The Marine Life Information Network

Information on the biology of species and the ecology of habitats found around the coasts and seas of the British Isles

Navigation

Mixed kelp with foliose red seaweeds, sponges and ascidians on sheltered tide-swept infralittoral rock

Distribution MapBIO Map Legend

Summary

UK and Ireland classification

Description

Bedrock and boulders, often in tide-swept areas, that are subject to scouring or periodic burial by sand, characterized by a canopy of mixed kelps such as Saccharina latissima, Laminaria hyperborea and Saccorhiza polyschides and the brown seaweed Desmarestia aculeata; there may also be an understorey of foliose seaweeds that can withstand scour such as Plocamium cartilagineum, Chondrus crispus, Dilsea carnosa, Metacallophyllis laciniata as well as the filamentous Heterosiphonia plumosa and the foliose brown seaweed Dictyota dichotoma. The perennial red seaweed Brongniartella byssoides re-grows in the summer months. The L. hyperborea stipes often support a growth of epiphytes, such as Delesseria sanguinea, Phycodrys rubens and Cryptopleura ramosa. The scour can reduce the rock surface to bare coralline crusts at times; sponge crusts and the colonial ascidian Botryllus schlosseri can also grow on the stipes and holdfasts. The faunal diversity on the rock is usually low and restricted to robust, low-profile animals such as the tube-building polychaete Spirobranchus triqueter, the barnacle Balanus crenatus, encrusting bryozoans such as Membranipora membranacea, the anthozoan Urticina felina, the starfish Asterias rubens and the urchin Echinus esculentus. Deeper sites support more hydroids and bryozoans, particularly Bugula spp. Where this biotope occurs in very shallow water Laminaria digitata may be found in combination with the other kelp species. Other species present only in shallow water include the red algae Corallina officinalis and the sand-binding alga Rhodothamniella floridula.

Depth range

0-5 m, 5-10 m

Additional information

-

Sensitivity reviewHow is sensitivity assessed?

Sensitivity characteristics of the habitat and relevant characteristic species

IR.MIR.KT.XKT & IR.MIR.KT.XKTX are defined by bedrock reefs and mixed substrata of boulders, cobbles, pebbles and gravel, typically found in strong tidal streams. The community is characterized by mixed kelp canopies of Laminaria hyperborea and Saccharina latissima (syn. Laminaria saccharina). Dense stands of the brown seaweed Halidrys siliquosa can occur within the kelp along with Dictyota dichotoma. Kelp stipes may also support prolific growths of foliose red seaweeds such as Phycodrys rubens, Membranoptera alata, Delesseria sanguinea and Plocamium cartilagineum. The dominance of kelp species can vary between sites however as substrata stability decreases, as in IR.MIR.KT.XKTX, Saccharina latissima becomes the more dominant canopy-forming species (Connor et al., 2004).

In undertaking this assessment of sensitivity, an account is taken of knowledge of the biology of all characterizing species in the biotope. There is an abundance of literature for regeneration of mono-specific Laminaria hyperborea beds, however at the time of writing there is limited research for the recovery of mixed kelp canopies. For this sensitivity assessment Laminaria hyperborea and Saccharina latissima are the primary foci of research, however, have been researched independently and inter-specific competition may influence recovery times. It is also recognized that the understorey red seaweed communities also define the biotope. Examples of important species groups are mentioned where appropriate.

Resilience and recovery rates of habitat

In favourable conditions, Laminaria hyperborea can recover following disturbance events reaching comparable plant densities and size to pristine Laminaria hyperborea beds within 2-6 years (Kain, 1979; Birkett et al., 1998b; Christie et al., 1998).  Holdfast communities may recover in 6 years (Birkett et al., 1998b). Full epiphytic community and stipe habitat complexity regeneration require over six years to recover (possibly 10 years).  These recovery rates were based on discrete kelp harvesting events and recurrent disturbance occurring frequently within 2-6 years of the initial disturbance is likely to lengthen recovery time (Birkett et al., 1998b, Burrows et al., 2014). 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 two years of clearance the blocks were dominated by Laminaria hyperborea.

Laminaria hyperborea has a heteromorphic life strategy, A vast number of zoospores (mobile asexual spores) are released into the water column between October-April (Kain & Jones, 1964). Zoospores settle onto rock substrata and develop into dioecious gametophytes (Kain, 1979) which, following fertilization, develop into sporophytes and mature within 1-6 years (Kain, 1979; Fredriksen et al., 1995; Christie et al., 1998). Laminaria hyperborea zoospores have a recorded dispersal range of approximately .200m (Fredriksen et al., 1995). However, zoospore dispersal is greatly influenced by water movements, and zoospore density and the rate of successful fertilization decreases exponentially with distance from the parental source (Fredriksen et al., 1995). Hence, recruitment following disturbance can be influenced by the proximity of mature kelp beds producing viable zoospores to the disturbed area (Kain, 1979, Fredriksen et al., 1995).

Other factors that are likely to influence the recovery of kelp biotopes is competitive interactions with the Invasive Non-Indigenous Species (INIS) Undaria pinnatifida (Smale et al., 2013; Brodie et al., 2014; Heiser et al., 2014). Undaria pinnatifida has received a large amount of research attention as an INIS which could out-compete UK kelp habitats (see Farrell & Fletcher, 2006; Thompson & Schiel, 2012, Brodie et al., 2014; Hieser et al., 2014). Undaria pinnatifida was first recorded in Plymouth Sound, UK in 2003 (NBN, 2015) subsequent surveys in 2011 have reported that Undaria pinnatifida is widespread throughout Plymouth Sound, colonizing rocky reef habitats. Where Undaria pinnatifida is present there was a significant decrease in the abundance of other Laminaria species, including Laminaria hyperborea (Heiser et al., 2014). In New Zealand, Thompson & Schiel (2012) observed that native fucoids could out-compete Undaria pinnatifida and re-dominate the substratum. However, Thompson & Schiel (2012) suggested the fucoid recovery of the substratum was partially due to an annual Undaria pinnatifida die back, which as noted by Heiser et al. (2014) did not occur in Plymouth sound, UK. Whether Undaria pinnatifida will out-compete native macroalgae in the UK is unknown. However, from 2003-2011 Undaria pinnatifida had increased throughout Plymouth sound, UK, becoming a visually dominant species at some locations within summer months (Hieser et al., 2014). At the time of writing there is limited evidence available to assess the ecological impacts of Undaria pinnatifida on Laminaria hyperborea associated communities.  Kelp biotopes are unlikely to fully recover until Undaria pinnatifida is fully removed from the habitat, which as stated above is unlikely to occur.

Saccharina latissima is a perennial kelp characteristic of wave sheltered sites of the North East Atlantic, distributed from northern Portugal to Spitzbergen, Svalbard (Birkett et al., 1998b; Conor et al., 2004; Bekby & Moy, 2011; Moy & Christie, 2012). Saccharina latissima is capable of reaching maturity within 15-20 months (Sjøtun, 1993) and has a life expectancy of 2-4 years (Parke, 1948). Maximum growth has been recorded in late winter early spring, in late summer and autumn growth rates slow (Parke, 1948; Lüning, 1979; Birkett et al., 1998b). The overall length of the sporophyte may not change during the growth season due to marginal (distal) erosion of the blade, but extension growth of the blade has been measured at 1.1 cm/day, with total length addition of over 2.25 m of tissue per year (Birkett et al., 1998b). Saccharina latissima has a heteromorphic life strategy.  Vast numbers of zoospores are released from sori located centrally on the blade between autumn and winter. Zoospores settle onto rock substrata and develop into dioecious gametophytes (Kain, 1979) which, following fertilization, germinate into juvenile sporophytes from winter-spring.  Kelp zoospores are expected to have a large dispersal range; however, zoospore density and the rate of successful fertilization decreases exponentially with distance from the parental source (Fredriksen et al., 1995). Hence, recruitment following disturbance can be influenced by the proximity of mature kelp beds producing viable zoospores to the disturbed area (Kain, 1979; Fredriksen et al., 1995).

The temperature isotherm of 19-20°C has been reported as limiting Saccharina latissima growth (Müller et al., 2009). Gametophytes can develop in ≤23°C (Lüning, 1990). However, Bolton & Lüning (1982) reported an experimental optimal temperature of 10-15°C for the growth of the Saccharina latissima sporophyte. Growth was inhibited by 50-70% at 20°C and, all experimental specimens completely disintegrated after seven days at 23°C. In the field, Saccharina latissima has however shown significant regional variation in its acclimation response to changing environmental conditions.  For example, Gerard & Dubois (1988) observed sporophytes of Saccharina latissima which were regularly exposed to ≥20°C could tolerate these high temperatures, whereas sporophytes from other populations which rarely experience ≥17°C showed 100% mortality after three weeks of exposure to 20°C. Therefore, the response of Saccharina latissima to a change in temperatures is likely to be locally variable.

In 2002, a large scale decline of Saccharina latissima was discovered on the Norwegian coast (Moy & Christie, 2012). A subsequent large survey was undertaken between 2004-2009 of 660 sites covering 34,000 km of south and west Norway to assess the decline of Saccharina latissima abundance and distribution (Moy & Christie, 2012). The survey indicated an 83% reduction of Saccharina latissima forests across the south Norwegian region of Skagerrak.  The west Norwegian coast was less affected but Saccharina latissima was either absent or very sparse at 38% of sites where it was expected to be abundant.  At all sites where Saccharina latissima was sparse a community of ephemeral macro-algae species was dominant and persisted throughout the study period (2004-2009).  Bekby & Moy (2011) modelled the regional decline which indicated a decline of 50.7% of Saccharina latissima from Skagerrak, Norway. Approximately 50% of Europe’s Saccharina latissima is found in Norway (Moy et al., 2006), therefore, despite large discrepancies between the two estimates of Saccharina latissima decline (50.7-83%) the results indicated a significant decline in Saccharina latissima across the region. Moy & Christie (2012) suggested the ephemeral filamentous macroalgae communities represented a stable state shift that had persisted throughout the study period (2004-2009).  Although no measurements were made, they suggested that the decline was due to low tidal movement and wave action in the worst affected areas combined with the impacts of dense human populations and increased land run-off multiple stressors such as eutrophication, increasing regional temperature, increased siltation and overfishing may also be acting synergistically to cause the observed habitat shift.

Resilience assessment. Of the two kelp species (Laminaria hyperborea and Saccharina latissima) that characterize IR.MIR.KT.XKT & IR.MIR.KT.XKTX, Laminaria hyperborea is the slowest to recover following disturbance. Laminaria hyperborea can regenerate from disturbance within a period of 1-6 years, and the associated community within 7-10 years. Saccharina latissima has reportedly a rapid recovery rate or re-generation time, following clearance of Strongylocentrotus droebachiensis from ‘urchin Barrens’ Saccharina latissima was a rapid colonizer appearing after a few weeks, and can reach maturity within 15-20 months (Birkett et al., 1998b). Due to comparatively slow growth rates, resilience estimates are based on Laminaria hyperborea, however, the recovery of Saccharina latissima and the understorey red seaweed is accounted for where relevant.  Resilience has therefore been assessed as ‘Medium’.

Climate Change Pressures

Use / to open/close text displayedResistanceResilienceSensitivity
Low Very Low High
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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

Laminaria hyperborea has an optimum temperature for growth of 15°C, and an upper temperature limit of 21°C (Bolton & Lüning, 1982). At 17°C gamete survival is reduced (Steinhoff et al., 2008) and gametogenesis is inhibited at 21°C (Dieck, 1992). Therefore, Laminaria hyperborea recruitment could be impaired at a sustained temperature increase above 17°C. However, sporophytes can tolerate slightly higher temperatures of 20°C. Temperature tolerances for Laminaria hyperborea are also seasonally variable and temperature changes are less tolerated in winter months than summer months (Birkett et al., 1998b).

There is evidence that climate change is already having an impact on Laminaria hyperborea populations in the English Channel. Poleward range expansion of the warm temperate Laminaria ochroleuca as a result of ocean warming has led to competition with Laminaria hyperborea in UK waters (Smale et al., 2015). Laminaria ochroleuca was not found in the UK last century.  But Laminaria ochroleuca has now increased its range to include the southwest of England (Smale et al., 2015) and the west coast of Ireland (Schoenrock et al., 2019).

During the 2013-2014 Northeast Atlantic storm season, the UK was subjected to some of the most intense storms recorded within the past five years. A study by Smale & Vance. (2015) investigated the impacts of the storms on kelp canopies along the south coast of the UK, findings indicated monospecific canopies of Laminaria hyperborea were unaffected by the storms. However, the storms significantly altered a mixed canopy study site, composed of Laminaria ochroleuca, Saccharina latissima and Laminaria hyperborea. Therefore, if climate change continues to change species composition within kelp forests resistance to storm disturbance could be altered.

Smale et al. (2015) found that Laminaria hyperborea suffered from much higher epiphytic loadings and lower productivity than its competitor Laminaria ochroleuca during the summer months, which reduced its competitive ability. The decreased competitive ability because of ocean warming corresponds to findings by Pessarrodona et al. (2018), who found a decrease in the size of Laminaria hyperborea plants along a north-south gradient in Scotland, with average maximum stipe lengths of over 150 cm, whereas in southern England they were less than 100 cm. Similarly, Smale et al. (2020b) observed clear differences between net primary productivity (NNP) and carbon standing stock of Laminaria hyperborea between the colder northern and warmer southern test sites in the UK, with NNP and standing stock being 1.5 and 2.5 times greater in the northern sites. Identifying ocean temperatures as a lively driver of productivity, with reduced NNP and standing stock observed in warmer waters (Smale et al., 2020b). 

The decrease in productivity in southern England suggests that Laminaria hyperborea is already growing at suboptimal temperatures. Assis et al. (2018) predicted that under the highest emission scenario (RCP 8.5) the biogeographic range of Laminaria hyperborea will move northwards, and this retreat would lead to the species being lost from approximately 30% of the coastline of the UK.

Saccharina latissima is a polar to temperate macroalgae distributed from Greenland to the coast of Portugal, and is the NW Atlantic, is found as far south as New York State, USA. At its southern distribution in New York, temperatures can regularly reach ≥20°C for six weeks or more during summer months (Gerard & Du Bois, 1988).

Saccharina latissima has an optimal growth temperature between 10- 15°C, with growth reducing by 50-70% at 20°C, and all experimental specimens disintegrating after seven days at 23°C (Bolton & Lüning, 1982). The temperature isotherm of 19-20°C has been reported as limiting Saccharina latissima growth (Müller et al., 2009). Temperature is an environmental factor controlling the development of the microscopic stages of Saccharina latissima, with crucial changes in survival, growth, and gametogenesis occurring within a few degrees of its upper thermal limits (Redmond, 2013). The optimal germination temperature for Saccharina latissima is between 2°C and 12°C, with gametophyte survival between 23-25°C (Müller et al., 2009). Germination rates drop at 22°C, with surviving gametophytes smaller than those grown at lower temperatures (Redmond, 2013). Park et al. (2017) observed reductions in the percentage of sporophytes produced at 15°C when compared to values produced at 5°C and 10°C. 

In the field, Saccharina latissima has shown significant regional variation in its acclimation response to changing environmental conditions.  For example, Gerard & Dubois (1988) observed that sporophytes of Saccharina latissima that were regularly exposed to ≥20°C tolerated these high temperatures, whereas sporophytes from other populations that rarely experience ≥17°C showed 100% mortality after 3 weeks of exposure to 20°C.

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

Assis et al. (2018) predicted that, under the highest emission scenario (RCP 8.5), the range of Saccharina latissima would move northwards, retreating from their southern-most locations, with a predicted loss of Saccharina latissima from the southwest coast of the UK. 

Many of the red algae species associated with the understorey turf can tolerate warm water temperatures. Corallina officinalis may tolerate between -4 to 28°C (Lüning, 1990), although when Colthart & Johansen (1973) exposed this species to a number of different temperatures, they found that growth was maintained at 18°C and ceased at 25°C. Abrupt temperature changes (10°C in California, Seapy & Littler 1984; 4.8 to 8.5°C, Hawkins & Hartnoll, 1985) resulted in dramatic declines. However, in both cases recovery was rapid, suggesting that the crustose bases survived. 

Sensitivity assessment. UK populations of Saccharina latissima are found in the middle of the species distribution and are known to be able to survive at higher temperatures than currently experienced around the UK. The ability to tolerate summer seawater temperatures of >20°C in populations at their southern geographic limit is thought to be a genetic adaptation (Gerard & Du Bois, 1988), and maybe crucial in the persistence of this species around the UK, as seawater temperatures rise. However, Laminaria hyperborea is already growing at suboptimal temperatures in the southern UK, based on evidence of decreased productivity comparative to Scotland (Pessarrodona et al., 2018; Smale et al., 2020b), and predictions have estimated Laminaria hyperborean to be lost from the UK by 2100 as a result of warming (Brodie et al., 2014). 

Under the middle emission scenario, a rise of 3°C could lead to maximum summer high temperatures of 22°C in the south of the UK. Populations of Saccharina latissima and the understorey community of mixed red seaweeds may be able to adapt to cope with a gradual rise in ocean temperatures of 3°C. However, this is above the upper thermal limit of 21°C for Laminaria hyperborea (Bolton & Lüning, 1982), and is likely to lead to loss of this species from the south of England. Furthermore, biomass and plant sizes are expected to decrease as waters warm, with Scottish Laminaria hyperborea stipe lengths decreasing to lengths observed in southern England, leading to a decline in carbon assimilation, productivity and habitat quality. Therefore, resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very Low’, as the loss is likely to be a long-term decline, due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming under this scenario.

For the high and extreme emission scenario where sea temperatures rise by 4-5°C to potential southern summer temperatures of 23-24°C by the end of this century Saccharina latissima and Laminaria hyperborea is likely to be lost from southern England. The northward retreat of the distribution of Laminaria hyperborea is expected to increase. Under the high emission scenario it is expected to be lost from 30% of the coastline around the UK (Assis et al., 2018), and under the extreme scenario even more is projected to be lost. Populations of Laminaria hyperborea that remain around the UK are predicted to become less productive. Therefore, under these scenarios, resistance is assessed as ‘Low’, and resilience is assessed as ‘Very Low’. Therefore, this biotope is assessed as ‘High’ sensitivity to ocean warming under this scenario.

Low Very Low High
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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

Laminaria hyperborea has an optimum temperature for growth of 15°C, and an upper temperature limit of 21°C (Bolton & Lüning, 1982). At 17°C gamete survival is reduced (Steinhoff et al., 2008) and gametogenesis is inhibited at 21°C (Dieck, 1992). Therefore, Laminaria hyperborea recruitment could be impaired at a sustained temperature increase above 17°C. However, sporophytes can tolerate slightly higher temperatures of 20°C. Temperature tolerances for Laminaria hyperborea are also seasonally variable and temperature changes are less tolerated in winter months than summer months (Birkett et al., 1998b).

There is evidence that climate change is already having an impact on Laminaria hyperborea populations in the English Channel. Poleward range expansion of the warm temperate Laminaria ochroleuca as a result of ocean warming has led to competition with Laminaria hyperborea in UK waters (Smale et al., 2015). Laminaria ochroleuca was not found in the UK last century.  But Laminaria ochroleuca has now increased its range to include the southwest of England (Smale et al., 2015) and the west coast of Ireland (Schoenrock et al., 2019).

During the 2013-2014 Northeast Atlantic storm season, the UK was subjected to some of the most intense storms recorded within the past five years. A study by Smale & Vance. (2015) investigated the impacts of the storms on kelp canopies along the south coast of the UK, findings indicated monospecific canopies of Laminaria hyperborea were unaffected by the storms. However, the storms significantly altered a mixed canopy study site, composed of Laminaria ochroleuca, Saccharina latissima and Laminaria hyperborea. Therefore, if climate change continues to change species composition within kelp forests resistance to storm disturbance could be altered.

Smale et al. (2015) found that Laminaria hyperborea suffered from much higher epiphytic loadings and lower productivity than its competitor Laminaria ochroleuca during the summer months, which reduced its competitive ability. The decreased competitive ability because of ocean warming corresponds to findings by Pessarrodona et al. (2018), who found a decrease in the size of Laminaria hyperborea plants along a north-south gradient in Scotland, with average maximum stipe lengths of over 150 cm, whereas in southern England they were less than 100 cm. Similarly, Smale et al. (2020b) observed clear differences between net primary productivity (NNP) and carbon standing stock of Laminaria hyperborea between the colder northern and warmer southern test sites in the UK, with NNP and standing stock being 1.5 and 2.5 times greater in the northern sites. Identifying ocean temperatures as a lively driver of productivity, with reduced NNP and standing stock observed in warmer waters (Smale et al., 2020b). 

The decrease in productivity in southern England suggests that Laminaria hyperborea is already growing at suboptimal temperatures. Assis et al. (2018) predicted that under the highest emission scenario (RCP 8.5) the biogeographic range of Laminaria hyperborea will move northwards, and this retreat would lead to the species being lost from approximately 30% of the coastline of the UK.

Saccharina latissima is a polar to temperate macroalgae distributed from Greenland to the coast of Portugal, and is the NW Atlantic, is found as far south as New York State, USA. At its southern distribution in New York, temperatures can regularly reach ≥20°C for six weeks or more during summer months (Gerard & Du Bois, 1988).

Saccharina latissima has an optimal growth temperature between 10- 15°C, with growth reducing by 50-70% at 20°C, and all experimental specimens disintegrating after seven days at 23°C (Bolton & Lüning, 1982). The temperature isotherm of 19-20°C has been reported as limiting Saccharina latissima growth (Müller et al., 2009). Temperature is an environmental factor controlling the development of the microscopic stages of Saccharina latissima, with crucial changes in survival, growth, and gametogenesis occurring within a few degrees of its upper thermal limits (Redmond, 2013). The optimal germination temperature for Saccharina latissima is between 2°C and 12°C, with gametophyte survival between 23-25°C (Müller et al., 2009). Germination rates drop at 22°C, with surviving gametophytes smaller than those grown at lower temperatures (Redmond, 2013). Park et al. (2017) observed reductions in the percentage of sporophytes produced at 15°C when compared to values produced at 5°C and 10°C. 

In the field, Saccharina latissima has shown significant regional variation in its acclimation response to changing environmental conditions.  For example, Gerard & Dubois (1988) observed that sporophytes of Saccharina latissima that were regularly exposed to ≥20°C tolerated these high temperatures, whereas sporophytes from other populations that rarely experience ≥17°C showed 100% mortality after 3 weeks of exposure to 20°C.

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

Assis et al. (2018) predicted that, under the highest emission scenario (RCP 8.5), the range of Saccharina latissima would move northwards, retreating from their southern-most locations, with a predicted loss of Saccharina latissima from the southwest coast of the UK. 

Many of the red algae species associated with the understorey turf can tolerate warm water temperatures. Corallina officinalis may tolerate between -4 to 28°C (Lüning, 1990), although when Colthart & Johansen (1973) exposed this species to a number of different temperatures, they found that growth was maintained at 18°C and ceased at 25°C. Abrupt temperature changes (10°C in California, Seapy & Littler 1984; 4.8 to 8.5°C, Hawkins & Hartnoll, 1985) resulted in dramatic declines. However, in both cases recovery was rapid, suggesting that the crustose bases survived. 

Sensitivity assessment. UK populations of Saccharina latissima are found in the middle of the species distribution and are known to be able to survive at higher temperatures than currently experienced around the UK. The ability to tolerate summer seawater temperatures of >20°C in populations at their southern geographic limit is thought to be a genetic adaptation (Gerard & Du Bois, 1988), and maybe crucial in the persistence of this species around the UK, as seawater temperatures rise. However, Laminaria hyperborea is already growing at suboptimal temperatures in the southern UK, based on evidence of decreased productivity comparative to Scotland (Pessarrodona et al., 2018; Smale et al., 2020b), and predictions have estimated Laminaria hyperborean to be lost from the UK by 2100 as a result of warming (Brodie et al., 2014). 

Under the middle emission scenario, a rise of 3°C could lead to maximum summer high temperatures of 22°C in the south of the UK. Populations of Saccharina latissima and the understorey community of mixed red seaweeds may be able to adapt to cope with a gradual rise in ocean temperatures of 3°C. However, this is above the upper thermal limit of 21°C for Laminaria hyperborea (Bolton & Lüning, 1982), and is likely to lead to loss of this species from the south of England. Furthermore, biomass and plant sizes are expected to decrease as waters warm, with Scottish Laminaria hyperborea stipe lengths decreasing to lengths observed in southern England, leading to a decline in carbon assimilation, productivity and habitat quality. Therefore, resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very Low’, as the loss is likely to be a long-term decline, due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming under this scenario.

For the high and extreme emission scenario where sea temperatures rise by 4-5°C to potential southern summer temperatures of 23-24°C by the end of this century Saccharina latissima and Laminaria hyperborea is likely to be lost from southern England. The northward retreat of the distribution of Laminaria hyperborea is expected to increase. Under the high emission scenario it is expected to be lost from 30% of the coastline around the UK (Assis et al., 2018), and under the extreme scenario even more is projected to be lost. Populations of Laminaria hyperborea that remain around the UK are predicted to become less productive. Therefore, under these scenarios, resistance is assessed as ‘Low’, and resilience is assessed as ‘Very Low’. Therefore, this biotope is assessed as ‘High’ sensitivity to ocean warming under this scenario.

Medium Very Low Medium
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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

Laminaria hyperborea has an optimum temperature for growth of 15°C, and an upper temperature limit of 21°C (Bolton & Lüning, 1982). At 17°C gamete survival is reduced (Steinhoff et al., 2008) and gametogenesis is inhibited at 21°C (Dieck, 1992). Therefore, Laminaria hyperborea recruitment could be impaired at a sustained temperature increase above 17°C. However, sporophytes can tolerate slightly higher temperatures of 20°C. Temperature tolerances for Laminaria hyperborea are also seasonally variable and temperature changes are less tolerated in winter months than summer months (Birkett et al., 1998b).

There is evidence that climate change is already having an impact on Laminaria hyperborea populations in the English Channel. Poleward range expansion of the warm temperate Laminaria ochroleuca as a result of ocean warming has led to competition with Laminaria hyperborea in UK waters (Smale et al., 2015). Laminaria ochroleuca was not found in the UK last century.  But Laminaria ochroleuca has now increased its range to include the southwest of England (Smale et al., 2015) and the west coast of Ireland (Schoenrock et al., 2019).

During the 2013-2014 Northeast Atlantic storm season, the UK was subjected to some of the most intense storms recorded within the past five years. A study by Smale & Vance. (2015) investigated the impacts of the storms on kelp canopies along the south coast of the UK, findings indicated monospecific canopies of Laminaria hyperborea were unaffected by the storms. However, the storms significantly altered a mixed canopy study site, composed of Laminaria ochroleuca, Saccharina latissima and Laminaria hyperborea. Therefore, if climate change continues to change species composition within kelp forests resistance to storm disturbance could be altered.

Smale et al. (2015) found that Laminaria hyperborea suffered from much higher epiphytic loadings and lower productivity than its competitor Laminaria ochroleuca during the summer months, which reduced its competitive ability. The decreased competitive ability because of ocean warming corresponds to findings by Pessarrodona et al. (2018), who found a decrease in the size of Laminaria hyperborea plants along a north-south gradient in Scotland, with average maximum stipe lengths of over 150 cm, whereas in southern England they were less than 100 cm. Similarly, Smale et al. (2020b) observed clear differences between net primary productivity (NNP) and carbon standing stock of Laminaria hyperborea between the colder northern and warmer southern test sites in the UK, with NNP and standing stock being 1.5 and 2.5 times greater in the northern sites. Identifying ocean temperatures as a lively driver of productivity, with reduced NNP and standing stock observed in warmer waters (Smale et al., 2020b). 

The decrease in productivity in southern England suggests that Laminaria hyperborea is already growing at suboptimal temperatures. Assis et al. (2018) predicted that under the highest emission scenario (RCP 8.5) the biogeographic range of Laminaria hyperborea will move northwards, and this retreat would lead to the species being lost from approximately 30% of the coastline of the UK.

Saccharina latissima is a polar to temperate macroalgae distributed from Greenland to the coast of Portugal, and is the NW Atlantic, is found as far south as New York State, USA. At its southern distribution in New York, temperatures can regularly reach ≥20°C for six weeks or more during summer months (Gerard & Du Bois, 1988).

Saccharina latissima has an optimal growth temperature between 10- 15°C, with growth reducing by 50-70% at 20°C, and all experimental specimens disintegrating after seven days at 23°C (Bolton & Lüning, 1982). The temperature isotherm of 19-20°C has been reported as limiting Saccharina latissima growth (Müller et al., 2009). Temperature is an environmental factor controlling the development of the microscopic stages of Saccharina latissima, with crucial changes in survival, growth, and gametogenesis occurring within a few degrees of its upper thermal limits (Redmond, 2013). The optimal germination temperature for Saccharina latissima is between 2°C and 12°C, with gametophyte survival between 23-25°C (Müller et al., 2009). Germination rates drop at 22°C, with surviving gametophytes smaller than those grown at lower temperatures (Redmond, 2013). Park et al. (2017) observed reductions in the percentage of sporophytes produced at 15°C when compared to values produced at 5°C and 10°C. 

In the field, Saccharina latissima has shown significant regional variation in its acclimation response to changing environmental conditions.  For example, Gerard & Dubois (1988) observed that sporophytes of Saccharina latissima that were regularly exposed to ≥20°C tolerated these high temperatures, whereas sporophytes from other populations that rarely experience ≥17°C showed 100% mortality after 3 weeks of exposure to 20°C.

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

Assis et al. (2018) predicted that, under the highest emission scenario (RCP 8.5), the range of Saccharina latissima would move northwards, retreating from their southern-most locations, with a predicted loss of Saccharina latissima from the southwest coast of the UK. 

Many of the red algae species associated with the understorey turf can tolerate warm water temperatures. Corallina officinalis may tolerate between -4 to 28°C (Lüning, 1990), although when Colthart & Johansen (1973) exposed this species to a number of different temperatures, they found that growth was maintained at 18°C and ceased at 25°C. Abrupt temperature changes (10°C in California, Seapy & Littler 1984; 4.8 to 8.5°C, Hawkins & Hartnoll, 1985) resulted in dramatic declines. However, in both cases recovery was rapid, suggesting that the crustose bases survived. 

Sensitivity assessment. UK populations of Saccharina latissima are found in the middle of the species distribution and are known to be able to survive at higher temperatures than currently experienced around the UK. The ability to tolerate summer seawater temperatures of >20°C in populations at their southern geographic limit is thought to be a genetic adaptation (Gerard & Du Bois, 1988), and maybe crucial in the persistence of this species around the UK, as seawater temperatures rise. However, Laminaria hyperborea is already growing at suboptimal temperatures in the southern UK, based on evidence of decreased productivity comparative to Scotland (Pessarrodona et al., 2018; Smale et al., 2020b), and predictions have estimated Laminaria hyperborean to be lost from the UK by 2100 as a result of warming (Brodie et al., 2014). 

Under the middle emission scenario, a rise of 3°C could lead to maximum summer high temperatures of 22°C in the south of the UK. Populations of Saccharina latissima and the understorey community of mixed red seaweeds may be able to adapt to cope with a gradual rise in ocean temperatures of 3°C. However, this is above the upper thermal limit of 21°C for Laminaria hyperborea (Bolton & Lüning, 1982), and is likely to lead to loss of this species from the south of England. Furthermore, biomass and plant sizes are expected to decrease as waters warm, with Scottish Laminaria hyperborea stipe lengths decreasing to lengths observed in southern England, leading to a decline in carbon assimilation, productivity and habitat quality. Therefore, resistance is assessed as ‘Medium’, and resilience is assessed as ‘Very Low’, as the loss is likely to be a long-term decline, due to the long-term nature of ocean warming. Therefore, this biotope is assessed as ‘Medium’ sensitivity to ocean warming under this scenario.

For the high and extreme emission scenario where sea temperatures rise by 4-5°C to potential southern summer temperatures of 23-24°C by the end of this century Saccharina latissima and Laminaria hyperborea is likely to be lost from southern England. The northward retreat of the distribution of Laminaria hyperborea is expected to increase. Under the high emission scenario it is expected to be lost from 30% of the coastline around the UK (Assis et al., 2018), and under the extreme scenario even more is projected to be lost. Populations of Laminaria hyperborea that remain around the UK are predicted to become less productive. Therefore, under these scenarios, resistance is assessed as ‘Low’, and resilience is assessed as ‘Very Low’. Therefore, this biotope is assessed as ‘High’ sensitivity to ocean warming under this scenario.

None Very Low High
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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

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

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

Under experimental conditions, Nepper-Davidson et al. (2019) exposed a northern (Denmark) population of Saccharina latissima to a simulated three-week heatwave of three different intensities; 18, 21 and 24°C. When exposed to heatwaves of 18 and 21°C there was a decrease in photosynthesis and growth. When a 24°C was simulated, 91% of sporophytes were dead within a week, and the fronds of the few survivors were disintegrating, so the experiment was terminated (Nepper-Davidsen et al., 2019). 

Simonson et al. (2015) investigated the impacts of four temperature treatments (11°C, 14°C, 18°C & 21°C) on Saccharina latissima tissue over three weeksHistological analysis showed temperature mediated tissue damage, including holes, splitting of the medulla, damage to the meristoderm and loss of differentiation between tissue layers at temperatures between 14-21°C. 

Laminaria hyperborea is a cold-temperate species of kelp with an optimum temperature for growth of 15°C, and an upper temperature limit of 21°C (Bolton & Lüning, 1982). Germination success can decrease by almost two thirds at temperatures as low as 17°C. Therefore, it is expected that similar to other kelp species, Laminaria hyperborea will be highly sensitive to marine heatwaves.

Sensitivity assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, this could lead to summer sea temperatures reaching up to 24°C in southern England. Laminaria hyperborea and Saccharina latissima are unlikely to survive a heatwave of this magnitude and likely to suffer severe mortality in the south. Although, in Scotland, where a significant portion of these biotopes occur, temperatures are not predicted to rise above 20°C, and therefore, Laminaria hyperborea and Saccharina latissima are likely to survive a heatwave of this magnitude.  However, the southern assemblages are likely to be impacted therefore, resistance has been assessed as ‘None’. As widespread mortality may lead to a lack of viable sporophytes for recruitment, resilience has been assessed as ‘Very low.’ This biotope IR.MIR.KT.XKT is assessed as having ‘High’ sensitivity to marine heatwaves under the middle emission scenario

Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, this could lead to the heatwave lasting the entire summer with temperatures reaching up to 26.5°C in southern England. Laminaria hyperborea and Saccharina latissima are unlikely to survive a heatwave of this magnitude, and as temperatures are likely to reach > 21°C in Scotland under this scenario, there is likely to be mortality throughout these species’ UK biogeographic distribution. Therefore, resistance has been assessed as ‘None’. As a further heatwave is likely to affect this habitat before full recovery (under the pressure benchmark definition), resilience has been assessed as ‘Very low.’ Therefore, this biotope is assessed as having ‘High’ sensitivity to marine heatwaves under the high emission scenario.

None Very Low High
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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

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

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

Under experimental conditions, Nepper-Davidson et al. (2019) exposed a northern (Denmark) population of Saccharina latissima to a simulated three-week heatwave of three different intensities; 18, 21 and 24°C. When exposed to heatwaves of 18 and 21°C there was a decrease in photosynthesis and growth. When a 24°C was simulated, 91% of sporophytes were dead within a week, and the fronds of the few survivors were disintegrating, so the experiment was terminated (Nepper-Davidsen et al., 2019). 

Simonson et al. (2015) investigated the impacts of four temperature treatments (11°C, 14°C, 18°C & 21°C) on Saccharina latissima tissue over three weeksHistological analysis showed temperature mediated tissue damage, including holes, splitting of the medulla, damage to the meristoderm and loss of differentiation between tissue layers at temperatures between 14-21°C. 

Laminaria hyperborea is a cold-temperate species of kelp with an optimum temperature for growth of 15°C, and an upper temperature limit of 21°C (Bolton & Lüning, 1982). Germination success can decrease by almost two thirds at temperatures as low as 17°C. Therefore, it is expected that similar to other kelp species, Laminaria hyperborea will be highly sensitive to marine heatwaves.

Sensitivity assessment. Under the middle emission scenario, if heatwaves occurred every three years, with a maximum intensity of 2°C for 80 days by the end of this century, this could lead to summer sea temperatures reaching up to 24°C in southern England. Laminaria hyperborea and Saccharina latissima are unlikely to survive a heatwave of this magnitude and likely to suffer severe mortality in the south. Although, in Scotland, where a significant portion of these biotopes occur, temperatures are not predicted to rise above 20°C, and therefore, Laminaria hyperborea and Saccharina latissima are likely to survive a heatwave of this magnitude.  However, the southern assemblages are likely to be impacted therefore, resistance has been assessed as ‘None’. As widespread mortality may lead to a lack of viable sporophytes for recruitment, resilience has been assessed as ‘Very low.’ This biotope IR.MIR.KT.XKT is assessed as having ‘High’ sensitivity to marine heatwaves under the middle emission scenario

Under the high emission scenario, if heatwaves occur every two years by the end of this century, reaching a maximum intensity of 3.5°C for 120 days, this could lead to the heatwave lasting the entire summer with temperatures reaching up to 26.5°C in southern England. Laminaria hyperborea and Saccharina latissima are unlikely to survive a heatwave of this magnitude, and as temperatures are likely to reach > 21°C in Scotland under this scenario, there is likely to be mortality throughout these species’ UK biogeographic distribution. Therefore, resistance has been assessed as ‘None’. As a further heatwave is likely to affect this habitat before full recovery (under the pressure benchmark definition), resilience has been assessed as ‘Very low.’ Therefore, this biotope is assessed as having ‘High’ sensitivity to marine heatwaves under the high emission scenario.

High High Not sensitive
Q: Medium
A: Medium
C: Medium
Q: High
A: High
C: High
Q: Medium
A: Medium
C: Medium

Increasing levels of CO2 in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop up to a further 0.35 units by the end of this century, dependent on emission scenario. Marine autotrophs will generally benefit from ocean acidification, through an increase in the availability of aqueous COfor photosynthesis (Koch et al., 2013). 

Most species of kelp, including Laminaria hyperborea, appear to be undersaturated in respect to carbon dioxide, although they can generally utilise HCO3 and have external carbonic anhydrase for extracellular dehydration of HCO3to CO2 (Koch et al., 2013). This was confirmed for Laminaria hyperborea by Olischläger et al. (2012) who found that ocean acidification at levels expected for the end of this century (700 µatm CO2; a value between the middle and high emission scenario) led to an increase in female gametogenesis and increasing net photosynthesis and growth of sporophytes. 

Research on other kelp species has revealed a positive or neutral effect of ocean acidification (Roleda et al., 2012, Fernández et al., 2015, Nunes et al., 2015, Iñiguez et al., 2016b, a), except for one study, which found that ocean acidification negatively impacted photosynthesis and growth in the southern hemisphere species, Ecklonia radiata (Britton et al., 2016).

Under experimental COenrichment at levels expected by the end of this century, germination rates in Saccharina latissima were the same as control samples but gametophyte size increased, suggesting a benefit for juvenile stages of this species (Roleda et al., 2012). Nunes et al. (2015) found that experimental exposure of adult Saccharina latissima to enhanced CO2 led to an increase in net primary production, while Gordillo et al. (2015) found that enhanced CO2 led to increased photosynthesis and growth. In contrast, Iñiguez et al. (2016) found no increase in carbon fixation under elevated CO2 conditions. Although contrasting in findings, these studies show that ocean acidification will not negatively impact Saccharina latissima.

Sensitivity assessment. Kelp forests live in a naturally variable pH habitat, with diel fluctuations of 0.3 - 0.45 pH units (Krause-Jensen et al., 2015, Britton et al., 2016), and boundary layer pH fluctuation of up to 0.8 units (Krause-Jensen et al., 2015). Laminaria hyperborea and Saccharina latissima are not expected to be impacted by ocean acidification at levels expected for the end of this century. Therefore, under both the middle and high emission scenario resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a score of ‘Not sensitive’.

High High Not sensitive
Q: Medium
A: Medium
C: Medium
Q: High
A: High
C: High
Q: Medium
A: Medium
C: Medium

Increasing levels of CO2 in the atmosphere have led to the average pH of sea surface waters dropping from 8.25 in the 1700s to 8.14 in the 1990s (Jacobson, 2005), with it expected to drop up to a further 0.35 units by the end of this century, dependent on emission scenario. Marine autotrophs will generally benefit from ocean acidification, through an increase in the availability of aqueous COfor photosynthesis (Koch et al., 2013). 

Most species of kelp, including Laminaria hyperborea, appear to be undersaturated in respect to carbon dioxide, although they can generally utilise HCO3 and have external carbonic anhydrase for extracellular dehydration of HCO3to CO2 (Koch et al., 2013). This was confirmed for Laminaria hyperborea by Olischläger et al. (2012) who found that ocean acidification at levels expected for the end of this century (700 µatm CO2; a value between the middle and high emission scenario) led to an increase in female gametogenesis and increasing net photosynthesis and growth of sporophytes. 

Research on other kelp species has revealed a positive or neutral effect of ocean acidification (Roleda et al., 2012, Fernández et al., 2015, Nunes et al., 2015, Iñiguez et al., 2016b, a), except for one study, which found that ocean acidification negatively impacted photosynthesis and growth in the southern hemisphere species, Ecklonia radiata (Britton et al., 2016).

Under experimental COenrichment at levels expected by the end of this century, germination rates in Saccharina latissima were the same as control samples but gametophyte size increased, suggesting a benefit for juvenile stages of this species (Roleda et al., 2012). Nunes et al. (2015) found that experimental exposure of adult Saccharina latissima to enhanced CO2 led to an increase in net primary production, while Gordillo et al. (2015) found that enhanced CO2 led to increased photosynthesis and growth. In contrast, Iñiguez et al. (2016) found no increase in carbon fixation under elevated CO2 conditions. Although contrasting in findings, these studies show that ocean acidification will not negatively impact Saccharina latissima.

Sensitivity assessment. Kelp forests live in a naturally variable pH habitat, with diel fluctuations of 0.3 - 0.45 pH units (Krause-Jensen et al., 2015, Britton et al., 2016), and boundary layer pH fluctuation of up to 0.8 units (Krause-Jensen et al., 2015). Laminaria hyperborea and Saccharina latissima are not expected to be impacted by ocean acidification at levels expected for the end of this century. Therefore, under both the middle and high emission scenario resistance is assessed as ‘High’, and resilience is assessed as ‘High’ leading to a score of ‘Not sensitive’.

Medium Very Low Medium
Q: Low
A: NR
C: NR
Q: High
A: High
C: High
Q: Low
A: Low
C: Low

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). Sea-level rise is expected to lead to substantial loss of intertidal habitats. Rocky shores backed by cliffs constitute about 80% of oceanic coastlines globally and in Britain, 42% of the coastline is hard rock, with many areas having cliffs behind the shore (Jackson & McIlvenny, 2011).

This biotope (IR.MIR.KT.XKT) occurs on sheltered, very sheltered and extremely sheltered infralittoral bedrock, boulders and cobbles (JNCC, 2015). Light availability and water turbidity are principal factors in determining kelp depth range (Birkett et al., 1998b), with laminarians being reported to be able to withstand light levels of up to 1% surface irradiance. 

Understanding how sea-level rise will affect tidal energy is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. 

Although the distribution of Laminaria hyperborea is positivity related to wave exposure (Pedersen et al., 2012), and Saccharina latissima is abundant at both turbid and deep sites (Gerard, 1990), this biotope (IR.MIR.KT.XKT) occurs at wave sheltered sites, so that an increase in wave exposure (e.g. to moderate or higher) is likely to result in modification of the community and loss of the biotope. 

Sensitivity assessment. An increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope, which is already constrained to shallow waters through limits to light availability. The biotope is recorded from 0 to 10 m in depth (JNCC, 2015). 

This biotope (IR.MIR.KT.XKT) may be able to expand its range and migrate landwards to compensate for sea-level rise, if not constrained by lack of tide-swept rock, or human-modified shorelines (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of this biotope will shrink substantially in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery, due to the increased depth, leading to a reduction in light availability for photosynthesis. 

There is likely to be considerable variation between sites, the relative contribution of wave surge and exposure to habitat suitability, and the depth range occupied by the biotope. Hence, it is difficult to assess the effect of the different sea-level rise scenarios. However, as the biotope (IR.MIR.KT.XKT) can occur from 0-10 m in depth, it is assumed at a sea-level rise of 50 cm, or 70 cm (middle to high emission scenarios) would have limited effect but that a 107 cm rise (the extreme emission scenario) might result in loss of some of the deeper extent of the biotope in some sites. Therefore, resistance is assessed as ‘High’ under the middle and high emission scenarios so that resilience is ‘High’ and sensitivity assessed as ‘Not sensitive’. But resistance may be ‘Medium’ under the extreme emission scenario so that resilience is ‘Very low’ and sensitivity assessed as ‘Medium’, albeit with ‘Low’ confidence.

High High Not sensitive
Q: Low
A: NR
C: NR
Q: High
A: High
C: High
Q: Low
A: Low
C: Low

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). Sea-level rise is expected to lead to substantial loss of intertidal habitats. Rocky shores backed by cliffs constitute about 80% of oceanic coastlines globally and in Britain, 42% of the coastline is hard rock, with many areas having cliffs behind the shore (Jackson & McIlvenny, 2011).

This biotope (IR.MIR.KT.XKT) occurs on sheltered, very sheltered and extremely sheltered infralittoral bedrock, boulders and cobbles (JNCC, 2015). Light availability and water turbidity are principal factors in determining kelp depth range (Birkett et al., 1998b), with laminarians being reported to be able to withstand light levels of up to 1% surface irradiance. 

Understanding how sea-level rise will affect tidal energy is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. 

Although the distribution of Laminaria hyperborea is positivity related to wave exposure (Pedersen et al., 2012), and Saccharina latissima is abundant at both turbid and deep sites (Gerard, 1990), this biotope (IR.MIR.KT.XKT) occurs at wave sheltered sites, so that an increase in wave exposure (e.g. to moderate or higher) is likely to result in modification of the community and loss of the biotope. 

Sensitivity assessment. An increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope, which is already constrained to shallow waters through limits to light availability. The biotope is recorded from 0 to 10 m in depth (JNCC, 2015). 

This biotope (IR.MIR.KT.XKT) may be able to expand its range and migrate landwards to compensate for sea-level rise, if not constrained by lack of tide-swept rock, or human-modified shorelines (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of this biotope will shrink substantially in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery, due to the increased depth, leading to a reduction in light availability for photosynthesis. 

There is likely to be considerable variation between sites, the relative contribution of wave surge and exposure to habitat suitability, and the depth range occupied by the biotope. Hence, it is difficult to assess the effect of the different sea-level rise scenarios. However, as the biotope (IR.MIR.KT.XKT) can occur from 0-10 m in depth, it is assumed at a sea-level rise of 50 cm, or 70 cm (middle to high emission scenarios) would have limited effect but that a 107 cm rise (the extreme emission scenario) might result in loss of some of the deeper extent of the biotope in some sites. Therefore, resistance is assessed as ‘High’ under the middle and high emission scenarios so that resilience is ‘High’ and sensitivity assessed as ‘Not sensitive’. But resistance may be ‘Medium’ under the extreme emission scenario so that resilience is ‘Very low’ and sensitivity assessed as ‘Medium’, albeit with ‘Low’ confidence.

High High Not sensitive
Q: Low
A: NR
C: NR
Q: High
A: High
C: High
Q: Low
A: Low
C: Low

Sea-level rise is occurring through a combination of thermal expansion and ice melt.  Sea levels have risen 1-3 mm/yr. in the last century (Cazenave & Nerem, 2004, Church et al., 2004, Church & White, 2006). Sea-level rise is expected to lead to substantial loss of intertidal habitats. Rocky shores backed by cliffs constitute about 80% of oceanic coastlines globally and in Britain, 42% of the coastline is hard rock, with many areas having cliffs behind the shore (Jackson & McIlvenny, 2011).

This biotope (IR.MIR.KT.XKT) occurs on sheltered, very sheltered and extremely sheltered infralittoral bedrock, boulders and cobbles (JNCC, 2015). Light availability and water turbidity are principal factors in determining kelp depth range (Birkett et al., 1998b), with laminarians being reported to be able to withstand light levels of up to 1% surface irradiance. 

Understanding how sea-level rise will affect tidal energy is fraught with uncertainty, although evidence appears to suggest that any alterations will be non-linear (Pickering et al., 2012, Li et al., 2016). Modelling potential outcomes of sea-level rise on the tidal and residual currents in the Bohai Sea, China showed effects were site-dependent, with energy either increasing or decreasing (Li et al., 2016). Similarly, Pickering et al. (2012) found a similar pattern around the UK for tidal amplitude. 

Although the distribution of Laminaria hyperborea is positivity related to wave exposure (Pedersen et al., 2012), and Saccharina latissima is abundant at both turbid and deep sites (Gerard, 1990), this biotope (IR.MIR.KT.XKT) occurs at wave sheltered sites, so that an increase in wave exposure (e.g. to moderate or higher) is likely to result in modification of the community and loss of the biotope. 

Sensitivity assessment. An increase in sea level height of 50, 70 and 107 cm could have severe repercussions for the extent of this biotope, which is already constrained to shallow waters through limits to light availability. The biotope is recorded from 0 to 10 m in depth (JNCC, 2015). 

This biotope (IR.MIR.KT.XKT) may be able to expand its range and migrate landwards to compensate for sea-level rise, if not constrained by lack of tide-swept rock, or human-modified shorelines (IPCC, 2019). If landward migration is not possible, it is expected that depth distribution of this biotope will shrink substantially in response to a 50, 70 or 107 cm sea-level rise, without the possibility of recovery, due to the increased depth, leading to a reduction in light availability for photosynthesis. 

There is likely to be considerable variation between sites, the relative contribution of wave surge and exposure to habitat suitability, and the depth range occupied by the biotope. Hence, it is difficult to assess the effect of the different sea-level rise scenarios. However, as the biotope (IR.MIR.KT.XKT) can occur from 0-10 m in depth, it is assumed at a sea-level rise of 50 cm, or 70 cm (middle to high emission scenarios) would have limited effect but that a 107 cm rise (the extreme emission scenario) might result in loss of some of the deeper extent of the biotope in some sites. Therefore, resistance is assessed as ‘High’ under the middle and high emission scenarios so that resilience is ‘High’ and sensitivity assessed as ‘Not sensitive’. But resistance may be ‘Medium’ under the extreme emission scenario so that resilience is ‘Very low’ and sensitivity assessed as ‘Medium’, albeit with ‘Low’ confidence.

Hydrological Pressures

Use / to open/close text displayedResistanceResilienceSensitivity
Low High Low
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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

The temperature isotherm of 19-20°C has been reported as limiting Saccharina lattisma growth (Müller et al., 2009). Gametophytes can develop in ≤23°C (Lüning, 1990). Optimal temperature for Saccharina latissima sporophyte growth was 10-15°C (Bolton & Lüning, 1982), while  reported  growth was inhibited by 50-70% at 20°C and all experimental specimens completely disintegrated after 7 days at 23°C.  In the field, Saccharina latissima has however shown significant regional variation in its acclimation response to changing environmental conditions.  For example Gerard & Dubois (1988) found Saccharina latissima sporophytes which were regularly exposed to ≥20°C could tolerate these high temperatures, whereas sporophytes from other populations which rarely experience ≥17°C showed 100% mortality after 3 weeks of exposure to 20°C.  Therefore, the response Saccharina latissima to a change in temperatures is likely to be locally variable.

Andersen et al. (2011) transplanted Saccharina latissima in the Skagerrak region, Norway and from 2006-2009. There was annual variation, however, high mortality occurred from August-November within each year of the experiment. In 2008 of the original 17 sporophytes 6 survived from March-September (approx. 65% mortality rate). All surviving sporophytes were heavily fouled by epiphytic organisms (estimated cover of 80 & 100%). Between 1960-2009, sea surface temperatures in the region have regularly exceeded 20°C and so has the duration which temperatures remain above 20°C. High sea temperatures has been linked to slow growth of Saccharina latissima which is likely to decrease the photosynthetic ability of, and increase the vulnerability of Saccharina latissima to epiphytic loading, bacterial and viral attacks (Anderson et al., 2011). These factors combined with establishment of annual filamentous algae in Skagerrak, Norway are likely to prevent the establishment of self sustaining populations in the area (Anderson et al., 2011; Moy & Christie, 2012).

IR.MIR.KT.XKT & IR.MIR.KT.XKTX is distributed throughout the UK (Connor et al., 2004). Northern to southern Sea Surface Temperature (SST) ranges from 8-16°C in summer and 6-13°C in winter (Beszczynska-Möller & Dye, 2013).

Sensitivity assessment. A 2°C increase for one year may impair Laminaria hyperborea recruitment processes and Saccharina latissima sporophyte growth but otherwise not affect the characterizing species.  A 5°C increase for one month combined with high UK summer temperatures is likely to affect Laminaria hyperborea sporophyte growth. Saccharina latissima populations that are not acclimated to >20°C may incur mass mortality within 3 weeks of exposure. Resistance has been assessed as ‘Low’, to reflect the potential mass mortality effect of sudden temperature increases on Saccharina latissima, and resilience as ‘High’. Sensitivity has been assessed as ‘Low’.

High High Not sensitive
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

Kain (1964) stated that Laminaria hyperborea sporophyte growth and reproduction could occur within a temperature range of 0-20°C. Upper and lower lethal temperatures have been estimated at between 1-2°C above or below the extremes of these ranges (Birkett et al., 1988). Saccharina lattissima has a lower temperature threshold for sporophyte growth at 0°C (Lüning, 1990). Subtidal red algae can survive at temperatures between -2 °C and 18-23 °C (Lüning, 1990; Kain & Norton, 1990).

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

Low Medium Medium
Q: High
A: High
C: High
Q: High
A: Low
C: High
Q: High
A: Low
C: NR

Lüning (1990) suggest that ‘kelps’ are stenohaline, their general tolerance to salinity as a phenotypic group covering 16-50 psu over a 24 hr period. Optimal growth probably occurs between 30-35 psu and growth rates are likely to be affected by periodic salinity stress. Birkett et al. (1998) suggested that long-term increases in salinity may affect Laminaria hyperborea growth and may result in loss of affected kelp, and therefore loss of the biotope.

Karsten (2007) tested the photosynthetic ability of Saccharina latissima under acute 2 and5 day exposure to salinity treatments ranging from 5-60 psu. A control experiment was also carried at 34 psu . Saccharina latissima showed high photosynthetic ability at >80% of the control levels between 25-55 psu. The affect of long-term salinity changes (>5 days) or salinity >60 PSU on Saccharina latissima’ photosynthetic ability was not tested.

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

Low Medium Medium
Q: High
A: High
C: High
Q: High
A: Low
C: High
Q: High
A: Low
C: High

Lüning (1990) suggest that ‘kelps’ are stenohaline, their general tolerance to salinity as a phenotypic group covering 16 - 50 psu over a 24 hr period. Optimal growth probably occurs between 30-35 psu and growth rates are likely to be affected by periodic salinity stress. Birkett et al. (1998) suggest that long-term changes in salinity may result in loss of affected kelp. Hopkin & Kain (1978) tested Laminaria hyperborea sporophyte growth at various low salinity treatments. The results showed that Laminaria hyperborea sporophytes could grow ‘normally’ at 19 psu, growth was reduced at 16 psu and did not grow at 7 psu.

Karsten (2007) tested the photosynthetic ability of Saccharina latissima under acute 2 and5 day exposure to salinity treatments ranging from 5-60 psu. A control experiment was also carried at 34 psu . Saccharina latissima showed high photosynthetic ability at >80% of the control levels between 25-55 psu. Hyposaline treatment of 10-20 psu led to a gradual decline of photosynthetic ability. After 2 days at 5 psu Saccharina latissima showed a significant decline in photosynthetic ability at approx. 30% of control. After 5 days at 5 psu Saccharina latissima specimens became bleached and showed signs of severe damage. The affect of long-term salinity changes (>5 days) or salinity >60 PSU on Saccharina latissima’ photosynthetic ability was not tested. The experiment was conducted on Saccharina latissima from the Arctic, and the authors suggest that at extremely low water temperatures (1-5°C) macroalgae acclimation to rapid salinity changes could be slower than at temperate latitudes. It is therefore possible that resident Saccharina latissima of the UK maybe be able to acclimate to salinity changes more effectively and quicker.

Sensitivity assessment. IR.MIR.KT.XKT & IR.MIR.KT.XKTX are recorded in both full and variable salinity (18-40) A decrease in one MNCR salinity scale to ‘Reduced Salinity’ (18-30 psu) may result in a decrease of Laminaria hyperborea sporophyte growth and Saccharina latissima. Resistance has been assessed as ‘Low’ and resilience as ‘Medium’. Therefore, sensitivity of this biotope to a decrease in salinity has been assessed as ‘Medium’.

High High Not sensitive
Q: Medium
A: High
C: High
Q: High
A: High
C: High
Q: Medium
A: High
C: High

Peteiro & Freire (2013) measured Saccharina latissima growth from 2 sites, the first had maximal water velocities of 0.3 m/sec and the second 0.1 m/sec. At site 1 Saccharina latissima had significantly larger biomass than at site 2 (16 kg/m to 12 kg/m respectively). Peteiro & Freire (2013) suggested that faster water velocities were beneficial to Saccharina latissima growth. However, Gerard & Mann (1979) found Saccharina latissima productivity is reduced in moderately strong tidal streams (≤1m/sec) when compared to weak tidal streams (<0.5 m/sec). Despite these results where the substratum is unstable Saccharina lattissima can become the dominant canopy forming kelp within tide swept conditions, as in IR.MIR.KT.XKTX (Connor et al., 2004).

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 & 0.3 m/s at the exposed and sheltered sites respectively. Despite the differences in wave exposure and water velocity there was no significant difference in Laminaria hyperborea growth between the exposed and sheltered sites. Therefore water flow was found to have no significant effect on Laminaria hyperborea growth at the observed range of water velocities.

Sensitivity assessment. IR.MIR.KT.XKT & IR.MIR.KT.XKTX are predominantly recorded from “Moderately strong” tidal streams (0.5-1.5 m/sec). Due to the range of tidal velocities that these biotopes are recorded within a change in flow of between 0.1-0.2m/sec would likely have no significant effect on Laminaria hyperborea or Saccharina latissima growth or productivity. Resistance has been assessed as ‘High’, resilience as ‘High’. Sensitivity has been assessed as ‘Not Sensitive’ at the benchmark level.

Low Medium Medium
Q: Low
A: NR
C: NR
Q: High
A: Low
C: High
Q: Low
A: Low
C: Low

IR.MIR.KT.XKT & IR.MIR.KT.XKTX are shallow water biotopes, recorded predominantly from 0-5 m BCD.  An increase in emergence will result in an increased risk of desiccation and mortality of the dominant kelp species (Laminaria hyperborea & Saccharina latissima). Removal of canopy forming kelps has also been shown to increase desiccation and mortality of the understorey macro-algae (Hawkins & Harkin, 1985). Several mobile species such as sea urchins, brittle stars and feather stars are likely to move away. However, providing that suitable substrata are present, the biotope is likely to re-establish further down the shore within a similar emergence regime to that which existed previously.

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

High High Not sensitive
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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. to 1.6 times greater between the exposed and sheltered sites. Maximal significant wave height (Hm0) was 3.67 & 2m at the exposed and sheltered sites. Maximal water velocity (Velrms) was 0.6 & 0.3m/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.

However, Pedersen et al. (2012) observed Laminaria hyperborea biomass, productivity and density increased with greater wave exposure.  At low wave exposure Laminaria hyperborea canopy forming plants were smaller, had lower densities and had higher mortality rates. At low wave exposure, high epiphytic loading on Laminaria hyperborea was suggested to impair photosynthesis, nutrient uptake, and increase the drag of the host Laminaria hyperborea during extreme storm events. The morphology of kelp stipe and blades vary in different water flows and wave exposures 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 et al., 1998). In addition, the lamina becomes narrower and thinner in strong currents (Sjøtun & Fredriksen, 1995).

Saccharina latissima is rarely found at wave exposed sites (Birkett et al., 1998). Saccharina latissima, if present, develops a short thick stipe and a short, narrow and tightly wrinkled blade (Birkett et al., 1998).

Sensitivity assessment. Wave exposure is one of the principal defining features of kelp biotopes, and changes in wave exposure are likely to alter the relative abundance of the kelp species, understorey community and hence the biotope. IR.MIR.KT.XKT & IR.MIR.KT.XKTX are recorded from wave sheltered sites, so that an increase in wave exposure (e.g. to moderate or higher) is likely to result in modification of the community and loss of the biotope. However a change in near shore significant wave height of 3-5% is unlikely to have any significant effect on IR.MIR.KT.XKT & IR.MIR.KT.XKTX. Resistance has been assessed as ‘High’, resilience as ‘High’ and sensitivity as ‘Not Sensitive’ at the benchmark level.

Chemical Pressures

Use / to open/close text displayedResistanceResilienceSensitivity
Not Assessed (NA) Not assessed (NA) Not assessed (NA)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

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

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 a,. (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.

Sporophytes of Saccharina latissima have a low intolerance to heavy metals, but the early life stages are more intolerant. The effects of copper, zinc and mercury on Saccharina latissima have been investigated by Thompson & Burrows (1984). They observed that the growth of sporophytes was significantly inhibited at 50 µg Cu /l, 1000 µg Zn/l and 50 µg Hg/l. Zoospores were found to be more intolerant and significant reductions in survival rates were observed at 25 µg Cu/l, 1000 µg Zn/l and 5 µg/l. Little is known about the effects of heavy metals on echinoderms. Bryan (1984) reported that early work had shown that echinoderm larvae were intolerant of heavy metals, e.g. the intolerance of larvae of Paracentrotus lividus to copper (Cu) had been used to develop a water quality assessment. Kinne (1984) reported developmental disturbances in Echinus esculentus exposed to waters containing 25 µg / l of copper (Cu). Sea-urchins, especially the eggs and larvae, are used for toxicity testing and environmental monitoring (reviewed by Dinnel et al. 1988). Taken together with the findings of Gomez & Miguez-Rodriguez (1999) above it is likely that echinoderms are intolerant of heavy metal contamination.

Not Assessed (NA) Not assessed (NA) Not assessed (NA)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

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

Laminaria hyperborea and Saccharina latissima fronds, being predominantly subtidal, would not come into contact with freshly released oil but only to sinking emulsified oil and oil adsorbed onto particles (Birkett et al., 1998). The mucilaginous slime layer coating of 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 hyperbora communities, showed no noticeable impacts of the Sea Empress oil spill and clean up (Rostron & Bunker, 1997). An assessment of holdfast fauna in Laminaria showed that although species richness and diversity decreased with increasing proximity to the Sea Empress oil spill, overall the holdfasts contained a reasonably rich and diverse fauna, even though oil was present in most samples (Sommerfield & Warwick, 1999). Laboratory studies of the effects of oil and dispersants on several red algae species, including Delesseria sanguinea (Grandy 1984; cited in Holt et al., 1995) concluded that they were all sensitive to oil/ dispersant mixtures, with little differences between adults, sporelings, diploid or haploid life stages. Holt et al. (1995) concluded that Delesseria sanguinea is probably generally sensitive of chemical contamination.

Not Assessed (NA) Not assessed (NA) Not assessed (NA)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

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). Saccharina latissima has also been found to be sensitive to antifouling compounds. Johansson (2009) exposed samples of Saccharina latissima to several antifouing compounds, observing chlorothalonil, DCOIT, dichlofluanid and tolylfluanid inhibited photosynthesis. Exposure to Chlorothalonil and tolylfluanid, was also found to continue inhibiting oxygen evolution after exposure had finished, and may cause irreversible damage.

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

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.

Not relevant (NR) Not relevant (NR) No evidence (NEv)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

No evidence

Not Assessed (NA) Not assessed (NA) Not assessed (NA)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

This pressure is Not assessed.

High High Not sensitive
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

Reduced oxygen concentrations can inhibit both photosynthesis and respiration in macroalgae (Kinne, 1977). Despite this, macroalgae are thought to buffer the environmental conditions of low oxygen, thereby acting as a refuge for organisms in oxygen depleted regions especially if the oxygen depletion is short term (Frieder et al., 2012). A rapid recovery from a state of low oxygen is expected if the environmental conditions are transient. If levels do drop below 4 mg/l negative effects on these organisms can be expected with adverse effects occurring below 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. In addition, IR.MIR.KT.XKT & IR.MIR.KT.XKTX are tide swept so that any deoxygenation would be highly localised and transient. Resistance has been assessed as ‘High’, Resilience as ‘High’. Sensitivity has been assessed as ‘Not sensitive’ at the benchmark level.

Not relevant (NR) Not relevant (NR) Not sensitive
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Conolly & Drew (1985) found Saccharina latissima sporophytes had relatively higher growth rates when in close proximity to a sewage outlet in St Andrews, UK when compared to other sites along the east coast of Scotland. At St Andrews nitrate levels were 20.22µM, which represents an approx 25% increase when compared to other comparable sites (approx 15.87 µM). Handå et al. (2013) also reported Saccharina latissima sporophytes grew approx 1% faster per day when in close proximity to Salmon farms, where elevated ammonium can be readily absorbed.  Read et al. (1983) reported after the installation of a new sewage treatment  works which reduced the suspended solid content of liquid effluent by 60% in the Firth of Forth, Saccharina latissima  became abundant where previously it had been absent. Bokn et al. (2003) conducted a nutrient loading experiment on intertidal fucoids. Within 3 years of the experiment no significant effect was observed in the communities, however 4-5 years into the experiment a shift occurred from perennials to ephemeral algae occurred. Although Bokn et al. (2003) focussed on fucoids the results could indicate that long term (>4 years) nutrient loading can result in community shift to ephemeral algae species. Disparities between the findings of the aforementioned studies are likely to be related to the level of organic enrichment however could also be time dependant.

Johnston & Roberts (2009) conducted a meta analysis, which reviewed 216 papers to assess how a variety of contaminants (including sewage and nutrient loading) affected 6 marine habitats (including subtidal reefs). A 30-50% 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 relative 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).

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

Sensitivity assessment. Although nutrients may not affect kelps directly, indirect effects such as turbidity may significantly affect photosynthesis. Furthermore nutrient enrichment may denude the associated community. However, the biotope is probably ‘Not sensitive’ (resistance is ‘High’ and resilience is ‘High) at the benchmark level (i.e. compliance with WFD criteria). 

Medium High Low
Q: High
A: High
C: High
Q: High
A: Medium
C: High
Q: High
A: High
C: High

Conolly & Drew (1985) found Saccharina latissima sporophytes had relatively higher growth rates when in close proximity to a sewage outlet in St Andrews, UK when compared to other sites along the east coast of Scotland. At St Andrews nitrate levels were 20.22µM, which represents an approx 25% increase when compared to other comparable sites (approx 15.87 µM). Handå et al. (2013) also reported Saccharina latissima sporophytes grew approx 1% faster per day when in close proximity to Norwegian Salmon farms, where elevated ammonium can be readily absorbed.  Read et al. (1983) reported after the installation of a new sewage treatment  works, which reduced the suspended solid content of liquid effluent by 60% in the Firth of Forth, Saccharina latissima became abundant where previously it had been absent. Bokn et al. (2003) conducted a nutrient loading experiment on intertidal fucoids. Within 3 years of the experiment no significant effect was observed in the communities, however 4-5 years into the experiment a shift occurred from perennials to ephemeral algae occurred. Although Bokn et al. (2003) focussed on fucoids the results could indicate that long term (>4 years) nutrient loading can result in community shift to ephemeral algae species. Disparities between the findings of the aforementioned studies are likely to be related to the level of organic enrichment however could also be time dependent.

Johnston & Roberts (2009) conducted a meta analysis, which reviewed 216 papers to assess how a variety of contaminants (including sewage and nutrient loading) affected 6 marine habitats (including subtidal reefs). A 30-50% 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 relative 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).

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).  Increases in ephemeral and opportunistic algae are associated with reduced numbers of perennial macrophytes (Fletcher, 1996).  Increased nutrients may also result in phytoplankton blooms that increase turbidity.

Sensitivity assessment. Although nutrients may not affect kelps directly, indirect effects such as turbidity may significantly affect photosynthesis. Furthermore, organic enrichment may denude the associated community. Resistance has therefore been assessed as ‘Medium’, resilience as ‘High’. Sensitivity has been assessed as ’Low’.

Physical Pressures

Use / to open/close text displayedResistanceResilienceSensitivity
None Very Low High
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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.

None Very Low High
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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

Not relevant (NR) Not relevant (NR) Not relevant (NR)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Not relevant

Not relevant (NR) Not relevant (NR) Not relevant (NR)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Not relevant to rock substrata.

None High Medium
Q: Low
A: NR
C: NR
Q: High
A: High
C: High
Q: Low
A: Low
C: Low

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

Sensitivity assessment. Abrasion by passing trawls or harvesting of macroalgae is likely remove a large proportion of the kelp biomass.  For example in kelp harvesting is likely to remove all the large canopy forming plants (Svendsen, 1972; Christie et al., 1998).  However, Saccharina latissima has been shown to be an early colonizer with the potential to recover rapidly (Kain, 1967; Leinaas & Christie, 1996). Therefore, resistance has been assessed as ‘None’, resilience as ‘High’, and sensitivity as ‘Low’.

Not relevant (NR) Not relevant (NR) Not relevant (NR)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Not relevant, please refer to pressure “Abrasion/disturbance of the substratum on the surface of the seabed”.

Low Medium Medium
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

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 one 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).

Sensitivity Assessment. Changes in water clarity are likely to affect photosynthetic rates and enable Saccharina latissima to compete more successfully with Laminaria hyperborea.  A decrease in turbidity is likely to support enhanced growth (and possible habitat expansion) and is therefore not considered in this assessment.  An increase in water clarity from clear to intermediate (10-100 mg/l) represents 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.

Medium High Low
Q: Low
A: NR
C: NR
Q: High
A: High
C: High
Q: Low
A: Low
C: Low

Smothering by sediment e.g. 5 cm material during a discrete event is unlikely to damage Saccharina latissima sporophytes but may affect holdfast fauna, gametophyte survival, interfere with zoospore settlement and therefore recruitment processes (Moy & Christie, 2012). Given the short life expectancy of Saccharina latissima (2-4 years-(Parke, 1948)), IR.MIR.KT.SlatT is likely to be dependent on annual Saccharina latissima recruitment (Moy & Christie, 2012). Given the microscopic size of the gametophyte, 5 cm of sediment could be expected to significantly inhibit growth. However, laboratory studies showed that kelp gametophytes can survive in darkness for between 6-16 months at 8°C and would probably survive smothering by a discrete event. Once returned to normal conditions the gametophytes resumed growth or maturation within one month (Dieck, 1993). Intolerance to this factor is likely to be higher during the peak periods of sporulation and/or spore settlement.

Dendrodoa grossularia is a small ascidian, capable of reaching a size of approx 8.5 mm (Miller, 1954) and is therefore likely to be inundated by deposition of 5 cm of sediment. If inundation is long lasting then the understorey community may be adversely affected. However, IR.MIR.KT.SlatT is found within strong-moderately strong (0.5-3 m/sec) and therefore deposited sediments are unlikely to remain for more than a few tidal cycles.

Sensitivity assessment. Resistance has been assessed as ‘Medium’, resilience as ‘High’. Sensitivity has been assessed as ‘Low’.

Low High Low
Q: Low
A: NR
C: NR
Q: High
A: High
C: High
Q: Low
A: Low
C: Low

Smothering by sediment e.g. 30 cm material during a discrete event is unlikely to damage Saccharina latissima sporophytes but may affect holdfast fauna, gametophyte survival, interfere with zoospore settlement and therefore recruitment processes (Moy & Christie, 2012). Given the short life expectancy of Saccharina latissima (2-4 years-(Parke, 1948)), IR.MIR.KT.SlatT is likely to be dependent on annual recruitment (Moy & Christie, 2012). 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 by a discrete event. Once returned to normal conditions the gametophytes resumed growth or maturation within 1 month (Dieck, 1993). Intolerance to this factor is likely to be higher during the peak periods of sporulation and/or spore settlement.

Dendrodoa grossularia is a small ascidian, capable of reaching a size of approx 8.5mm (Miller, 1954) and is therefore likely to be inundated by deposition of 30 cm of sediment. If inundation is long lasting then the understorey community may be adversely affected. However, IR.MIR.KT.SlatT is found within strong-moderately strong (0.5-3m/sec) and therefore deposited sediments are likely to be cleared rapidly, but inundation is likely to cause mortality in the understorey community.

Sensitivity assessment. Resistance has been assessed as ‘Low’, and resilience as ‘High’. Sensitivity has been assessed as ‘Low’.

Not Assessed (NA) Not assessed (NA) Not assessed (NA)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Not assessed.

Not relevant (NR) Not relevant (NR) No evidence (NEv)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

No evidence

Not relevant (NR) Not relevant (NR) Not relevant (NR)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Not relevant

Low Medium Medium
Q: Low
A: NR
C: NR
Q: Low
A: NR
C: NR
Q: Low
A: Low
C: Low

There is no evidence to suggest that anthropogenic light sources would affect macro-algal growth. Shading of the biotope (e.g. by construction of a pontoon, pier etc) could adversely affect the biotope in areas where the water clarity is also low, and tip the balance to shade tolerant species, resulting in the loss of the biotope directly within the shaded area, or a reduction in laminarian abundance from forest to park type biotopes.

Sensitivity assessment. Resistance is probably 'Low', with a 'Medium' resilience and a sensitivity of 'Medium', albeit with 'low' confidence due to the lack of direct evidence.

Not relevant (NR) Not relevant (NR) Not relevant (NR)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

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.

Not relevant (NR) Not relevant (NR) Not relevant (NR)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Not relevant. Collision from grounding vessels is addressed under abrasion above.

Not relevant (NR) Not relevant (NR) Not relevant (NR)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

Not relevant

Biological Pressures

Use / to open/close text displayedResistanceResilienceSensitivity
Not relevant (NR) Not relevant (NR) No evidence (NEv)
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR
Q: NR
A: NR
C: NR

No evidence.

Low Very Low High
Q: Low
A: NR
C: NR
Q: High
A: High
C: High
Q: Low
A: Low
C: Low

Competition with invasive macroalgae may be a potential threat to this biotope.  Potential invasives include Undaria pinnatifida, Sargassum muticum and Codium fragile spp. tormentosoides. In Nova Scotia, Codium fragile spp. tormentosoides competes successfully with native kelps for space including Laminaria digitata, exploiting gaps within the kelp beds.  Once established, the algal mat created by Codium fragile spp. tormentosoides prevents re-colonization by other macro-algae (Scheibling & Gagnon, 2006).

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 ppt.  Although fertilization does not occur below 15 ppt 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 (Staeher 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 is the clear waters of California.  However, it is a poor competitor under low light and only develops dense canopies in shallow areas (Engelen et al., 2015).

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

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

Undaria pinnatifida (Wakame or Asian kelp) is a large brown seaweed and an Invasive Non-Indigenous Species (INIS) that could out-compete native UK kelp species (see Farrell & Fletcher, 2006; Thompson & Schiel, 2012; Brodie et al., 2014; Hieser et al., 2014; Arnold et al., 2016; Epstein & Smale, 2017; Epstein & Smale, 2018; Kraan, 2017; Epstein et al., 2019a,b; Tidbury, 2020).  Undaria pinnatifida originates from Japan but is 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).  It has since proliferated along UK coastlines.  One year after its discovery at the Queen Anne Battery marina, Plymouth, it had become a major fouling plant on pontoons (Minchin & Nunn, 2014).  Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound.

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

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

Undaria pinnatifida has a wide physiological niche meaning it can occur in both coastal and estuarine environments showing tolerance for varying salinities, turbidity and siltation (Heiser et al., 2014; Epstein & Smale, 2018).  Undaria pinnatifida has a preference for sites sheltered with low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018).  In natural habitats, Undaria pinnatifida was not recorded if the wave fetch was greater than 642 km but increased in abundance and cover in very sheltered sites (Epstein & Smale, 2018).  In St Malo, France, there was evidence that Undaria pinnatifida 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 Laminaria hyperborea was significantly less abundant at sites with the presence Undaria pinnatifida, with only ca 0.5 Laminaria hyperborea individuals per m2 present compared to ca 8 individuals per m2 at sites without the presence Undaria pinnatifida. However, the results from their correlation study only showed that the species were not found together (pers. comm., Epstien 2021). Whereas, 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). However, Epstein et al. (2019b) also noted that Laminaria hyperborea had a non-significant positive relationship with Undaria pinnatifida due to low densities of Laminaria hyperborea across the study area, resulting in insufficient data.

Epstein et al. (2019b) reported that Undaria pinnatifida biomass was negatively related to Saccharina latissima in both intertidal and subtidal habitats. This was only statistically significant in subtidal habitats, which suggested that there was some competition between the two species (Epstein et al., 2019b). Heiser et al. (2014) surveyed 17 sites within Plymouth Sound, UK, and found that Saccharina latissima was significantly more abundant at sites with Undaria pinnatifida with ca 5 Saccharina latissima individuals present per m², compared to ca 0.5 Saccharina latissima individuals per m² present at sites without Undaria pinnatifida.

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

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

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

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.  The above evidence suggests that Undaria pinnatifida can co-exist with Saccharina latissima and Laminaria hyperborea.  For example, within natural habitats, it can co-exist with native kelp species within its depth range (-1 to 4 m), as shown in Plymouth Sound, UK.  A dense kelp canopy may restrict or slow the proliferation of Undaria pinnatifida, however, there has been mixed evidence on its colonization with Laminaria hyperborea beds and in some areas, a lower abundance of Laminaria hyperborea may result in increased Undaria pinnatifida growth.

This Laminaria hyperborea and/or Saccharina latissima dominated biotope (IR.MIR.KT.XKT) is found in the shallow sublittoral (0-10 m, JNCC, 2015) sheltered from wave action but structured by strong tidal streams and scour.  The abundance of the perennial Laminaria hyperborea suggests that scour is localised or intermittent rather than seasonal.  The evidence above suggests that Undaria prefers sheltered conditions, with a low tidal flow and is unlikely to be completive under the conditions that characterize this biotope.  However, However, Sargassum muticum prefers wave sheltered shallow sites in the sublittoral fringe.  It was reported to out-compete and replace Saccharina latissima in the Limfjorden, and achieve maximum abundance at 1-4 m (Staehr et al., 2000; Engelen et al., 2015).  But Strong & Dring (2011) concluded that Sargassum was not a threat to Saccharina latissima in the Dorn, Strangford Lough where it coexisted and grew better in mixed stands.  Hence, competition with Sargassum is probably site-specific and dependent on local conditions.  No evidence of the effects of Sargassum on Laminaria hyperborea beds was found.

Therefore, resistance is assessed as ‘Low’ to represent colonization by Sargassum of the shallow (0-5 m) Saccharina latissima dominated examples of the biotope and the possible loss of Saccharina.  Even if the two species co-exist, the invasion may result in a change in the classification of the biotope and the structure of the understorey macroalgae (Staehr et al., 2000).  Sargassum is unlikely to compete in deeper (5-10 m) examples of the biotope and/or examples dominated by Laminaria hyperborea.  Recovery after invasion by Sargassum, although rapid, would require direct intervention (removal).  Hence, resilience is probably ‘Very low’ so sensitivity is assessed as ‘High’.  Overall, confidence is assessed as ‘Low’ due to evidence of variation and site-specific nature of competition between native kelps and both Undaria pinnatifida and Sargassum muticum.

Medium High Low
Q: Medium
A: High
C: Medium
Q: Low
A: NR
C: NR
Q: Low
A: Low
C: Low

Laminaria hyperborea and Saccharina latissima may be infected by the microscopic brown alga Streblonema aecidioides. Infected algae show symptoms of Streblonema disease, i.e. alterations of the blade and stipe ranging from dark spots to heavy deformations and completely crippled thalli (Peters & Scaffelke, 1996). Infection can reduce growth rates of host algae.

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

Low High Low
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

There has been recent commercial interest in Saccharina lattisma as a consumable called ‘sea vegetables’ (Birkett et al., 1998). Laminaria hyperborea is also extracted on a commercial scale in southern Norway, primarily for alagnate (Werner & Kraan, 2004).

Commercial Laminaria hyperborea trawling occurs in Norway, during which Christie et al. (1998) reports all large canopy forming sporophytes are removed, sub-canopy sporophytes and understorey community however remain intact. Saccharina latissima is commercially cultivated, however typically sporophytes are matured on ropes (Handå et al., 2013) and not directly extracted from the seabed. Thus evidence to assess the resistance of Saccharina latissima to direct harvesting is limited.

Sensitivity assessment. Resistance has been assessed as ‘None’, Resilience as ‘Medium’. Sensitivity has been assessed as ‘Medium’.

None High Medium
Q: High
A: High
C: High
Q: High
A: High
C: High
Q: High
A: High
C: High

Incidental/accidental removal of Laminaria hyperborea and Saccharina latissima is likely to cause similar effects to that of direct harvesting; as such the same evidence has been used for both pressure assessments. There has been recent commercial interest in Saccharina latissima as a consumable called ‘sea vegetable’’ (Birkett et al., 1998). Laminaria hyperborea is also extracted on a commercial scale in southern Norway, primarily for alginates (Werner & Kraan, 2004).

Commercial Laminaria hyperborea trawling occurs in Norway, during which Christie et al. (1998) reports all large canopy forming sporophytes are removed, sub-canopy sporophytes and understorey community however remain intact. Saccharina latissima is commercially cultivated, however typically sporophytes are matured on ropes (Handå et al., 2013) and not directly extracted from the seabed. Thus evidence to assess the resistance of Saccharina latissima to direct harvesting is limited.

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

Bibliography

  1. Andersen, G.S., Steen, H., Christie, H., Fredriksen, S. & Moy, F.E., 2011. Seasonal patterns of sporophyte growth, fertility, fouling, and mortality of Saccharina latissima in Skagerrak, Norway: implications for forest recovery. Journal of Marine Biology, 2011, Article ID 690375, 8 pages.

  2. Andrew, N.L. & Viejo, R.M., 1998. Ecological limits to the invasion of Sargassum muticum in northern Spain. Aquatic Botany, 60 (3), 251-263. DOI https://doi.org/10.1016/S0304-3770(97)00088-0

  3. Arafeh-Dalmau, N., Montaño-Moctezuma, G., Martínez, J.A., Beas-Luna, R., Schoeman, D.S. & Torres-Moye, G., 2019. Extreme Marine Heatwaves Alter Kelp Forest Community Near Its Equatorward Distribution Limit. Frontiers in Marine Science, 6 (499). DOI https://doi.org/10.3389/fmars.2019.00499

  4. Arnold, M., Teagle, H., Brown, M.P. & Smale, D.A., 2016. The structure of biogenic habitat and epibiotic assemblages associated with the global invasive kelp Undaria pinnatifida in comparison to native macroalgae. Biological Invasions, 18 (3), 661-676. DOI https://doi.org/10.1007/s10530-015-1037-6

  5. Assis, J., Araújo, M.B. & Serrão, E.A., 2018. Projected climate changes threaten ancient refugia of kelp forests in the North Atlantic. Global Change Biology, 24 (1), e55-e66. DOI https://doi.org/10.1111/gcb.13818

  6. Assis, J., Lucas, A.V., Bárbara, I. & Serrão, E.Á., 2016. Future climate change is predicted to shift long-term persistence zones in the cold-temperate kelp Laminaria hyperborea. Marine Environmental Research, 113, 174-182. DOI https://doi.org/10.1016/j.marenvres.2015.11.005

  7. Assis, J., Serrão, E.A., Claro, B., Perrin, C. & Pearson, G.A., 2014. Climate-driven range shifts explain the distribution of extant gene pools and predict future loss of unique lineages in a marine brown alga. Molecular Ecology, 23 (11), 2797-2810. DOI https://doi.org/10.1111/mec.12772

  8. Beszczynska-Möller, A., & Dye, S.R., 2013. ICES Report on Ocean Climate 2012. In ICES Cooperative Research Report, vol. 321 pp. 73.

  9. Birkett, D.A., Maggs, C.A., Dring, M.J. & Boaden, P.J.S., 1998b. Infralittoral reef biotopes with kelp species: an overview of dynamic and sensitivity characteristics for conservation management of marine SACs. Natura 2000 report prepared by Scottish Association of Marine Science (SAMS) for the UK Marine SACs Project., Scottish Association for Marine Science. (UK Marine SACs Project, vol VI.), 174 pp. Available from: http://ukmpa.marinebiodiversity.org/uk_sacs/pdfs/reefkelp.pdf

  10. Bower, S.M., 1996. Synopsis of Infectious Diseases and Parasites of Commercially Exploited Shellfish: Bald-sea-urchin Disease. [On-line]. Fisheries and Oceans Canada. [cited 26/01/16]. Available from: http://www.dfo-mpo.gc.ca/science/aah-saa/diseases-maladies/bsudsu-eng.html

  11. Breeman, A.M., 1990. Expected Effects of Changing Seawater Temperatures on the Geographic Distribution of Seaweed Species. In Beukema, J.J., et al. (eds.). Expected Effects of Climatic Change on Marine Coastal Ecosystems, Dordrecht: Springer Netherlands, pp. 69-76. DOI: https://doi.org/10.1007/978-94-009-2003-3_9

  12. Brennan, G., Kregting, L., Beatty, G.E., Cole, C., Elsäßer, B., Savidge, G. & Provan, J., 2014. Understanding macroalgal dispersal in a complex hydrodynamic environment: a combined population genetic and physical modelling approach. Journal of The Royal Society Interface, 11 (95), 20140197.

  13. Britton, D., Cornwall, C.E., Revill, A.T., Hurd, C.L. & Johnson, C.R., 2016. Ocean acidification reverses the positive effects of seawater pH fluctuations on growth and photosynthesis of the habitat-forming kelp, Ecklonia radiata. Scientific reports, 6 (1), 26036. DOIhttps://doi.org/10.1038/srep26036

  14. Britton-Simmons, K.H., 2004. Direct and indirect effects of the introduced alga Sargassum muticum on benthic, subtidal communities of Washington State, USA. Marine Ecology Progress Series, 277, 61-78. DOI https://doi.org/10.3354/meps277061

  15. Brodie J., Williamson, C.J., Smale, D.A., Kamenos, N.A., Mieszkowska, N., Santos, R., Cunliffe, M., Steinke, M., Yesson, C. & Anderson, K.M., 2014. The future of the northeast Atlantic benthic flora in a high CO2 world. Ecology and Evolution, 4 (13), 2787-2798. DOI  https://doi.org/10.1002/ece3.1105

  16. Brodie J., Williamson, C.J., Smale, D.A., Kamenos, N.A., Mieszkowska, N., Santos, R., Cunliffe, M., Steinke, M., Yesson, C. & Anderson, K.M., 2014. The future of the northeast Atlantic benthic flora in a high CO2 world. Ecology and Evolution, 4 (13), 2787-2798. DOI  https://doi.org/10.1002/ece3.1105

  17. Bryan, G.W., 1984. Pollution due to heavy metals and their compounds. In Marine Ecology: A Comprehensive, Integrated Treatise on Life in the Oceans and Coastal Waters, vol. 5. Ocean Management, part 3, (ed. O. Kinne), pp.1289-1431. New York: John Wiley & Sons.

  18. Burrows, M.T., Smale, D., O’Connor, N., Rein, H.V. & Moore, P., 2014. Marine Strategy Framework Directive Indicators for UK Kelp Habitats Part 1: Developing proposals for potential indicators. Joint Nature Conservation Comittee,  Peterborough. Report no. 525.

  19. Casas, G., Scrosati, R. & Piriz, M.L., 2004. The invasive kelp Undaria pinnatifida (Phaeophyceae, Laminariales) reduces native seaweed diversity in Nuevo Gulf (Patagonia, Argentina). Biological Invasions, 6 (4), 411-416.

  20. Castric-Fey, A., Girard, A. & L'Hardy-Halos, M.T., 1993. The Distribution of Undaria pinnatifida (Phaeophyceae, Laminariales) on the Coast of St. Malo (Brittany, France). Botanica Marina, 36 (4), 351-358. DOI https://doi.org/10.1515/botm.1993.36.4.351

  21. Cazenave, A. & Nerem, R.S., 2004. Present-day sea-level change: Observations and causes. Reviews of Geophysics, 42 (3). DOI https://doi.org/10.1029/2003rg000139

  22. Chamberlain, Y.M., 1996. Lithophylloid Corallinaceae (Rhodophycota) of the genera Lithophyllum and Titausderma from southern Africa. Phycologia, 35, 204-221.

  23. Christie, H., Fredriksen, S. & Rinde, E., 1998. Regrowth of kelp and colonization of epiphyte and fauna community after kelp trawling at the coast of Norway. Hydrobiologia, 375/376, 49-58.

  24. Church, J.A. & White, N.J., 2006. A 20th century acceleration in global sea-level rise. Geophysical Research Letters, 33 (1). DOI https://doi.org/10.1029/2005gl024826

  25. Church, J.A., White, N.J., Coleman, R., Lambeck, K. & Mitrovica, J.X., 2004. Estimates of the Regional Distribution of Sea Level Rise over the 1950–2000 Period. Journal of Climate, 17 (13), 2609-2625. DOI https://doi.org/10.1175/1520-0442(2004)017

  26. Cole, S., Codling, I.D., Parr, W. & Zabel, T., 1999. Guidelines for managing water quality impacts within UK European Marine sites. Natura 2000 report prepared for the UK Marine SACs Project. 441 pp., Swindon: Water Research Council on behalf of EN, SNH, CCW, JNCC, SAMS and EHS. [UK Marine SACs Project.]. Available from: http://ukmpa.marinebiodiversity.org/uk_sacs/pdfs/water_quality.pdf

  27. Colhart, B.J., & Johanssen, H.W., 1973. Growth rates of Corallina officinalis (Rhodophyta) at different temperatures. Marine Biology, 18, 46-49.

  28. Connor, D.W., Dalkin, M.J., Hill, T.O., Holt, R.H.F. & Sanderson, W.G., 1997a. Marine biotope classification for Britain and Ireland. Vol. 2. Sublittoral biotopes. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06., Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06.

  29. Dauvin, J.C., Bellan, G., Bellan-Santini, D., Castric, A., Francour, P., Gentil, F., Girard, A., Gofas, S., Mahe, C., Noel, P., & Reviers, B. de., 1994. Typologie des ZNIEFF-Mer. Liste des parametres et des biocoenoses des cotes francaises metropolitaines. 2nd ed. Secretariat Faune-Flore, Museum National d'Histoire Naturelle, Paris (Collection Patrimoines Naturels, Serie Patrimoine Ecologique, No. 12). Coll. Patrimoines Naturels, vol. 12, Secretariat Faune-Flore, Paris.

  30. Davies, C.E. & Moss, D., 1998. European Union Nature Information System (EUNIS) Habitat Classification. Report to European Topic Centre on Nature Conservation from the Institute of Terrestrial Ecology, Monks Wood, Cambridgeshire. [Final draft with further revisions to marine habitats.], Brussels: European Environment Agency.

  31. Dayton, P.K. & Tegner, M.J., 1984. Catastrophic storms, El-Nino, and patch stability in a southern-california kelp community. Science, 224 (4646), 283-285.

  32. Dayton, P.K., Tegner, M.J., Parnell, P.E. & Edwards, P.B., 1992. Temporal and spatial patterns of disturbance and recovery in a kelp forest community. Ecological Monographs, 62, 421-445.

  33. De Bettignies, T., de Bettignies, F., Bartsch, I., Bekkby, T., Boiffin, A., Casado de Amezúa, P., Christie, H., Edwards, H., Fournier, N., García, A., Gauthier, L., Gillham, K., Halling, C., Harrald, M., Hennicke, J., Hernández, S., Kilnäs, M., Martinez, B., Mieszkowska, N., Moore, P., Moy, F., Mueller, M., Norderhaug, K.M., Ó Cadhla, O., Parry, M., Ramsay, K., Robertson, M., Russel, T., Serrão, E., Smale, D., Sousa Pinto, I., Steen, H., Street, M., Walday, M., Werner, T. & La Rivière, M., 2021. Background Document for Kelp Forests. OSPAR Commission, London, OSPAR 788/2021, 66 pp. Available from: https://www.ospar.org/documents?v=46796

  34. De Leij, R., Epstein, G., Brown, M.P. & Smale, D.A., 2017. The influence of native macroalgal canopies on the distribution and abundance of the non-native kelp Undaria pinnatifida in natural reef habitats. Marine Biology, 164 (7). DOI https://doi.org/10.1007/s00227-017-3183-0

  35. Devlin, M.J., Barry, J., Mills, D.K., Gowen, R.J., Foden, J., Sivyer, D. & Tett, P., 2008. Relationships between suspended particulate material, light attenuation and Secchi depth in UK marine waters. Estuarine, Coastal and Shelf Science, 79 (3), 429-439.

  36. Dieck, T.I., 1992. North Pacific and North Atlantic digitate Laminaria species (Phaeophyta): hybridization experiments and temperature responses. Phycologia, 31, 147-163.

  37. Dieck, T.I., 1993. Temperature tolerance and survival in darkness of kelp gametophytes (Laminariales: Phaeophyta) - ecological and biogeographical implications. Marine Ecology Progress Series, 100, 253-264.

  38. Dommasnes, A., 1968. Variation in the meiofauna of Corallina officinalis with wave exposure. Sarsia, 34, 117-124.

  39. Edwards, A., 1980. Ecological studies of the kelp Laminaria hyperborea and its associated fauna in south-west Ireland. Ophelia, 9, 47-60.

  40. Elner, R.W. & Vadas, R.L., 1990. Inference in ecology: the sea urchin phenomenon in the northwest Atlantic. American Naturalist, 136, 108-125.

  41. Engelen, A.H., Leveque, L., Destombe, C. & Valer, M., 2011. Spatial and temporal patterns of recovery of low intertidal Laminaria digitata after experimental spring and autumn removal. Cahiers De Biologie Marine, 52 (4), 441-453.

  42. Engelen, A.H., Serebryakova, A., Ang, P., Britton-Simmons, K., Mineur, F., Pedersen, M. F., & Toth, G., 2015. Circumglobal invasion by the brown seaweed Sargassum muticum. Oceanography and Marine Biology: An Annual Review, 53, 81-126.

  43. Epstein, G. & Smale, D.A., 2017. Undaria pinnatifida: A case study to highlight challenges in marine invasion ecology and management. Ecology and Evolution, 7 (20), 8624-8642. DOI https://doi.org/10.1002/ece3.3430

  44. Epstein, G. & Smale, D.A., 2018. Environmental and ecological factors influencing the spillover of the non-native kelp, Undaria pinnatifida, from marinas into natural rocky reef communities. Biological Invasions, 20 (4), 1049-1072. DOI https://doi.org/10.1007/s10530-017-1610-2

  45. Epstein, G., Foggo, A. & Smale, D.A., 2019a. Inconspicuous impacts: Widespread marine invader causes subtle but significant changes in native macroalgal assemblages. Ecosphere, 10 (7). DOI https://doi.org/10.1002/ecs2.2814

  46. Epstein, G., Hawkins, S.J. & Smale, D.A., 2019b. Identifying niche and fitness dissimilarities in invaded marine macroalgal canopies within the context of contemporary coexistence theory. Scientific Reports, 9. DOI https://doi.org/10.1038/s41598-019-45388-5

  47. Erwin, D.G., Picton, B.E., Connor, D.W., Howson, C.M., Gilleece, P. & Bogues, M.J., 1990. Inshore Marine Life of Northern Ireland. Report of a survey carried out by the diving team of the Botany and Zoology Department of the Ulster Museum in fulfilment of a contract with Conservation Branch of the Department of the Environment (N.I.)., Ulster Museum, Belfast: HMSO.

  48. Farrell, P. & Fletcher, R., 2006. An investigation of dispersal of the introduced brown alga Undaria pinnatifida (Harvey) Suringar and its competition with some species on the man-made structures of Torquay Marina (Devon, UK). Journal of Experimental Marine Biology and Ecology, 334 (2), 236-243.

  49. Fernández, P.A., Roleda, M.Y. & Hurd, C.L., 2015. Effects of ocean acidification on the photosynthetic performance, carbonic anhydrase activity and growth of the giant kelp Macrocystis pyrifera. 124 (3), 293-304. DOI https://doi.org/10.1007/s11120-015-0138-5

  50. Fletcher, R. & Farrell, P., 1998. Introduced brown algae in the North East Atlantic, with particular respect to Undaria pinnatifida (Harvey) Suringar. Helgolander Meeresuntersuchungen, 52 (3-4), 259-275.

  51. Fletcher, R.L., 1996. The occurrence of 'green tides' - a review. In Marine Benthic Vegetation. Recent changes and the Effects of Eutrophication (ed. W. Schramm & P.H. Nienhuis). Berlin Heidelberg: Springer-Verlag. [Ecological Studies, vol. 123].

  52. Fredriksen, S., Sjøtun, K., Lein, T.E. & Rueness, J., 1995. Spore dispersal in Laminaria hyperborea (Laminariales, Phaeophyceae). Sarsia, 80 (1), 47-53.

  53. Frieder, C., Nam, S., Martz, T. & Levin, L., 2012. High temporal and spatial variability of dissolved oxygen and pH in a nearshore California kelp forest. Biogeosciences, 9 (10), 3917-3930.

  54. Frölicher, T.L., Fischer, E.M. & Gruber, N., 2018. Marine heatwaves under global warming. Nature, 560 (7718), 360-364. DOI https://doi.org/10.1038/s41586-018-0383-9

  55. Gerard, V.A., 1990. Ecotypic differentiation in the kelp Laminaria saccharina: Phase-specific adaptation in a complex life cycle. Marine Biology, 107 (3), 519-528. DOI https://doi.org/10.1007/bf01313437

  56. Gommez, J.L.C. & Miguez-Rodriguez, L.J., 1999. Effects of oil pollution on skeleton and tissues of Echinus esculentus L. 1758 (Echinodermata, Echinoidea) in a population of A Coruna Bay, Galicia, Spain. In Echinoderm Research 1998. Proceedings of the Fifth European Conference on Echinoderms, Milan, 7-12 September 1998, (ed. M.D.C. Carnevali & F. Bonasoro) pp. 439-447. Rotterdam: A.A. Balkema.

  57. Gordillo, F.J.L., Aguilera, J., Wiencke, C. & Jiménez, C., 2015. Ocean acidification modulates the response of two Arctic kelps to ultraviolet radiation. Journal of Plant Physiology, 173, 41-50. DOI https://doi.org/10.1016/j.jplph.2014.09.008

  58. Gordillo, F.J.L., Dring, M.J. & Savidge, G., 2002. Nitrate and phosphate uptake characteristics of three species of brown algae cultured at low salinity. Marine Ecology Progress Series, 234, 111-116.

  59. Gorman, D., Bajjouk, T., Populus, J., Vasquez, M. & Ehrhold, A., 2013. Modeling kelp forest distribution and biomass along temperate rocky coastlines. Marine Biology, 160 (2), 309-325.

  60. Grandy, N., 1984. The effects of oil and dispersants on subtidal red algae. Ph.D. Thesis. University of Liverpool.

  61. Hammer, L., 1972. Anaerobiosis in marine algae and marine phanerograms. In Proceedings of the Seventh International Seaweed Symposium, Sapporo, Japan, August 8-12, 1971 (ed. K. Nisizawa, S. Arasaki, Chihara, M., Hirose, H., Nakamura V., Tsuchiya, Y.), pp. 414-419. Tokyo: Tokyo University Press.

  62. Harkin, E., 1981. Fluctuations in epiphyte biomass following Laminaria hyperborea canopy removal. In Proceedings of the Xth International Seaweed Symposium, Gø teborg, 11-15 August 1980 (ed. T. Levring), pp.303-308. Berlin: Walter de Gruyter.

  63. Hawkins, S.J. & Harkin, E., 1985. Preliminary canopy removal experiments in algal dominated communities low on the shore and in the shallow subtidal on the Isle of Man. Botanica Marina, 28, 223-30.

  64. Hayward, P.J. 1988. Animals on seaweed. Richmond, Surrey: Richmond Publishing Co. Ltd. [Naturalists Handbooks 9].

  65. Heiser, S., Hall-Spencer, J.M. & Hiscock, K., 2014. Assessing the extent of establishment of Undaria pinnatifida in Plymouth Sound Special Area of Conservation, UK. Marine Biodiversity Records, 7, e93.

  66. Hiscock, K. & Mitchell, R., 1980. The Description and Classification of Sublittoral Epibenthic Ecosystems. In The Shore Environment, Vol. 2, Ecosystems, (ed. J.H. Price, D.E.G. Irvine, & W.F. Farnham), 323-370. London and New York: Academic Press. [Systematics Association Special Volume no. 17(b)].

  67. Hiscock, K., 1983. Water movement. In Sublittoral ecology. The ecology of shallow sublittoral benthos (ed. R. Earll & D.G. Erwin), pp. 58-96. Oxford: Clarendon Press.

  68. Hofmann, L.C., Straub, S. & Bischof, K., 2013. Elevated CO2 levels affect the activity of nitrate reductase and carbonic anhydrase in the calcifying rhodophyte Corallina officinalis. Journal of Experimental Botany, 64 (4), 899-908. DOI https://doi.org/10.1093/jxb/ers369

  69. Holt, T.J., Jones, D.R., Hawkins, S.J. & Hartnoll, R.G., 1995. The sensitivity of marine communities to man induced change - a scoping report. Countryside Council for Wales, Bangor, Contract Science Report, no. 65.

  70. Hopkin, R. & Kain, J.M., 1978. The effects of some pollutants on the survival, growth and respiration of Laminaria hyperborea. Estuarine and Coastal Marine Science, 7, 531-553.

  71. Iñiguez, C., Carmona, R., Lorenzo, M.R., Niell, F.X., Wiencke, C. & Gordillo, F.J.L., 2016. Increased temperature, rather than elevated CO2, modulates the carbon assimilation of the Arctic kelps Saccharina latissima and Laminaria solidungula. 163 (12), 248. DOI https://doi.org/10.1007/s00227-016-3024-6

  72. Iñiguez, C., Carmona, R., Lorenzo, M.R., Niell, F.X., Wiencke, C. & Gordillo, F.J.L., 2016a. Increased CO2 modifies the carbon balance and the photosynthetic yield of two common Arctic brown seaweeds: Desmarestia aculeata and Alaria esculenta. Polar Biology, 39 (11), 1979-1991. DOI https://doi.org/10.1007/s00300-015-1724-x

  73. Irvine, L. M. & Chamberlain, Y. M., 1994. Seaweeds of the British Isles, vol. 1. Rhodophyta, Part 2B Corallinales, Hildenbrandiales. London: Her Majesty's Stationery Office.

  74. Jackson, A.C. & McIlvenny, J., 2011. Coastal squeeze on rocky shores in northern Scotland and some possible ecological impacts. Journal of Experimental Marine Biology and Ecology, 400 (1), 314-321. DOI https://doi.org/10.1016/j.jembe.2011.02.012
  75. Jacobson, M.Z., 2005. Studying ocean acidification with conservative, stable numerical schemes for nonequilibrium air-ocean exchange and ocean equilibrium chemistry. Journal of Geophysical Research: Atmospheres, 110 (D7). DOI https://doi.org/10.1029/2004jd00522

  76. James, K, 2017. A review of the impacts from invasion by the introduced kelp Undaria pinnatifida. Waikato Regional Council Technical Report 2016/40, Institute of Marine Science, University of Auckland, Hamilton, 40 pp. Available from: https://www.waikatoregion.govt.nz/assets/WRC/WRC-2019/TR201640.pdf

  77. JNCC (Joint Nature Conservation Committee), 2022.  The Marine Habitat Classification for Britain and Ireland Version 22.04. [Date accessed]. Available from: https://mhc.jncc.gov.uk/

  78. JNCC (Joint Nature Conservation Committee), 2022.  The Marine Habitat Classification for Britain and Ireland Version 22.04. [Date accessed]. Available from: https://mhc.jncc.gov.uk/

  79. JNCC (Joint Nature Conservation Committee), 1999. Marine Environment Resource Mapping And Information Database (MERMAID): Marine Nature Conservation Review Survey Database. [on-line] http://www.jncc.gov.uk/mermaid

  80. Jones, C.G., Lawton, J.H. & Shackak, M., 1994. Organisms as ecosystem engineers. Oikos, 69, 373-386.

  81. Jones, D.J., 1971. Ecological studies on macro-invertebrate communities associated with polluted kelp forest in the North Sea. Helgolander Wissenschaftliche Meersuntersuchungen, 22, 417-431.

  82. Jones, L.A., Hiscock, K. & Connor, D.W., 2000. Marine habitat reviews. A summary of ecological requirements and sensitivity characteristics for the conservation and management of marine SACs. Joint Nature Conservation Committee, Peterborough. (UK Marine SACs Project report.). Available from: http://www.ukmarinesac.org.uk/pdfs/marine-habitats-review.pdf

  83. Jones, N.S. & Kain, J.M., 1967. Subtidal algal recolonisation following removal of Echinus. Helgolander Wissenschaftliche Meeresuntersuchungen, 15, 460-466.

  84. Kain, J.M., 1964. Aspects of the biology of Laminaria hyperborea III. Survival and growth of gametophytes. Journal of the Marine Biological Association of the United Kingdom, 44 (2), 415-433.

  85. Kain, J.M. & Svendsen, P., 1969. A note on the behaviour of Patina pellucida in Britain and Norway. Sarsia, 38, 25-30.

  86. Kain, J.M., 1971a. Synopsis of biological data on Laminaria hyperborea. FAO Fisheries Synopsis, no. 87.

  87. Kain, J.M., 1975a. Algal recolonization of some cleared subtidal areas. Journal of Ecology, 63, 739-765.

  88. Kain, J.M., 1979. A view of the genus Laminaria. Oceanography and Marine Biology: an Annual Review, 17, 101-161.

  89. Kain, J.M., 1987. Photoperiod and temperature as triggers in the seasonality of Delesseria sanguinea. Helgolander Meeresuntersuchungen, 41, 355-370.

  90. Kain, J.M., & Norton, T.A., 1990. Marine Ecology. In Biology of the Red Algae, (ed. K.M. Cole & Sheath, R.G.). Cambridge: Cambridge University Press.

  91. Kain, J.M., Drew, E.A. & Jupp, B.P., 1975. Light and the ecology of Laminaria hyperborea II. In Proceedings of the Sixteenth Symposium of the British Ecological Society, 26-28 March 1974. Light as an Ecological Factor: II (ed. G.C. Evans, R. Bainbridge & O. Rackham), pp. 63-92. Oxford: Blackwell Scientific Publications.

  92. Karsten, U., 2007. Research note: salinity tolerance of Arctic kelps from Spitsbergen. Phycological Research, 55 (4), 257-262.

  93. Kawai, H., Sasaki, H., Maeda, Y. & Arai, S., 2001. Morphology, life history, and molecular phylogeny of Chorda rigida, sp. nov. (Laminariales, Phaeophyceae) from the sea of Japan and the genetic diversity of Chorda FilumJournal of Phycology, 37 (1), 130-142. DOI https://doi.org/10.1046/j.1529-8817.1999.014012130.x

  94. Kinne, O., 1977. International Helgoland Symposium "Ecosystem research": summary, conclusions and closing. Helgoländer Wissenschaftliche Meeresuntersuchungen, 30(1-4), 709-727.

  95. Kitching, J., 1941. Studies in sublittoral ecology III. Laminaria forest on the west coast of Scotland; a study of zonation in relation to wave action and illumination. The Biological Bulletin, 80 (3), 324-337

  96. Koch, M., Bowes, G., Ross, C. & Zhang, X.-H., 2013. Climate change and ocean acidification effects on seagrasses and marine macroalgae. Global Change Biology, 19 (1), 103-132. DOI https://doi.org/10.1111/j.1365-2486.2012.02791.x

  97. Kraan, S., 2017. Undaria marching on; late arrival in the Republic of Ireland. Journal of Applied Phycology, 29 (2), 1107-1114. DOI https://doi.org/10.1007/s10811-016-0985-2

  98. Krause-Jensen, D., Duarte, C.M., Hendriks, I.E., Meire, L., Blicher, M.E., Marbà, N. & Sejr, M.K., 2015. Macroalgae contribute to nested mosaics of pH variability in a subarctic fjord. Biogeosciences, 12 (16), 4895-4911. DOI https://doi.org/10.5194/bg-12-4895-2015

  99. Kregting, L., Blight, A., Elsäßer, B. & Savidge, G., 2013. The influence of water motion on the growth rate of the kelp Laminaria hyperborea. Journal of Experimental Marine Biology and Ecology, 448, 337-345.

  100. Kruuk, H., Wansink, D. & Moorhouse, A., 1990. Feeding patches and diving success of otters, Lutra lutra, in Shetland. Oikos, 57, 68-72.

  101. Lang, C. & Mann, K., 1976. Changes in sea urchin populations after the destruction of kelp beds. Marine Biology, 36 (4), 321-326.

  102. Lein, T.E., Sjøtun, K. & Wakili, S., 1991. Mass-occurrence of a brown filamentous endophyte in the lamina of the kelp Laminaria hyperborea (Gunnerus) Foslie along the southwestern coast of Norway. Sarsia, 76 (3), 187-193. DOI https://doi.org/10.1080/00364827.1991.10413474

  103. Leinaas, H.P. & Christie, H., 1996. Effects of removing sea urchins (Strongylocentrotus droebachiensis): stability of the barren state and succession of kelp forest recovery in the east Atlantic. Oecologia, 105(4), 524-536.

  104. Li, Y., Zhang, H., Tang, C., Zou, T. & Jiang, D., 2016. Influence of Rising Sea Level on Tidal Dynamics in the Bohai Sea. 74 (SI), 22-31. DOI https://doi.org/10.2112/si74-003.1

  105. Lobban, C.S. & Harrison, P.J., 1997. Seaweed ecology and physiology. Cambridge: Cambridge University Press.

  106. Lüning, K., 1990. Seaweeds: their environment, biogeography, and ecophysiology: John Wiley & Sons.

  107. Müller, R., Laepple, T., Bartsch, I. & Wiencke, C., 2009. Impact of oceanic warming on the distribution of seaweeds in polar and cold-temperate waters. Botanica Marina, 52 (6), 617-638.

  108. Macleod, A., Cottier-Cook, E., Hughes, D. & Allen, C., 2016. Investigating the impacts of marine invasive non-native species. Natural England Commissioned Report NECR223, Natural England, 58 pp. Available from: https://pureadmin.uhi.ac.uk/ws/portalfiles/portal/3729569/NECR223_edition_1.pdf

  109. Mann, K.H., 1982. Kelp, sea urchins, and predators: a review of strong interactions in rocky subtidal systems of eastern Canada, 1970-1980. Netherlands Journal of Sea Research, 16, 414-423.

  110. Martin, S. & Hall-Spencer, J.M., 2017. Effects of Ocean Warming and Acidification on Rhodolith / Maerl Beds. In Riosmena-Rodriguez, R., Nelson, W., Aguirre, J. (ed.) Rhodolith / Maerl Beds: A Global Perspective, Switzerland: Springer Nature, pp. 55-85. [Coastal Research Library, 15].

  111. Miller III, H.L., Neale, P.J. & Dunton, K.H., 2009. Biological weighting functions for UV inhibtion of photosynthesis in the kelp Laminaria hyperborea (Phaeophyceae) 1. Journal of Phycology, 45 (3), 571-584.

  112. Minchin, D. & Nunn, J., 2014. The invasive brown alga Undaria pinnatifida (Harvey) Suringar, 1873 (Laminariales: Alariaceae), spreads northwards in Europe. Bioinvasions Records, 3 (2), 57-63. DOI http://dx.doi.org/10.3391/bir.2014.3.2.01

  113. Moore, P.G., 1973a. The kelp fauna of north east Britain I. Function of the physical environment. Journal of Experimental Marine Biology and Ecology, 13, 97-125.

  114. Moore, P.G., 1973b. The kelp fauna of north east Britain. II. Multivariate classification: turbidity as an ecological factor. Journal of Experimental Marine Biology and Ecology, 13, 127-163.

  115. Moore, P.G., 1978. Turbidity and kelp holdfast Amphipoda. I. Wales and S.W. England. Journal of Experimental Marine Biology and Ecology, 32, 53-96.

  116. Moore, P.G., 1985. Levels of heterogeneity and the amphipod fauna of kelp holdfasts. In The Ecology of Rocky Coasts: essays presented to J.R. Lewis, D.Sc. (ed. P.G. Moore & R. Seed), 274-289. London: Hodder & Stoughton Ltd.

  117. Moy, F.E. & Christie, H., 2012. Large-scale shift from sugar kelp (Saccharina latissima) to ephemeral algae along the south and west coast of Norway. Marine Biology Research, 8 (4), 309-321.

  118. NBN, 2015. National Biodiversity Network 2015(20/05/2015). https://data.nbn.org.uk/

  119. Nepper-Davidsen, J., Andersen, D.T. & Pedersen, M.F., 2019. Exposure to simulated heatwave scenarios causes long-term reductions in performance in Saccharina latissima. Marine Ecology Progress Series, 630, 25-39
  120. Nichols, D., 1981. The Cornish Sea-urchin Fishery. Cornish Studies, 9, 5-18.

  121. Norderhaug, K., 2004. Use of red algae as hosts by kelp-associated amphipods. Marine Biology, 144 (2), 225-230.

  122. Norderhaug, K.M. & Christie, H.C., 2009. Sea urchin grazing and kelp re-vegetation in the NE Atlantic. Marine Biology Research, 5 (6), 515-528.

  123. Norderhaug, K.M., Christie, H. & Fredriksen, S., 2007. Is habitat size an important factor for faunal abundances on kelp (Laminaria hyperborea)? Journal of Sea Research, 58 (2), 120-124.

  124. Nordheim, van, H., Andersen, O.N. & Thissen, J., 1996. Red lists of Biotopes, Flora and Fauna of the Trilateral Wadden Sea area, 1995. Helgolander Meeresuntersuchungen, 50 (Suppl.), 1-136.

  125. Norton, T.A., 1992. Dispersal by macroalgae. British Phycological Journal, 27, 293-301.

  126. Norton, T.A., Hiscock, K. & Kitching, J.A., 1977. The Ecology of Lough Ine XX. The Laminaria forest at Carrigathorna. Journal of Ecology, 65, 919-941.

  127. Nunes, J., McCoy, S.J., Findlay, H.S., Hopkins, F.E., Kitidis, V., Queirós, A.M., Rayner, L. & Widdicombe, S., 2015. Two intertidal, non-calcifying macroalgae (Palmaria palmata and Saccharina latissima) show complex and variable responses to short-term CO2 acidification. ICES Journal of Marine Science, 73 (3), 887-896. DOI https://doi.org/10.1093/icesjms/fsv081

  128. Olischläger, M., Bartsch, I., Gutow, L. & Wiencke, C., 2012. Effects of ocean acidification on different life-cycle stages of the kelp Laminaria hyperborea (Phaeophyceae). Botanica Marina, vol. 55 pp. 511
  129. Park, J., Kim, J., Kong, J.-A., Depuydt, S., Brown, M. & Han, T., 2017. Implications of rising temperatures for gametophyte performance of two kelp species from Arctic waters. Botanica Marina, 60. DOI http://doi.org/10.1515/bot-2016-0103

  130. Pedersen, M., 2015. Temperature effects on the kelp Saccharina latissimaASLO,  Grenada, Spain,  pp.

  131. Pedersen, M.F., Nejrup, L.B., Fredriksen, S., Christie, H. & Norderhaug, K.M., 2012. Effects of wave exposure on population structure, demography, biomass and productivity of the kelp Laminaria hyperborea. Marine Ecology Progress Series, 451, 45-60.

  132. Penfold, R., Hughson, S., & Boyle, N., 1996. The potential for a sea urchin fishery in Shetland. http://www.nafc.ac.uk/publish/note5/note5.htm, 2000-04-14

  133. Pessarrodona, A., Moore, P.J., Sayer, M.D.J. & Smale, D.A., 2018. Carbon assimilation and transfer through kelp forests in the NE Atlantic is diminished under a warmer ocean climate. Global Change Biology, 24 (9), 4386-4398. DOI https://doi.org/10.1111/gcb.14303

  134. Philippart, C.J., Anadón, R., Danovaro, R., Dippner, J.W., Drinkwater, K.F., Hawkins, S.J., Oguz, T., O'Sullivan, G. & Reid, P.C., 2011. Impacts of climate change on European marine ecosystems: observations, expectations and indicators. Journal of Experimental Marine Biology and Ecology, 400 (1), 52-69.

  135. Pickering, M.D., Wells, N.C., Horsburgh, K.J. & Green, J.A.M., 2012. The impact of future sea-level rise on the European Shelf tides. Continental Shelf Research, 35, 1-15. DOI https://doi.org/10.1016/j.csr.2011.11.011

  136. Price, J.H., Irvine, D.E. & Farnham, W.F., 1980. The shore environment. Volume 2: Ecosystems. London Academic Press.

  137. Raffaelli, D.G.  & Hawkins, S.J., 1999. Intertidal Ecology 2nd edn.. London: Kluwer Academic Publishers.

  138. Redmond, S. 2013. Effects of Increasing Temperature and Ocean Acidification on the Microstages of two Populations of Saccharina latissima in the Northwest Atlantic. Master of Science,  University of Connecticut.

  139. Rinde, E. & Sjøtun, K., 2005. Demographic variation in the kelp Laminaria hyperborea along a latitudinal gradient. Marine Biology, 146 (6), 1051-1062.

  140. Rogers-Bennett, L. & Catton, C.A., 2019. Marine heatwave and multiple stressors tip bull kelp forest to sea urchin barrens. Scientific Reports, 9 (1), 15050. DOI https://doi.org/10.1038/s41598-019-51114-y

  141. Roleda, M.Y., Morris, J.N., McGraw, C.M. & Hurd, C.L., 2012. Ocean acidification and seaweed reproduction: increased CO2 ameliorates the negative effect of lowered pH on meiospore germination in the giant kelp Macrocystis pyrifera (Laminariales, Phaeophyceae). Global Change Biology, 18 (3), 854-864. DOI https://doi.org/10.1111/j.1365-2486.2011.02594.x

  142. Rostron, D.M. & Bunker, F. St P.D., 1997. An assessment of sublittoral epibenthic communities and species following the Sea Empress oil spill. A report to the Countryside Council for Wales from Marine Seen & Sub-Sea Survey., Countryside Council for Wales, Bangor, CCW Sea Empress Contact Science, no. 177.

  143. Scheibling, R.E. & Gagnon, P., 2006. Competitive interactions between the invasive green alga Codium fragile ssp tomentosoides and native canopy-forming seaweeds in Nova Scotia (Canada). Marine Ecology Progress Series, 325, 1-14.

  144. Schiel, D.R. & Foster, M.S., 1986. The structure of subtidal algal stands in temperate waters. Oceanography and Marine Biology: an Annual Review, 24, 265-307.

  145. Schoenrock, K.M., O’Callaghan, T., O’Callaghan, R. & Krueger-Hadfield, S.A., 2019. First record of Laminaria ochroleuca Bachelot de la Pylaie in Ireland in Béal an Mhuirthead, county Mayo. Marine Biodiversity Records, 12 (1), 9. DOI https://doi.org/10.1186/s41200-019-0168-3

  146. Seapy , R.R. & Littler, M.M., 1982. Population and Species Diversity Fluctuations in a Rocky Intertidal Community Relative to Severe Aerial Exposure and Sediment Burial. Marine Biology, 71, 87-96.

  147. SEEEC (Sea Empress Environmental Evaluation Committee), 1998. The environmental impact of the Sea Empress oil spill. Final Report of the Sea Empress Environmental Evaluation Committee, 135 pp., London: HMSO.

  148. Sheppard, C.R.C., Bellamy, D.J. & Sheppard, A.L.S., 1980. Study of the fauna inhabiting the holdfasts of Laminaria hyperborea (Gunn.) Fosl. along some environmental and geographical gradients. Marine Environmental Research, 4, 25-51.

  149. Simonson, E., Scheibling, R. & Metaxas, A., 2015. Kelp in hot water: I.Warming seawater temperature induces weakening and loss of kelp tissue. Marine Ecology Progress Series, 537. DOI http://doi.org/10.3354/meps11438

  150. Sivertsen, K., 1997. Geographic and environmental factors affecting the distribution of kelp beds and barren grounds and changes in biota associated with kelp reduction at sites along the Norwegian coast. Canadian Journal of Fisheries and Aquatic Sciences, 54, 2872-2887.

  151. Sjøtun, K., Christie, H. & Helge Fosså, J., 2006. The combined effect of canopy shading and sea urchin grazing on recruitment in kelp forest (Laminaria hyperborea). Marine Biology Research, 2 (1), 24-32.

  152. Sjøtun, K. & Schoschina, E.V., 2002. Gametophytic development of Laminaria spp. (Laminariales, Phaeophyta) at low temperatures. Phycologia, 41, 147-152.

  153. Smale, D.A. & Vance, T., 2015. Climate-driven shifts in species’ distributions may exacerbate the impacts of storm disturbances on North-east Atlantic kelp forests. Marine and Freshwater Research, 67 (1), 65-74. DOI https://doi.org/10.1071/MF14155

  154. Smale, D.A., 2020. Impacts of ocean warming on kelp forest ecosystems. New Phytologist, 225, 1447-1454. DOI https://doi.org/10.1111/nph.16107

  155. Smale, D.A., Burrows, M.T., Moore, P., O'Connor, N. & Hawkins, S.J., 2013. Threats and knowledge gaps for ecosystem services provided by kelp forests: a northeast Atlantic perspective. Ecology and evolution, 3 (11), 4016-4038.

  156. Smale, D.A., Epstein, G., Hughes, E., Mogg, A.O.M. & Moore, P.J., 2020. Patterns and drivers of understory macroalgal assemblage structure within subtidal kelp forests. Biodiversity and Conservation, 29 (14), 4173-4192. DOI https://doi.org/10.1007/s10531-020-02070-x

  157. Smale, D.A., Pessarrodona, A., King, N., Burrows, M.T., Yunnie, A., Vance, T. & Moore, P., 2020b. Environmental factors influencing primary productivity of the forest-forming kelp Laminaria hyperborea in the northeast Atlantic. Scientific Reports, 10 (1), 12161. DOI https://doi.org/10.1038/s41598-020-69238-x

  158. Smale, D.A., Wernberg, T., Oliver, E.C.J., Thomsen, M., Harvey, B.P., Straub, S.C., Burrows, M.T., Alexander, L.V., Benthuysen, J.A., Donat, M.G., Feng, M., Hobday, A.J., Holbrook, N.J., Perkins-Kirkpatrick, S.E., Scannell, H.A., Sen Gupta, A., Payne, B.L. & Moore, P.J., 2019. Marine heatwaves threaten global biodiversity and the provision of ecosystem services. Nature Climate Change, 9 (4), 306-312. DOI https://doi.org/10.1038/s41558-019-0412-1

  159. Smale, D.A., Wernberg, T., Yunnie, A.L. & Vance, T., 2014. The rise of Laminaria ochroleuca in the Western English Channel (UK) and comparisons with its competitor and assemblage dominant Laminaria hyperborea. Marine ecology.

  160. Smale, D.A., Wernberg, T., Yunnie, A.L.E. & Vance, T., 2015. The rise of Laminaria ochroleuca in the Western English Channel (UK) and comparisons with its competitor and assemblage dominant Laminaria hyperborea. Marine Ecology, 36 (4), 1033-1044. DOI https://doi.org/10.1111/maec.12199

  161. Smith, B.D., 1985. Recovery following experimental harvesting of Laminaria longicruris and Laminaria digitata in Southwestern Nova Scotia. Helgolander Meeresuntersuchungen, 39(1), 83-101.

  162. Smith, J.E. (ed.), 1968. 'Torrey Canyon'. Pollution and marine life. Cambridge: Cambridge University Press.

  163. Somerfield, P.J. & Warwick, R.M., 1999. Appraisal of environmental impact and recovery using Laminaria holdfast faunas. Sea Empress, Environmental Evaluation Committee., Countryside Council for Wales, Bangor, CCW Sea Empress Contract Science, Report no. 321.

  164. South, G.H. & Burrows, E.M., 1967. Studies on marine algae of the British Isles. 5. Chorda filum (l.) Stckh. British Phycological Bulletin, 3 , 379-402.

  165. Staehr, P.A., Pedersen, M.F., Thomsen, M.S., Wernberg, T. & Krause-Jensen, D., 2000. Invasion of Sargassum muticum in Limfjorden (Denmark) and its possible impact on the indigenous macroalgal community. Marine Ecology Progress Series, 207, 79-88. DOI https://doi.org/10.3354/meps207079

  166. Steneck, R.S., Graham, M.H., Bourque, B.J., Corbett, D., Erlandson, J.M., Estes, J.A. & Tegner, M.J., 2002. Kelp forest ecosystems: biodiversity, stability, resilience and future. Environmental conservation, 29 (04), 436-459.

  167. Steneck, R.S., Vavrinec, J. & Leland, A.V., 2004. Accelerating trophic-level dysfunction in kelp forest ecosystems of the western North Atlantic. Ecosystems, 7 (4), 323-332.

  168. Strong, J.A. & Dring, M.J., 2011. Macroalgal competition and invasive success: testing competition in mixed canopies of Sargassum muticum and Saccharina latissima. Botanica Marina, 54 (3), 223-229.

  169. Teagle, H., Hawkins, S. J., Moore, P. J. & Smale, D. A., 2017. The role of kelp species as biogenic habitat formers in coastal marine ecosystems. Journal of Experimental Marine Biology and Ecology, 492, 81-98. DOI https://doi.org/10.1016/j.jembe.2017.01.017

  170. Thompson, G.A. & Schiel, D.R., 2012. Resistance and facilitation by native algal communities in the invasion success of Undaria pinnatifida. Marine Ecology, Progress Series, 468, 95-105.

  171. Tidbury, H, 2020. Wakame (Undaria pinnatifida). GB Non-native Species Rapid Risk Assessment., 15 pp. Available from: http://www.nonnativespecies.org/index.cfm?pageid=143

  172. Vadas, R.L. & Elner, R.W., 1992. Plant-animal interactions in the north-west Atlantic. In Plant-animal interactions in the marine benthos, (ed. D.M. John, S.J. Hawkins & J.H. Price), 33-60. Oxford: Clarendon Press. [Systematics Association Special Volume, no. 46].

  173. Vadas, R.L., Johnson, S. & Norton, T.A., 1992. Recruitment and mortality of early post-settlement stages of benthic algae. British Phycological Journal, 27, 331-351.

  174. Valentine, J. P. & Johnson, C. R., 2003. Establishment of the introduced kelp Undaria pinnatifida in Tasmania depends on disturbance to native algal assemblages. Journal of Experimental Marine Biology and Ecology, 295 (1), 63-90. DOI https://doi.org/10.1016/S0022-0981(03)00272-7

  175. Van den Hoek, C., 1982. The distribution of benthic marine algae in relation to the temperature regulation of their life histories. Biological Journal of the Linnean Society, 18, 81-144.

  176. Van den Hoek, C., Mann, D.G. & Jahns, H.M., 1995. Algae: an introduction to phycology: Cambridge University Press.

  177. Vaz-Pinto, F., Rodil, I.F., Mineur, F., Olabarria, C. & Arenas, F., 2014. Understanding biological invasions by seaweeds. In Pereira, L. & Neto, J.M. (eds.). Marine algae: biodiversity, taxonomy, environmental assessment and biotechnology. Boca Raton, Florida: CRC Press, pp. 140-177.

  178. Viejo, R.M., Arrontes, J. & Andrew, N.L., 1995. An Experimental Evaluation of the Effect of Wave Action on the Distribution of Sargassum muticum in Northern Spain. , 38 (1-6), 437-442. DOI https://doi.org/10.1515/botm.1995.38.1-6.437

  179. Vost, L.M., 1983. The influence of Echinus esculentus grazing on subtidal algal communities. British Phycological Journal, 18, 211.

  180. Whittick, A., 1983. Spatial and temporal distributions of dominant epiphytes on the stipes of Laminaria hyperborea (Gunn.) Fosl. (Phaeophyta: Laminariales) in S.E. Scotland. Journal of Experimental Marine Biology and Ecology, 73, 1-10.

  181. Wiens, J.J., 2016. Climate-Related Local Extinctions Are Already Widespread among Plant and Animal Species. PLOS Biology, 14 (12), e2001104. DOI https://doi.org/10.1371/journal.pbio.2001104
  182. Wood, E. (ed.), 1988. Sea Life of Britain and Ireland. Marine Conservation Society. IMMEL Publishing, London

  183. Wotton, D.M., O'Brien, C., Stuart, M.D. & Fergus, D.J., 2004. Eradication success down under: heat treatment of a sunken trawler to kill the invasive seaweed Undaria pinnatifida. Marine Pollution Bulletin, 49 (9), 844-849.

  184. Yarish, C., Penniman, C.A. & Egan, B., 1990. Growth and reproductibe responses of Laminaria longicruris (Laminariales, Phaeophyta) to nutrient enrichment. Hydrobiologia, 204, 505-511.

  185. Yildiz, G., Hofmann Laurie, C., Bischof, K. & Dere, Ş., 2013. Ultraviolet radiation modulates the physiological responses of the calcified rhodophyte Corallina officinalis to elevated CO2Botanica Marina, vol. 56 pp. 161

Citation

This review can be cited as:

Stamp, T.E. & Williams, E.., Lloyd, K.A., & Mardle, M.J., 2021. Mixed kelp with foliose red seaweeds, sponges and ascidians on sheltered tide-swept infralittoral rock. In Tyler-Walters H. Marine Life Information Network: Biology and Sensitivity Key Information Reviews, [on-line]. Plymouth: Marine Biological Association of the United Kingdom. [cited 27-01-2023]. Available from: https://www.marlin.ac.uk/habitat/detail/1043

 Download PDF version


Last Updated: 30/03/2021