovered by one-third by the end of the study period. It was suggested that 80% of the new canopy consisted of understorey plants that had survived the harvesting, and that the resilience of this biotope is dependent on the rate of harvesting (Steen et al., 2016).

Recurrent disturbance on a timescale shorter than the 2 to 6-year recovery period could prolong the recovery. Kain (1975) cleared sublittoral blocks of Laminaria hyperborea at different times of the year for several years. The first colonizers and succession communities differed between blocks and the time of year they were cleared. However, within 2 years of clearance, the blocks were dominated by Laminaria hyperborea. Leinaas & Christie (1996) also observed Laminaria hyperborea recolonizing urchin barrens following urchin removal. The substratum was initially colonized by filamentous macroalgae and Saccharina latissima. However, Laminaria hyperborea dominated the community after 2 to 4 years.

Laminaria hyperborea forests subjected to regular harvesting support different associated communities compared to unharvested, preserved forests (Leclerc et al., 2015). Macroalgal species richness was consistently higher at the harvested site across all parts of the kelp and on the surrounding rock. Sessile fauna richness was slightly higher on the stipes and surrounding rock at the harvested site, but lower on the holdfast compared to the preserved site. In contrast, mobile fauna richness and density were generally greater on all parts of the kelp in the preserved site, although both were higher on the surrounding rock at the harvested site.

Following disturbance, or in areas experiencing frequent disturbance occurs, Laminaria hyperborea recruitment may be affected by interspecific competition with Non-Indigenous Invasive Species or ephemeral algae (Brodie et al., 2014; Smale et al., 2013). However, evidence for this is limited and thus not included in this assessment.

Sensitivity Assessment

Resistance has been assessed as ‘None’, Resilience as ‘Medium’, and sensitivity as ‘Medium’.

Removal of non-target species

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

Evidence

Both Laminaria hyperborea and Saccharina latissima are commercially harvested (see evidence in Removal of Target Species pressure above, and references therein), so industries targeting one may accidentally impact the other in areas where this biotope occurs. Incidental/accidental removal of Laminaria hyperborea and Saccharina latissima is likely to cause similar effects to that of direct harvesting. It is assumed that incidental non-targeted catch (e.g. by trawls or dredges) could mobilise sediment, remove large kelp species, overturn boulders and cobbles, bury smaller seaweeds, and cause high mortality within the affected area. Therefore, the defining features of this biotope are expected to have no resistance to this pressure.

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).

Sensitivity Assessment

Resistance has been assessed as ‘None’, resilience as ‘Medium’ 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 1887-1890 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., 2011; Helmer et al., 2019; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015). It ranges from the Baltic Sea, the Kattegat and Skagerrak, the North Sea coasts of the UK, Germany, and Belgium, through the English Channels and into the Irish sea coasts of Ireland and south Wales with records in east and west Scotland, Northern Ireland, northwest France, Spain and south into the Mediterranean (NBN, 2023; OBIS, 2026).

Abundances at its northern and southern extremes may be low but densities in UK and France are often over 1000 individuals/m² and it may carpet the seafloor in the Solent and Essex. In the UK, it was reported to reach abundances of >1000 individuals/m² (max. 2,748 individuals/m²) in the Milford Harbour Waterway (Bohn et al., 2012), 84 individuals/m² in Portsmouth, 174 individuals/m² in Langstone and 306 individuals/m² in Chichester harbours in 2017 (Helmer et al., 2019). In France, it has been reported to reach >4,700 individuals/m² in the Bay of Marennes-Oleron, France, 11.6 tonnes/ha in 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 to 100% cover), aided by local oyster farming and shellfish dredging (Blanchard, 2009). However, in Arcachon Bay, France, Crepidula fornicata was limited to only 155 tonnes in 1999 and 312 tonnes in 2011 (De Montaudouin et al., 2001, 2018). It was confined to muddy sediments, which accounted for only 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 oysters 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 recorded from rock, artificial substrata, and Sabellaria alveolata reefs (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Helmer et al., 2019; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015; Tillin et al., 2020).

In the eastern Solent harbours of Portsmouth, Langstone, and Chichester, 75% to 98% of Crepidula larvae settled on dead Crepidula shells, while approx. 4% settled on stone, 2.5% on live Crepidula, 0.3% oyster shell, 0.6% cockle shell, 0.3% winkle shell and 0.1% periwinkle shell (Preston et al., 2020). 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 but, 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 functions 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. (2011) recorded Crepidula off the Isle of Wight in the English Channel, at approx. 60 m on rough ground in areas of high tidal flow. Tillin et al. (2020) suggested that the effect of oscillatory wave meditated 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 approx. 160 m in depth but it most common in the shallow subtidal and low water springs (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 2026; 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 level (approx. 1.8 m) in the Milford Harbour Waterway (Bohn et al., 2015).

The density of Crepidula populations in the northern Europe (Germany, Denmark, and Norway) is significantly lower (<100 individuals/m²) than in southern waters. Thieltges et al. (2004) reported that the population of Crepidula was affected strongly by cold winters in the Wadden Sea. The winters of 2001 and 2003 resulted in approx. 56 to 64% mortality of intertidal Crepidula and up to 97% on one mussel bed, compared to only 11 to 14% in southern areas without frost. Crepidula almost vanished from the Wadden Sea after the 1978/79 winter and took ten years to recover due to moderate winters which regularly affected the population. Similarly, 25% mortality was observed in Crepidula populations on the south coast of the UK after the extreme 1962/63 winter (Crisp, 1964, Bohn et al., 2012). Thieltges et al. (2003) suggested that global warming way allow Crepidula populations become more abundant in northern Europe. Valdizan et al. (2011) noted higher water temperatures between 2000 to 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 favourable breeding temperatures and increased phytoplankton (Valdizan et al., 2011).

Nehls et al. (2006) noted that the decline in mussel (Mytilus edulis) beds in the Wadden Sea was due to mild winters that favoured non-native oysters (Magallana gigas) and slipper limpets, which coexisted with the mussels.

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 (Hinz et al., 2011). However, Hinz et al. (2011) reported that Crepidula fornicata only dominated one assemblage (with an average of 181 individuals per trawl) on gravel substratum with boulders. Bohn et al. (2015) noted that Crepidula occurred at low density or was absent in areas dominated by boulders, and Bohn et al. (2013a, 2013b, 2015) and Preston et al. (2020) showed that while Crepidula could settle on slate panels or ‘stone’ it preferred shell, especially that of conspecifics.

Sensitivity Assessment

At present, there is insufficient evidence to assess this biotope’s sensitivity to Crepidula fornicata invasion.

The carpet sea squirt, Didemnum vexillum

Evidence

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

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

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

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

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

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 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 saltmarshes, 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 to 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 to 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 approx. 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 Botrylloides sp. and was similar to that of Alcyonium digitatum. Alcyonium 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 was associated with the period of greater larval production, and implied that fragmentation aided dispersal.

The oscillatory nature of wave-mediated water flow (wave orbital velocities) combined with wave pressure in the lacerating zone, where breaking wave causes multidirectional strong water movement (Hiscock, 1983), would probably dislodge and break up Didemnum vexillum colonies, prevent them from forming suffocating mats, and restrict the colonies to crevices and overhangs. However, it is unclear if moderately wave exposed conditions would be adequate to prevent Didemnum vexillum from developing extensive mats in the summer months when wave action is typically reduced. Hitchin (2012) suggested that the presence of Didemnum vexillum in Whitstable, Kent was contrary to its then known habitat preferences.

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 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. 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). 

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).

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 individuals/m², little of the underlying substratum remains visible (Herbert et al., 2016). These reefs can stabilize the sediment surface locally (Troost, 2010). When such reefs are formed or, particularly when the species colonizes soft sediments such as mud or sand, it can change and affect local communities, by creating hard substrata for mobile species, which might not otherwise be present before the invasion (Padilla, 2010). However, Hansen et al. (2023) suggested that no immediate ecosystem risk is observed where the Pacific oyster occurs sporadically.

Settlement requires hard substrata, including rock, bedrock, chalk, bare boulders, cobbles and pebbles and shells (Kochmann 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 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 & Jensen, 2013).

Although shorelines comprised of mainly 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; 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 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 to 35 PSU (Troost, 2010; Tillin et al., 2020). Growth of Pacific oysters can occur between 10 to 30 PSU (Troost, 2010).

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

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

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

Dense macroalgal cover is unsuitable for the Magallana gigas (Herbert et al., 2012, 2016; Tillin et al., 2020), being rarely found under macroalgal cover in Northern Ireland, absent from exposed bedrock or large boulders with macroalgae cover in the Solway Firth, Scotland, and absent in Poole Harbour where there was competition with macroalgae (Kochmann 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 kelp canopy. Therefore, this biotope is assessed as ‘Not Sensitive’ to this pressure, but with ‘Low’ confidence due to the lack of direct evidence.

Wireweed, Sargassum muticum

Evidence

Sargassum muticum is a circumglobal invasive species (Engelen et al., 2015). It is recorded (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 if 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 outcompete leathery, canopy-forming macroalgae such as Saccharina latissima, Halidrys siliquosa, and Fucus spp. and, to a lesser degree, understorey species such as Codium fragile, Chondrus crispus and Dictyota dichotoma in Limfjorden, Denmark between 1984 and 1997 (Staehr et al., 2000; Engelen et al., 2015; De Bettignies et al., 2021). The invasion in Limfjorden had 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 to 0.5 m and 4 to 6 m, in the turbid waters of the Limfjorden. Limfjorden is wave sheltered but wave exposure has been reported to restrict the growth and survival of Sargassum muticum (Staehr et al., 2000). Viejo et al. (1995) reported that Sargassum muticum transplanted to wave exposed shores in Spain experienced >80% breakages within a month and that the growth of undamaged plants was significantly lower than that of plants on sheltered shores. Similarly, Andrew & Viejo (1998) noted that Sargassum muticum was restricted to intertidal rockpools in wave exposed sites in the Bay of Biscay. 

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 to 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 the 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, the 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 evidence above suggests that Sargassum muticum prefers wave sheltered, shallow sites in the sublittoral fringe. No evidence of the effects of Sargassum on Laminaria hyperborea beds was found, but Sargassum muticum can compete with and coexist with Saccharina latissima, depending on local conditions. For example, Sargassum muticum outcompeted Saccharina latissima in the Limfjorden but coexisted in the Dorn in Strangford Lough. Therefore, competition with Sargassum is probably site-specific and dependent on local conditions. Since Sargassum muticum can occupy the upper limit of this biotope’s depth range (0 to 5 m), resistance is assessed as ‘Low’ for shallow examples of the biotope and is probably ‘High’ for examples below 5 m. Recovery after invasion by Sargassum would require direct removal, so resilience is assessed as ‘Very low’. Hence, the sensitivity of shallow examples 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

Undaria pinnatifida (Wakame or Asian kelp) is a large brown seaweed and an Invasive Non-Indigenous Species (INIS) that could outcompete native UK kelp species (see Farrell & Fletcher, 2006; Thompson & Schiel, 2012; Brodie et al., 2014; 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 established currently on the coastlines of New Zealand, Australia, Northern France, Spain, Italy, the UK, Portugal, Belgium, Holland, Argentina, Mexico, and the USA (De Leij et al., 2017). Undaria pinnatifida was first recorded in the UK in the Hamble Estuary in 1994 (Macleod et al., 2016) and has since proliferated along UK coastlines. One year after its discovery at the Queen Anne Battery marina, Plymouth, it became a major fouling plant on pontoons (Minchin & Nunn, 2014). Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound. 

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

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

Undaria pinnatifida has a wide physiological niche meaning it can occur in both coastal and estuarine environments showing tolerance for varying salinities, turbidity and siltation (Heiser et al., 2014; Epstein & Smale, 2018). Undaria pinnatifida has a greater preference for sites sheltered with low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018). In natural habitats, Undaria pinnatifida was not recorded if the wave fetch was greater than 642 km but increased in abundance and cover in very sheltered sites (Epstein & Smale, 2018). 

In St Malo, France, there was evidence that Undaria pinnatifida coexisted with Laminaria hyperborea under certain conditions (Castric-Fey et al., 1993). Epstein & Smale (2018) also observed that Undaria pinnatifida was relatively common (abundance of >70 individuals per 25 m transect) at three sites in Devon, UK (Jennycliff, Bovisand and Beacon Cove) where Laminaria spp. were abundant (40 to 79%) or superabundant (>80%), which suggested that Undaria pinnatifida could coexist within refugia amongst areas with dense Laminaria spp. 

In Plymouth Sound, UK, Heiser et al. (2014) observed that Laminaria hyperborea was significantly less abundant at sites with the presence of Undaria pinnatifida, with only approx. 0.5 Laminaria hyperborea individuals per m² present compared to approx. 8 individuals per m² at sites without the presence of Undaria pinnatifida. However, the results from their correlation study only showed that the species were not found together (pers. comm., Epstein, 2021). Whereas exclusion and succession experiments on reefs tell us that Laminaria spp. exclude Undaria pinnatifida, not the other way around. Epstein & Smale (2018) reported that in Devon, UK, persistent, dense, and intact Laminaria spp. canopies in rocky reef habitats exerted a strong influence over the presence/absence, abundance, and percentage cover of Undaria pinnatifida. A dense canopy of native kelp restricts the proliferation of Undaria pinnatifida and disturbance of the canopy is often the key to the recruitment of Undaria pinnatifida. Epstein et al. (2019b) reported that Undaria pinnatifida density and biomass were significantly negatively correlated with the sum of all Laminaria spp. in Plymouth, UK. The evidence indicated that native Laminaria spp. canopies in the UK inhibited Undaria pinnatifida and that Undaria pinnatifida was opportunistic but competitively inferior (Farrell & Fletcher, 2006; Heiser et al., 2014; Minchin & Nunn, 2014; De Leij et al., 2017; Epstein & Smale, 2018; Epstein et al., 2019b). However, Epstein et al. (2019b) also noted that Laminaria hyperborea had a non-significant positive relationship with Undaria pinnatifida due to low densities of Laminaria hyperborea across the study area, resulting in insufficient data. 

In Plymouth Sound (UK), Epstein et al. (2019b) found that within its depth range (+1 to –4 m), Undaria pinnatifida coexisted with seven species of canopy-forming brown macroalgae, including Laminaria hyperborea. De Leij et al. (2017) found that natural habitats with dense native macroalgal canopies, such as Laminaria hyperborea had more resistance to Undaria pinnatifida invasion than disturbed or sparse canopies, due to limited space and light availability for Undaria recruits. However, the dense canopies will not prevent the invasion of Undaria, as sporophytes were still recorded within dense Laminaria canopies, and this suggests that canopy disturbance is not always required. 

Undaria pinnatifida was successfully eradicated on a sunken ship in Chatham 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.

Sensitivity Assessment

Undaria pinnatifida has the potential to colonize and coexist in refugia within Laminaria sp.-dominated habitats, especially in shallow examples of their biotopes that are within its depth range (1 to 4 m) and sheltered from wave action. A dense kelp canopy may restrict or slow the proliferation of Undaria pinnatifida but there is mixed evidence of its colonization with Laminaria hyperborea beds and in some areas. A lower abundance of Laminaria hyperborea may result in increased Undaria pinnatifida growth. Moreover, the above evidence suggests that Undaria pinnatifida can both compete with and coexist with Saccharina latissima, depending on local conditions. For example, Undaria pinnatifida can outcompete competitive native species like Saccharina latissima in artificial habitats, such as in Torquay Marina, but within natural habitats, it can coexist with native kelp species within its depth range (-1 to 4 m), as shown in Plymouth Sound, UK. Furthermore, the degree of sheltering from wave exposure that defines this biotope is also suitable for Undaria colonization. Therefore, resistance is assessed as ‘Low’ for shallow examples of the biotope, i.e. above 5 m in depth, while resistance is probably ‘High’ in examples below 5 m. Recovery after invasion by Undaria would require direct removal, so resilience is ‘Very low’. Hence, sensitivity 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. While not considered a traditional invasive species, its northward expansion into the UK has led to competition with Laminaria hyperborea. In Plymouth Sound, southwest UK, estimates of Laminaria ochroleuca standing stock are now comparable to those of Laminaria hyperborea (Taylor-Robinson et al., 2024; also see Smale et al., 2016 for standing stock of Laminaria hyperborea).

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.

There is contrasting evidence on the relative resilience of Laminaria ochroleuca and Laminaria hyperborea 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, Smale & Vance (2015) found that Laminaria hyperborea was highly resistant to severe storms in the UK during the 2013 to 2014 winter season. 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. Given this conflicting evidence, it remains unclear whether Laminaria ochroleuca biotopes could displace Laminaria hyperborea 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., 2015). 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).

The introduction of Laminaria ochroleuca into Laminaria hyperborea forests can have negative impacts on biodiversity. Kelp stipe assemblages differ significantly between the two species 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 >3600 more than Laminaria ochroleuca stipe assemblages. Therefore, the proliferation of Laminaria ochroleuca could reduce available habitat space for epibionts that are associated with Laminaria hyperborea biotopes.

Sensitivity Assessment

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) suggests that 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., 2015; Taylor-Robinson et al., 2024). Its greater stipe length could reduce light availability for smaller kelps, and when combined with elevated temperatures, could create unfavourable conditions for the persistence and recovery of native species. Laminaria ochroleuca does form mixed forests with Laminaria hyperborea in moderately sheltered to exposed shores, but it has physiological and morphological advantages that could allow it to proliferate if Laminaria hyperborea and/or Saccharina latissima density was reduced. Therefore, resistance to Laminaria ochroleuca is assessed as ‘Low’ based on the evidence of Laminaria ochroleuca rivalling Laminaria hyperborea in Plymouth Sound, southwest UK. 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’.