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

IR.LIR.K.LhypLoch and IR.LIR.K.LhypR.Loch are defined by two kelp species. Laminaria hyperborea has a northern distribution (Birkett et al., 1998b), while Laminaria ochroleuca has a southern distribution. IR.LIR.K.LhypR.Loch and IR.LIR.K.LhypLoch are recorded exclusively in the southwest UK, where summer sea temperatures range from 12 to 16°C, and winter temperatures range from 8 to 13°C (Beszczynska-Möller & Dye, 2013).

Laminaria hyperborea

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). Laminaria hyperborea is near the southern limit of its range in Portugal and the southwest UK, 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). According to OBIS (2026), Laminaria hyperborea occurs in temperatures ranging from 5 to 15°C, with most observations 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. An increase in 5°C in the southwest UK would likely affect Laminaria hyperborea recruitment processes and limit sporophyte growth. 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).

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). 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). However, 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 indicated 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 hours) and long (24:0 hours) 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).

Laminaria ochroleuca

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). In the UK, Laminaria ochroleuca is more abundant in southwest England than in higher and colder latitudes (Smale & Moore, 2017). 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 native kelps such as Laminaria hyperborea.

Laminaria ochroleuca has a thermal optimum for spore development between 12 to 18°C (Izquierdo et al., 2002). Spores have a maximum development temperature of 23 to 24°C and temperatures above 25°C will likely cause high mortality (Pereira et al., 2011).

Long-term climate-related change

Between 1977 and 2007, many cool-water macroalgae including Laminaria hyperborea and Laminaria ochroleuca have almost disappeared and been replaced by warm-water macroalgae on the west coast of Asturias, northern Spain (Fernández, 2016).

Some evidence suggests that Laminaria ochroleuca may have a competitive advantage over Laminaria hyperborea due to its tolerance of warmer waters. 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 in 1997 (Barrientos et al., 2025). In 2023, only eight of the 21 remaining dense forests still had the same canopy-forming species as they did in 1997. Dominance in these sites had shifted from Laminaria hyperborea to Laminaria ochroleuca, and the former species 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). Sea surface temperatures overall increased by around 0.01°C to 0.02°C per year in this region and across this period. Under elevated temperatures (15 and 18°C, compared to 12°C control), Laminaria ochroleuca shows improved growth and photosynthetic efficiency (Hargrave et al., 2017).

Globally, Laminaria hyperborea and Laminaria ochroleuca have experienced a range contraction of 14 and 10% between the 1980s and 2010s, respectively (Casado-Amezúa et al., 2019). The range loss for Laminaria hyperborea is estimated to continue to up to 39.34% under the most extreme greenhouse gas emissions projections (Assis et al., 2016).

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). Predictions show that Laminaria ochroleuca will continue to experience a northward range expansion as far as the Faroe Islands by 2090-2100 under the RCP8.5 emission scenario (Assis et al., 2018). According to OBIS (2026), Laminaria ochroleuca occurs in sea surface temperatures ranging from 10 to 20°C, with most records coming from the 10 to 15°C range.

King et al. (2018) found that the maximum temperature at which heat shock protein production no longer kept pace with thermal stress (T_peak) in Laminaria ochroleuca was 24°C, while protein production stopped completely at 28°C (T_off). Photosynthetic efficiency remained mostly stable across daily 1-hour exposures at 24°C, with only a slight decline on the fourth day, followed by a full recovery over the next three days. However, photosynthetic efficiency declined in the 28°C treatment, followed by a gradual recovery over the next three days. The most severe decline was in the 32°C treatment, where photosynthetic efficiency dropped by roughly half after the first exposure and continued to decline overnight before the second exposure. Kelp discs were non-viable by the end of the second exposure, with no subsequent recovery. Notably, this experiment was designed to reflect summer tidal pool temperatures, so these conditions are unlikely to be directly relevant to these infralittoral biotopes. Nevertheless, the evidence suggests that Laminaria ochroleuca should be able to tolerate increases in temperature at the benchmark level. However, temperatures beyond the benchmark may lead to negative physiological effects and/or mortality.

Laminaria ochroleuca responses to heat shock experiments can vary between locations. Pereira et al. (2015) collected Laminaria ochroleuca individuals from France and Portugal and subjected them to different temperatures in a laboratory. Kelps were kept at 15°C for five days for acclimation and were then subjected to heat shock treatments of varying temperatures. Heat shock treatment temperatures were 22.5, 25, 27.5 and 30°C, each with a ramping up period of 30 minutes from the 15°C baseline. These treatments were applied for one hour every day for four consecutive days, with photosynthetic efficiency measurements taken before and after each heat shock exposure. Photosynthetic efficiency in the French Laminaria ochroleuca was stable across the experiment in the control and 22.5°C treatments, increased across the four days in the 25 and 27.5°C, and decreased across the 30°C treatment with no recovery between exposures. Similar patterns were observed in the Portuguese individuals, apart from the 30°C which again showed an overall decline between each exposure, but also full recovery between exposures. Relative growth rate (%/day) declined with each level of temperature in the French individuals, with negative growth rates (necrosis resulting in size decrease) seen in the 25, 27.5 and 30°C treatments. However, growth rates remained positive in the Portuguese individuals with no significant differences between temperature treatments. This suggests that Laminaria ochroleuca populations in Portugal are more resistant to heat shock than those from the colder waters of northern France.

King et al. (2024) investigated whether pre-exposure to marine heatwaves (MHWs) affected the critical thermal maxima (CT_max) of six marine foundation species including Laminaria ochroleuca. They exposed six individuals to temperatures of 14, 18, and 22°C for control, moderate MHW and extreme MHW treatments, respectively, for 28 days. This was then followed by a 1°C reduction each day until temperatures returned to 14°C. A ramping up period in which temperatures were increased by 2°C every day to a maximum of 30°C was then used to determine CT_max, with photosynthetic efficiency measured each day as a proxy for physiological health. Mortality was determined when photosynthetic efficiency reached zero. In the control treatment, i.e. no pre-exposure to an MHW, overall photosynthetic efficiency began to decline at 28°C when one of the six individuals died. The rest of the individuals then died the following day at 30°C. Prior exposure to MHW conditions reduced thermal tolerance, with effects becoming more pronounced under the more severe treatment. In the moderate MHW treatment, photosynthetic efficiency began to decline at 20°C, one individual died at 24°C, half of the individuals had died by 28°C, and all individuals were dead by 30°C. The extreme MHW treatment showed the strongest response, with photosynthetic efficiency already declining at the start of the ramping up period, one mortality occurring at 18°C, and all individuals dead by 24°C. Estimated CT_max estimates were 23, 28.3, and 29.6°C for the extreme MHW, moderate MHW, and control treatments, respectively. This suggests that resistance could be diminished after repeated warming events. Similarly, Leathers et al. (2026, in review) also found that CT_max can depend on thermal history. In their CT_max trial, where the temperature was increased by 2°C each day, all Laminaria ochroleuca individuals that were pre-exposed the extreme MHW simulation (22°C for 28 days) died at 22 to 24°C, compared to full mortality at 30°C in the control treatment at 14°C with no pre-exposure to an MHW.

The duration and intensity of MHWs have an interactive effect on Laminaria ochroleuca (Leathers et al., 2024). Individuals were exposed to short (14-day) and long (28-day) simulations at control (14°C), moderate (18°C), and extreme (22°C) temperatures, followed by cooling by 1°C/day and a five-day recovery period once temperatures returned to 14°C. Photosynthetic efficiency was largely maintained under short control and moderate MHW treatments, but declined under extreme warming, falling from 0.74 to 0.38 after short extreme MHW recovery and from approximately 0.73 to 0.07 after the long extreme MHW. Relative growth rate remained positive in control and moderate treatments (indicating tissue growth) but declined in extreme treatments, from approximately 0.6 to -0.76%/day after the short MHW and from approximately 0 to -2.26%/day after the long MHW. Bleaching occurred only under extreme MHWs, reaching a bleaching index of approx. 3 after short MHW recovery and 6 (total bleaching of the tissue) after long-MHW recovery (Leathers et al., 2024).

Bass et al. (2023) showed that Laminaria ochroleuca responses to 28-day marine heatwaves (MHWs) can differ between seasons and levels of light. In the spring, Laminaria ochroleuca showed in increase in biomass across all temperature treatments (10°C control, 12°C moderate MHW, and 14°C extreme MHW) in the high-light treatment (100 μmol photons/m²/s). However, in the low-light treatment (10 μmol photons/m²/s), Laminaria ochroleuca showed almost no growth across all temperature treatments. In the summer, high-light treatment yielded some growth in the control treatment (18°C), but almost no growth occurred in either of the MHW treatments. Finally, no growth occurred in the control treatment combined with low light, while the moderate MHW (20°C) showed a slight reduction in biomass, and the extreme MHW (22°C) showed a significant reduction in biomass. Photosynthetic efficiency was mostly unaffected in all treatment combinations in the spring. Conversely, it decreased in all temperature treatments combined with high light in the summer. The smallest response was seen in the moderate MHW treatment, followed by an intermediate response in the control, and the strongest decline in the extreme MHW. In the low-light treatment, Laminaria ochroleuca in the control and moderate MHW treatments were minimally affected, while the extreme MHW caused a significant decline in photosynthetic efficiency (Bass et al., 2023).

Responses to marine heatwaves can also vary between populations. Laminaria ochroleuca gametophytes from northern France, northern Spain, Morocco, and Italy, each experiencing different thermal conditions, showed differences in survival, reproductive success, and photosynthesis under 11-day MHW simulations (Strasser et al., 2022). Gametophytes from the northernmost population (France), where sea temperatures are lower, experienced increased mortality at 23°C compared to the control (17°C), whereas the other populations only showed mortality at 27°C. During recovery from the 23 and 25°C treatments, reproductive success was reduced in the Spanish and Moroccan populations. In contrast, after exposure to 27°C, these same populations showed higher reproductive success than the other populations. Moreover, photosynthetic efficiency decreased across the experiment period in all treatments, even the control. In the control, photosynthetic efficiency continued to decline throughout the 17-day recovery period. In the recovery period following the 23 and 25°C treatments, photosynthetic efficiency showed a strong recovery in the French and Italian populations, and a weak recovery in the Moroccan population. Moreover, in the 27°C treatment, photosynthetic efficiency recovered fully in all populations, indicating that higher temperatures enable recovery from photosynthetic decline. Strasser et al. (2022) suggested that this surprising response could be due to an interaction between irradiance and temperature resulting in the production of new cells with increased resilience. They also concluded that responses to thermal stress are influenced by genetic differences between populations.

On the coast of Asturias, northern Spain, Laminaria hyperborea, Laminaria ochroleuca, and Saccorhiza polyschides were dominant species between the early 1980s and early 1990s but were found to have mostly disappeared by surveys conducted between 2009 and 2023 (Rico et al., 2026). While the authors did not formally test relationships between warming and kelp loss, they combined their kelp abundance data with MHW records from Izquierdo et al. (2022). They noted that the decline of Laminaria ochroleuca in 2007 and the disappearance of Laminaria hyperborea in 2009 followed a period in 2006 characterized by 10 maximum temperature anomalies which exceeded the 90^th percentile threshold by up to 2°C.

Other characterizing species

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).

Sensitivity Assessment

While UK populations of Laminaria ochroleuca appear to be resistant to increases in sea surface temperature at the pressure benchmark level, Laminaria hyperborea will likely experience some mortality and diminished physiological function. The loss of a portion of Laminaria hyperborea could result in a change in biotope from IR.LIR.K.LhypR.Loch to the Laminaria ochroleuca-dominated IR.LIR.K.LhypLoch biotope. If Laminaria hyperborea was reduced or lost in the Laminaria ochroleuca-dominated IR.LIR.K.LhypLoch, the biotope’s diversity would we reduced but probably still recognised.

Therefore, the resistance of both biotopes is assessed as ‘Medium’. Resilience is deemed to be ‘High’ in LhypLoch due to Laminaria ochroleuca’s competitive advantage over Laminaria hyperborea, i.e. its greater thermal tolerance and greater stipe length which would limit Laminaria hyperborea recruitment. Therefore, the sensitivity of LhypLoch is assessed as ‘Low’. However, in LhypRLoch, the delayed recovery of Laminaria hyperborea abundance would decrease the biotope resilience. Hence, it is assessed as ‘Medium’ and sensitivity as ‘Medium’.

Temperature decrease (local)

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

Evidence

IR.LIR.K.LhypLoch and IR.LIR.K.LhypR.Loch are defined by two kelp species. Laminaria hyperborea has a northern distribution (Birkett et al., 1998b), while Laminaria ochroleuca has a southern distribution. IR.LIR.K.LhypR.Loch and IR.LIR.K.LhypLoch are recorded exclusively in the southwest UK, where summer sea temperatures range from 12 to 16°C, and winter temperatures range from 8 to 13°C (Beszczynska-Möller & Dye, 2013).

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.

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 this range (Birkett et al., 1988). According to OBIS (2026), Laminaria hyperborea occurs in temperatures ranging from 5 to 15°C, with most observations coming from the 10 to 15°C range.

Laminaria ochroleuca has a reported temperature optimum for spore development between 12 to 18°C (Izquierdo et al., 2002), Laminaria ochroleuca gametophyte development can occur in temperatures as low 5°C (Lüning, 1990). Recent modelling has also predicted winter temperature in the UK to strongly influence the distribution, and hence spread, of Laminaria ochroleuca in the UK (Yesson et al., 2015). Pereira et al. (2011) examined the development of early Laminaria ochroleuca life stages and found that Laminaria ochroleuca had relatively low fecundity at 10°C and may be maladapted to environmental conditions within the UK and vulnerable to local extinctions. In the UK, Laminaria ochroleuca is more abundant in southwest England than other regions than in higher and colder latitudes (Smale & Moore, 2017). This is also corroborated by records from OBIS (2026), which show that this species occurs in sea surface temperatures ranging from 10 to 20°C, with most records coming from the 10 to 15°C range.

King et al. (2018) found that one-hour exposures to temperatures of 0 and 4°C (to simulate winter tide pool conditions) had no significant effect on the photosynthetic efficiency of Laminaria ochroleuca. Photosynthetic efficiency declined slightly over the 4-day experiment with repeated exposures, but values did not differ significantly from those in the control group at 15°C.

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), maintenance of sea temperatures above 13°C may affect recruitment success.

Sensitivity Assessment

IR.LIR.K.LhypLoch and IR.LIR.K.LhypR.Loch are recorded exclusively in the UK southwest, where the average Sea Temperature (ST) range from 12 to 16°C, and winter 8 to 13°C (Beszczynska-Möller & Dye, 2013). Little evidence is available for an acute cold shock in both Laminaria ochroleuca and Laminaria hyperborea. A decrease of 5°C for one month during winter could negatively affect Laminaria ochroleuca fecundity, spore, and gametophyte development and hence recruitment success. A decrease of 2°C for one year could negatively affect fecundity and spore development. A decrease in temperature may cause local extinctions of Laminaria ochroleuca and/or increase the dominance of either monospecific Laminaria hyperborea (as in IR.MIR.KR.Lhyp) or mixed Saccharina latissima canopies (IR.LIR.K.LhypSlat), dependant on the wave exposure.

Evidence for the effects of this pressure at the benchmark level is lacking, so this assessment is based on the known distribution and physiological limits of both species. Therefore, confidence in the assessment is ‘Low’. Since Laminaria ochroleuca’s distribution is limited by low winter temperatures, and their photosynthetic efficiency declines slightly on a timescale much shorter than that of the pressure benchmark, a precautionary resistance of ‘Low’ is given, since the loss of Laminaria ochroleuca would shift this biotope towards a Laminaria hyperborea-dominated biotope. Resilience is ‘High’ if temperatures return to normal, so the biotope’s sensitivity is assessed as ‘Low’.

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

Lüning (1990) suggest that kelps are stenohaline, their general tolerance to salinity as a phenotypic group covering 16 to 50 PSU over a 24-hr period. Optimal growth probably occurs between 30 to 35 PSU (MNCR category - 'Full' salinity) and growth rates are likely to be affected by periodic salinity stress.

According to OBIS (2026), Laminaria hyperborea occurs in salinities ranging from 25 to 40 PSU, with most records coming from the 30 to 35 PSU range. 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.

Laminaria ochroleuca grows in the Messina strait, Mediterranean where ambient salinity has been measured at 38.5 PSU (Sheppard et al., 1978). Laminaria ochroleuca occurs in salinities ranging from 30 to 40 PSU, with slightly more records from the 30 to 35 PSU range. Laminaria ochroleuca has also been found growing in Portuguese intertidal rock pools (Barradas et al., 2011). As a result of high evaporation rates, rainfall and land run-off, rock pools and the organisms that reside within them are likely to experience short-term high variability in salinity (Reed & Russell, 1978). The effects of long-term (e.g. ≥ 1 year) salinity changes on Laminaria ochroleuca are unknown.

Sensitivity Assessment

An increase in salinity at the benchmark level would result in hypersaline conditions in these normally fully saline biotopes. Therefore, resistance has been assessed as ‘Low’ for both biotopes. Resilience and sensitivity are ‘High’ and ‘Low’, respectively, for the Laminaria ochroleuca-dominated biotope LhypLoch, while the resilience and sensitivity of the Laminaria hyperborea-dominated biotope LhypR.Loch are both ‘Medium’. Confidence in these assessments is ‘Low’ due to the lack of direct evidence.

Salinity decrease (local)

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

Evidence

Lüning (1990) suggest that kelps are stenohaline, their general tolerance to salinity as a phenotypic group covering 16 to 50 PSU over a 24-hr period. Optimal growth probably occurs between 30 to 35 PSU (MNCR category - 'Full' salinity) and growth rates are likely to be affected by periodic salinity stress.

According to OBIS (2026), Laminaria hyperborea occurs in salinities ranging from 25 to 40 PSU, with most records coming from the 30 to 35 PSU range. 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.

Hopkin & Kain (1978) tested Laminaria hyperborea sporophyte growth at various low salinity levels. The results showed that Laminaria hyperborea sporophytes grew normally at 19 PSU, was reduced at 16 PSU, and did not occur at 7 PSU. 

Laminaria ochroleuca grows in the Messina strait, Mediterranean where ambient salinity has been measured at 38.5 PSU (Sheppard et al., 1978). Laminaria ochroleuca occurs in salinities ranging from 30 to 40 PSU, with slightly more records from the 30 to 35 PSU range. Laminaria ochroleuca has also been found growing in Portuguese intertidal rock pools (Barradas et al., 2011). As a result of high evaporation rates, rainfall and land run-off, rock pools and the organisms that reside within them are likely to experience short-term high variability in salinity (Reed & Russell, 1978). The effects of long-term (e.g. ≥ 1 year) salinity changes on Laminaria ochroleuca are unknown.

Sensitivity Assessment

A decrease in one MNCR salinity scale from Full Salinity (30 to 40 PSU) to Reduced Salinity (18 to 30 PSU) may result in a decrease of Laminaria hyperborea sporophyte growth. Laminaria hyperborea may also be outcompeted by low salinity tolerant species e.g. Saccharina latissima (Karsten, 2007). However, the effects of long-term (e.g. ≥ 1 year) salinity changes on Laminaria ochroleuca are unknown, so the sensitivity assessment for this species is based on the salinity levels their known distribution. Resistance is assessed as ‘Low’ for both biotopes. Resilience and sensitivity are ‘High’ and ‘Low’, respectively, for the Laminaria ochroleuca-dominated biotope LhypLoch, while the resilience and sensitivity of the Laminaria hyperborea-dominated biotope LhypR.Loch are both ‘Medium’. Confidence in the assessment of LhypLoch is ‘Low’ due to the lack of direct evidence.

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

Kregting et al. (2013) measured Laminaria hyperborea blade growth and stipe elongation from an exposed and a sheltered site in Strangford Lough, Ireland, from March 2009-April 2010. Maximal significant wave height (Hm0) was 3.67 & 2m at the exposed and sheltered sites, and maximal water velocity (Velrms) was 0.6 & 0.3m/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. Therefore, water flow was found to have no significant effect on Laminaria hyperborea growth at the observed range of water velocities.

Biotope structure is, however, different between wave exposed and sheltered sites. Pedersen et al. (2012) observed Laminaria hyperborea biomass, productivity and density increased with an increase in wave exposure. At low wave exposure, Laminaria hyperborea canopy forming plants were smaller, had lower densities and had higher mortality rates than at exposed sites. At low wave exposure Pedersen et al. (2012) suggested that high epiphytic loading on Laminaria hyperborea impaired light conditions, nutrient uptake, and increased the drag on the host Laminaria hyperborea during extreme storm events.

The morphology of the stipe and blade of kelps vary with water flow.  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). However, the stipe of Laminaria hyperborea is relatively stiff and can snap in strong currents. Laminaria hyperborea is usually absent from areas of high wave action or strong currents, although it is found  in the Menai Strait, Wales, where tidal velocities can exceed 4 m/s (NBN, 2015) and in tidal rapids in Norway (J. Jones, pers. comm.)  Laminaria hyperborea growth can persist in very strong tidal streams (>3 m/s).

Increase water flow rate may also remove or inhibit grazers including Patella pellucida and Echinus esculentus and remove epiphytic algae growth (Pedersen et al., 2012). The associated algal flora and suspension feeding faunal populations change significantly with different water flow regimes. Increased water flow rates may reduce the understorey epiflora, to be replaced by an epifauna dominated community (e.g. sponges, anemones and polyclinid ascidians) as in the biotope IR.HIR.KFaR.LhypFa. The composition of the holdfast fauna may also change, e.g. energetic or sheltered water movements favour different species of amphipods (Moore, 1985).

IR.HIR.KFaR.LhypR, IR.HIR.KFaR.LhypFa, IR.MIR.KR.Lhyp, and their associated sub-biotopes are found within strong (1.5-3 m/s)-moderate (0.5-1.5 m/s) tidal streams. A change in peak mean spring bed flow velocity which does not result in a change in tidal streams above or below 0.5-3 m/s is not likely to affect the dominance of Laminaria hyperborea within the community, but may cause changes in the understorey community. The prominent understorey filter feeding community within IR.HIR.KFaR.LhypFa is reliant on high water movement. A decrease in tidal streams may result in a decline of filter feeding fauna and an increase in red seaweeds within the understorey community or vice versa with an increase in tidal streams A decrease in tidal flow within this range may also decrease urchin dislodgment and increase urchin grazing. An increase in urchin grazing may cause a decline in the understorey community abundance and diversity (as in IR.MIR.KR.Lhyp.GzFt/Pk and IR.MIR.KR.LhypPar).

Sensitivity assessment. A change in peak mean spring bed flow velocity of between 0.1m/s to 0.2m/s for more than 1 year is not likely to affect the dominance of Laminaria hyperborea, however, subtle differences in tidal regime may influence the understorey community. Resistance to the pressure is considered ‘High’, and resilience ‘High’. Hence, the sensitivity of this biotope to changes in peak mean spring bed velocity has been assessed as ‘Not Sensitive’.

However, if peak mean spring bed flow velocity changes but remains within 0.5-3 m/s Laminaria hyperborea is likely to remain the dominant habitat but the understorey community may be affected; directly by a change in water velocity or through increased grazing pressure.

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

The upper limit of the Laminaria hyperborea bed is determined by wave action and water flow, desiccation, and competition from the more emergence resistant Laminaria digitata. Laminaria hyperborea exposed at extreme low water are very intolerant of desiccation, the most noticeable effect being bleaching of the frond and subsequent death of the meristem and loss of the plant. An increase in wave exposure (see below- water flow), as a result of increased emergence, has been found to exclude Laminaria hyperborea from shallow waters due to dislodgement of the sporophyte or snapping of the stipe (Birket et al., 1998). Hence, an increase in emergence is likely to lead to mortality of exposed Laminaria hyperborea and the associated habitat.

An increase in water depth/decreased emergence (at the benchmark level) may increase the upper depth restriction of Laminaria hyperborea forest biotope variants. However, limited light availability at depth will decrease the lower extent of Laminaria hyperborea, and may, therefore, result in a shift from forest to park biotope variants at depth. Further increases in depth will cause a community shift to that characterized by circalittoral faunal species, however, this is beyond the scope of the benchmark.

Sensitivity assessment. Resistance to the pressure is considered ‘Low’, and resilience ‘Medium’. The sensitivity of this biotope to changes in tidal emergence has been assessed as ‘Medium’.

Wave exposure changes (local)

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

Evidence

Laminaria hyperborea and Laminaria ochroleuca coexist in IR.LIR.K.LhypR.Loch and IR.LIR.K.LhypLoch biotopes in the southwest UK. The former exists on exposed shores and has a higher abundance of Laminaria hyperborea relative to Laminaria ochroleuca. The latter exists on moderately exposed or sheltered shores and has a higher relative abundance of Laminaria ochroleuca. There is conflicting evidence on the resistance and resilience of each species to extreme increases in wave action, such as storm events. 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 after two years. 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.

Kregting et al. (2013) measured Laminaria hyperborea blade growth and stipe elongation from an exposed and a sheltered site in Strangford Lough, Ireland from March 2009-April 2010. Wave exposure was found to be between 1.1 and 1.6 times greater between the exposed and sheltered sites. Maximal significant wave height (Hm0) was 3.67 & 2 m at the exposed and sheltered sites. Maximal water velocity (Velrms) was 0.6 & 0.3 m/s at the exposed and sheltered sites. Despite the differences in wave exposure and water velocity, there was no significant difference in Laminaria hyperborea growth between the exposed and sheltered site.

Biotope structure is, however, different between wave exposed and sheltered sites. Pedersen et al. (2012) observed Laminaria hyperborea biomass, productivity and density increased with an increase in wave exposure. At low wave exposure, Laminaria hyperborea canopy forming plants were smaller, had lower densities and had higher mortality rates than at exposed sites. At low wave exposure high epiphytic loading on Laminaria hyperborea was theorised to impair light conditions, nutrient uptake, and increase the drag of the host Laminaria hyperborea during extreme storm events.

Sensitivity Assessment

Wave exposure significantly effects the relative abundance of each of the kelp species in both Laminaria ochroleuca biotopes (LhypR.Loch and LhypLoch) because Laminaria ochroleuca is less tolerant of wave exposure than Laminaria hyperborea.

LhypR.Loch

Laminaria hyperborea is ‘Abundant’ while Laminaria ochroleuca is only ‘Common’ in HIR.KFaR.LhypR.Loch, which occurs in Very exposed and Exposed wave conditions (JNCC, 2022). However, JNCC (2022) note that the biotope commonly occurs below exposed kelp forest biotopes (LhypR.Ft) because Laminaria ochroleuca is less tolerant of wave exposure than Laminaria hyperborea, in agreement with the above evidence. Therefore, an increase in wave exposure (by one category) to extremely exposed may reduce the Laminaria ochroleuca abundance further resulting in an Laminaria hyperborea dominated kelp biotope typical of wave exposed conditions (e.g., LhypR.Ft). However, a reduction in wave exposure from exposed to moderately exposed may allow the Laminaria ochroleuca abundance to increase, resulting in a biotope similar to LIR.K.LhypLoch. The change to LhypLoch would be gradual and probably take longer than the one-year period (defined within the benchmark), but an increase in wave exposure may remove a significant proportion of the Laminaria ochroleuca relatively quickly. Hence, resistance is assessed as ‘Low’. Nevertheless, resilience is probably ‘High’, and sensitivity is assessed as ‘Low’, albeit with ‘Low’ confidence.

LhypLoch

Laminaria ochroleuca is ‘Abundant’ while Laminaria hyperborea is only ‘Occasional’ in LIR.K.LhypLoch, which occurs in moderately exposed and sheltered wave conditions (JNCC, 2022). JNCC (2022) note that Laminaria ochroleuca can form dense forest (LhypLoch) below exposed kelp forest biotopes (LhypR.Ft) but above park (LhypR.Ft). A decrease in wave exposure (by one category) to very sheltered may allow opportunistic kelps (e.g. Saccharina) to increase in abundance but the longer lived Laminaria species would probably remain during the one-year period (defined within the benchmark). However, an increase in wave exposure (by one category) to exposed may reduce the Laminaria ochroleuca abundance resulting in an Laminaria hyperborea dominated mixed biotope (LhypR.Loch) or a kelp biotope typical of wave exposed conditions (e.g., LhypR.Ft). An increase in wave exposure may remove a proportion of the Laminaria ochroleuca relatively quickly. Hence, resistance is assessed as ‘Medium’. Nevertheless, resilience is probably ‘High’, and sensitivity is assessed as ‘Low’, albeit with ‘Low’ confidence.

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.

Laminaria ochroleuca significantly reduces copper (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).

Arsenic has been detected in concentrations as high as 99mg/kg in Laminaria ochroleuca samples from China (Yu et al., 2024).

Bryan (1984) suggested that the general order for heavy metal toxicity in seaweeds is: Organic Hg > inorganic Hg > Cu > Ag > Zn > Cd > Pb. Cole et al. (1999) reported that Hg was very toxic to macrophytes. Similarly, Hopkin & Kain (1978) demonstrated sub-lethal effects of heavy metals on Laminaria hyperborea gametophytes and sporophytes, including reduced growth and respiration. 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). Echinus esculentus recruitment is likely to be impaired by heavy metal contamination due to the intolerance of its larvae. Echinus esculentus are long-lived and poor recruitment may not reduce grazing pressure in the short-term. 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.

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 fronds, being almost exclusively subtidal, would not come into contact with freshly released oil, but only to sinking emulsified oil and oil adsorbed onto particles (Birket et al., 1998). The mucilaginous slime layer coating of laminarians 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 hyperborea 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. Overall the red algae are likely to be highly intolerant to hydrocarbon contamination. 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.

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.

O'Brian & Dixon (1976) suggested that red algae were the most sensitive group of macrophytes to oil and dispersant contamination (see Smith, 1968). 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). 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. 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 insensitive 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 have been shown to inhibiting both photosynthesis and respiration in macroalgae (Kinne, 1977). Despite this, macroalgae are thought to buffer the environmental conditions of low oxygen, thereby acting as a refuge for organisms in oxygen depleted regions especially if the oxygen depletion is short-term (Frieder et al., 2012).  A rapid recovery from a state of low oxygen is expected if the environmental conditions are transient. If levels do drop below 4 mg/l negative effects on these organisms can be expected with adverse effects occurring below 2mg/l (Cole et al., 1999).

Sensitivity Assessment. Reduced oxygen levels are likely to inhibit photosynthesis and respiration but not cause a loss of the macroalgae population directly.  However, small invertebrate epifauna may be lost, causing a reduction in species richness. Therefore a resistance of ‘High’ is recorded.  Resilience is likely to be ‘High’, and the biotopes is probably ‘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

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

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 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 nutrient enrichment since healthy populations are found at ends of sublittoral 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 stipe production but with concomitant decreases in species numbers and diversity (Fletcher, 1996).

Increased nutrients may result in phytoplankton blooms that increase turbidity. Increased nutrients may favour sea urchins, e.g. Echinus esculentus, due their ability to absorb dissolved organics, and result in increased grazing pressure leading to loss of understorey epiflora/fauna, decreased kelp recruitment and possibly '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.

Sensitivity Assessment

The above evidence suggests that increased nutrients may benefit Laminaria hyperborea kelp beds but alter the associated community. No evidence was found on the effects of nutrient enrichment on Laminaria ochroleuca. However, in extreme cases, phytoplankton blooms and turbidity as a result of nutrient enrichment may attenuate light and be detrimental to the kelps that define these biotopes. Hence, resistance is assessed as ‘Medium’ for both biotopes under assessment to represent the potential loss of the associated community. Resilience and sensitivity are ‘High’ and ‘Low’, respectively, for the Laminaria ochroleuca-dominated biotope LhypLoch, while the resilience and sensitivity of the Laminaria hyperborea-dominated biotope LhypR.Loch are both ‘Medium’.

Organic enrichment

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

Evidence

Holt et al. (1995) suggest that Laminaria hyperborea may be tolerant of organic enrichment since healthy populations are found at ends of sublittoral 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). Increase 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. Therefore, although nutrients may not affect kelps directly, indirect effects such as turbidity may significantly affect the structure of Laminaria hyperborea biotopes.

Sensitivity assessment. Resistance to the pressure is considered 'Medium', and resilience 'High'. The sensitivity of this biotope to organic enrichment is assessed as 'Low'.

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 Laminaria hyperborea would not be able to tolerate (Birket et al., 1998). 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 on bedrock habitats

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 rock substrata.

Abrasion / disturbance of the surface of the substratum or seabed

Benchmark. Damage to surface features (e.g. species and physical structures within the habitat) (Surface abrasion/disturbance pressure definition).

Evidence

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).

Barradas et al. (2011) scraped all macro-algae from intertidal Portuguese rock pools, including the dominant canopy forming Laminaria ochroleuca and understorey algae. Small (mean length 2.02cm) Laminaria ochroleuca recruits appeared 1 month following removal at a mean density of 40 recruits/m². Four months after removal sporophytes had an average length of 14.91cm and an average density of 82 recruits per m². Barradas et al. (2011) noted a lack of recruitment in natural adjacent Laminaria ochroleuca populations and theorized the rapid colonization of Laminaria ochroleuca was caused by latent microscopic spores on the underlying rock which grew rapidly when the Laminaria ochroleuca canopy was removed. Barradas et al. (2011) did not comment on the length of time for Laminaria ochroleuca to reach similar size and density to that of pre-treatment nor the likely environmental conditions which spores could tolerate and for how long, however, the results in Barradas et al. (2011) demonstrate Laminaria ochroleuca can recover from disturbance rapidly.

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 due to passing bottom gear has the potential to remove a significant proportion of the standing kelp population, similar to direct harvesting. Therefore, resistance in probably ‘Low’. Laminaria ochroleuca has been shown to recovery rapidly (within 1 to 2 years) following complete kelp canopy removal, so the LhypLoch resilience is ‘High’, and sensitivity is ‘Low’. Laminaria hyperborea has been shown can recover within 2 to 6 and the associated community 7 to >10 years (Birkett et al., 1998), so resilience is ‘Medium’ and sensitivity is ‘Medium’.

Penetration or disturbance of the substratum subsurface

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

Evidence

Not Relevant, please refer to pressure “Abrasion/disturbance of the substratum or seabed

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 kelp species can grow and it has been reported that laminarians grow down to 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 6-7 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 Laminaria hyperborea 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 (Birkett et al. 1998b; Lüning, 1990).

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 Laminaria ochroleuca and decrease Laminaria hyperborea abundance and density (see sub-biotope- IR.MIR.KR.Lhyp.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 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 out-competed 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).

Díez et al., (2003) studied subtidal vegetation distribution patterns in relation to environmental conditions (pollution, wave exposure, sedimentation, substratum slope and depth) in northern Spain. The results showed that Laminaria ochroleuca only occurred at sites without sedimentation loading. 

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 significant decline and resistance to this pressure is assessed as ‘Low’. Resilience to this pressure is probably ‘Medium’ at the benchmark.  Hence, this biotope is assessed as having a sensitivity of ‘Medium ‘to this pressure.

Smothering and siltation rate changes (light)

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

Evidence

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 kelp species can grow and it has been reported that laminarians grow down to 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 6-7 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 Laminaria hyperborea 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 (Birkett et al. 1998b; Lüning, 1990).

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 Laminaria ochroleuca and decrease Laminaria hyperborea abundance and density (see sub-biotope- IR.MIR.KR.Lhyp.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 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 out-competed 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).

Díez et al., (2003) studied subtidal vegetation distribution patterns in relation to environmental conditions (pollution, wave exposure, sedimentation, substratum slope and depth) in northern Spain. The results showed that Laminaria ochroleuca only occurred at sites without sedimentation loading. 

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 significant decline and resistance to this pressure is assessed as ‘Low’. Resilience to this pressure is probably ‘Medium’ at the benchmark.  Hence, this biotope is assessed 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. 30 cm of material) during a discrete event is unlikely to damage Laminaria hyperborea plants but is likely to affect gametophyte survival, holdfast communities, epiphytic community at the base of the stipe, and interfere with zoospore settlement. Given the microscopic size of the gametophyte, 30 cm 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 within 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.

If clearance of deposited sediment occurs rapidly then understorey communities are expected to recover quickly. If inundation is long lasting then the understorey epifauna/flora may be adversely affected, e.g. suspension or filter feeding fauna and/or algal species.  While this  biotope occurs in high to moderate energy habitats (due to water flow or wave action) deposition of 30 cm of sediment represents a large volume of material that would likely remain for a number of tidal cycles and is expected to damage understorey flora/fauna as well as juvenile Laminaria hyperborea.

Sensitivity assessment. Resistance to the pressure is considered ‘Medium’, and resilience ‘High’. The sensitivity of this biotope to heavy deposition of up to 30cm of fine material added to the seabed in a single discreet event is assessed as ‘Low’.

Litter

Benchmark. The introduction of man-made objects able to cause physical harm (surface, water column, seafloor or strandline) (Litter pressure definition). 

Evidence

Not assessed.

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT (Electromagnetic pressure definition).

Evidence

Evidence on the effect of electromagnetic fields (EMFs) on benthic organisms is still severely lacking. No studies examining the effect of EMFs on macroalgae were found. Some studies have investigated the effect of anthropogenically induced EMFs on benthic invertebrates at intensities ranging between 2 nT and 40 mT, which is often much higher than in-situ measurements from subsea cables. While some report changes to behaviour, physiology, reproduction, development, immunology, cytotoxicity and orientation, others demonstrate no effect from exposure to the EMF (Albert et al., 2020; Hutchison et al., 2020), depending on the study species and duration and intensity of exposure. No studies investigating the effect of EMFs at the population or community level for benthic organisms were found.

Sensitivity Assessment

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

Underwater noise changes

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

Evidence

Not relevant

Introduction of light or shading

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

Evidence

Light availability is a key environmental factor influencing the distribution, morphology, and productivity of kelps. 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 in Laminaria hyperborea. 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/s 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 10°C, 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 5°C and 0°C 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 0°C but increased slightly with the polar night treatment. In the same time frame, photosynthetic efficiency declined significantly at 5°C 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 hours) and long (24:0 hours) 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.

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.

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 kelps. 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’ and ‘Medium’ for LhypLoch and LhypR.Loch, respectively, 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

Recently, Laminaria ochroleuca has been found to hybridise with Laminaria digitata in Europe (De Clercq, 2025, cited in Peeters et al., 2026, abstract only). Hybridised kelps showed improved growth rates across all 18-day marine heatwave (MHW) treatments (12°C control, 18, 22, and 24°C) compared to Laminaria digitata, and survived temperatures above 22°C whilst Laminaria digitata underwent severe bleaching. Peeters et al. (2026, abstract only) concluded that hybrids with Laminaria ochroleuca mothers (i.e., female gametes were provided by Laminaria ochroleuca as opposed to Laminaria digitata) show increased resistance to extreme temperatures.

Sensitivity Assessment

No evidence of genetic modification of the characteristic kelps was found. Recent (2026) evidence suggests that the non-native Laminaria ochroleuca may hybridize with Laminaria digitata, resulting in temperature tolerant hybrids. However, no evidence was found on the potential population effects of hybrids in UK waters. Therefore, the evidence is currently ‘Insufficient’ to form the basis of an assessment.

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

Galls on the blade of Laminaria hyperborea and spot disease are associated with the endophyte Streblonema sp. although the causal agent is unknown (bacteria, virus or endophyte). The resultant damage to the blade and stipe may increase losses in storms. The endophyte inhibits spore production and, therefore, recruitment and recoverability (Lein et al., 1991).

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

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale (targeted removal pressure definition).

Evidence

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. Within the study area trawling removed all large canopy-forming adult Laminaria hyperborea, however, sub-canopy recruits were unaffected. Within 2 to 3 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 within the same time period. Christie et al. (1998) suggested that kelp habitats were relatively resistant to direct disturbance 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 occurring at a smaller time scale than the recovery period of 2 to 6 years (stated above) could extend recovery time. Kain (1975) cleared sublittoral blocks of Laminaria hyperborea at different times of the year for several years. The first colonizers and succession community differed between blocks and at what time of year the blocks were cleared however within 2 years of clearance the blocks were dominated by Laminaria hyperborea. Leinaas & Christie (1996) also observed Laminaria hyperborea recolonization of “urchin barrens”, following removal of urchins. The substratum was initially colonized by filamentous macroalgae and Saccharina latissima however after 2 to 4 years Laminaria hyperborea dominated the community.

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 were recurrent rapid disturbance occurs Laminaria hyperborea recruitment could also be affected by interspecific competitive interactions with Invasive Non-Indigenous Species  or ephemeral algal species (Brodie et al., 2014; Smale et al., 2013), however, evidence for this is limited and thus not included within this assessment. No evidence was found for the harvesting of Laminaria ochroleuca.

Sensitivity Assessment

Although there was no evidence of Laminaria ochroleuca being targeted for harvesting, trawling for Laminaria hyperborea in this biotope would not only remove the Laminaria hyperborea, resulting in a pure Laminaria ochroleuca stand, but would also decrease Laminaria ochroleuca abundance. Since neither kelps have any defence against these harvesting methods, resistance for both LhypR.Loch and LhypLoch biotopes is assessed as ‘None’. Resilience and sensitivity for LhypR.Loch are both ‘Medium’, while for LhypLoch they are ‘High’ and ‘Medium’, respectively.

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

Incidental/accidental removal of Laminaria hyperborea from extraction of other marine resources, e.g. fisheries or aggregates, is likely to cause similar effects to that of direct harvesting of Laminaria hyperborea; hence, the same evidence has been used for both pressure assessments.

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

Recurrent disturbance occurring at a smaller time scale than the recovery period of 2-6 years (stated above) could extend recovery time. Kain (1975) cleared sublittoral blocks of Laminaria hyperborea at different times of the year for several years. The first colonizers and succession community differed between blocks and at what time of year the blocks were cleared however within 2 years of clearance the blocks were dominated by Laminaria hyperborea. Leinaas & Christie (1996) also observed Laminaria hyperborea re-colonization of “urchin barrens”, following removal of urchins. The substratum was initially colonized by filamentous macroalgae and Saccharina latissima however after 2-4 years Laminaria hyperborea dominated the community.

Following disturbance or in areas were recurrent rapid disturbance occurs Laminaria hyperborea recruitment could also be affected by interspecific competitive interactions with Invasive Non-Indigenous Species or ephemeral algal species (Brodie et al., 2014; Smale et al., 2013), however, evidence for this is limited and thus not included within this assessment.

Sensitivity assessment. Resistance to the pressure is considered ‘Low’, and resilience ‘Medium’. The sensitivity of this biotope to damage to seabed surface features is assessed 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, 2024; OBIS, 2026).

Abundances at its northern and southern extremes may be low but densities in UK and France are often over 1000 /m² and it may carpet the seafloor in the Solent and Essex. In the UK, it was reported to reach abundances of >1000 /m² (max. 2,748 /m²) in the Milford Harbour Waterway (Bohn et al., 2012), 84 /m² in Portsmouth, 174 /m² in Langstone and 306 /m² in Chichester harbours in 2017 (Helmer et al., 2019). In France, it has been reported to reach >4,700 /m² in the Bay of Marennes-Oleron, 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 ~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 ~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 ~160 m in depth but is most common in the shallow subtidal and low water springs (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 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 ~0.7 m) and was absent from high tidal level (~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 / m²) than in southern waters. Thieltges et al. (2004) reported that the population of Crepidula was affected strongly by cold winters in the Wadden Sea. The winters of 2001 and 2003 resulted in ca 56-64% mortality of intertidal Crepidula and up to 97% on one mussel bed, compared to only 11 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 (Magellana gigas) and slipper limpets, which co-existed 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 suggest that mixed stands of Laminaria hyperborea and Laminaria ochroleuca are sensitive to colonization by Crepidula fornicata.

The carpet sea squirt, Didemnum vexillum

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions 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 have led to the widespread misidentification of Didemnum vexillum and it is often recorded as Didemnum spp. Hence, the native range of the species is not known conclusively (Lambert, 2009; Stefaniak et al., 2012; McKenzie et al., 2017; Holt, 2024). However, molecular data and limited historical evidence have suggested that the species may be native to Japan with its native range possibly extending into continental Asia and north-western Pacific (Stefaniak et al., 2012; Tillin et al., 2020; Holt, 2024). Previously unrecorded populations of a colonial ascidian have been recently identified as Didemnum vexillum (Tillin et al., 2020).

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

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

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

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

This species requires suitable hard substrata for successful settlement and the establishment of colonies. It can grow quickly and can establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems by its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock gravel, pebble, cobble, or boulders (Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent, and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020). It is therefore likely that the substrate in this biotope is suitable for Didemnum vexillum colonisation. In addition, the depth range at which these mixed kelp biotopes are found (0 to 20 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 ca 8 m/s or (more likely) human activity could cause fragmentation of tendrils.

Reinhart et al. (2012) noted that the tensile strength of Didemnum vexillum was an order of magnitude higher than Botrylloides sp. and was similar to that of Alyconium digitatum. Alyconium digitatum is reported from sheltered to very wave exposed conditions, but in the sublittoral. Reinhart et al. (2012) also suggested that seasonal changes in the condition of Didemnum vexillum reduced the tensile strength of colonies and 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 membranacea 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. Didemnum prefer sheltered conditions so the wave exposed conditions that characterise IR.HIR.KFaR.LhypR.Loch may mitigate its abundance. 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 ind./m², little of the underlying substratum remains visible (Herbert et al., 2016). These reefs can stabilize the sediment surface locally (Troost, 2010). When such reefs are formed or, particularly when the species colonizes soft sediments such as mud or sand, it can change and affect local communities, by creating hard substrata for mobile species, which might not otherwise be present before the invasion (Padilla, 2010). However, Hansen et al. (2023) suggested that no immediate ecosystem risk is observed where the Pacific oyster occurs sporadically.

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

Populations of Magallana gigas have been found from wave-exposed rocky shores to wave-sheltered soft sediment environments and it has been described as a habitat generalist (Troost, 2010; Kochmann, 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, 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 to 3 m below chart datum (Tillin et al., 2020).

Bergstrom et al. (2021) suggested the optimal depth in the Skagerrak is around 0.5 m, although presence is documented down to 5 m. In Lim Bay (Adriatic Sea), 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 to 50 m wide (Kochmann et al., 2013; Cook et al., 2014). It has been suggested that recruitment is enhanced, and abundances are higher in wave-sheltered conditions (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, 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, resistance is ‘High’, resilience is ‘High’ (by default), and sensitivity is assessed as ‘Not Sensitive’ to this pressure.

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 as long as temperatures are over 8°C for at least four months of the year (Engelen et al., 2015). However, its distribution is limited by the availability of hard substratum (e.g. stones >10 cm) and light (Staehr et al., 2000; Strong & Dring 2011; Engelen et al., 2015). It is most abundant between 1 and 3 m below mean water. But it has been recorded at 18 m or 30 m in the clear waters of California. However, it is a poor competitor under low light and only develops dense canopies in shallow areas (Engelen et al., 2015).

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

Experimental manipulation of subtidal algal canopies in San Juan Islands, Washington State, USA, showed that Sargassum muticum reduced the abundance of native macroalgae, including the kelp Laminaria bongardiana due to shading. However, experimental removal of Sargassum resulted in the recovery of native species within about one year (Britton-Simmons, 2004; Engelen et al., 2015). The negative effects of Sargassum muticum on native macroalgae are mainly due to completion to light, rather than changes in nutrient availability, sedimentation, or water flow (Britton-Simmons, 2004; Engelen et al., 2015). Cosson (1999) reported a significant decline in Laminaria digitata at two sites between 1983 and 1997 on the coast of Normandy, France, due to an increase in Sargassum muticum abundance in the same areas. For example, on the Grandcamp rocks, Laminaria digitata has almost disappeared while Sargassum muticum had covered 80% of the lower intertidal and subtidal zone in summer. However, no evidence was found on the effects of Sargassum muticum in native Laminaria hyperborea beds. 

Sensitivity Assessment

The evidence suggests that Sargassum muticum prefers wave sheltered, shallow sites. IR.HIR.KFaR.LhypR.Loch occurs on exposed infralittoral rock from depths of 5 to 20 m. There is no evidence of Sargassum muticum having negative effects on either of the characterising kelp species, and therefore it is unlikely to compete with them and negatively affect the biotope. However, IR.LIR.K.LhypLoch occurs on moderately exposed or sheltered infralittoral rock at depths from 0 to 20 m. While there is no evidence of Sargassum muticum displacing either characterizing species, the conditions of this biotope are suitable for invasion, especially if the Laminaria canopy is reduced by another pressure. IR.HIR.KFaR.LhypR.Loch resistance is assessed as ‘High’, resilience as ‘High’ by default (as there is nothing to recover from), and sensitivity as ‘Not Sensitive’. However, IR.LIR.K.LhypLoch is given a precautionary resistance of ‘Low’, a resilience of ‘Very Low’ (since recovery would require the full removal of Sargassum muticum), and a sensitivity of ‘High’.

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; Heiser 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 had become a major fouling plant on pontoons (Minchin & Nunn, 2014). Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound.

Undaria pinnatifida seems to settle better on artificial substrata (e.g. floats, marinas, or piers) than on natural rocky shores among local kelps (Vaz-Pinto et al., 2014). It is found predominantly in low intertidal to shallow subtidal habitats (Epstein et al., 2019b) and is significantly more abundant on artificial substrata compared to natural rocky substrata (Heiser et al., 2014; Epstein & Smale, 2018). James (2017) suggested that Undaria pinnatifida could out-compete native species on artificial substrata (such as marinas and wharf structures). De Leij et al. (2017) suggested that in natural substrata, Undaria pinnatifida can be inhibited by the presence of native competitors, such as large perennial species. The dense macroalgae canopies formed by native kelps result in limited space and light availability for Undaria pinnatifida recruits. However, it will not always completely prevent the assimilation of Undaria pinnatifida (De Leij et al., 2017; Epstein & Smale, 2018).

Undaria pinnatifida species behaves as a winter annual and recruitment occur 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 Plymouth Sound (UK), Epstein et al. (2019b) found that within its depth range (+1 to –4 m), Undaria pinnatifida co-existed with seven species of canopy-forming brown macroalgae, including Laminaria ochroleuca and Laminaria hyperborea. This may be due to the higher tidal water flow in the subtidal fringe than in the lower subtidal zone, as Undaria pinnatifida may have lower fitness than Laminaria ochroleuca in lower velocity subtidal waters (Epstein et al., 2019b). De Leij et al. (2017) found that in natural habitats where kelps, such as Laminaria hyperborea and Laminaria ochroleuca, formed dense native macroalgal canopies there was more resistance to Undaria pinnatifida invasion resulting in low abundance and cover than found disturbed or sparse canopies. This is due to limited space and light availability for Undaria pinnatifida recruits. However, the dense canopies will not prevent invasion of Undaria pinnatifida as sporophytes were still recorded within dense Laminaria canopies, suggesting that canopy disturbance is not always required.

In St Malo, France, there was evidence that Undaria pinnatifida co-existed 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 (40-79%) or superabundant (>80%), which suggested that Undaria pinnatifida could co-exist within refugia amongst areas with dense Laminaria spp.

In Plymouth Sound, UK, Heiser et al. (2014) observed that both Laminaria ochroleuca and Laminaria hyperborea were significantly less abundant at sites with the presence Undaria pinnatifida. Only ca 2.5 Laminaria ochroleuca individuals per m² were present compared to ca 5.5 individuals per m² at sites without Undaria pinnatifida and ca 0.5 Laminaria hyperborea individuals per m² were present compared to ca 8 individuals per m² at sites without Undaria pinnatifida. However, the results from their correlation study only showed that the species were not found together (pers. comm., Epstein 2021). Exclusion and succession experiments on reefs tell us that Laminaria spp. exclude Undaria pinnatifida, not the other way round. 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 restricted 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 implied 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).

Epstein et al. (2019b) found that Laminaria ochroleuca had a strong negative correlation with Undaria pinnatifida. 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 and resultant lack of data.

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

Sensitivity Assessment

Undaria pinnatifida has the potential to colonize and co-exist 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. However, 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. The degree of wave exposure and the depth range (5 to 20 m) of IR.LIR.K.LhypR.Loch makes Undaria pinnatifida invasion unlikely. Therefore, resistance is ‘High’, resilience is ‘High’ by default (nothing to recover from), and sensitivity is assessed as ‘Not Sensitive’. In contrast, IR.LIR.K.LhypLoch occurs in low to moderate energy environments with at depths ranging from 0 to 20 m, which overlaps with conditions that are suitable for Undaria pinnatifida. Therefore, as a precaution, resistance is assessed as ‘Low’, resilience as ‘Very Low’ as recovery would require the removal of Undaria pinnatifida, and sensitivity 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

Falkenbergia rufolanosa, the fast-growing invasive tetrasporophyte stage of Asparagopsis armata, can proliferate in degraded kelp habitats. However, healthy stands of Laminaria ochroleuca appear to resist its establishment or expansion (Alvite et al., 2026). Surveys in Ría de Vigo (northwestern Spain) from 2020 to 2024 showed that Falkenbergia rufolanosa only occurred on degraded reefs where the mean number of kelp adults per 0.25 m² was less than 1. This remained true despite the overall decrease in kelp abundance across the study period.

Sensitivity Assessment

While there is evidence of Laminaria ochroleuca resistance to Falkenbergia rufolanosa, there is Insufficient evidence to assess the sensitivity of these biotopes to this pressure due to the lack of evidence regarding Laminaria hyperborea.