Mixed Laminaria hyperborea and Laminaria ochroleuca forest on exposed infralittoral rock

Summary

UK and Ireland classification

Description

Mixed Laminaria hyperborea and Laminaria ochroleuca forest on upper infralittoral moderately exposed or sheltered rock is restricted to the coast of Cornwall and the Isles of Scilly. Superficially, the Laminaria ochroleuca biotope looks similar to a moderately exposed Laminaria hyperborea forest (MIR.Lhyp.Ft), containing a similar suite of foliose and filamentous red algae beneath the canopy. Unlike Laminaria hyperborea, however, Laminaria ochroleuca has a smooth stipe so it lacks dense assemblages of epiphytic seaweeds. Laminaria ochroleuca occurs across a wide range of wave exposures (in common with Laminaria hyperborea) and consequently, it occurs at low abundance in other kelp biotopes (sheltered through to exposed) that occur between Dorset to Lundy. In such cases, records should be considered as regional variations of the usual kelp biotopes. Records should only be assigned to this biotope when the canopy is dominated by Laminaria ochroleuca alone, or by a mixture of both Laminaria hyperborea and Laminaria ochroleuca (though the latter is usually in greater abundance). This biotope is similar to the mixed Laminaria hyperborea and Laminaria ochroleuca biotope found on exposed coasts (EIR.LhypR.Loch), though the latter generally occurs in slightly deeper water (often below pure Laminaria hyperborea - EIR.LhypR.Ft) as Laminaria ochroleuca is less tolerant of strong wave action at its northern distributional limit. Both Laminaria ochroleuca biotopes are common on the Brittany and Normandy coasts. (Information taken from JNCC, 2022). 

Depth range

5-10 m, 10-20 m

Additional information

-

Listed By

Sensitivity reviewHow is sensitivity assessed?

Sensitivity characteristics of the habitat and relevant characteristic species

At high densities, Laminaria hyperborea forms a canopy over infralittoral rock. Beneath the canopy an understorey community grows, typically defined by a red seaweed turf although faunal species dominate in tide swept and/or wave surged conditions. Grazing by the urchins Echinus esculentus and Paracentrotus lividus can also define the biotope and reduce the biomass of Laminaria hyperborea and understorey flora. The abundance of Laminaria hyperborea is determined by light availability, which decreases with an increase in water depth. Therefore, depth and water clarity determines the density of Laminaria and hence the distribution of kelp forest (high density kelp) and park (low density kelp) sub-biotopes.

Kelp biotopes are a major source of primary productivity, and support magnified secondary productivity within North Atlantic coastal waters (Smale et al., 2013, Brodie et al., 2014). In Scotland, alone kelp biotopes are estimated to cover 8000km2 (Walker, 1953), and account for ca 45% of primary production in UK coastal waters (Smale et al., 2013). Therefore kelp biotopes, of which Laminaria hyperborea is dominant within UK subtidal rocky reefs (Birkett et al., 1998), make a substantial contribution to coastal primary production in the UK (Smale et al., 2013). Laminaria hyperborea is grazed directly by species such as Patella pellucida, however approximately 80% of primary production is consumed as detritus or dissolved organic material (Krumhansl, 2012) which is both retained within and transported out of the parent kelp forest, providing valuable nutrition to potentially low productivity habitats such as sandy beaches (Smale et al., 2013).

Laminaria hyperborea also acts as an ecosystem engineer (Jones et al., 1996; Smale et al., 2013) by altering; light levels (Sjøtun et al., 2006), physical disturbance (Connell, 2003), sedimentation rates (Eckman et al., 1989) and water flow (Smale et al., 2013), profoundly altering the physical environment for fauna and flora in close proximity. Laminaria hyperborea biotopes increase the three-dimensional complexity of unvegetated rock (Norderhaug, 2004, Norderhaug et al., 2007, Norderhaug & Christie, 2011, Gorman et al., 2012; Smale et al., 2013) and support high local diversity, abundance and biomass of epi/benthic species (Smale et al., 2013), and serve as a nursery ground for a number of commercial important species, e.g. Gadidae (The taxonomic family that contains many commercially important marine fish species, including the Atlantic Cod and Pollack) (Rinde et al., 1992).

In undertaking this assessment of sensitivity, an account is taken of knowledge of the biology of all characterizing species/taxa in the biotope. For this sensitivity assessment Laminaria hyperborea is the primary focus of research, however, it is recognized that the understorey community, typically red seaweeds, also define the biotope. Examples of important species groups are mentioned where appropriate.

Resilience and recovery rates of habitat

A number of review and experimental publications have assessed the recovery of Laminaria hyperborea kelp beds and the associated community. If environmental conditions are favourable 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 6 years (possibly 10 years). These recovery rates were based on discrete kelp harvesting events.  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 2 years of clearance the blocks were dominated by Laminaria hyperborea (Fletcher et al., 2006).

In south Norway, Laminaria hyperborea forests are harvested, which results in large scale removal of the canopy-forming kelps.  Cristie et al., (1998) found that in south Norwegian Laminaria hyperborea beds a pool of small (<25 cm) understorey Laminaria hyperborea plants persist beneath the kelp canopy for several years. The understorey Laminaria hyperborea sporophytes had fully re-established the canopy at a height of 1m within 2-6 years after kelp harvesting. Within 1 year following harvesting, and each successive year thereafter, a pool of Laminaria hyperborea recruits had re-established within the understorey beneath the kelp canopy. Cristie et al., (1998) suggested that Laminaria hyperborea bed re-establishment from understorey recruits (see above) inhibits the colonization of other kelps species and furthers the dominance of Laminaria hyperborea within suitable habitats, stating that Laminaria hyperborea habitats are relatively resilient to disturbance events.

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 ~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).

Laminaria hyperborea biotopes are partially reliant on low (or no) populations of sea urchins, primarily the species; Echinus esculentus, Paracentrotus lividus and Strongylocentrotus droebachiensis, which graze directly on macroalgae, epiphytes and the understorey community.  Multiple authors (Steneck et al., 2002; Steneck et al., 2004; Rinde & Sjøtun, 2005; Norderhaug & Christie, 2009; Smale et al., 2013) have reported dense aggregations of sea urchins to be a principal threat to Laminaria hyperborea biotopes of the North Atlantic. Intense urchin grazing creates expansive areas known as “urchin barrens”, in which a shift can occur from Laminaria hyperborea dominated biotopes to those characterized by coralline encrusting algae, with a resultant reduction in biodiversity (Leinaas & Christie, 1996; Steneck et al., 2002; Norderhaug & Christie, 2009). Continued intensive urchin grazing pressure on Laminaria hyperborea biotopes can inhibit the Laminaria hyperborea recruitment (Sjøtun et al., 2006) and cause urchin barrens to persist for decades (Cristie et al., 1998; Stenneck et al., 2004; Rinde & Sjøtun, 2005). The mechanisms that control sea urchin aggregations are poorly understood but have been attributed to anthropogenic pressure on top down urchin predators (e.g. cod or lobsters). While these theories are largely unproven a number of studies have shown that removal of urchins from grazed areas coincides with kelp re-colonization (Leinaas & Christie, 1996; Norderhaug & Christie, 2009). Leinaas & Christie, (1996) removed Strongylocentrotus droebachiensis from “urchin barrens” and observed a succession effect, in which the substratum was initially colonized by filamentous macroalgae and Saccharina latissima.  However, after 2-4 years Laminaria hyperborea dominated the community.

Reports of large scale urchin barrens within the North East Atlantic are generally limited to regions of the North Norwegian and Russian Coast (Rinde & Sjøtun, 2005, Norderhaug & Christie, 2009). Within the UK, urchin grazed biotopes (IR.MIR.KR.Lhyp.GzFt/Pk, IR.HIR.KFaR.LhypPar, IR.LIR.K.LhypSlat.Gz & IR.LIR.K.Slat.Gz) are generally localised to a few regions in North Scotland and Ireland (Smale et al., 2013; Stenneck et al., 2002; Norderhaug & Christie 2009; Connor et al., 2004). IR.MIR.KR.Lhyp.GzFt/Pk, IR.HIR.KFaR.LhypPar, IR.LIR.K.LhypSlat.Gz & IR.LIR.K.Slat.Gz are characterized by a canopy-forming kelp. However, urchin grazing decreases the abundance and diversity of understorey species. In the Isle of Man. Jones & Kain (1967) observed low Echinus esculentus grazing pressure can control the lower limit of Laminaria hyperborea and remove Laminaria hyperborea sporelings and juveniles. Urchin abundances in “urchin barrens” have been reported as high as 100 individuals/m2 (Lang & Mann, 1978). Kain (1967) reported urchin abundances of 1-4/m2 within experimental plots of the Isle of Man. Therefore, while “urchin barrens” are not presently an issue within the UK, relatively low urchin grazing has been found to control the depth distribution of Laminaria hyperborea, negatively impact on Laminaria hyperborea recruitment and reduce the understorey community abundance and diversity.

Other factors that are likely to influence the recovery of Laminaria hyperborea biotopes is competitive interactions with Invasive Non-Indigenous Species  (INIS), e.g. Undaria pinnatifida (Smale et al., 2013; Brodie et al., 2014; Heiser et al., 2014), and/or the Lusitanian kelp Laminaria ochroleuca (Brodie et al., 2014; Smale et al., 2015). A predicted sea temperature rise in the North and Celtic seas of between 1.5-5°C over the next century (Philippart et al., 2011) is likely to create northward range shifts in many macroalgal species, including Laminaria hyperborea. Laminaria hyperborea is a northern (Boreal) kelp species, thus increases in seawater temperature is likely to affect the resilience and recoverability of Laminaria hyperborea biotopes with southerly distributions in the UK (Smale et al., 2013; Stenneck et al., 2002). Evidence suggests that the Lusitanian kelp Laminaria ochroleuca (Smale et al., 2015), and the INIS Undaria pinnatifida (Heiser et al., 2014) are competing with Laminaria hyperborea along the UK south coast and may displace Laminaria hyperborea from some subtidal rocky reef habitats. The wider ecological consequences of Laminaria hyperborea’ competition with Laminaria ochroleuca and Undaria pinnatifida are however as of yet unknown.

Resilience assessment. The evidence suggests that beds of mature Laminaria hyperborea can regenerate from disturbance within a period of 1-6 years, and the associated community within 7-10 years. However, other factors such as competitive interactions with Laminaria ochroleuca and Undaria pinnatifida may limit recovery of Laminaria hyperborea biotopes following disturbance. Also, urchin grazing pressure is shown to limit Laminaria hyperborea recruitment and reduce the diversity and abundance of the understorey community and may limit habitat recovery following disturbance. The recovery of Laminaria hyperborea biotopes to disturbance from commercial harvesting in south Norway suggests that Laminaria hyperborea beds and the associated community could recover from a significant loss of canopy cover within 10 years, resilience has therefore been assessed as Medium.

Please note* as in Northern Norway urchin grazing pressure could extend recovery/resilience of the Laminaria hyperborea biotopes >25 years, If intensive urchin grazing (as seen in Northern Norway) occurs in the UK resilience would be re-assessed as Very Low. However, because of the limited/localised incidence of urchin grazing within the UK, urchin grazing on large scales (as in Northern Norway) has not been included in this general resilience assessment. The introduction of Invasive Non-Indigenous Species  (INIS) will also inhibit the recovery of Laminaria hyperborea biotopes for an indeterminate amount of time, in these cases, resilience would need to be re-assessed as Very Low. Another factor that is beyond the scope of this sensitivity assessment is the presence of multiple concurrent synergistic or cumulative effects, which Smale et al. (2013) suggest could be a more damaging than the individual pressures.

Hydrological Pressures

Use [show more] / [show less] to open/close text displayed

ResistanceResilienceSensitivity
Temperature increase (local) [Show more]

Temperature increase (local)

Benchmark. A 5°C increase in temperature for one month, or 2°C for one year. Further detail

Evidence

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., 1988). Above 17°C 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., 1998).

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

Laminaria hyperborea has a geographic range from mid-Portugal to Northern Norway (Birket et al., 1998), and a mid range within southern Norway (60°-65° North)(Kain, 1971). The average seawater temperature for southern Norway in October is 12-13°C (Miller et al., 2009), and average annual sea temperature, from 1970 to 2014 is 8°C (Beszczynska-Möller & Dye, 2013). Against the pressure benchmark, the available information suggests that Laminaria hyperborea recruitment processes may be affected and associated red algae communities may decline.

Sensitivity assessment. Overall, a chronic change (2°C for a year) outside the normal range for a year may reduce recruitment and growth, resulting in a minor loss in the population of kelp, especially in winter months or in southern examples of the biotope. However, an acute change (5°C for a month; e.g. from thermal effluent) may result in loss of abundance of kelp or extent of the bed, especially in winter. Therefore, resistance to the pressure is considered 'Medium', and resilience 'Medium'. The sensitivity of this biotope to increases in temperature has been assessed as 'Medium'.

Medium
High
High
High
Help
Medium
High
High
High
Help
Medium
High
High
High
Help
Temperature decrease (local) [Show more]

Temperature decrease (local)

Benchmark. A 5°C decrease in temperature for one month, or 2°C for one year. Further detail

Evidence

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., 1988). Subtidal red algae can survive at temperatures between -2 °C and 18-23 °C (Lüning, 1990; Kain & Norton, 1990).

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

Sensitivity assessment. Resistance to the pressure is considered ‘High’, and resilience ‘High’. The sensitivity of this biotope to decreases in temperature has been assessed as ‘Not Sensitive’.

 

High
High
High
High
Help
High
High
High
High
Help
Not sensitive
High
High
High
Help
Salinity increase (local) [Show more]

Salinity increase (local)

Benchmark. A increase in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

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 (MNCR category- 'Full' salinity) 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.

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

Low
Low
NR
NR
Help
Medium
High
Medium
High
Help
Medium
Low
NR
NR
Help
Salinity decrease (local) [Show more]

Salinity decrease (local)

Benchmark. A decrease in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

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 (MNCR category-Full Salinity) and growth rates are likely to be affected by periodic salinity stress. Birkett et al. (1998b) suggest that long-term changes in salinity may result in loss of affected kelp and, therefore, loss of this biotope.

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. A decrease in one MNCR salinity scale from 'Full' salinity (30-40psu) to 'Reduced' salinity (18-30 psu) would result in a decrease of Laminaria hyperborea sporophyte growth. Laminaria hyperborea may also be out-competed by low salinity tolerant species e.g. Saccharina latissma (Karsten, 2007), or the Invasive Non-Indigenous Species Undaria pinnatifida (Burrows et al., 2014).

If salinity was returned to 'Full' salinity (30-40 psu) Laminaria hyperborea could out-compete Saccharina latissma and re-establish community dominance in 2-4 years (Kain, 1975; Leinaas & Christie, 1996), however full habitat structure may take over 10 years to recover (Birkett et al., 1998; Cristie et al., 1998). The ability of Laminaria hyperborea to out-compete Undaria pinnatifida within the UK is, however, unknown (Heiser et al., 2014), and as such interspecific interaction between Laminaria hyperborea and Undaria pinnatifida is not included within this sensitivity assessment.

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

Low
Medium
Medium
Medium
Help
Medium
High
Medium
High
Help
Medium
Medium
Medium
Medium
Help
Water flow (tidal current) changes (local) [Show more]

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s to 0.2 m/s for more than one year. Further detail

Evidence

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

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

The morphology of the stipe and blade of kelps vary with water flow.  In wave exposed areas, for example, Laminaria hyperborea develops a long and flexible stipe and this is probably a functional adaptation to strong water movement (Sjøtun, 1998). In addition, the lamina becomes narrower and thinner in strong currents (Sjøtun & Fredriksen, 1995). However, the stipe of Laminaria hyperborea is relatively stiff and can snap in strong currents. Laminaria hyperborea is usually absent from areas of high wave action or strong currents, although it is found  in the Menai Strait, Wales, where tidal velocities can exceed 4 m/s (NBN, 2015) and in tidal rapids in Norway (J. Jones, pers. comm.)  Laminaria hyperborea growth can persist in very strong tidal streams (>3 m/s).

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

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

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

 

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

High
High
High
High
Help
High
High
High
High
Help
Not sensitive
High
High
High
Help
Emergence regime changes [Show more]

Emergence regime changes

Benchmark.  1) A change in the time covered or not covered by the sea for a period of ≥1 year or 2) an increase in relative sea level or decrease in high water level for ≥1 year. Further detail

Evidence

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

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

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

Low
Low
NR
NR
Help
Medium
High
Low
High
Help
Medium
Low
NR
NR
Help
Wave exposure changes (local) [Show more]

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year. Further detail

Evidence

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

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

The morphology of the stipe and blade of kelps vary with water flow. In wave exposed areas, for example, Laminaria hyperborea develops a long and flexible stipe and this is probably a functional adaptation to strong water movement (Sjøtun, 1998). In addition, the lamina becomes narrower and thinner in strong currents (Sjøtun & Fredriksen, 1995). However, the stipe of Laminaria hyperborea is relatively stiff and can snap in strong currents. Laminaria hyperborea is usually absent from areas of extreme wave action and can be replaced by Alaria esculenta. In extreme wave exposures, Alaria esculenta can dominate the shallow sublittoral to a depth of 15m (Birkett et al., 1998b).

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

IR.HIR.KFaR.LhypR, IR.HIR.KFaR.LhypFa, IR.MIR.KR.Lhyp, and their associated sub-biotopes are found between extremely exposed to moderate wave exposure. Changes in local wave height above or below that experienced in extremely exposed to moderately exposed sites will affect the dominance of Laminaria hyperborea. Smaller changes in local wave height have the potential to cause changes to the understorey community. The prominent understorey filter feeding community within IR.HIR.KFaR.LhypFa is reliant on wave surge currents. A decrease in wave surge may result in a decline of filter feeding fauna and an increase in red seaweeds within the understorey community or vice versa. A decrease in local wave height may also decrease the chance of urchins being dislodged (removed) from biotopes found at sites with traditionally high wave exposure and may, therefore, increase urchin grazing. An increase in urchin grazing may cause a decline in the understorey community abundance and diversity (as in IR.MIR.KR.Lhyp.GzFt/Pk and IR.MIR.KR.LhypPar).

Sensitivity assessment. A change in nearshore significant wave height >3% but <5% is, however, unlikely to have a significant effect. Resistance to the pressure is considered ‘High’, and resilience ‘High’. Hence, the sensitivity of this biotope to changes in local wave height has been assessed as ‘Not Sensitive’.

High
High
High
High
Help
High
High
High
High
Help
Not sensitive
High
High
High
Help

Chemical Pressures

Use [show more] / [show less] to open/close text displayed

ResistanceResilienceSensitivity
Transition elements & organo-metal contamination [Show more]

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

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 al. (1999) reported that Hg was very toxic to macrophytes. Similarly, Hopkin & Kain (1978) demonstrated sub-lethal effects of heavy metals on Laminaria hyperborea gametophytes and sporophytes, including reduced growth and respiration. Sheppard et al., (1980) noted that increasing levels of heavy metal contamination along the west coast of Britain reduced species number and richness in holdfast fauna, except for suspension feeders which became increasingly dominant. Gastropods may be relatively tolerant of heavy metal pollution (Bryan, 1984). Echinus esculentus recruitment is likely to be impaired by heavy metal contamination due to the intolerance of its larvae. Echinus esculentus are long-lived and poor recruitment may not reduce grazing pressure in the short-term. Although macroalgae species may not be killed, except by high levels of contamination, reduced growth rates may impair the ability of the biotope to recover from other environmental disturbances.

Not Assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Hydrocarbon & PAH contamination [Show more]

Hydrocarbon & PAH contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

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

Laminaria hyperborea fronds, being almost exclusively subtidal, would not come into contact with freshly released oil, but only to sinking emulsified oil and oil adsorbed onto particles (Birket et al., 1998). The mucilaginous slime layer coating of laminarians may protect them from smothering by oil. Hydrocarbons in solution reduce photosynthesis and may be algicidal. However, Holt et al. (1995) reported that oil spills in the USA and from the 'Torrey Canyon' had little effect on kelp forests. Similarly, surveys of subtidal communities at a number sites between 1-22.5m below chart datum, including Laminaria hyperborea communities, showed no noticeable impacts of the Sea Empress oil spill and clean up (Rostron & Bunker, 1997). An assessment of holdfast fauna in Laminaria showed that although species richness and diversity decreased with increasing proximity to the Sea Empress oil spill, overall the holdfasts contained a reasonably rich and diverse fauna, even though oil was present in most samples (Sommerfield & Warwick, 1999). Laboratory studies of the effects of oil and dispersants on several red algae species, including Delesseria sanguinea (Grandy 1984; cited in Holt et al., 1995) concluded that they were all sensitive to oil/ dispersant mixtures, with little differences between adults, sporelings, diploid or haploid life stages. Holt et al., (1995) concluded that Delesseria sanguinea is probably generally sensitive of chemical contamination. Overall the red algae are likely to be highly intolerant to hydrocarbon contamination. Loss of red algae is likely to reduce the species richness and diversity of the biotope and the understorey may become dominated by encrusting corallines; however, red algae are likely to recover relatively quickly.

Not Assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Synthetic compound contamination [Show more]

Synthetic compound contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

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

O'Brian & Dixon (1976) suggested that red algae were the most sensitive group of macrophytes to oil and dispersant contamination (see Smith, 1968). Although Laminaria hyperborea sporelings and gametophytes are intolerant of atrazine (and probably other herbicides) overall they may be relatively tolerant of synthetic chemicals (Holt et al., 1995). Laminaria hyperborea survived within >55m from the acidified halogenated effluent discharge polluting Amlwch Bay, Anglesey, albeit at low density. These specimens were greater than 5 years of age, suggesting that spores and/or early stages were more intolerant (Hoare & Hiscock, 1974). Patella pellucida was excluded from Amlwch Bay by the pollution and the species richness of the holdfast fauna decreased with proximity to the effluent discharge; amphipods were particularly intolerant although polychaetes were the least affected (Hoare & Hiscock, 1974). The richness of epifauna/flora decreased near the source of the effluent and epiphytes were absent from Laminaria hyperborea stipes within Amlwch Bay. The red alga Phyllophora membranifolia was also tolerant of the effluent in Amlwch Bay. Smith (1968) also noted that epiphytic and benthic red algae were intolerant of dispersant or oil contamination due to the Torrey Canyon oil spill; only the epiphytes Crytopleura ramosa and Spermothamnion repens and some tufts of Jania rubens survived together with Osmundea pinnatifida, Gigartina pistillata and Phyllophora crispa from the sublittoral fringe. Delesseria sanguinea was probably to most intolerant since it was damaged at depths of 6m (Smith, 1968). Holt et al., (1995) suggested that Delesseria sanguinea is probably generally sensitive of chemical contamination. Although Laminaria hyperborea may be relatively insensitive to synthetic chemical pollution, evidence suggests that grazing gastropods, amphipods and red algae are sensitive. Loss of red algae is likely to reduce the species richness and diversity of the biotope and the understorey may become dominated by encrusting corallines; however, red algae are likely to recover relatively quickly.

Not Assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Radionuclide contamination [Show more]

Radionuclide contamination

Benchmark. An increase in 10µGy/h above background levels. Further detail

Evidence

No evidence

No evidence (NEv)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
No evidence (NEv)
NR
NR
NR
Help
Introduction of other substances [Show more]

Introduction of other substances

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed.

Not Assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
De-oxygenation [Show more]

De-oxygenation

Benchmark. Exposure to dissolved oxygen concentration of less than or equal to 2 mg/l for one week (a change from WFD poor status to bad status). Further detail

Evidence

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

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

High
High
Medium
High
Help
High
High
High
High
Help
Not sensitive
High
Medium
High
Help
Nutrient enrichment [Show more]

Nutrient enrichment

Benchmark. Compliance with WFD criteria for good status. Further detail

Evidence

Holt et al. (1995) suggest that Laminaria hyperborea may be tolerant of nutrient enrichment since healthy populations are found at ends of sublittoral untreated sewage outfalls in the Isle of Man. Increased nutrient levels e.g. from sewage outfalls, has been associated with increases in abundance, primary biomass and Laminaria hyperborea stipe production but with concomitant decreases in species numbers and diversity (Fletcher, 1996).

Increased nutrients may result in phytoplankton blooms that increase turbidity (see above). Increased nutrients may favour sea urchins, e.g. Echinus esculentus, due their ability to absorb dissolved organics, and result in increased grazing pressure leading to loss of understorey epiflora/fauna, decreased kelp recruitment and possibly 'urchin barrens'. Therefore, although nutrients may not affect kelps directly, indirect effects such as turbidity, siltation and competition may significantly affect the structure of the biotope.

However, this biotope is considered to be 'Not sensitive' at the pressure benchmark, that assumes compliance with good status as defined by the WFD.

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not sensitive
NR
NR
NR
Help
Organic enrichment [Show more]

Organic enrichment

Benchmark. A deposit of 100 gC/m2/yr. Further detail

Evidence

Holt et al. (1995) suggest that Laminaria hyperborea may be tolerant of organic enrichment since healthy populations are found at ends of sublittoral untreated sewage outfalls in the Isle of Man. Increased nutrient levels e.g. from sewage outfalls, has been associated with increases in abundance, primary biomass and Laminaria hyperborea stipe production but with concomitant decreases in species numbers and diversity (Fletcher, 1996). Increase in ephemeral and opportunistic algae are associated with reduced numbers of perennial macrophytes (Fletcher, 1996). Increased nutrients may also result in phytoplankton blooms that increase turbidity. Therefore, although nutrients may not affect kelps directly, indirect effects such as turbidity may significantly affect the structure of Laminaria hyperborea biotopes.

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

Medium
Medium
Medium
Medium
Help
High
High
Medium
High
Help
Low
Medium
Medium
Medium
Help

Physical Pressures

Use [show more] / [show less] to open/close text displayed

ResistanceResilienceSensitivity
Physical loss (to land or freshwater habitat) [Show more]

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site. Further detail

Evidence

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

None
High
High
High
Help
Very Low
High
High
High
Help
High
High
High
High
Help
Physical change (to another seabed type) [Show more]

Physical change (to another seabed type)

Benchmark. Permanent change from sedimentary or soft rock substrata to hard rock or artificial substrata or vice-versa. Further detail

Evidence

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

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

None
High
High
High
Help
Very Low
High
High
High
Help
High
High
High
High
Help
Physical change (to another sediment type) [Show more]

Physical change (to another sediment type)

Benchmark. Permanent change in one Folk class (based on UK SeaMap simplified classification). Further detail

Evidence

Not relevant on bedrock habitats

Not relevant (NR)
Low
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Habitat structure changes - removal of substratum (extraction) [Show more]

Habitat structure changes - removal of substratum (extraction)

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

Evidence

Not Relevant to rock substrata.

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Abrasion / disturbance of the surface of the substratum or seabed [Show more]

Abrasion / disturbance of the surface of the substratum or seabed

Benchmark. Damage to surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

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

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

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

 

Low
High
High
High
Help
Medium
High
High
High
Help
Medium
High
High
High
Help
Penetration or disturbance of the substratum subsurface [Show more]

Penetration or disturbance of the substratum subsurface

Benchmark. Damage to sub-surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

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

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Changes in suspended solids (water clarity) [Show more]

Changes in suspended solids (water clarity)

Benchmark. A change in one rank on the WFD (Water Framework Directive) scale e.g. from clear to intermediate for one year. Further detail

Evidence

Suspended Particle Matter (SPM) concentration has a linear relationship with sub-surface light attenuation (Kd) (Devlin et al., 2008). An increase in SPM results in a decrease in sub-surface light attenuation. Light availability and water turbidity are principal factors in determining kelp depth range (Birkett et al., 1998). Light penetration influences the maximum depth at which kelp species can grow and it has been reported that laminarians grow down to depths at which the light levels are reduced to 1 percent of incident light at the surface. Maximal depth distribution of laminarians, therefore, varies from 100 m in the Mediterranean to only 6-7 m in the silt-laden German Bight. In Atlantic European waters, the depth limit is typically 35 m. In very turbid waters the depth at which Laminaria hyperborea is found may be reduced, or in some cases excluded completely (e.g. Severn Estuary), because of the alteration in light attenuation by suspended sediment (Birkett et al. 1998b; Lüning, 1990).

Laminaria spp. show a decrease of 50% photosynthetic activity when turbidity increases by 0.1/m (light attenuation coefficient =0.1-0.2/m; Staehr & Wernberg, 2009). An increase in water turbidity will likely affect the photosynthetic ability of Laminaria hyperborea and Laminaria ochroleuca and decrease Laminaria hyperborea abundance and density (see sub-biotope- IR.MIR.KR.Lhyp.Pk). Kain (1964) suggested that early Laminaria hyperborea gametophyte development could occur in the absence of light. Furthermore, observations from south Norway found that a pool of Laminaria hyperborea recruits could persist growing beneath Laminaria hyperborea canopies for several years, indicating that sporophyte growth can occur in light-limited environments (Christe et al., 1998). However in habitats exposed to high levels of suspended silts Laminaria hyperborea is out-competed by Saccharina latissima, a silt tolerant species, and thus, a decrease in water clarity is likely to decrease the abundance of Laminaria hyperborea in the affected area (Norton, 1978).

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

Sensitivity Assessment. An increase in water clarity from clear to intermediate (10-100mg/l) represent a change in light attenuation of ca 0.67-6.7 Kd/m, and is likely to result in a greater than 50% reduction in photosynthesis of Laminaria spp. Therefore, the dominant kelp species will probably suffer a significant decline and resistance to this pressure is assessed as ‘Low’. Resilience to this pressure is probably ‘Medium’ at the benchmark.  Hence, this biotope is assessed as having a sensitivity of ‘Medium ‘to this pressure.

Low
High
High
High
Help
Medium
High
High
High
Help
Medium
High
High
High
Help
Smothering and siltation rate changes (light) [Show more]

Smothering and siltation rate changes (light)

Benchmark. ‘Light’ deposition of up to 5 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

Smothering by sediment (e.g. 5 cm of material) during a discrete event is unlikely to damage Laminaria hyperborea sporophytes but is likely to affect gametophyte survival as well as holdfast fauna, and interfere with zoospore settlement.  Given the microscopic size of the gametophyte, 5 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 one month (Dieck, 1993). Intolerance to this factor is likely to be higher during the peak periods of sporulation and/or spore settlement.

If inundation is long lasting then the understorey epifauna/flora may be adversely affected, e.g. suspension or filter feeding fauna and/or algal species.  This biotope occurs in high wave exposures and, therefore, deposited sediments are unlikely to remain for more than a few tidal cycles, except in the deepest of rock-pools. Therefore, the effects of depositing 5 cm of fine sediment in a discrete event are likely to be transient.

Sensitivity assessment. Resistance to the pressure is considered ‘High’, and resilience ‘High’. The sensitivity of this biotope to light deposition of up to 5 cm of fine material added to the seabed in a single discreet event is assessed as ‘Not sensitive’.

High
Low
NR
NR
Help
High
High
High
High
Help
Not sensitive
Low
NR
NR
Help
Smothering and siltation rate changes (heavy) [Show more]

Smothering and siltation rate changes (heavy)

Benchmark. ‘Heavy’ deposition of up to 30 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

Smothering by sediment (e.g. 30 cm of material) during a discrete event is unlikely to damage Laminaria hyperborea plants but is likely to affect gametophyte survival, holdfast communities, epiphytic community at the base of the stipe, and interfere with zoospore settlement. Given the microscopic size of the gametophyte, 30 cm of sediment could be expected to significantly inhibit growth. However, laboratory studies showed that gametophytes can survive in darkness for between 6 - 16 months at 8 °C and would probably survive smothering within a discrete event. Once returned to normal conditions the gametophytes resumed growth or maturation within 1 month (Dieck, 1993). Intolerance to this factor is likely to be higher during the peak periods of sporulation and/or spore settlement.

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

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

Medium
Low
NR
NR
Help
High
Low
NR
NR
Help
Low
Low
NR
NR
Help
Litter [Show more]

Litter

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

Evidence

Not assessed.

Not Assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Not assessed (NA)
NR
NR
NR
Help
Electromagnetic changes [Show more]

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT. Further detail

Evidence

No evidence

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
No evidence (NEv)
NR
NR
NR
Help
Underwater noise changes [Show more]

Underwater noise changes

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

Evidence

Not relevant

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Introduction of light or shading [Show more]

Introduction of light or shading

Benchmark. A change in incident light via anthropogenic means. Further detail

Evidence

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.

Low
Low
NR
NR
Help
Medium
Low
NR
NR
Help
Medium
Low
NR
NR
Help
Barrier to species movement [Show more]

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion. Further detail

Evidence

Not relevant. This pressure is considered applicable to mobile species, e.g. fish and marine mammals rather than seabed habitats. Physical and hydrographic barriers may limit the dispersal of spores.  But spore dispersal is not considered under the pressure definition and benchmark.

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Death or injury by collision [Show more]

Death or injury by collision

Benchmark. Injury or mortality from collisions of biota with both static or moving structures due to 0.1% of tidal volume on an average tide, passing through an artificial structure. Further detail

Evidence

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

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Visual disturbance [Show more]

Visual disturbance

Benchmark. The daily duration of transient visual cues exceeds 10% of the period of site occupancy by the feature. Further detail

Evidence

Not relevant

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help

Biological Pressures

Use [show more] / [show less] to open/close text displayed

ResistanceResilienceSensitivity
Genetic modification & translocation of indigenous species [Show more]

Genetic modification & translocation of indigenous species

Benchmark. Translocation of indigenous species or the introduction of genetically modified or genetically different populations of indigenous species that may result in changes in the genetic structure of local populations, hybridization, or change in community structure. Further detail

Evidence

No evidence regarding the genetic modification or effects of translocation of native kelp populations was found.

Not relevant (NR)
NR
NR
NR
Help
Not relevant (NR)
NR
NR
NR
Help
No evidence (NEv)
NR
NR
NR
Help
Introduction or spread of invasive non-indigenous species [Show more]

Introduction or spread of invasive non-indigenous species

Benchmark. The introduction of one or more invasive non-indigenous species (INIS). Further detail

Evidence

Competition with invasive macroalgae may be a potential threat to native kelp biotopes. Potential invasives include Undaria pinnatifida, Sargassum muticum and Codium fragile (de Bettignies et al., 2021). In Nova Scotia, Codium fragile 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 prevents re-colonization by other macroalgae (Scheibling et al., 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 in the clear waters of California. However, it is a poor competitor under low light and only develops dense canopies in shallow areas (Engelen et al., 2015).

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

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 completion to light, rather than changes in nutrient availability, sedimentation or water flow (Britton-Simmons, 2004; Engelen et al., 2015). Cosson (1999) reported a significant decline in Laminaria digitata at two sites between 1983 and 1997 on the coast of Normandy, France, due to an increase in Sargassum muticum abundance in the same areas. For example, on the Grandcamp rocks, Laminaria digitata has almost disappeared while Sargassum muticum had covered 80% of the lower intertidal and subtidal zone in summer. However, no evidence was found on the effects of Sargassum muticum in native Laminaria hyperborea beds. 

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) and has since proliferated along UK coastlines. One year after its discovery at the Queen Anne Battery marina, Plymouth, it had become a major fouling plant on pontoons (Minchin & Nunn, 2014). Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound.

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

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

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

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

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

In Plymouth Sound, UK, Heiser et al. (2014) observed that both Laminaria ochroleuca and Laminaria hyperborea were significantly less abundant at sites with the presence Undaria pinnatifida. Only ca 2.5 Laminaria ochroleuca individuals per m2 were present compared to ca 5.5 individuals per m2 at sites without Undaria pinnatifida and ca 0.5 Laminaria hyperborea individuals per m2 were present compared to ca 8 individuals per m2 at sites without Undaria pinnatifida. However, the results from their correlation study only showed that the species were not found together (pers. comm., Epstein 2021). Exclusion and succession experiments on reefs tell us that Laminaria spp. exclude Undaria pinnatifida, not the other way round. Epstein & Smale (2018) reported that in Devon, UK, persistent, dense, and intact Laminaria spp. canopies in rocky reef habitats exerted a strong influence over the presence/absence, abundance, and percentage cover of Undaria pinnatifida. A dense canopy of native kelp restricted the proliferation of Undaria pinnatifida and disturbance of the canopy is often the key to the recruitment of Undaria pinnatifida. Epstein et al. (2019b) reported that Undaria pinnatifida density and biomass were significantly negatively correlated with the sum of all Laminaria spp in Plymouth, UK. The evidence indicated that native Laminaria spp. canopies in the UK inhibited Undaria pinnatifida and implied that Undaria pinnatifida was opportunistic but competitively inferior (Farrell & Fletcher, 2006; Heiser et al., 2014; Minchin & Nunn, 2014; De Leij et al., 2017; Epstein & Smale, 2018; Epstein et al., 2019b).

Epstein et al. (2019b) found that Laminaria ochroleuca had a strong negative correlation with Undaria pinnatifida. However, Epstein et al. (2019b) also noted that Laminaria hyperborea had a non-significant positive relationship with Undaria pinnatifida due to low densities of Laminaria hyperborea across the study area and resultant lack of data.

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

Sensitivity Assessment. The above evidence suggests that Undaria pinnatifida can co-exist with Laminaria ochroleuca 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 of 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 dominated biotope (IR.HIR.KFaR.LhypR.Loch) is found within the infralittoral zone with moderately strong to weak tidal streams and is exposed to wave action. The evidence above suggests that 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 between 1 and 4 m (Staehr et al., 2000; Engelen et al., 2015). However, no evidence of the effects of Sargassum on Laminaria hyperborea beds was found. Therefore, competition with Sargassum is probably site-specific and dependent on local conditions, so it is highly unlikely that Sargassum will be able to colonize and survive due to the depth and high degree of wave exposure that characterizes this biotope.  

Undaria pinnatifida has the potential to colonize and co-exist in refugia within Laminaria sp. dominated habitats that are within its shallow depth range (+ 1 to – 4 m) and sheltered from wave action. However, like Sargassum muticum it is highly unlikely that Undaria pinnatifida will be able to colonize or survive at the depth and the of degree wave exposure that characterizes this biotope. Therefore, resistance to Sargassum or Undaria is assessed as ‘High’, resilience as 'High', and sensitivity is assessed as ‘Not Sensitive’. Overall, confidence is assessed as ‘Low’ due to evidence of variation and the site-specific nature of competition between native kelps, Undaria pinnatifida and Sargassum. 

High
Low
NR
NR
Help
High
High
High
High
Help
Not sensitive
Low
NR
NR
Help
Introduction of microbial pathogens [Show more]

Introduction of microbial pathogens

Benchmark. The introduction of relevant microbial pathogens or metazoan disease vectors to an area where they are currently not present (e.g. Martelia refringens and Bonamia, Avian influenza virus, viral Haemorrhagic Septicaemia virus). Further detail

Evidence

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

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

Medium
Low
NR
NR
Help
High
High
Low
High
Help
Low
Low
NR
NR
Help
Removal of target species [Show more]

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

Christie et al. (1998) observed Laminaria hyperborea habitat regeneration following commercial Laminaria hyperborea trawling in south Norway. Within the study area trawling removed all large canopy-forming adult Laminaria hyperborea, however, sub-canopy recruits were unaffected. Within 2-3 years of harvesting, a new canopy had formed 1m off the seabed. The associated holdfast communities recovered in 6 years however the epiphytic stipe community did not fully recover within the same time period. Christie et al. (1998) suggested that kelp habitats were relatively resistant to direct disturbance of Laminaria hyperborea canopy.

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

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

Sensitivity assessment. Resistance to the pressure is considered ‘None’, and resilience ‘Medium’. The sensitivity of this biotope to damage to seabed surface features is assessed as ‘Medium’.

None
High
High
High
Help
Medium
High
High
High
Help
Medium
High
High
High
Help
Removal of non-target species [Show more]

Removal of non-target species

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

Evidence

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

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

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

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

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

Low
High
High
High
Help
Medium
High
High
High
Help
Medium
High
High
High
Help

Bibliography

  1. 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

  2. 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

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

  4. 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

  5. Bishop, G.M., 1985. Aspects of the reproductive ecology of the sea urchin Echinus esculentus L. Ph.D. thesis, University of Exeter, UK.

  6. Boney, A.D., 1971. Sub-lethal effects of mercury on marine algae. Marine Pollution Bulletin, 2, 69-71.

  7. 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

  8. 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

  9. 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

  10. 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.

  11. 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.

  12. 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.

  13. 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

  14. 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.

  15. 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

  16. 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.

  17. Crisp, D.J. (ed.), 1964. The effects of the severe winter of 1962-63 on marine life in Britain. Journal of Animal Ecology, 33, 165-210.

  18. 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.

  19. 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.

  20. 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.

  21. 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

  22. 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

  23. 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.

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

  25. 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.

  26. Dinnel, P.A., Pagano, G.G., & Oshido, P.S., 1988. A sea urchin test system for marine environmental monitoring. In Echinoderm Biology. Proceedings of the Sixth International Echinoderm Conference, Victoria, 23-28 August 1987, (R.D. Burke, P.V. Mladenov, P. Lambert, Parsley, R.L. ed.), pp 611-619. Rotterdam: A.A. Balkema.

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

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

  29. 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.

  30. 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

  31. 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

  32. 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

  33. 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

  34. 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.

  35. 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.

  36. 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.

  37. 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].

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

  39. 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.

  40. 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.

  41. 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.

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

  43. 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.

  44. 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.

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

  46. 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.

  47. 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)].

  48. Hoare, R. & Hiscock, K., 1974. An ecological survey of the rocky coast adjacent to the effluent of a bromine extraction plant. Estuarine and Coastal Marine Science, 2 (4), 329-348.

  49. 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.

  50. 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.

  51. 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/

  52. 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

  53. Johansson ,G., Eriksson, B.K., Pedersen, M. & Snoeijs, P., 1998. Long term changes of macroalgal vegetation in the Skagerrak area. Hydrobiologia, 385, 121-138.

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

  55. 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.

  56. 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://ukmpa.marinebiodiversity.org/uk_sacs/pdfs/marine-habitats-review.pdf

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

  58. 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.

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

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

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

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

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

  64. 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.

  65. 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.

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

  67. Kinne, O. (ed.), 1984. Marine Ecology: A Comprehensive, Integrated Treatise on Life in Oceans and Coastal Waters.Vol. V. Ocean Management Part 3: Pollution and Protection of the Seas - Radioactive Materials, Heavy Metals and Oil. Chichester: John Wiley & Sons.

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

  69. 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

  70. 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

  71. 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.

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

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

  74. 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

  75. 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.

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

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

  78. 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

  79. 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.

  80. 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.

  81. 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

  82. 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.

  83. 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.

  84. 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.

  85. 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.

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

  87. Nichols, D., 1981. The Cornish Sea-urchin Fishery. Cornish Studies, 9, 5-18.

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

  89. 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.

  90. 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.

  91. 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.

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

  93. 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.

  94. O'Brien, P.J. & Dixon, P.S., 1976. Effects of oils and oil components on algae: a review. British Phycological Journal, 11, 115-142.

  95. 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.

  96. 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

  97. 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.

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

  99. Rietema, H., 1993. Ecotypic differences between Baltic and North Sea populations of Delesseria sanguinea and Membranoptera alata. Botanica Marina, 36, 15-21.

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

  101. 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.

  102. 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.

  103. Sheppard, C.R.C. & Bellamy, D.J., 1974. Pollution of the Mediterranean around Naples. Marine Pollution Bulletin, 5, 42-44.

  104. 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.

  105. 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.

  106. 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.

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

  108. 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.

  109. 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.

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

  111. 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.

  112. 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

  113. 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.

  114. 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.

  115. Stock, J.H., 1988. Lamippidae (Copepoda : Siphonostomatoida) parasitic in Alcyonium. Journal of the Marine Biological Association of the United Kingdom, 68 (2), 351-359.

  116. 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.

  117. 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

  118. 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.

  119. 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

  120. 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].

  121. 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.

  122. 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.

  123. 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.

  124. 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

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

  126. 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.

  127. Wilkinson, M., 1995. Information review on the impact of kelp harvesting. Scottish Natural Heritage Review, no. 34, 54 pp.

  128. 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.

Citation

This review can be cited as:

Stamp, T.E., Tyler-Walters, H.,, Burdett, E.G. & Lloyd, K.A., 2023. Mixed Laminaria hyperborea and Laminaria ochroleuca forest on exposed 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 23-04-2024]. Available from: https://www.marlin.ac.uk/habitat/detail/28

 Download PDF version


Last Updated: 26/10/2023