|Researched by||Dr Heidi Tillin||Refereed by||Admin|
Shallow and relatively smal rockpools throughout the eulittoral zone on very exposed to exposed shores, characterized by a covering of encrusting coralline algae on which Corallina officinalis forms a dense turf. The bottom of these pools can be covered in coarse gravel and cobbles. In south and west Ireland these coralline pools may be dominated by the sea urchin Paracentrotus lividus and the seaweed diversity is generally low due to the grazing pressure of Paracentrotus lividus, the top shells Gibbula cineraria and Gibbula umbilicalis, and winkles such as Littorina littorea. Within the pools, pits and crevices are often occupied by the anemone such as Actinia equina and Anemonia viridis and small individuals of the mussel Mytilus edulis. The siphonous green seaweed Codium spp. can also be present along with the wrack Himanthalia elongata and the brown seaweed Leathesia difformis and the filamentous red seaweed Ceramium spp. The barnacle Semibalanus balanoides is either absent or occurs at low abundance in these rockpools, presumably due to the grazing pressure on the larval stage and the predation pressure from the whelk Nucella lapillus. Soft bedrock, such as limestone, allows P.aracentrous lividus to bore into the rock (JNCC, 2015)
The biotope description and characterizing species are taken from Connor et al. (2004). The biotope describes shallow and relatively small rockpools throughout the eulittoral zone on very exposed to exposed shores, characterized by a covering of encrusting coralline algae on which Corallina officinalis forms a dense turf. The coralline crusts and turfs of Corallina officinalis are considered the key characterizing and structuring species within the biotope and the sensitivity assessments focus on these species (although information on coralline crusts is limited). The bottom of these pools can be covered in coarse gravel and cobbles. In south and west Ireland these coralline pools may be dominated by the sea urchin Paracentrotus lividus and the seaweed diversity is generally low due to the grazing pressure of Paracentrotus lividus, the top shells Gibbula cineraria and Gibbula umbilicalis, and winkles such as Littorina littorea. Grazing is a key factor structuring the biotope and the loss of grazers would alter the algal assemblage within the pools (Robles, 1982,; Albrecht, 1998; Palacín et al., 1998; Sala et al., 1998). The sensitivity of grazers is considered within the sensitivity assessments with particular emphasis on littorinids and Paracentrotus lividus. Within the pools, pits and crevices are often occupied by the anemone such as Actinia equina and Anemonia viridis and small individuals of the mussel Mytilus edulis. The siphonous green seaweed Codium spp. can also be present along with the wrack Himanthalia elongata and the brown seaweed Leathesia difformis and the filamentous red seaweed Ceramium spp. These common rocky shore species while contributing to species diversity and ecological function within the biotope are not considered to be key to the biotope and are therefore not specifically considered within the assessments. Soft bedrock, such as limestone, allows Paracentrotus lividus to bore into the rock and is therefore a key component of the biotope and is assessed where the pressure may lead to alterations in the substratum.
Where this biotope is impacted by pressures, recovery of the key structuring and characterizing species Corallina officinalis will require either regrowth from surviving holdfast or basal crusts or recolonization by propagules. The crustose holdfast or base is perennial and grows apically (continuous growth at tips), similar to encrusting corallines such as Lithothamnia sp. The basal crust may grow continuously until stimulated to produce fronds (Littler & Kauker 1984; Colhart & Johanssen 1973). Littler & Kauker (1984) suggest that the crustose bases are an adaptation to resist grazing and desiccation whereas the fronds are adapted for higher primary productivity and reproduction. The basal crusts are tougher than the upright fronds (requiring a pressure of 94 g/mm2 to penetrate, compared to 43 g/mm2 respectively). Regeneration of the basal crusts provides a more rapid route to recovery than recolonization. Experiments in the intertidal in southern California found that areas scraped back to crusts recovered four times more rapidly than sterilised plots where the crusts were removed (Littler & Kauker, 1994).
In culture Corallina officinalis fronds exhibited an average growth rate of 2.2 mm/month at 12 and 18 oC. Growth rate was only 0.2 mm/month at 6 oC and no growth was observed at 25 oC (Colhart & Johanssen 1973). Similarly Blake & Maggs (2003) observed much higher growth rates of 2 mm/month over 6 months starting from September in Corralina officinalis grown in Strangford Lough (Northern Ireland) at 5 and 10m depth, these rates are similar to those observed by Andrake & Johansen, (1980) in winter in New Hampshire. The evidence for growth rate suggests that to achieve a height of 10 cm the turf would be at least 4 years old (probably older as higher temperatures appear to slow growth. A lower level turf of, for example 5 cm, could theoretically be achieved within 2 years.
Where the bases are removed, recovery will depend on recolonization. Areas that are cleared during the reproductive period have the potential to be rapidly colonized. Corallina officinalis was shown to settle on artificial substances within 1 week of their placement in the intertidal in New England summer (Harlin & Lindbergh, 1977). However, settlement plates laid out in the autumn were not recolonized until the next spring. In the lower rocky intertidal in southern California dominated by Corallina officinalis with foliose overstorey algae present, Littler & Kauker (1984) experimentally cleared plots and followed the recovery for 12 months. Some areas were scraped allowing the basal crusts to remain whereas others were completely sterilised (removal of all material and surfaces then scorched with a blow torch to remove bases). In scraped plots up to 15% cover of Corallina officinalis fronds returned within 3 months after removal of fronds and all other epiflora/fauna (Littler & Kauker, 1984) while in sterilized plots (all basal crusts removed) appearance of articulated fronds occurred 6 months following clearance. At the end of the 12 month observation period Corallina officinalis cover had increased to approximately 18% in plots where basal crusts remained and to approximately 10% in sterilised plots. Similarly Bamber & Irving (1993) reported that new plants grew back in scraped transects within 12 months, although the resistant crustose bases were probably not removed.
Once established turfs of Corallina spp. can persist for a long time, surveys of rocky intertidal ledges at Hinkley point, Somerset in England have found that the patches mapped in the 1980s (Bamer & Irving, 1993) had not changed position when resurveyed 18 years later (Burdon et al., 2009). It has been speculated but not definitively demonstrated that turf-forming algae and canopy forming algae may represent alternate stable states on temperate rocky shores and a shift in balance to the alternate state may prevent recovery. Some potential mechnaisms for inhibition of canopy forming species are space pre-emption by turfs that prevent recruitment of taller algae (Perkol-Finkel & Airoldi, 2010, Kennellyy 1987) due to the coverage of suitable rock surfaces and the presence of sediments within the turf (Airoldi, 2003). Clearance experiments on rocky, intertidal shores in Southern California (Sousa, 1979) found that Ulva species which have a longer reproductive season could colonize cleared areas preventing the establishment of perennial red algae. However grazing by crabs removed the green algae (Sousa, 1979), highlighting the potential importance of grazers, particularly littorinids, to the reestablishment of this biotope.
Although ubiquitous in marine coastal systems little is understood about the taxonomy, biology and ecology of this taxa (Littler & Littler, 2013). Coralline crust is a generic term that in UK biotopes refers to nongeniculate (crustose) species from the family Corallinacea that could include Lithophyllum incrustans which is noted to form thick crusts in tidepools, especially in the south west (Adey & Adey, 1973), Lithothamnion spp. and Phymatolithon spp. Throughout the sensitivity assessments the term coralline crust is used to refer to the Corallinacea within the biotope. Due to the lack of evidence for species the assessments are generic, although species specific information is presented where available.
A number of papers by Edyvean & Ford (1984a & b; 1986;1987) describe aspects of reproduction and growth of encrusting coralline, Lithophyllum incrustans. Studies by Edyvean & Forde (1987) in populations of Lithophyllum incrustans in Pembroke south-west Wales suggest that reproduction occurs on average early in the third year. Reproduction may be sexual or asexual. Populations release spores throughout the year but abundance varies seasonally, with the populations studied in Cullercoats Bay, and Lannacombe Bay (North East and South West England, respectively) producing less spores in the summer. Spore release is initiated by changes in temperature or salinity (see relevant pressure information) at low tide so that spore dispersal is restricted to within the tide pool enhancing local recruitment. Spore survival is extremely low with only a tiny proportion of spores eventually recruiting to the adult population (Edyvean & Ford, 1986). The spores are released from structures on the surface called conceptacles, these are formed annually and subsequently buried by the new layer of growth. Plants can be aged by counting the number of layers of conceptacles. Edyvean & Ford (1984a) found that the age structure of populations sampled from Orkney (Scotland) Berwick (northern England) and Devon (England) were similar, mortality seemed highest in younger year classes with surviving individuals after the age of 10 years appear relatively long-lived (up to 30 years). In St Mary’s Northumberland, the population was dominated by the age 6-7 year classes (Edyvean & Ford, 1984a). Growth rates were highest in young plants measured at Pembroke (south-west Wales) with an approximate increase in diameter of plants of 24 mm in year class 0 and 155 mm in year 1 and slowing towards an annual average horizontal growth rate of 3mm/year (Edyvean & Ford, 1987).
Some repair of damaged encrusting coralline occurs through vegetative growth. Chamberlain (1996) observed that although Lithophyllum incrustans was quickly affected by oil during the Sea Empress spill, recovery occurred within about a year. The oil was found to have destroyed about one third of the thallus thickness but regeneration occurred from thallus filaments below the damaged area. Recolonization by propagules is an important recovery mechanism, Airoldi (2000) observed that encrusting coralline algae recruited rapidly on to experimentally cleared subtidal rock surfaces in the Mediterranean Sea, reaching up to 68% cover in 2 months. As encrusting corallines are sensitive to desiccation (Dethier, 1994) it should be noted that these subtidal habitats are probably more favourable for recruitment, growth and survival than intertidal rock pools.
In some areas the urchin Paracentrotus lividus is a key characterizing species of this biotope (Connor et al., 2004). In rock pools generally low mortalities from predators are observed in the first four years (Crapp & Willis, 1975) although mortality increases as the individual ages, the maximum longevity may exceed 13 years (Delmas, 1992). In some instances studied populations have been relatively stable but significant annual variations in populations are more usual (Boudouresque & Verlaque, 2001). For example, over 40 years of monitoring populations in Loch Hyne have varied by four orders of magnitude, mostly in the absence of fishing (Andrew et al., 2002 and references therein). Adult urchins are mobile and during the daily activity peak may travel 40 cm/hour, although movement may be limited in areas with high water movements and in very exposed conditions adults may remain in burrows (Boudoresque & Verlaque, 2001). Migration of adults may therefore provide a mechanism of recovery for impacted populations. In areas where Paracentrotus lividus occurred in high densities and was experimentally removed, migration by foraging adults had led to a recovery in 2 months (Kempf, 1962), however, in low density populations very little migration was observed (Palacin et al., 1997).
The reproductive cycle varies between localities but typically there are one or two seasonal spawning peaks (although low-level larval production may occur year round) these are influenced by a number of factors including food supply, temperature and day length (Boudouresque & Valerque, 2001 and references therein). The planktonic lifespan of the larvae is of 20–40 days (Pedrotti, 1993), allowing this species to disperse over great distances. Despite annual supply of larvae, successful recruitment may be episodic and populations may show missing size cohorts, or those that are under or over-represented (Delmas, 1992). Episodic recruitment accounts for some of the population variations observed (Boudouresque & Verlaque, 2001).
Resilience assessment. No direct evidence was found for the age of individual Corallina crusts, longevity of turfs or the time to recover from basal crusts or sterilised plots to a full dense cover. New crustose bases may recruit and develop quickly but the formation of new fronds from these bases and recovery of original cover may take longer. Once a coralline turf has developed it will probably be colonized by epiphytic invertebrates such as harpacticoids, amphipods and isopods relatively quickly from the surrounding area. Therefore, the biotope would be recognizable once the coralline turf has regrown, which is likely to be quite rapid if the resistant crustose bases remain. The clearance experiments by Littler & Kauker (1984) suggest that recovery of a dense turf cover whether basal crusts remained or were totally removed would require more than 2 years. Presumably, as crusts can grow in all directions percentage cover is not a linear function and that gap closure would speed up with greater cover.
Recruitment of associated species of red algae is probably equally rapid, and once the algal turf has developed most of the epiphytic invertebrates would colonize quickly. The ephemeral green algae associated with the biotope are opportunist colonizers of gaps and would be expected to recover within a year. Littorinds and Paracentrotus lividus could recover through migration but where populations are removed over a larger area recolonization by larvae would be required. Recruitment through larvae may be episodic and recovery to former population structure may require > 2 years. More detailed information on recovery of associated species can be found in the information for biotopes where these are the key characterizing species. Resilience of the biotope is assessed as ‘High’ where resistance is ‘High’ (no significant impact) or ‘Medium’ (where <25 % of Paracentrotus lividus and Corallina officinalis fronds are removed and bases remain) based on regrowth from the basal crusts and vegetative growth from surrounding turfs and repair, and migration or recolonization of associated species. Where resistance is ‘Low’ or ‘None’ then resilience is assessed as ‘Medium’, between 2 -10 years but towards the lower end of that range for Corallina officinalis and Paracentrotus lividus and the associated species. Where perturbations have a large spatial footprint with widespread removal of crusts and grazers over a large area then the development of an alternate state emerging with dominance by canopy forming algae is a possibility. In such an instance recovery could take much longer and depend on active management or further perturbations. No evidence was found however to determine when such shifts might occur.
NB: The resilience and the ability to recover from human induced pressures is a combination of the environmental conditions of the site, the frequency (repeated disturbances versus a one off event) and the intensity of the disturbance. Recovery of impacted populations will always be mediated by stochastic events and processes acting over different scales including, but not limited to, local habitat conditions, further impacts and processes such as larval-supply and recruitment between populations. Full recovery is defined as the return to the state of the habitat that existed prior to impact. This does not necessarily mean that every component species has returned to its prior condition, abundance or extent but that the relevant functional components are present and the habitat is structurally and functionally recognizable as the initial habitat of interest. It should be noted that the recovery rates are only indicative of the recovery potential.
Species found in the intertidal are exposed to extremes of high and low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period of time during the tidal cycle. In winter, air temperatures are colder than the sea, conversely in summer air temperatures are much warmer than the sea. Species that occur in this intertidal biotope are therefore generally adapted to tolerate a range of temperatures, although the timing of site-specific factors such as low tides will influence local acclimation.
The key characterizing species, Corallina officinalis has a cosmopolitan distribution (Guiry & Guiry, 2015) and throughout its range experiences wide variation in temperatures (although local populations may be acclimated to the prevailing thermal regime). Littler & Kauker (1984) suggested that the crustose bases of Corallina officinalis are more resistant of desiccation or heating than fronds. Severe damage was noted in Corallina officinalis fronds as a result of desiccation during unusually hot and sunny weather in summer 1983. An abrupt increase in temperature of 10°C caused by the hot, dry 'Santa Anna' winds (between January -and February) in Santa Cruz, California resulted in die back of several species of algae exposed at low tide (Seapy & Littler, 1982). Lüning (1990) reported that Corallina officinalis from Helgoland survived one week exposure to temperatures between 0°C and 28°C. Encrusting corallines are also considered to have a high tolerance for increases in temperature at the pressure benchmark as these are widespread species within the intertidal. Latham (2008) investigated the effects of temperature stress on Corallina officinalis through laboratory tests on samples collected in the Autumn in Devon, England from rock pools. Samples were kept at 15°C for three days and then exposed to temperatures of 5, 15, 20, 25 and 30°C (the normal range of temperature experienced was suggested to be between 5 and 15°C). At 35°C the Corallina was completely bleached after 3 days with a sample kept at 30°C beginning to bleach. After 7 days (the end of the experiment) the sample kept at 30°C was partially bleached. Samples kept at 5, 15, 20 and 25°C showed little change in chemicals produced in reaction to thermal stress and no bleaching suggesting the temperatures in that range had not induced stress reactions.
In an exceptionally hot summer (1983, with an increase of between 4.8 and 8.5 °C) Hawkins & Hartnoll (1985) observed no temperature bleaching of adult Himanthalia elongata (although some buttons were bleached) or other canopy forming species. However, understorey red algae showed more signs of damage with bleached Corallina officinalis and ‘lithothamnia’ observed around the edges of pools due to dessication. Occasional damaged specimens of Palmaria palmata, Osmundea pinnatifida and Mastocarpus stellatus were observed.
Boudouresque & Verlaque, (2001), reviewed the ecology of Paracentrotus lividus. They suggest that the species is particularly common in regions where winter water temperatures range from 10- 15°C and summer temperatures from 18-25°C. Its northern limits correspond to 8°C winter and 28°C summer isotherms. In the English Channel, lower and upper lethal temperatures are 4°C to 29°C. Spawning may be initiated by temperature changes and seawater temperatures 13°C- 16°C may initiate spawning, where there are two spawning periods the first would occur as the temperature rises to a critical level and the second when it falls to that level although length of daylight may also be a cue (Boudouresque & Verlaque, 2001 and references therein).
Littorinids are also key grazers within this biotope biotope. Laboratory studies suggest that adults of these species can tolerate temperature increases. The median upper lethal temperature limit in laboratory tests on Littorina littorea, Littorina saxatilis was approximately 35°C (Davenport & Davenport, 2005). The smaller species associated with the Corallina officinalis may be protected within fronds and accumulated sediments from changes in temperature although no direct evidence was found to assess sensitivity of these to increased temperatures.
Sensitivity assessment. Based on the global distribution of Corallina officinalis and the experiments by Latham (2008) which approximate to the pressure benchmark more than the observations of extreme events (Seapy & Littler, 1982, Hawkins & Hartnoll, 1985) it is suggested that Corallina officinalis would not be sensitive to either an acute or chronic increase in temperature at the pressure benchmark. Littler & Littler (1984) suggest that the basal crustose stage is adaptive as resisters of sand scour and wave shearing as well as physiological stressors such as desiccation and heating. Where these survive any increases in temperature above the pressure benchmark they would provide a mechanism for biotope recovery. The key grazing species Paracentrotus lividus and the littorinids have a broad temperature tolerance. The sensitivity of the biotope is based on the key characterizing Corallina turf and Paracentrotus lividus and littorinid grazers, but it should be noted that many of the associated species are considered to have ‘High’ resistance to changes in temperature at the pressure benchmark. Resistance of the biotope to both acute and chronic increases in temperature is considered to be 'High' and resilience is 'High' (by default) and the biotope is considered to be 'Not sensitive'. It should be noted that the timing of acute increases would alter the degree of impact and hence sensitivity. An acute change occurring on the hottest day of the year and exceeding thermal tolerances would lead to mortality.
Many intertidal species are tolerant of freezing conditions as they are exposed to extremes of low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period of time during the tidal cycle. In winter air temperatures are colder than the sea, conversely in summer, air temperatures are much warmer than the sea. Species that occur in the intertidal are therefore generally adapted to tolerate a range of temperatures, with the width of the thermal niche positively correlated with the height of the shore (Davenport & Davenport, 2005).
Under extremely low temperatures, components of the community demonstrate tolerance. Lüning (1990) reported that Corallina officinalis from Helgoland survived 0°C when exposed for one week. New Zealand specimens were found to tolerate -4°C (Frazer et al., 1988). Lüning (1990) suggested that most littoral algal species were tolerant of cold and freezing. For example, the photosynthetic rate of Chondrus crispus recovered after 3hrs at -20°C but not after 6 hrs (Dudgeon et al., 1990). The photosynthetic rate of Mastocarpus stellatus higher on the shore fully recovered from 24 hrs at -20°C.
A series of experiments indicated that median lower lethal temperature tolerances for Littorina saxatilis and Littorina littorea were -16.4 and -13°C respectively (Davenport & Davenport, 2005).
Boudouresque & Verlaque, (2001), reviewed the ecology of Paracentrotus lividus. They suggest that the species is particularly common in regions where winter water temperatures range from 10- 15°C and summer temperatures from 18-25°C. Its northern limits correspond to 8°C winter and 28°C summer isotherms. In the English Channel, lower and upper lethal temperatures are 4°C to 29°C. As this species is generally subtidal its tolerance for colder temperatures is predicted to be lower than other intertidal species (Davenport & Davenport, 2005). Populations in rockpool may be exposed to lethal temperatures in winter in intertidal rock pools.
Sensitivity assessment. Based on the characterizing Corallina turf and associated species, this biotope is considered to have ‘High’ resistance and ‘High resilience (by default) to this pressure and is therefore considered to be ‘Not sensitive’. The timing of changes and seasonal weather could result in greater impacts on species. An acute decrease in temperature coinciding with unusually low winter temperatures may exceed thermal tolerances and lead to mortalities of the associated species although this would not alter the character of the biotope. Where this biotope is defined by the presence of the urchin Paracentrotus lividus, sensitivity to this pressure is considered greater where acute decreases in temperature occur in winter. Resistance of the biotope defined by Paracentrotus lividus is therefore assessed as 'Low' and resilience as 'Medium', sensitivity is therefore assessed as 'Medium'. The more precautionary assessment is presented in the sensitivity assessment table.
Local populations may be acclimated to the prevailing salinity regime and may therefore exhibit different tolerances to other populations subject to different salinity conditions and therefore caution should be used when inferring tolerances from populations in different regions. This biotope is found in full (30-35 ppt) salinity (Connor et al., 2004). Biotopes found in the intertidal will naturally experience fluctuations in salinity where evaporation increases salinity and inputs of rainwater expose individuals to fresh water. Species found in the intertidal are therefore likely to have some form of behavioural or physiological adaptations to changes in salinity.
The characterizing species Corallina officinalis and crustose corallines are found in rockpools, such as this biotope, where salinities may fluctuate markedly during exposure to the air. Kinne (1971a) cites maximal growth rates for Corallina officinalis between 33 and 38 psu in Texan lagoons. Edyvean & Ford (1984b) suggest that populations of Lithophyllum incrustans are affected by temperature changes and salinity and that temperature and salinity ‘shocks’ induce spawning but no information on thresholds was provided (Edyvean & Ford, 1984b). Populations of Lithophyllum incrustans were less stable in rockpools with a smaller volume of water that were more exposed to temperature and salinity changes due to lower buffering capacity. Sexual plants (or the spores that give rise to them) were suggested to be more susceptible than asexual plants to extremes of local environmental variables (temperature, salinity etc.) as they occur with greater frequency at sites where temperature and salinity were more stable (Edyvean & Ford, 1984b).
Although present in Mediterranean coastal lagoons and Atlantic estuaries (Boudouresque & Verlaque, 2001) Paracentrotus lividus is sensitive to high salinities. Upper lethal salinities are 39-44 ppt for long-term exposure (Pastor, 1971; Allain, 1975; Le Gall et al., 1989). Young Littorina littorea inhabit rock pools where salinity may increase above 35 psu. Thus, the associated species may be able to tolerate some increase in salinity. Mytilus edulis is found in a wide range of salinities from variable salinity areas and mussels in rock pools are likely to experience hypersaline conditions on hot days. Newell (1979) recorded salinities as high as 42 psu in intertidal rock pools, suggesting that Mytilus edulis can tolerate high salinities.
Sensitivity assessment. No direct evidence was found to assess sensitivity to this pressure. Although some increases in salinity may be tolerated by the associated species present these are generally short-term and mitigated during tidal inundation. This biotope is considered, based on the salinity tolerance of Paracentrotus lividus and the distribution of Corallina officinalis on the mid to lower shore to be sensitive to a persistent increase in salinity to > 40 ppt. Resistance is therefore assessed as ‘Low’ and recovery as ‘Medium’ (following restoration of usual salinity). Sensitivity is therefore assessed as ‘Medium’.
Biotopes found in the intertidal will naturally experience fluctuations in salinity where evaporation increases salinity and inputs of rainwater expose individuals to fresh water. Species found in the intertidal are therefore likely to have some form of behavioural or physiological adaptations to changes in salinity. As this biotope is present in full salinity, the assessed change at the pressure benchmark is a reduction in salinity to a reduced regime (18-30 ppt).
Edyvean & Ford (1984b) suggest that populations of Lithophyllum incrustans are affected by temperature changes and salinity and that temperature and salinity ‘shocks’ induce spawning but no information on thresholds was provided (Edyvean & Ford, 1984b). Populations of Lithophyllum incrustans were less stable in tide pools with a smaller volume of water that were more exposed to temperature and salinity changes due to lower buffering capacity. Sexual plants (or the spores that give rise to them) were suggested to be more susceptible than asexual plants to extremes of local environmental variables (temperature, salinity etc.) as they occur with greater frequency at sites where temperature and salinity were more stable (Edyvean & Ford, 1984b).
In the Baltic, Corallina officinalis is confined to deeper waters as surface salinity decreases (Kinne, 1971a) suggesting that full salinity is required in the long term although short-term fluctuations may be tolerated (although the thresholds of this tolerance are not clear). Kinne (1971a) cites maximal growth rates for Corallina officinalis between 33 and 38 psu in Texan lagoons, so that a decrease in salinity at the pressure benchmark would be predicted to lead to reduced growth. Although present in Mediterranean coastal lagoons and Atlantic estuaries (Boudouresque & Verlaque, 2001) Paracentrotus lividus is sensitive to low salinities. Lower lethal salinities are 15-20 ppt for long-term exposure (Allain, 1975; Pastor. 1971; Le Gall et al., 1989). An abrupt drop in salinity of 7 psu following an exceptional high rainfall event resulted in the estimated loss of 50% of Paracentrotus lividus from a brackish lagoon (salinity typically 30-38 psu) in Urbinu, Corsica (Fernandez et al., 2006). This event was coupled with higher turbidity’s and siltation so that the effects cannot be ascribed to salinity changes alone.
Based on occurrence in estuaries it is clear that some of the species associated with this biotope have a high tolerance for this pressure. However, it should be noted that local populations may be acclimated to the prevailing salinity regime and may therefore exhibit different tolerances to other populations subject to different salinity conditions so that caution should be used when inferring tolerances from populations in different regions. Ulva species are considered to be a very euryhaline species, tolerant of extreme salinities ranging from 0 psu to 136 psu (Reed & Russell, 1979). Although some variation in salinity tolerance between populations of Ulva intestinalis have been found indicating that plants have some adaptation to the local salinity regime (Reed & Russell, 1979). Littorina littorea is found in waters of full, variable and reduced salinities (Connor et al., 2004) and so populations are considered tolerant of decreases in salinity at the pressure benchmark. Mytilus edulis is found in a wide range of salinities from variable salinity areas (18-35 ppt) such as estuaries and intertidal areas, to areas of more constant salinity (Connor et al., 2004). Mytilus edulis was recorded to grow in a dwarf form in the Baltic sea where the average salinity was 6.5psu (Riisgård et al., 1993).
Prolonged reduction in salinity, e.g. from full to reduced (18-30 ppt), is likely to reduce the species richness of the biotope due to loss of some intolerant invertebrates from the assemblage associated with the Corallina officinalis turf.
Sensitivity assessment. Although some daily changes in salinity may be experienced these will be mitigated during tidal inundation. This biotope is considered to be sensitive to a decrease in saliniyt at the pressure benchmark, based on the salinity tolerances of Paracentrotus lividus and Corallina officinalis distribution and the evidence from Kinne (1971). Resistance is therefore assessed as ‘Low’ and recovery as ‘Medium’ (following restoration of usual salinity). Sensitivity is therefore assessed as ‘Medium’.
Corallina officinalis dominated biotopes are found in a range of flow rates from 'moderately strong' (0.5-1.5 m/s) to very 'weak' negligible) (Connor et al., 2004; Dommasnes, 1969) and the rock pool habitat of this biotope will provide some protection from water flows. Moderate water movement is beneficial to seaweeds as it carries a supply of nutrients and gases to the plants and removes waste products. However, if flow becomes too strong, plants may become displaced. Additionally, an increase to stronger flows may inhibit settlement of spores and remove adults or germlings. However, Corallina officinalis have a compact, turf forming growth which reduce water flow through turbulence and friction and are probably resistant to displacement by an increase in water flow. The coralline crusts characterizing this biotope are securely attached and as these are flat are subject to little or no drag.
The shallow burrows created by Paracentrotus lividus also provide some protection from dislodgement. Changes in water flow at the pressure benchmark may result in increased or decreased sediment deposition; these are not considered to alter the character of the biotope but may alter species richness of the small invertebrates associated with the turf.
Sensitivity assessment. The biotope is found across a range of flow rates, mid-range populations are considered to have 'High' resistance to a change in water flow at the pressure benchmark (although see sediment supply caveats). Resilience is assessed as ‘High’, by default, and the biotope is considered ‘Not sensitive’.
Emergence regime is a key factor structuring this (and other) intertidal biotopes although it should be noted that Corallina officinalis may occur at a range of shore heights depending on local conditions such as the degree of wave action (Dommasnes, 1969), shore topography, run-off and degree of shelter from canopy forming macroalgae. The rock pool habitat will provide some protection from emergence but increased exposure would lead to an increase in emergence which may lead to shallow pools drying out, greater exposure to air temperatures with corresponding fluctuations in water temperature and changes in salinity following evaporation or dilution by rainfall. Changes in emergence may therefore lead to species replacement and the development of a biotope more typical of the changed shore level.
This biotope is considered sensitive to increased emergence as the key characterizing Corallina officinalis is sensitive to desiccation (Dommasnes, 1969) and is generally not found on open rock unless protected by algal canopies or where the surfaces are damp or wet. At Hinkley Point (Somerset, England), for example, seawater run-off from deep pools high in the intertidal supports dense turfs of Corallina spp. lower on the shore (Bamber & Irving, 1993). Fronds are highly intolerant of desiccation and do not recover from a 15% water loss, which might occur within 40-45 minutes during a spring tide in summer (Wiedemann, 1994). Bleached corallines were observed 15 months after the 1964 Alaskan earthquake which elevated areas in Prince William Sound by 10 m. Similarly, increased exposure to air caused by upward movement of 15 cm due to nuclear tests at Armchitka Island, Alaska adversely affected Corallina pilulifera (Johansen, 1974). During an unusually hot summer, Hawkins & Hartnoll (1985) observed damaged Corallina officinalis and other red algae. Littler & Littler, (1984) suggest that the basal crustose stage is adaptive, allowing individuals to survive periods of physical stress as well as physiological stress such as desiccation and heating. The basal crust stage may persist for extended periods with frond regrowth occurring when conditions are favourable.
Occurrence of encrusting coralline algae seems to be critically determined by exposure to air and sunlight. Colonies survive in damp conditions under algal canopies or in pools but not on open rock where desiccation effects are important. Increased emergence leading to drying out of shallow pools would reduce habitat suitability for this group. Spore release by the crusting coralline Lithophyllum incrustans is triggered by small changes in salinity and temperature and therefore changes in emergence may alter patterns in reproduction and recruitment (see relevant pressures for further information). However, this species does occur both high and low in the intertidal (Edyvean & Ford, 1986) and presumably such impacts are limited.
Paracentrotus lividus is typically a sublittoral species (Boudouresque & Verlaque, 2001) and the upper limits of Paracentrotus lividus on a shore are limited by desiccation. The rock pool habitat provides some protection but increases in emergence that lead to the drying of shallow pools, greater exposure to air temperatures or increased salinity would negatively impact this species.
Mobile epifauna are likely to relocate to more suitable habitats. Species such as Littorina littorea that are found throughout the intertidal zone are adapted to tolerate desiccation to some extent. For example: littorinids can seal their shell using the operculum while limpets clamp tightly to rock reducing water loss.
Corallina officinalis, Paracentrotus lividus and many of the associated species are found subtidally. Decreased emergence is likely to lead to the habitat the biotope is found in becoming more suitable for the lower shore species generally found below the biotope, leading to replacement by, for example, a kelp dominated biotope with red algae and Corallina officinalis and Paracentrotus lividus surviving under the canopy. Although the key species would be expected to survive the biotope would have altered from the description.
Sensitivity assessment. Emergence is a key factor structuring the distribution of on the shore. Resistance to increased emergence is assessed as ‘Medium’ as the coralline crusts, Corallina officinalis and Paracentrotus lividus are intolerant of desiccation but are generally protected by the pool environment. Recovery is assessed as ‘High’ and sensitivity is therefore assessed as ‘Low’. An increase in emergence repositioning the biotope at an upper shore level would be likely to significantly affect the species composition due to increased temperatures, greater variation in salinity and greater sensitivity.
A decrease in emergence would mean that shallow rockpools would be at less risk of desiccation. In addition, depending on the nature of the surrounding bedrock, the rockpool may become slightly deeper. As a result, it is possible that species diversity could increase but, on the whole, biotopes at the upper and mid-shore are likely to be tolerant of a decrease in emergence at the benchmark level. Pre-emption of space by Corallina officinalis and other red algae may reduce the establishment of lower shore species including kelps, resistance is therefore assessed as ‘Medium’ to decreased emergence and recovery as ‘High’ so that sensitivity is assessed as ‘Low’.
Colonies of Lithophyllum incrustans appear to thrive in conditions exposed to strong water movement. Irvine & Chamberlain (1994) observed that this species is best developed on wave exposed shores. Corallina officinalis dominated biotopes are found in a range of flow rates from 'moderately strong' (0.5-1.5 m/s) to very 'weak' negligible) (Connor et al., 2004; Dommasnes, 1969) and the rock pool habitat of this biotope will provide some protection from water flows. Moderate water movement is beneficial to seaweeds as it carries a supply of nutrients and gases to the plants and removes waste products. However, if flow becomes too strong, plants may become displaced. Additionally, an increase to stronger flows may inhibit settlement of spores and remove adults or germlings. However, Corallina officinalis have a compact, turf forming growth which reduce water flow through turbulence and friction and are probably resistant to displacement by an increase in water flow. The shallow burrows created by Paracentrotus lividus also provide some protection from dislodgement. Changes in water flow at the pressure benchmark may result in increased or decreased sediment deposition, these are not considered to alter the character of the biotope but may alter species richness of the small invertebrates associated with the turf.
Sensitivity assessment. Wave action rather than water flow is a more significant source or water movement than currents in this biotope. Corallina officinalis biotopes are found across a range of flow rates, mid-range populations are considered to have 'High' resistance to a change in water flow at the pressure benchmark (although see sediment supply caveats). Resilience is assessed as ‘High’, by default, and the biotope is considered ‘Not sensitive’.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Contamination by non-synthetic chemicals, at levels greater than the pressure benchmark may adversely impact the biotope. Little information was found concerning the effects of heavy metals on turf forming and encrusting coralline algae. However, Bryan (1984) suggested that the general order for heavy metal toxicity in seaweeds is: organic Hg> inorganic Hg > Cu > Ag > Zn> Cd>Pb. Most of the information available suggests that the associated adult gastropod molluscs are rather tolerant of heavy-metal toxicity (Bryan, 1984). Winkles may absorb metals from the surrounding water by absorption across the gills or from their diet, and evidence from experimental studies on Littorina littorea suggest that diet is the most important source (Bryan et al., 1983). The species has been suggested as a suitable bioindicator species for some heavy metals in the marine environment. Bryan et al. (1983) suggested that the species is a reasonable indicator for Ag, Cd, Pb and perhaps As. In the Fal estuary Patella vulgata occurs at, or just outside, Restronguet Point, at the end of the creek where metal concentrations are in the order: Zinc (Zn) 100-2000 µg/l, copper (Cu) 10-100µg/l and cadmium (Cd) 0.25-5µg/l (Bryan & Gibbs, 1983). However, in the laboratory Patella vulgata was found to be intolerant of small changes in environmental concentrations of Cd and Zn by Davies (1992). At concentrations of 10µg/l pedal mucus production and levels of activity were both reduced, indicating a physiological response to metal concentrations. Exposure to Cu at a concentration of 100 µg/l for one week resulted in progressive brachycardia (slowing of the heart beat) and the death of limpets. Zn at a concentration of 5500 µg/l produced the same effect (Marchan et al., 1999).
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Contamination by non-synthetic chemicals, at levels greater than the pressure benchmark may adversely impact the biotope. Where exposed to direct contact with fresh hydrocarbons, encrusting coralline algae appear to have a high intolerance. Crump et al. (1999) described "dramatic and extensive bleaching" of 'Lithothamnia' following the Sea Empress oil spill. Observations following the Don Marika oil spill (K. Hiscock, pers. comm.) were of rockpools with completely bleached coralline algae. However, Chamberlain (1996) observed that although Lithophyllum incrustans was affected in a short period of time by oil during the Sea Empress spill, recovery occurred within about a year. The oil was found to have destroyed about one third of the thallus thickness but regeneration occurred from thallus filaments below the damaged area.
Following the Torrey Canyon oil spill in 1967, oil and detergent dispersants affected high shore specimens of Corallina officinalis more than low shore specimens. Plants in deep pools were afforded some initial protection, although probably later affected by contaminated runoff. In areas of heavy spraying, however, Corallina officinalis was killed. (Smith, 1968). Intolerance to hydrocarbon pollution has been assessed to be high, as key structural and important characterizing coralline algal species will be lost and the biotope not be recognized in their absence.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Contamination by synthetic chemicals, at levels greater than the pressure benchmark may adversely impact the biotope. Smith (1968) reported that oil and detergent dispersants from the Torrey Canyon spill affected high water plants of Corallina officinalis more than low shore plants and some plants were protected in deep pools. In areas of heavy spraying, however, Corallina officinalis was killed (Smith, 1968). Regrowth of fronds had begun within two months after spraying ceased (Smith, 1968). O'Brien & Dixon (1976) suggested that red algae were the most sensitive group of algae to oil or dispersant contamination, possibly due to the susceptibility of phycoerythrins to destruction.
Cole et al. (1999) suggested that herbicides were (not surprisingly) very toxic to algae and macrophytes. Hoare & Hiscock (1974) noted that with the exception of Phyllophora species, all red algae including encrusting coralline forms, were excluded from the vicinity of an acidified halogenated effluent discharge in Amlwch Bay, Anglesey and that intertidal populations of Corallina officinalis occurred in significant amounts only 600 m east of the effluent. Chamberlain (1996) observed that although Lithophyllum incrustans was quickly affected by oil during the Sea Empress spill, recovery occurred within about a year. The oil was found to have destroyed about one third of the thallus thickness but regeneration occurred from thallus filaments below the damaged area.
Most pesticides and herbicides were suggested to be very toxic for invertebrates, especially crustaceans (amphipods isopods, mysids, shrimp and crabs) and fish (Cole et al., 1999).
|No evidence (NEv)||Not relevant (NR)||No evidence (NEv)|
No evidence was found to assess this pressure at the benchmark. Algae bioaccumulate radionuclides (with extent depending on the radionuclide and the algae species). Adverse effects have not been reported at low levels.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
Over geological timescales periods of increased nutrient availability have experienced increases in the distribution of crustose coralline species at the expense of corals (Littler & Littler, 2013), suggesting that this group have some tolerance for enhanced nutrient levels. Overall, Littler and Littler (2013) suggest that corallines as a group can tolerate both low and elevated levels of nutrients. The key characterizing Corallina officinalis and the associated green algae species have been identified worldwide as species that occur in areas subject to increased nutrient input within the vicinity of sewage outfalls and at intermediately polluted sites (Littler & Murray, 1975; May, 1985; Brown et al., 1990; Belgrove et al., 1997 and 2010). For example Kindig and Littler (1980) demonstrated that Corallina officinalis var. chilensis in South California showed equivalent or enhanced health indices, highest productivity and lowest moralities (amongst the species examined) when exposed to primary or secondary sewage effluent. Grazers in the biotope may benefit from increased availability of food resources, due to enhanced growth.
The review by Boudouresque & Verlaque, (2001) considered that Paracentrotus lividus are insensitive to nutrient enrichment as dense populations are found in polluted areas. Density of Paracentrotus lividus increased towards a sewage outlet in Cortiou cove, France (Boudouresque & Verlaque, 2001). Paracentrotus lividus cultivated adjacent to salmon pens exposed to high levels of nutrients thrived (Cook & Kelly, 2007).
Atalah & Crowe (2010) added nutrients to rockpools occupied by a range of algae including encrusting corallines, turfs of Mastocarpus stellatus, Chondrus crispus and Corallina officinalis and green and red filamentous algae. The invertebrates present were mostly Patella ulyssiponensis, the winkle Littorina littorea and the flat top shell Gibbula umbilicalis. Nitrogen and phosphorous enhancement was via the addition of fertilisers, as either 40 g/l or 20 g/l. The treatments were applied for seven month and experimental conditions were maintained every two weeks. The experimental treatments do not directly relate to the pressure benchmark but indicate some general trends in sensitivity. Nutrients had no significant effect on the cover of Corallina officinalis. The cover of green filamentous algae was significantly increased both by reduced grazing and increased nutrients, although the effect size was synergistically magnified by the combined effect of grazer removal and nutrients. Nutrient enrichment caused an absolute increase in the average cover of green filamentous algae of 19% (±3.9 S.E.) with respect to the control treatments while the cover of red turfing algae was not affected by nutrient addition (Atalah & Crowe, 2010).
Sensitivity assessment. The pressure benchmark is relatively protective and may represent a reduced level of nutrient enrichment in previously polluted areas. Due to the tolerance of high levels of nutrient input demonstrated generally by Corallina turfs e.g. Bellgrove et al., (2010) and Atalah & Crowe, (2010) and Paracentrotus lividus (Boudouresque & Verlaque, 2001; Cook & Kelly, 2007). resistance to this pressure is assessed as ‘High’ and resilience as ‘High’ so that the biotope is assessed as ‘Not sensitive’. Where Corallina dominated biotopes have replaced canopy forming species in enriched areas it is not clear whether a change to the benchmark would lead to a shift in biotope type. Once established the presence of Corallina spp. and other turf forming species may limit recruitment by taller species (Bellgrove et al., 2010). No evidence was found to support an assessment of this indirect effect and it is not presented within the table.
No direct evidence was found to assess this pressure. Organic enrichment may lead to eutrophication with adverse environmental effects including deoxygenation, algal blooms and changes in community structure (see nutrient enrichment and de-oxygenation). Where the biotopes occur in tide swept or wave exposed areas (Connor et al., 2004) water movements will disperse organic matter reducing the level of exposure. The key characterizing species Corallina officinalis has been noted to increase in abundance and may form extensive turfs within the vicinity of sewage outfalls and at intermediately polluted sites (Littler & Murray, 1975; May, 1985; Brown et al., 1990; Bellgrove et al., 2010). Corallina elongata and the crusting coralline Lithophyllum incrustans were present at sites dominated by Ulva spp. in the Mediterranean exposed to high levels of organic pollution from domestic sewage (Arévalo et al., 2007). As turf forming algae Corallina spp. trap large amounts of sediment and are therefore not considered sensitive to sedimentation. The turfs host a variety of associated species and deposit feeders amongst these would be able to consume inputs of organic matter.
The review by Boudouresque & Verlaque, (2001) considered that Paracentrotus lividus are insensitive to nutrient enrichment as dense populations are found in polluted areas. Density of Paracentrotus lividus increased towards a sewage outlet in Cortiou cove, France (Boudouresque & Verlaque, 2001). Paracentrotus lividus cultivated adjacent to salmon pens exposed to high levels of particulate organic matter thrived (Cook & Kelly, 2007).
Sensitivity assessment. Based on resistance to sedimentation, exposure to wave action, the presence of detrital consumers and the persistence of turfs and Paracentrotus lividus in areas subject to sewage inputs resistance is assessed as ‘High’ and resilience as ‘High’ (by default). The biotope is therefore considered to be ‘Not sensitive’ to this pressure at the benchmark.
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.
This biotope is characterized by the hard rock substratum to which the characterizing encrusting corallines and Corallina turf and associated species, can firmly attach to, in the case of Paracentrotus lividus on limestone, burrow into. Changes to a free draining sediment type or to an impervious artificial substratum that was free draining would significantly alter the character of the biotope through the loss of the pool habitat. A change to an artificial substratum could also impact the development of this biotope as species may have settlement preferences for particular surface textures. Corallina officinalis shows optimal settlement on finely rough artificial substrata (0.5 - 1 mm surface particle diameter). Although spores will settle and develop as crustose bases on smooth surfaces, fronds were only initiated on rough surfaces. Corallina officinalis settled on artificial substrata within one week in the field in summer months in New England (Harlin & Lindbergh 1977). However, in the laboratory fronds can grow from bases attached to smooth surfaces (Wiedeman pers comm. Previous MarLIN review).
Paracentrotus lividus occurs in a range of habitats on solid rocks, boulders and in meadows of sea grasses but is uncommon on sand and other sediment types (Boudouresque & Verlaque, 2001). Paracentrotus lividus burrows into sandstone, limestone basalt and granite but not hard slate (Boudouresque & Verlaque, 2001), other artificial hard substratum may not be suitable for burrows.
Artificial hard substratum may also differ in character from natural hard substratum, so that replacement of natural surfaces with artificial may lead to changes in the biotope through changes in species composition, richness and diversity (Green et al., 2012; Firth et al., 2013) or the presence of non-native species (Bulleri & Airoldi, 2005).
Sensitivity assessment. The biotope is considered to have 'No' resistance to this pressure based on a change to a soft sediment substratum or a hard, free draining substratum. Recovery is ‘very Low’ as the change at the benchmark is permanent. Sensitivity is therefore assessed as ‘High’.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Not relevant to biotopes occurring on bedrock.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
The species characterizing this biotope are epifauna or epiflora occurring on rock and would be sensitive to the removal of the habitat. However, extraction of rock substratum is considered unlikely and this pressure is considered to be ‘Not relevant’ to hard substratum habitats.
The species characterizing this biotope occur on the rock and therefore have no shelter from abrasion at the surface. Littler & Littler, 1984) suggest that the basal crustose stage of Corallina officinalis is adaptive to resist sand scour and wave shearing (as well as physiological stressors such as desiccation and heating). The base is much tougher than the fronds shown by experiments that have demonstrated that the base has nearly twice the mechanical resistance (measured by penetration) of fronds (Littler & Kauker, 1984).
In general, studies show that Corallina and other turf forming algae appear to be relatively resistant to single events and low levels of trampling. Brosnan & Crumrine (1994), for example, found that in experimentally trampled plots the cover of foliose and canopy forming species declined while turf forming algae were relatively resistant. Similarly, a comparison of rocky intertidal ledges that received different amounts of visitors in Dorset, England, found that Corallina officinalis were present on both heavily visited and less visited ledges suggesting that this species has some resistance to trampling (Pinn & Rodgers, 2005). Povey & Keough (1991) in Mornington Peninsula, Australia investigated the effects of sustained trampling on intertidal coralline algal mats where upright branching Corallina spp. formed a turf with other red algae with sand and encrusting coralline algae between turfs. The experimental strips were 2 m long and 0.5 m wide. The percentage cover of upright Corallina spp. was significantly affected by 25 passages of a strip per day after 12 and 33 days. The algae appeared flattened and were shorter (1-2 cm high) compared with the low intensity and control plots (3-4 cm high). However low intensity trampling within a strip (2 passages/ day) did not significantly affect the coralline turf.
Brown & Taylor (1999) also found that higher intensities of trampling damaged turfs. Moderate (50 steps per 0.09 sq. metres) or more trampling on intertidal articulated coralline algal turf in New Zealand reduced turf height by up to 50%, and weight of sand trapped within turf to about one third of controls. This resulted in declines in densities of the meiofaunal community within two days of trampling. Although the community returned to normal levels within 3 months of trampling events, it was suggested that the turf would take longer to recover its previous cover (Brown & Taylor 1999). Similarly, Schiel & Taylor (1999) noted that trampling had a direct detrimental effect on coralline turf species on the New Zealand rocky shore. At one site coralline bases were seen to peel from the rocks (Schiel & Taylor 1999), however, this was probably due to increased desiccation caused by loss of the algal canopy.
Schiel & Taylor (1999) reported the death of encrusting corallines one month after trampling due to removal of their protective canopy of fucoids by trampling (10 -200 tramples where one trample equals one transect walked by one person). A higher proportion of corallines died back in spring treatments presumably due to the higher levels of desiccation stress expected at this time of year. However, encrusting corallines increased within the following year and cover returned to control levels within 21 months (Schiel & Taylor, 1999). Mechanical abrasion from scuba divers was also reported to impact encrusting corallines, with cover of Lithophyllum stictaeforme greater in areas where diving was forbidden than visited areas (abundance, 6.36 vs 1.4; it is presumed this refers to proportion of cover, although this is not clear from the text, Guarinieri et al., 2012).
Dethier (1994) experimentally manipulated surface abrasion on a range of encrusting algae including Lithophyllum impressum. Crusts were brushed with either a nylon or steel brush for 1 minute a month for 24 months. Unbrushed controls grew by approximately 50% where the cover of nylon brushed crusts and steel brushed crusts decreased by approximately 25% and 40% respectively (interpreted from figures in Dethier, 1994). In laboratory tests on chips of Lithophyllum impressum brushing with a steel brush for 1 minute once a week for 3 weeks, resulted in no cover loss of two samples while a third ‘thinned and declined’ (Dethier, 1994).
The burrows of Paracentrotus lividus may provide some protection from abrasion. No direct evidence was found to assess sensitivity. MacDonald et al. (1996) assessed benthic species sensitivity to fishing disturbance by ‘scoring’ each species ability to withstand the physical impact of a single fishing disturbance and recovery potential assuming no further fishing disturbance occurred. These authors classified the larger urchin slow growing Echinus esculentus as being ‘very fragile’. Kaiser et al. (2000) reported that Echinus esculentus were less abundant in areas subject to high trawling disturbance in the Irish Sea. This species was reported to suffer badly as a result of impact with passing scallop or queen scallop dredges (Bradshaw et al., 2000; Hall-Spencer & Moore, 2000a), it is likely that when directly exposed to abrasion Paracentrotus lividus would be similarly impacted.
No evidence was found for the sensitivity of the small invertebrates associated with the coralline turf but abrasion could displace and damage these. The littorinids and top shells that occur in this biotope, have some protection from hard shells but abrasion may damage and kill individuals Povey & Keough (1991) in experiments on shores in Mornington peninsula, Victoria, Australia, found that in single step experiments less than 5% of littorinids were crushed in single step experiments (Povey & Keough, 1991).In sites with mobile cobbles and boulders increased scour results in lower densities of Littorina spp. compared with other, local sites with stable substratum (Carlson et al., 2006).
Sensitivity assessment. The impact of surface abrasion will depend on the footprint, duration and magnitude of the pressure. Based on evidence from the step experiments and the relative robustness of the encrusting corallines and Corallina officinalis turf and associated species, resistance, to a single abrasion event is assessed as ‘Medium’ and recovery as ‘High’, so that sensitivity is assessed as ‘Low’. Resistance and resilience will be lower (and hence sensitivity greater) to abrasion events that exert a greater crushing force and remove the bases than the trampling examples the assessment is based on). The sensitivity of the key characterizing species Paracentrotus lividus is likely to be greater when directly exposed to abrasion. Resistance is therefore assessed as ‘Low’ and recovery as ‘Medium’ so that the sensitivity of the biotope defined by this species is assessed as ‘Medium’.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
The species characterizing this biotope group are epifauna and epiflora occurring on rock which is resistant to subsurface penetration. The assessment for abrasion at the surface only is therefore considered to equally represent sensitivity to this pressure.
Intertidal biotopes will only be exposed to this pressure when submerged during the tidal cycle and thus have limited exposure. Siltation, which may be associated with increased suspended solids and the subsequent deposition of these is assessed separately (see siltation pressures). In general, increased suspended particles reduce light penetration and increase scour and deposition. They may enhance food supply to filter or deposit feeders (where the particles are organic in origin) or decrease feeding efficiency (where the particles are inorganic and require greater filtration efforts).
Increases in the cover of sediment trapping, turf forming algae at the expense of canopy forming species has been observed worldwide in temperate systems and has been linked to increased suspended solids linked to human activities worldwide (Airoldi, 2003). Corallina species accumulate more sediment than any other alga (Hicks, 1985). Hence an increase in suspended sediment is likely to accumulate in the coralline turf. A significant increase may result in smothering (see above). An accumulation of sediment within the turf may attract more sediment dwelling interstitial invertebrates such as nematodes, harpacticoids and polychaetes, although in more wave exposed locations accumulation of sediment is likely to be minimal. Increased suspended sediment may also result in increased scour, which may adversely affect foliose red algae, and interfere with settling spores and recruitment if the factor is coincident with their major reproductive period. However, coralline algae, especially the crustose forms are thought to be resistant of sediment scour (Littler & Kauker, 1984), and will probably not be adversely affected at the benchmark level.
No direct evidence was found for the sensitivity of Paracentrotus lividus to changes in suspended solids. However this species has been reported to increase in abundance at sites subject to organic pollution where levels of particulate organic matter are high. Density of Paracentrotus lividus increased towards a sewage outlet in Cortiou cove (Boudouresque & Verlaque, 2001, references therein). Paracentrotus lividus cultivated adjacent to salmon pens that received high levels of particulate organic matter thrived (Cook & Kelly, 2007).
This community is unlikely to be dependent on suspended sediment. Although accumulated sediment within coralline turf habitats is likely to increase the species diversity of the epiphytic fauna, in very wave exposed locations, accumulated sediment in the habitat is likely to be minimal. A reduction in suspended sediment will probably reduce the risk of scour, and reduce food availability for the few suspension feeding species in the biotope (e.g. barnacles and spirorbids present), although effects are not likely to be lethal.
The biotope occurs in shallow waters where light attenuation due to increases in turbidity is probably low. Red algae and coralline algae especially are known to be shade tolerant and are common components of the understorey on seaweed dominated shores. Limited shading from suspended sediments is therefore not considered to negatively affect this genus. Therefore, a decrease in light intensity is unlikely to adversely affect the biotope. An increase in light intensity is unlikely to adversely affect the biotope as plants can acclimate to different light levels.
Sensitivity assessment. The exposure of biotope to suspended sediments in the water column will be limited to immersion periods, and wave action will reduce accumulation. The biotope is considered to be ‘Not sensitive’ to a reduction in suspended solids, although this may reduce food supply to the barnacles and other filter feeders that occur in this biotope. An increase in suspended solids may lead to some sub-lethal abrasion of fronds however, evidence globally indicates that increase suspended solids favour the turf-forming algae that characterize this biotope (Airoldi, 2003). Resistance is therefore assessed as ‘High’ and resilience as ‘High’ (by default) so that the biotope is considered to be ‘Not sensitive’.
Increased abundance of algal turfs worldwide has been linked to sediment perturbations although not all the pathways and mechanisms of these effects are clear (see review by Airoldi, 2003). However, even the most tolerant of organisms would eventually suffer from inhibition and mortality following smothering although the thresholds for these effects have not been identified (Airoldi, 2003). Corallina officinalis and others within the genus (e.g. Corallina pinnatifolia and Corallina vancouveriensis) are found on shores subject to high rates of sedimentation that are periodically disturbed by sand burial and scour (Stewart, 1989). Coralline turfs also trap sediments within the turf. The amount of sediment present and the associated fauna varies naturally depending on local conditions such as wave exposure (Dommasnes, 1969). On intertidal shores in southern California the amount of sediment trapped within turfs of Corallina spp. varied seasonally (from < 5 mm to >4.5 cm) and was closely related to species composition and the structure of the turf. Airoldi (2003) identified a number of morphological, physiological and life history traits that conferred high levels of tolerance to sedimentation. Those shared by Corallina spp are the regeneration of upright fronds from a perennial basal crust resistant to burial and scour, calcified thalli, apical meristems, large reproductive ouptuts, lateral vegetative growth and slow growth rates (Airoldi, 2003).
In a review of the effects of sedimentation on rocky coast assemblages, Airoldi (2003) outlined the evidence for the sensitivity of coralline algae to sedimentation. The reported results are contradictory with some authors suggesting that coralline algae are negatively affected by sediments while others report that encrusting corallines are often abundant or even dominant in a variety of sediment impacted habitats (Airoldi, 2003 and references therein). Crustose corallines have been reported to survive under a turf of filamentous algae and sediment for 58 days (the duration of experiment) in the Galapagos (species not identified, Kendrick, 1991). The crustose coralline Hydrolithon reinboldii, has also been reported to survive deposition of silty sediments on subtidal reefs off Hawaii (Littler, 1973).
Atalah & Crowe (2010) added sediment to rockpools in controlled experiments that appear to be very similar to this biotope. The rockpools were occupied by a range of algae including encrusting corallines, turfs of Mastocarpus stellatus, Chondrus crispus and Corallina officinalis and green and red filamentous algae. The invertebrates present were mostly Patella ulyssiponensis, the winkle Littorina littorea and the flat top shell Gibbula umbilicalis. Sediment treatment involved the addition of a mixture of coarse and fine sand of either 300 mg/cm2/month or 600 mg/cm2 every 15 days (the depth of sediment was not reported). The experimental treatments do not directly relate to the pressure benchmark but indicate some general trends in sensitivity. In the pools the chronic addition of both levels of sediment led to a significant decrease in grazers and crustose coralline algae also decreased. Sedimentation had no significant effect on the cover of green filamentous algae (Ulva sp.) but led to an increase in the mean cover of red turfing algae (Mastocarpus stellatus and Chondrus crispus and Corallina officinalis) from 11.7% (±1.0 S.E.) in controls to 26.1% (±4.7 S.E.) in sedimented assemblages, but there were no differences between the two levels of sedimentation. The cover of red filamentous algae (Ceramium spp. Gelidium spp.) was also significantly increased in the sedimentation experiments. The experimental results support the general trend of greater sensitivity of grazers and encrusting corallines to sedimentation than turf-forming algae.
No direct evidence was found to support this assessment for Paracentrotus lividus. Last et al., (2011) found that another epifaunal urchin Psammechinus miliaris is moderately tolerant of shorter term (12 days) burial events, with less than 25% mortality of all buried specimens. Survivorship was partly due to the re-emergence of many specimens, even from depths of up to 7cm, particularly when buried under coarse sediment. After 12 days of burial, mortality in the specimens that remained buried was high. Percentage mortality increased with progressively finer sediment fractions (Last et al., 2011).
Sensitivity assessment. Where sediments are added to tide pools they may be removed rapidly in wave exposed environments where pools are flushed or they may remain in-situ increasing the level of exposure. These effects are site specific and will mediate the level of impact. Within the biotope the associated urchins and encrusting corallines are probably the most sensitive element, littorinids may also be negatively impacted but his will depend on the character of the deposit. Where grazing is reduced then the algal assemblage is likely to change. The encrusting corallines may be sensitive to sedimentation but the level of sensitivity is difficult to ascertain from the evidence base. The turf-forming species are generally tolerant of some sedimentation but the threshold in relation to the benchmark is not easily determined, especially as this biotope occurs in rockpools which may increase the exposure. The encrusting corallines are considered to have ‘Low’ resistance to this pressure and ‘Medium’ resilience, while the algal turf and Paracentrotus lividus is considered to have ‘Medium’ resistance and ‘High’ resilience. The more precautionary assessment of ‘Medium’ sensitivity is presented in the sensitivity assessment table.
The available evidence for siltation pressures is outlined for the ‘light’ deposition pressure. At the pressure benchmark ‘heavy deposition’ represents a considerable thickness of deposit and may fill shallow pools. Complete burial of algal turf and encrusting corallines and associated animals would occur. Removal of the sediments by wave action and tidal currents would result in considerable scour. The effect of this pressure is likely to be mediated by the length of exposure to the deposit. Resistance is assessed as ‘Low’ to ‘None’ as the impact is likely to be significant and would almost certainly result in the loss of grazers and a high proportion of the encrusting corallines. Resilience is assessed as ‘Medium’ and sensitivity is assessed as ‘Medium’.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
|No evidence (NEv)||Not relevant (NR)||No evidence (NEv)|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Coralline crusts and Corallina officinalis are shade tolerant algae, often occurring under a macralgal canopy that reduces light penetration. These species can acclimate to different levels of light intensity and quality and encrusting corallines can occur in deeper water than other algae where light penetration is limited. Samples of Lithophyllum impressum suspended from a raft and shaded (50-75% light reduction) continued to grow over two years (Dethier, 1994). In areas of higher light levels, the fronds and bases may be lighter in colour due to bleaching (Colhart & Johansen, 1973). Other red algae in the biotope are flexible with regard to light levels and can also acclimate to different light levels. Canopy removal experiments in a rocky sub tidal habitat in Nova Scotia, Canada by Schmidt and Scheibling (2007) did not find a shift in understorey macraoalgal turfs (dominated by Corallina officinalis, Chondrus crispus and Mastocarpus stellatus) to more light-adapted species over 18 months.
Bright light and shading elicit well studied reactions in echinoderms. In echinoids shading results in the 'shadow reaction' in which the pedicellariae and spines are pointed in the direction of the shade in a defensive reaction. Echinoids move away from bright light and seek out crevices and / or cover themselves with debris such as shells and drift algae, the 'covering reaction' (see Boolootian (1966) for discussion). An increase in light may therefore lead to more covering by Paracentrotus lividus.
Sensitivity assessment. As the key structuring and characterizing species colonize a broad range of light environments from intertidal to deeper sub tidal and shaded understorey habitats the biotope is considered to have ‘High’ resistance and, by default, ‘High’ resilience and therefore is ‘Not sensitive’ to this pressure.
Barriers that reduce the degree of tidal excursion may alter larval supply to suitable habitats from source populations. Conversely the presence of barriers may enhance local population supply by preventing the loss of larvae from enclosed habitats. Barriers and changes in tidal excursion are not considered relevant to the characterizing Corallina officinalis as species dispersal is limited by the rapid rate of settlement and vegetative growth from bases rather than reliance on recruitment from outside of populations. Other species associated with the biotope such as Paracentrotus lividus produce large numbers of larvae capable of long distance transport and survival, resistance to this pressure is assessed as 'High' and resilience as 'High' by default. This biotope is therefore considered to be 'Not sensitive'.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Not relevant’ to seabed habitats. NB. Collision by grounding vessels is addressed under ‘surface abrasion.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Bright light and shading elicit well studied reactions in echinoderms. In echinoids shading results in the 'shadow reaction' in which the pedicellariae and spines are pointed in the direction of the shade in a defensive reaction. Echinoids move away from bright light and seek out crevices and / or cover themselves with debris such as shells and drift algae, the 'covering reaction' (see Boolootian (1966) for discussion). Movement of boats is unlikely to be noticed, if echinoids react to the approach of divers and snorkelers at closer proximity, the reaction is likely to be short lived and insignificant.
Paracentrotus lividus cultivation occurs in some of its range. The key characterizing species Paracentrotus lividus is cultivated commercially from hatchery reared individuals, this practice is relatively recent and no evidence was found to assess potential impact of gene flow between wild and cultivated populations. Due to the wide dispersal of the larval there appears to be much gene flow and mixing between populations with little genetic variation (Duran et al., 2004; Luri et al., 2007).
Commercial cultivation of Mytilus edulis (which occurs in low densities in this biotope) involves the collection of juvenile mussel ‘seed’ or spat (newly settled juveniles ca 1-2 cm in length) from wild populations, with subsequent transportation around the UK for re-laying in suitable habitats. As the seed is harvested from wild populations from various locations the gene pool will not necessarily be decreased by translocations. Movement of mussel seed has the potential to transport pathogens and non-native species (see sensitivity assessments for Mytilus edulis bed biotopes). A review by Svåsand et al. (2007) concluded that there was a lack of evidence distinguishing between different Mytilus edulis populations to accurately assess the impacts of hybridisation with the congener Mytilus galloprovincialis and in particular how the gene flow may be affected by aquaculture. Therefore, it cannot be confirmed whether farming will have an impact on the genetics of wild sindividuals beyond a potential for increased hybridisation.
Sensitivity assessment. No direct evidence was found regarding the potential for negative impacts of translocated mussel seed on wild Mytilus edulis populations. While it is possible that translocation of mussel seed could lead to genetic flow between cultivated beds and local wild populations, there is currently no evidence to assess the impact (Svåsand et al., 2007). Gene flow between populations of Paracentrotus lividus is high due to the long range dispersal and there is little genetic differentiation between populations. Resistance of this species and the biotope it defines is therefore assessed as ‘High’, resilience is ‘High’ (by default) and the biotope is ‘Not sensitive’.
Within rockpools temperature and salinity may be subject to rapid variation and these conditions will largely exclude species that are adapted to more stable, subtidal environments. Invasive non-indigenous species (INIS) that can alter habitats (ecological engineers), or outcompete native macroalgae for space and other resources such as light and nutrients, are the most likely species to negatively affect this biotope. Space pre-emption by encrusting corallines and the crustose bases of the macroalgae forming the turf, as well as the trapped sediment within the turf, may prevent settlement of INIS until disturbance events create gaps for invasion. However, in the Mediterranean crustose corallines and algal turfs facilitate attachment of Caulerpa racemosa by providing a more complex substratum than bare rock (Bulleri & Benedetti-Cecchi, 2008).
Algal species which may have overlapping habitat requirements include the green seaweed Codium fragile ssp. tormentosoides (now renamed as Codium fragile fragile) and the red seaweed Heterosiphonia japonica, neither of these have so far been recorded in nuisance densities (Sweet, 2011j). Beneath a canopy of the invasive Codium fragile ssp. tomentosoides on subtidal rocky shores in Nova Scotia, Corallina officinalis was the dominant species comprising 78-80% of the turf biomass, while Chondrus crispus and Mastocarpus stellatus comprised 18% (Schmidt & Scheibling, 2007). The biomass of Corallina officinalis was similar to those under a canopy of the native Laminaria species on the same shore (Laminaria longicruris and Laminaria digitata), suggesting little negative effect on turf forming algae. Wireweed, Sargassum muticum, grows best on sheltered shores and in rockpools (Sewell, 2011c), where this species dominates rockpools the character of the biotope will be different to the description and the biotope would be reclassified as LR.FLR.Rkp.FK.Sar. The red seaweeds Heterosiphonia japonica and Neosiphonia harveyi may also occur in this biotope but again no impacts have been reported.
The red seaweed Grateloupia turuturu occurs on the lower shore in pools. No ecosystem impacts have been reported in Great Britain; however this large, fast-growing seaweed may have the potential to displace native seaweed species and shade neighbouring species. In North America this species is a major competitor of Chondrus crispus which provides an important winter food source for littorinids and other invertebrates. As Grateloupia turuturu dies-back in the winter, the dispolacement of other species may therefore affect grazers (Sweet, 2011g).
The tunicates Didemnum vexillum and Asterocarpa humilis, the hydroid Schizoporella japonica and the bryozoan Watersipora subatra (Bishop, 2012c, Bishop, 2015a and b; Wood, 2015) are currently only recorded from artificial hard substratum in the UK and it is not clear what their established range and impacts in the UK would be. Didemnum vexillum occurs in tide pools in other areas where it has become established (Bishop, 2012c) and can have substantial effects on communities, similarly the tunicates Corella eumyota and Botrylloides violaceus can smother rock habitats (Bishop, 2011b and 2012b). A significant potential INIS is the Pacific oyster Magallana gigas, as its distribution and environmental tolerances are considered to overlap with this biotope and this reef forming species can alter habitat structure. This species may also affect the grazers present in the biotope. No evidence was found for effects on rock pools, although on the Mediterranean coast Magallana gigas is cultivated in micro-tidal lagoons and has established wild populations (Miossec et al., 2009, cited from Herbert et al., 2012). In the Wadden Sea and North Sea, Magallana gigas overgrows mussel beds in the intertidal zone (Diederich, 2005, 2006; Kochmann et al., 2008), although larvae did show preference for settling on conspecifics before the mussels, and struggled to settle on mussels with a fucoid covering. It has been observed that mussel beds in the Wadden Sea that are adjacent to oyster farms were quickly converted to oyster beds (Kochmann et al., 2008). Dense aggregations of Magallana gigas on a former mussel bed showed increased abundance and biomass of Littorina littorea in the Wadden Sea (Markert et al., 2010). However, Eschweiler and Buschbaum (2011) found that juvenile Littorina littorea could carry Magallana gigas and Crepidula fornicata as epibionts. Body dry weight of snails without oyster overgrowth was twice as high compared to winkles covered with oysters. Also crawling speed of snails with oyster epigrowth was significantly slowed down and about ten times lower than in unfouled periwinkles. Additionally, oyster epibionts caused a strong decrease in reproductive output. In laboratory experiments, egg production of fouled Littorina littorea was about 100-fold lower than in affected individuals. Field surveys in different years and habitats demonstrated that up to 10% of individuals occurring on epibenthic bivalve beds and up to 25% of snails living on sand flats may be fouled by Magallana gigas.
The non-native crab Hemigrapsus sanguineus has recently been recorded in the UK (Sweet & Sewell, 2014) and has the potential to be a significant predator of intertidal invertebrates. Significant reductions in common shore crab abundance and mussel density have been reported where the Asian shore crab has achieved high densities in mainland Europe (Sweet & Sewell, 2014). In Rye, New York, declines of approximately 80% of Littorina littorea in the intertidal were reported to coincide with an expansion of the Hemigrapsus sanguineus population (Kraemer et al., 2007). This crab occurs on exposed shores and may therefore occur in this biotope when established. If predation of littorinids was significantly increased this could impact the algal composition and abundance of this biotope by altering the level of grazing pressure.
Sensitivity assessment. Little evidence was found to assess the impact of INIS on this biotope and much of the evidence comes from intertidal habitats in other countries. The conversion of this biotope to a Magallana gigas reef would present a significantly negative impact. Replacement of red algal turfs by other similar species may lead to some subtle effects on local ecology but at low abundances the biotope would still be recognisable from the description. An increase in Sargassum muticum would lead to reclassification to the specific biotope. Based on Magallana gigas and Sargassum muticum, resistance to this pressure is assessed as ‘Low’. The biotope will only recover if these species are removed, either through active management or natural processes. To recognise that recovery may be prolonged, resilience is assessed as ‘Very Low’ and sensitivity is therefore assessed as ‘High’.
Several coralline and non-coralline species are epiphytic on Corallina officinalis. Irvine and Chamberlain (1994) cite tissue destruction caused by Titanoderma corallinae. However, no information on pathogenic organisms in the UK was found. In Rhodophycota, viruses have been identified by means of electron microscopy (Lee, 1971) and they are probably widespread. However, nothing is known of their effects on growth or reproduction in red algae and experimental transfer from an infected to an uninfected specimen has not been achieved (Dixon & Irvine, 1977). Corallina officinalis, like many other algal species has been demonstrated to produce antibacterial substances (Taskin et al., 2007).
Diseased encrusting corallines were first observed in the tropics in the early 1990’s when the bacterial pathogen Coralline Lethal Orange Disease (CLOD) was discovered (Littler & Littler, 1995). All species of articulated and crustose species tested to date are easily infected by CLOD and it has been increasing in occurrence at sites where first observed and spreading through the tropics. Another bacterial pathogen causing a similar CLOD disease has been observed with a greater distribution and a black fungal pathogen first discovered in American Samoa has been dispersing (Littler & Littler, 1998). An unknown pathogen has also been reported to lead to white ‘target-shaped’ marks on corallines, again in the tropic (Littler et al., 2007). No evidence was found that these are impacting temperate coralline habitats.
Sea urchins are susceptible to 'bald-sea-urchin disease', which causes lesions, loss of spines, tube feet, pedicellariae, destruction of the upper layer of skeletal tissue and death. The disease has been associated with mass mortalities of Paracentrotus lividus in the French Mediterranean with the estimated loss of between 10 and 74% of populations re-surveyed (Boudouresque et al., 1980). An investigation by Becker et al. (2008) concluded that opportunistic bacteria in the water column are responsible for the symptoms of this disease.
Other species associated with this biotope such as littorinids and other algae also experience low levels of infestation by pathogens but mass-mortalities have not been recorded. For example, parasitism by trematodes may cause sterility in Littorina littorea. Littorina littorea are also parasitized by the boring polychaete, Polydora ciliata and Cliona sp, which weakens the shell and increases crab predation. Outbreaks of the shellfish pathogen Martelia spp. may cause widespread mortality of Mytilus edulis (Mainwaring et al., 2014) but populations within the UK have not been significantly impacted.
Sensitivity assessment. Based on the available evidence the characterizing coralline crusts and Corallina officinalis biotope is considered to have ‘High’ resistance and hence ‘High’ resilience and is classed as ‘Not sensitive’ at the pressure benchmark. However, where the biotope is defined by the presence of Paracentrotus lividus, an outbreak of bacterial pathogens may lead to mass mortalities and subsequent changes in algal abundance and composition. In that instance biotope resistance is assessed as ‘Low’ and resilience as ‘Medium’, sensitivity is therefore assessed as ‘Medium’. The assessment based on Paracentrotus lividus is presented in the sensitivity assessment table.
Direct, physical impacts from harvesting are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of target species on this biotope. The key characterizing and structuring species Corallina officinalis is collected for medical purposes; the fronds are dried and converted to hydroxyapatite and used as bone forming material (Ewers et al., 1987). It is also sold as a powder for use in the cosmetic industry. Some species present in the biotope may also be targeted. The blue mussel Mytilus edulis is too small and patchy in this biotope to be targeted for commercial harvesting. However, some, unregulated recreational hand-gathering of this species may occur.
The key characterizing species Paracentrotus lividus is harvested commercially in Europe and in Ireland. Populations in Ireland have been depleted (Byrne, 1990; Moylan, 1997) and although some variation in recruitment may contribute the probable cause is overfishing (Andrews et al., 2002). As individuals are large, easily visible in the pools they inhabit and have limited mobility collectors will have little difficulty removing populations of suitably sized individuals from pools. Recently Himanthalia elongata, Fucus serratus and other brown seaweeds have colonized new areas of Lough Hyne (Northern Ireland) following the mass mortality of the Paracentrotus lividus (Trowbridge et al., 2013), demonstrating the effects that targeted removal of large numbers of this species could have on biotope character.
Littorinids are one of the most commonly harvested species of the rocky shore. Large scale removal of Littorina littorea may allow a proliferation of opportunistic green algae, such as Ulva, on which it preferentially feeds. Experiments designed to test the effects of harvesting by removing individuals at Strangford Lough found that there was no effect of experimental treatments (either harvesting or simulated disturbance) on Littorina littorea abundance or body size over a 12 week period (Crossthwaite et al., 2012). This suggests that these animals are generally abundant and highly mobile; thus, animals that were removed were quickly replaced by dispersal from surrounding, un-harvested areas. However, long-term exploitation, as inferred by background levels of harvest intensity, did significantly influence population abundance and age structure (Crossthwaite et al., 2012). A broadscale study of harvesting in Ireland using field studies and interviews with wholesalers and pickers did suggest that some areas were over harvested but the lack of background data and quantitative records make this assertion difficult to test (Cummins et al., 2002). Similarly experimental grazer removal (manual removal of all gastropods, including Littorina littorea, in pool and a 1 m surrounding perimeter) caused strong and highly significant changes in assemblage structure in rockpools that contained red turf forming algae mainly due to an increase in the cover of green filamentous algae and a decrease in cover of live crustose coralline algae (25.40%) (Atalah & Crowe, 2010). Red and green algae may also be collected, Mastocarpus stellatus, for example is harvested to produce carrageen.
Sensitivity assessment. Collection of the key characterizing species would significantly alter the character and structure of the biotope and result in the loss of species inhabiting the turf. Collection of the urchin Paracentrotus lividus and/or littorinid grazers may allow red and green algae to increase in abundance and density, resulting in the biotope differing from the description. However, these algae may also be subject to harvesting limiting their dominance. The resistance of this biotope to targeted harvesting of characterizing and associated species is ‘Low’ as the species are all relatively large, conspicuous and easily collected. Resilience is assessed ‘Medium’ and sensitivity is assessed as ‘Medium’.
Incidental removal of the key characterizing species and associated species would alter the character of the biotope. The biotope is characterized by dense turfs of Corallina officinalis, these provide habitat and attachment surfaces for epiphytic species and where these trap sediments also provide a habitat for associated species. The loss of the turf due to incidental removal as by-catch would therefore alter the character of the habitat and result in the loss of species richness. The ecological services such as primary and secondary production provided by the characterizing and associated species would also be lost.
Removal of grazers, including the key characterizing Paracentrotus lividus could result in changes in algal composition with an increase in macroalgal abundance and/or species present, leading to reclassification of the biotope. Recently, Himanthalia elongata, Fucus serratus and other brown seaweeds have colonized new areas of Lough Hyne (Northern Ireland) following the mass mortality of Paracentrotus lividus (Trowbridge et al., 2013). Similarly, experimental grazer removal (manual removal of all gastropods in pool and a 1m surrounding perimeter) caused strong and highly significant changes in assemblage structure in rockpools that contained red turf forming algae mainly due to an increase in the cover of green filamentous algae and a decrease in cover of live crustose coralline algae (25.40%) (Atalah & Crowe, 2010).
Sensitivity assessment. Removal of a large percentage of the characterizing turf and encrusting coralline algae species and grazers would alter the character of the biotope, species richness and ecosystem function. Resistance is therefore assessed as ‘Low’ and recovery as ‘Medium’, so that sensitivity is assessed as 'Medium'.
Adey, W.H. & Adey, P.J., 1973. Studies on the biosystematics and ecology of the epilithic crustose corallinacea of the British Isles. British Phycological Journal, 8, 343-407.
Airoldi, L., 2003. The effects of sedimentation on rocky coast assemblages. Oceanography and Marine Biology: An Annual Review, 41,161-236
Airoldi, L., 2000. Responses of algae with different life histories to temporal and spatial variability of disturbance in subtidal reefs. Marine Ecology Progress Series, 195 (8), 81-92.
Airoldi, L. & Hawkins, S.J., 2007. Negative effects of sediment deposition on grazing activity and survival of the limpet Patella vulgata. Marine Ecology Progress Series, 332, 235-240.
Albrecht, A. & Reise, K., 1994. Effects of Fucus vesiculosus covering intertidal mussel beds in the Wadden Sea. Helgoländer Meeresuntersuchungen, 48 (2-3), 243-256.
Albrecht, A.S., 1998. Soft bottom versus hard rock: Community ecology of macroalgae on intertidal mussel beds in the Wadden Sea. Journal of Experimental Marine Biology and Ecology, 229 (1), 85-109.
Allain, J.-Y., 1975. Structure des populations de Paracentrotus lividus (Lamarck)(Echinodermata, Echinoidea) soumises a la peche sur les cotes nord de Bretagne. Revue des Travaux de l'Institut des Pêches Maritimes, 39 (2), 171-212.
Almada-Villela P.C., 1984. The effects of reduced salinity on the shell growth of small Mytilus edulis L. Journal of the Marine Biological Association of the United Kingdom, 64, 171-182.
Andrake, W. & Johansen, H.W., 1980. Alizarin red dye as a marker for measuring growth in Corallina officinalis L. (Corallinaceae, Rhodophyta). Journal of Phycology, 16 (4), 620-622.
Andrew, N., Agatsuma, Y., Ballesteros, E., Bazhin, A., Creaser, E., Barnes, D., Botsford, L., Bradbury, A., Campbell, A. & Dixon, J., 2002. Status and management of world sea urchin fisheries. Oceanography and Marine Biology an Annual Review, 40, 343-425.
Arévalo, R., Pinedo, S. & Ballesteros, E., 2007. Changes in the composition and structure of Mediterranean rocky-shore communities following a gradient of nutrient enrichment: descriptive study and test of proposed methods to assess water quality regarding macroalgae. Marine Pollution Bulletin, 55 (1), 104-113.
Atalah, J. & Crowe, T.P., 2010. Combined effects of nutrient enrichment, sedimentation and grazer loss on rock pool assemblages. Journal of Experimental Marine Biology and Ecology, 388 (1), 51-57.
Bailey, J., Parsons, J. & Couturier, C., 1996. Salinity tolerance in the blue mussel, Mytilus edulis. Rep. Report no. 0840-5417, Aquaculture Association of Canada, New Brunswick, Canada
Bamber, R.N. & Irving, P.W., 1993. The Corallina run-offs of Bridgewater Bay. Porcupine Newsletter, 5, 190-197.
Barnes, D. K. A., Steele, S., Maguire, D. & Turner, J., 1999. Population dynamics of the urchin Paracentrotus lividus at Lough Hyne, Ireland. In Echinoderm research, M. D. C. Carnevali & F. Bonasoro (eds). Rotterdam: Balkema, 427– 431.
Barnes, H., Finlayson, D.M. & Piatigorsky, J., 1963. The effect of desiccation and anaerobic conditions on the behaviour, survival and general metabolism of three common cirripedes. Journal of Animal Ecology, 32, 233-252.
Becker, P.T., Egea, E. & Eeckhaut, I., 2008. Characterization of the bacterial communities associated with the bald sea urchin disease of the echinoid Paracentrotus lividus. Journal of Invertebrate Pathology, 98 (2), 136-147.
Bellgrove, A., Clayton, M.N. & Quinn, G., 1997. Effects of secondarily treated sewage effluent on intertidal macroalgal recruitment processes. Marine and Freshwater Research, 48 (2), 137-146.
Bellgrove, A., McKenzie, P.F., McKenzie, J.L. & Sfiligoj, B.J., 2010. Restoration of the habitat-forming fucoid alga Hormosira banksii at effluent-affected sites: competitive exclusion by coralline turfs. Marine Ecology Progress Series, 419, 47-56.
Bishop J. 2011b.Orange-tipped sea squirt, Corella eumyota. Great Britain Non-native Species Secretariat. [cited 16/06/2015]. Available from: http://www.nonnativespecies.org
Bishop, J. 2012c. Carpet Sea-squirt, Didemnum vexillum.Great Britain Non-native Species Secretariat [On-line]. [cited 16/06/2015]. Available from: <http://www.nonnativespecies.org
Bishop, J. 2012b. Botrylloides violaceus. Great Britain Non-native Species Secretariat. [On-line] [cited 16/06/2015]. Available from:
Bishop, J. 2015b. Watersipora subatra. Great Britain Non-native Species Secretariat. [On-line][cited 16/06/2015]. Available from: <http://www.nonnativespecies.org
Bishop, J. 2015a. Compass sea squirt, Asterocarpa humilis. Great Britain Non-native Species Secretariat. [On-line] [cited 16/06/2015]. Available from: <http://www.nonnativespecies.org
Blake, C. & Maggs, C.A., 2003. Comparative growth rates and internal banding periodicity of maerl species (Corallinales, Rhodophyta) from northern Europe. Phycologia, 42 (6), 606-612.
Boney, A.D., 1971. Sub-lethal effects of mercury on marine algae. Marine Pollution Bulletin, 2, 69-71.
Boolootian, R.A.,1966. Physiology of Echinodermata. (Ed. R.A. Boolootian), pp. 822-822. New York: John Wiley & Sons.
Boudouresque, C.F. & Verlaque, M., 2001. Ecology of Paracentrotus lividus. Developments in Aquaculture and Fisheries Science, 32, 177-216.
Boudouresque, C.-F., Nedelec, H. & Shepherd, S., 1980. The decline of a population of the sea urchin Paracentrotus lividus in the Bay of Port-Cros (Var, France). Scientific Reports of Port-Cros National Park, 6-243.
Bowman, R.S. & Lewis, J.R., 1977. Annual fluctuations in the recruitment of Patella vulgata L. Journal of the Marine Biological Association of the United Kingdom, 57, 793-815.
Bowman, R.S., 1981. The morphology of Patella spp. juveniles in Britain, and some phylogenetic inferences. Journal of the Marine Biological Association of the United Kingdom, 61, 647-666.
Bradshaw, C., Veale, L.O., Hill, A.S. & Brand, A.R., 2000. The effects of scallop dredging on gravelly seabed communities. In: Effects of fishing on non-target species and habitats (ed. M.J. Kaiser & de S.J. Groot), pp. 83-104. Oxford: Blackwell Science.
Brawley, S.H., 1992b. Mesoherbivores. In Plant-animal interactions in the marine benthos (ed. D.M John, S.J. Hawkins & J.H. Price), pp. 235-263. Oxford: Clarendon Press. [Systematics Association Special Volume, no. 46.]
Brosnan, D.M. & Crumrine, L.L., 1994. Effects of human trampling on marine rocky shore communities. Journal of Experimental Marine Biology and Ecology, 177, 79-97.
Brown, P.J. & Taylor, R.B., 1999. Effects of trampling by humans on animals inhabiting coralline algal turf in the rocky intertidal. Journal of Experimental Marine Biology and Ecology, 235, 45-53.
Brown, V., Davies, S. & Synnot, R., 1990. Long-term monitoring of the effects of treated sewage effluent on the intertidal macroalgal community near Cape Schanck, Victoria, Australia. Botanica Marina, 33 (1), 85-98.
Bryan, G.W. & Gibbs, P.E., 1983. Heavy metals from the Fal estuary, Cornwall: a study of long-term contamination by mining waste and its effects on estuarine organisms. Plymouth: Marine Biological Association of the United Kingdom. [Occasional Publication, no. 2.]
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.
Bryan, G.W., Langston, W.J., Hummerstone, L.G., Burt, G.R. & Ho, Y.B., 1983. An assessment of the gastropod Littorina littorea (L.) as an indicator of heavy metal contamination in United Kingdom estuaries. Journal of the Marine Biological Association of the United Kingdom, 63, 327-345.
Bulleri, F. & Airoldi, L., 2005. Artificial marine structures facilitate the spread of a non‐indigenous green alga, Codium fragile ssp. tomentosoides, in the north Adriatic Sea. Journal of Applied Ecology, 42 (6), 1063-1072.
Bulleri, F. & Benedetti-Cecchi, L., 2008. Facilitation of the introduced green alga Caulerpa racemosa by resident algal turfs: experimental evaluation of underlying mechanisms. Marine Ecology Progress Series, 364, 77-86.
Burdin, K.S. & Bird, K.T., 1994. Heavy metal accumulation by carrageenan and agar producing algae. Botanica Marina, 37, 467-470.
Burdon, D., Dawes, O., Eades, R., Leighton, A., Musk, M. & Thompson, S.2009. BEEMS WP6 Intertidal Studies; Hinkley Survey-Report to Cefas. Institute of Estuarine and Coastal Studies, University of Hull.
Byrne, M., 1990. Annual reproductive cycles of the commercial sea urchin Paracentrotus lividus from an exposed intertidal and a sheltered subtidal habitat on the west coast of Ireland. Marine Biology, 104 (2), 275-289.
Carlson, R.L., Shulman, M.J. & Ellis, J.C., 2006. Factors Contributing to Spatial Heterogeneity in the Abundance of the Common Periwinkle Littorina Littorea (L.). Journal of Molluscan Studies, 72 (2), 149-156.
Carvalho, G.R., 1989. Microgeographic genetic differentiation and dispersal capacity in the intertidal isopod, Jaera albifrons Leach. In Proceedings of the 23rd European Marine Biology Symposium, Swansea, 5-9 September 1988. Reproduction, Genetics and Distribution of Marine Organisms (ed. J.S. Ryland & P.A. Tyler), pp. 265-271. Denmark: Olsen & Olsen.
Chamberlain, Y.M., 1996. Lithophylloid Corallinaceae (Rhodophycota) of the genera Lithophyllum and Titausderma from southern Africa. Phycologia, 35, 204-221.
Chandrasekara, W.U. & Frid, C.L.J., 1998. A laboratory assessment of the survival and vertical movement of two epibenthic gastropod species, Hydrobia ulvae, (Pennant) and Littorina littorea (Linnaeus), after burial in sediment. Journal of Experimental Marine Biology and Ecology, 221, 191-207.
Choat, J.H. & Kingett, P.D., 1982. The influence of fish predation on the abundance cycles of an algal turf invertebrate fauna. Oecologia, 54, 88-95.
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.], http://www.ukmarinesac.org.uk/
Colhart, B.J., & Johanssen, H.W., 1973. Growth rates of Corallina officinalis (Rhodophyta) at different temperatures. Marine Biology, 18, 46-49.
Connor, D.W., Allen, J.H., Golding, N., Howell, K.L., Lieberknecht, L.M., Northen, K.O. & Reker, J.B., 2004. The Marine Habitat Classification for Britain and Ireland. Version 04.05. Joint Nature Conservation Committee, Peterborough. www.jncc.gov.uk/MarineHabitatClassification.
Connor, D.W., Brazier, D.P., Hill, T.O., & Northen, K.O., 1997b. Marine biotope classification for Britain and Ireland. Vol. 1. Littoral biotopes. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 229, Version 97.06., Joint Nature Conservation Committee, Peterborough, JNCC Report No. 230, Version 97.06.
Cook, E. & Kelly, M., 2007. Enhanced production of the sea urchin Paracentrotus lividus in integrated open-water cultivation with Atlantic salmon Salmo salar. Aquaculture, 273 (4), 573-585.
Coull, B.C. & Wells, J.B.J., 1983. Refuges from fish predation: experiments with phytal meiofauna from the New Zealand rocky intertidal. Ecology, 64, 1599-1609.
Crapp, G. & Willis, M., 1975. Age determination in the sea urchin Paracentrotus lividus (Lamarck), with notes on the reproductive cycle. Journal of Experimental Marine Biology and Ecology, 20 (2), 157-178.
Crisp, D.J. & Mwaiseje, B., 1989. Diversity in intertidal communities with special reference to the Corallina officinalis community. Scientia Marina, 53, 365-372.
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.
Crossthwaite, S.J., Reid, N. & Sigwart, J.D., 2012. Assessing the impact of shore-based shellfish collection on under-boulder communities in Strangford Lough. Report prepared by the Natural Heritage Research Partnership (NHRP) between Quercus, Queen’s University Belfast and the Northern Ireland Environment Agency (NIEA) for the Research and Development Series No. 13/03.
Crump, R.G., Morley, H.S., & Williams, A.D., 1999. West Angle Bay, a case study. Littoral monitoring of permanent quadrats before and after the Sea Empress oil spill. Field Studies, 9, 497-511.
Cummins, V., Coughlan, S., McClean, O., Connolly, N., Mercer, J. & Burnell, G., 2002. An assessment of the potential for the sustainable development of the edible periwinkle, Littorina littorea, industry in Ireland.Report by the Coastal and Marine Resources Centre, Environmental Research Institute, University College Cork.
Davenport, J. & Davenport, J.L., 2005. Effects of shore height, wave exposure and geographical distance on thermal niche width of intertidal fauna. Marine Ecology Progress Series, 292, 41-50.
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.
Davies, M.S., 1992. Heavy metals in seawater: effects on limpet pedal mucus production. Water Research, 26, 1691-1693.
Davies, S.P., 1970. Physiological ecology of Patella IV. Environmental and limpet body temperatures. Journal of the Marine Biological Association of the United Kingdom, 50 (04), 1069-1077.
Delany, J., Myers, A.A. & McGrath, D., 1998. Recruitment, immigration and population structure of two coexisting limpet species in mid-shore tidepools, on the west coast of Ireland. Journal of Experimental Marine Biology and Ecology, 221, 221-230.
Dethier, M.N., 1994. The ecology of intertidal algal crusts: variation within a functional group. Journal of Experimental Marine Biology and Ecology, 177 (1), 37-71.
Diederich, S., 2005. Differential recruitment of introduced Pacific oysters and native mussels at the North Sea coast: coexistence possible? Journal of Sea Research, 53 (4), 269-281.
Diederich, S., 2006. High survival and growth rates of introduced Pacific oysters may cause restrictions on habitat use by native mussels in the Wadden Sea. Journal of Experimental Marine Biology and Ecology, 328 (2), 211-227.
Dixon, P.S. & Irvine, L.M., 1977. Seaweeds of the British Isles. Volume 1 Rhodophyta. Part 1 Introduction, Nemaliales, Gigartinales. London: British Museum (Natural History) London.
Dodds, W.K. & Gudder, D.A., 1992. The ecology of Cladophora. Journal of Phycology, 28, 415-427.
Dommasnes, A., 1968. Variation in the meiofauna of Corallina officinalis with wave exposure. Sarsia, 34, 117-124.
Dommasnes, A., 1969. On the fauna of Corallina officinalis L. in western Norway. Sarsia, 38, 71-86.
Dudgeon, S.R., Davison, I.R. & Vadas, R.L., 1990. Freezing tolerance in the intertidal red algae Chondrus crispus and Mastocarpus stellatus: relative importance of acclimation and adaptation. Marine Biology, 106, 427-436.
Duran, S., Palacin, C., Becerro, M.A., Turon, X. & Giribet, G., 2004. Genetic diversity and population structure of the commercially harvested sea urchin Paracentrotus lividus (Echinodermata, Echinoidea). Molecular Ecology, 13 (11), 3317-3328.
Ebere, A.G. & Akintonwa, A., 1992. Acute toxicity of pesticides to Gobius sp., Palaemonetes africanus, and Desmocaris trispimosa. Bulletin of Environmental Contamination and Toxicology, 49, 588-592.
Edyvean, R. & Ford, H., 1986. Spore production by Lithophyllum incrustans (Corallinales, Rhodophyta) in the British Isles. British Phycological Journal, 21 (3), 255-261.
Edyvean, R. & Ford, H., 1987. Growth rates of Lithophyllum incrustans (Corallinales, Rhodophyta) from south west Wales. British Phycological Journal, 22 (2), 139-146.
Edyvean, R.G.J. & Ford, H., 1984a. Population biology of the crustose red alga Lithophyllum incrustans Phil. 2. A comparison of populations from three areas of Britain. Biological Journal of the Linnean Society, 23 (4), 353-363.
Edyvean, R.G.J. & Ford, H., 1984b. Population biology of the crustose red alga Lithophyllum incrustans Phil. 3. The effects of local environmental variables. Biological Journal of the Linnean Society, 23, 365-374.
Ekaratne, S.U.K. & Crisp, D.J., 1984. Seasonal growth studies of intertidal gastropods from shell micro-growth band measurements, including a comparison with alternative methods. Journal of the Marine Biological Association of the United Kingdom, 64, 183-210.
Eschweiler, N. & Buschbaum, C., 2011. Alien epibiont (Crassostrea gigas) impacts on native periwinkles (Littorina littorea). Aquatic Invasions, 6 (3), 281-290.
Evans, R.G., 1948. The lethal temperatures of some common British littoral molluscs. The Journal of Animal Ecology, 17, 165-173.
Ewers, R., Kasperk, C. & Simmons, B., 1987. Biologishes Knochenimplantat aus Meeresalgen. Zahnaerztliche Praxis, 38, 318-320.
Fernandez, C., Pasqualini, V., Boudouresque, C.-F., Johnson, M., Ferrat, L., Caltagirone, A. & Mouillot, D., 2006. Effect of an exceptional rainfall event on the sea urchin (Paracentrotus lividus) stock and seagrass distribution in a Mediterranean coastal lagoon. Estuarine, Coastal and Shelf Science, 68 (1), 259-270.
Firth, L., Thompson, R., Bohn, K., Abbiati, M., Airoldi, L., Bouma, T., Bozzeda, F., Ceccherelli, V., Colangelo, M. & Evans, A., 2014. Between a rock and a hard place: Environmental and engineering considerations when designing coastal defence structures. Coastal Engineering, 87, 122-135.
Fish, J.D. & Fish, S., 1996. A student's guide to the seashore. Cambridge: Cambridge University Press.
Foster, B.A., 1970. Responses and acclimation to salinity in the adults of some balanomorph barnacles. Philosophical Transactions of the Royal Society of London, Series B, 256, 377-400.
Frazer, A.W.J., Brown, M.T. & Bannister, P., 1988. The frost resistance of some littoral and sub-littoral algae from southern New Zealand. Botanica Marina, 31, 461-464.
Fretter, V. & Graham, A., 1994. British prosobranch molluscs: their functional anatomy and ecology, revised and updated edition. London: The Ray Society.
Gallagher, M.C., Davenport, J., Gregory, S., McAllen, R. & O'Riordan, R., 2015. The invasive barnacle species, Austrominius modestus: Its status and competition with indigenous barnacles on the Isle of Cumbrae, Scotland. Estuarine, Coastal and Shelf Science, 152, 134-141.
Gomes-Filho, J., Hawkins, S., Aquino-Souza, R. & Thompson, R., 2010. Distribution of barnacles and dominance of the introduced species Elminius modestus along two estuaries in South-West England. Marine Biodiversity Records, 3, e58.
Grahame, J., & Hanna, F.S., 1989. Factors affecting the distribution of the epiphytic fauna of Corallina officinalis (L.) on an exposed rocky shore. Ophelia, 30, 113-129.
Grandy, N., 1984. The effects of oil and dispersants on subtidal red algae. Ph.D. Thesis. University of Liverpool.
Green, D., Chapman, M. & Blockley, D., 2012. Ecological consequences of the type of rock used in the construction of artificial boulder-fields. Ecological Engineering, 46, 1-10.
Grenon, J.F. & Walker, G., 1981. The tenacity of the limpet, Patella vulgata L.: an experimental approach. Journal of Experimental Marine Biology and Ecology, 54, 277-308.
Gressner, F. & Schramm, W., 1971. Salinity - Plants. In Marine Ecology. A comprehensive, integrated treatise on life in oceans and coastal waters. Vol. 1 Environmental Factors. Part 2. (ed. O. Kinne), pp. 705-820. London: John Wiley & Sons.
Guarnieri, G., Terlizzi, A., Bevilacqua, S. & Fraschetti, S., 2012. Increasing heterogeneity of sensitive assemblages as a consequence of human impact in submarine caves. Marine Biology, 159 (5), 1155-1164.
Guiry, M.D. & Blunden, G., 1991. Seaweed Resources in Europe: Uses and Potential. Chicester: John Wiley & Sons.
Guiry, M.D. & Guiry, G.M. 2015. AlgaeBase [Online], National University of Ireland, Galway [cited 30/6/2015]. Available from: http://www.algaebase.org/
Hagerman, L., 1968. The ostracod fauna of Corallina officinalis L. in western Norway. Sarsia, 36, 49-54.
Hall-Spencer, J.M. & Moore, P.G., 2000a. Impact of scallop dredging on maerl grounds. In Effects of fishing on non-target species and habitats. (ed. M.J. Kaiser & S.J., de Groot) 105-117. Oxford: Blackwell Science.
Harlin, M.M., & Lindbergh, J.M., 1977. Selection of substrata by seaweed: optimal surface relief. Marine Biology, 40, 33-40.
Hawkins, S.J. & Harkin, E., 1985. Preliminary canopy removal experiments in algal dominated communities low on the shore and in the shallow subtidal on the Isle of Man. Botanica Marina, 28, 223-30.
Hawkins, S.J. & Hartnoll, R.G., 1985. Factors determining the upper limits of intertidal canopy-forming algae. Marine Ecology Progress Series, 20, 265-271.
Hawkins, S.J., Proud, S.V., Spence, S.K. & Southward, A.J., 1994. From the individual to the community and beyond: water quality, stress indicators and key species in coastal systems. In Water quality and stress indicators in marine and freshwater ecosystems: linking levels of organisation (individuals, populations, communities) (ed. D.W. Sutcliffe), 35-62. Ambleside, UK: Freshwater Biological Association.
Herbert, R.J.H., Roberts, C., Humphreys, J., & Fletcher, S. 2012. The Pacific oyster (Crassostra gigas) in the UK: economic, legal and environmental issues associated with its cultivation, wild establishment and exploitation. Available from: http://www.dardni.gov.uk/pacific-oysters-issue-paper.pdf
Hicks, G.R.F. & Coull, B.C., 1983. The ecology of marine meiobenthic harpacticoid copepods. Oceanography and Marine Biology: an Annual Review, 21, 67-175.
Hicks, G.R.F., 1980. Structure of phytal harpacticoid copepod assemblages and the influence of habitat complexity and turbidity. Journal of Experimental Marine Biology and Ecology, 44, 17-192.
Hicks, G.R.F., 1985. Meiofauna associated with rocky shore algae. In The Ecology of Rocky Coasts: essays presented to J.R. Lewis, D.Sc., (ed. P.G. Moore & R. Seed, ed.). pp. 36-56. London: Hodder & Stoughton Ltd.
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.
Holt, T.J., Hartnoll, R.G. & Hawkins, S.J., 1997. The sensitivity and vulnerability to man-induced change of selected communities: intertidal brown algal shrubs, Zostera beds and Sabellaria spinulosa reefs. English Nature, Peterborough, English Nature Research Report No. 234.
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.
Holt, T.J., Rees, E.I., Hawkins, S.J. & Seed, R., 1998. Biogenic reefs (Volume IX). An overview of dynamic and sensitivity characteristics for conservation management of marine SACs. Scottish Association for Marine Science (UK Marine SACs Project), 174 pp.
Hong, J. & Reish, D.J., 1987. Acute toxicity of cadmium to eight species of marine amphipod and isopod crustaceans from southern California. Bulletin of Environmental Contamination and Toxicology, 39, 884-888.
Huff, T.M. & Jarett, J.K., 2007. Sand addition alters the invertebrate community of intertidal coralline turf. Marine Ecology Progress Series, 345, 75-82.
Irvine, L. M. & Chamberlain, Y. M., 1994. Seaweeds of the British Isles, vol. 1. Rhodophyta, Part 2B Corallinales, Hildenbrandiales. London: Her Majesty's Stationery Office.
Irvine, L.M., 1983. Seaweeds of the British Isles vol. 1. Rhodophyta Part 2A. Cryptonemiales (sensu stricto), Palmariales, Rhodymeniales. London: British Museum (Natural History).
Iuri, V., Patti, F. & Procaccini, G., 2007. Phylogeography of the sea urchin Paracentrotus lividus (Lamarck)(Echinodermata: Echinoidea): first insights from the South Tyrrhenian Sea. Biodiversity in Enclosed Seas and Artificial Marine Habitats: Springer, pp. 77-84.
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
Johansen, W.H., 1974. Articulated coralline algae. Oceanography and Marine Biology: an Annual Review, 12, 77-127.
Johansson ,G., Eriksson, B.K., Pedersen, M. & Snoeijs, P., 1998. Long term changes of macroalgal vegetation in the Skagerrak area. Hydrobiologia, 385, 121-138.
Kain, J.M., 1975a. Algal recolonization of some cleared subtidal areas. Journal of Ecology, 63, 739-765.
Kaiser, M.J., Ramsay, K., Richardson, C.A., Spence, F.E. & Brand, A.R., 2000. Chronic fishing disturbance has changed shelf sea benthic community structure. Journal of Animal Ecology, 69, 494-503.
Kempf, M., 1962. Recherches d’écologie comparée sur Paracentrotus lividus (Lmk.) et Arbacia lixula (L.). Recueil des Travaux de la Station Marine d'Endoume, 25 (39), 47-115.
Kendrick, G.A., 1991. Recruitment of coralline crusts and filamentous turf algae in the Galapagos archipelago: effect of simulated scour, erosion and accretion. Journal of Experimental Marine Biology and Ecology, 147 (1), 47-63
Kennelly, S.J., 1987. Inhibition of kelp recruitment by turfing algae and consequences for an Australian kelp community. Journal of Experimental Marine Biology and Ecology, 112 (1), 49-60.
Kindig, A.C., & Littler, M.M., 1980. Growth and primary productivity of marine macrophytes exposed to domestic sewage effluents. Marine Environmental Research, 3, 81-100.
Kinne, O. (ed.), 1971a. Marine Ecology: A Comprehensive, Integrated Treatise on Life in Oceans and Coastal Waters. Vol. 1 Environmental Factors, Part 2. Chichester: John Wiley & Sons.
Kitching, J. & Ebling, F., 1961. The ecology of Lough Ine. The Journal of Animal Ecology, 373-383.
Kochmann, J., Buschbaum, C., Volkenborn, N. & Reise, K., 2008. Shift from native mussels to alien oysters: differential effects of ecosystem engineers. Journal of Experimental Marine Biology and Ecology, 364 (1), 1-10.
Kraemer, G.P., Sellberg, M., Gordon, A. & Main, J., 2007. Eight-year record of Hemigrapsus sanguineus (Asian shore crab) invasion in western Long Island sound estuary. Northeastern Naturalist, 14 (2), 207-224.
Last, K.S., Hendrick V. J, Beveridge C. M & Davies A. J, 2011. Measuring the effects of suspended particulate matter and smothering on the behaviour, growth and survival of key species found in areas associated with aggregate dredging. Report for the Marine Aggregate Levy Sustainability Fund,
Latham, H., 2008. Temperature stress-induced bleaching of the coralline alga Corallina officinalis: a role for the enzyme bromoperoxidase. Bioscience Horizons, 1-10
Le Gall, P. & Bucaille, D., 1989. Sea urchins production by inland farming, in: De Pauw, N. et al. Ed.) (1989). Aquaculture a biotechnology in progress: volume 1. pp. 53-59.
Lee, R.E., 1971. Systemic viral material in the cells of the freshwater alga Sirodotia tenuissima (Holden) Skuja. Journal of Cell Science, 8, 623-631.
Lewis, J.R., 1964. The Ecology of Rocky Shores. London: English Universities Press.
Littler, M. & Littler, D., 1998. An undescribed fungal pathogen of reef-forming crustose corraline algae discovered in American Samoa. Coral Reefs, 17 (2), 144-144.
Littler, M. & Littler, D.S. 2013. The nature of crustose coralline algae and their interactions on reefs. Smithsonian Contributions to the Marine Sciences, 39, 199-212
Littler, M. & Murray, S., 1975. Impact of sewage on the distribution, abundance and community structure of rocky intertidal macro-organisms. Marine Biology, 30 (4), 277-291.
Littler, M. M., & Littler, D. S. 1984. Relationships between macroalgal functional form groups and substrata stability in a subtropical rocky-intertidal system. Journal of Experimental Marine Biology and Ecology, 74(1), 13-34.
Littler, M.M., 1973. The population and community structure of Hawaiian fringing-reef crustose Corallinaceae (Rhodophyta, Cryptonemiales). Journal of Experimental Marine Biology and Ecology, 11 (2), 103-120.
Littler, M.M. & Littler, D.S., 1995. Impact of CLOD pathogen on Pacific coral reefs. Science, 267, 1356-1356.
Littler, M.M., & Kauker, B.J., 1984. Heterotrichy and survival strategies in the red alga Corallina officinalis L. Botanica Marina, 27, 37-44.
Littler, M.M., Littler, D.S. & Brooks, B.L. 2007. Target phenomena on south Pacific reefs: strip harvesting by prudent pathogens? Reef Encounter, 34, 23-24
Littler, M.M., Murray, S.N. & Arnold, K.E., 1979. Seasonal variations in net photosynthetic performance and cover of intertidal macrophytes. Aquatic Botany, 7, 35-46.
Long, J.D., Cochrane, E. & Dolecal, R., 2011. Previous disturbance enhances the negative effects of trampling on barnacles. Marine Ecology Progress Series, 437, 165-173.
Lüning, K., 1990. Seaweeds: their environment, biogeography, and ecophysiology: John Wiley & Sons.
Luther, G., 1976. Bewuchsuntersuchungen auf Natursteinsubstraten im Gezeitenbereich des Nordsylter Wattenmeeres: Algen. Helgoländer Wissenschaftliche Meeresuntersuchungen, 28 (3-4), 318-351.
MacDonald, D.S., Little, M., Eno, N.C. & Hiscock, K., 1996. Disturbance of benthic species by fishing activities: a sensitivity index. Aquatic Conservation: Marine and Freshwater Ecosystems, 6 (4), 257-268.
Maggs, C.A. & Hommersand, M.H., 1993. Seaweeds of the British Isles: Volume 1 Rhodophycota Part 3A Ceramiales. London: Natural History Museum, Her Majesty's Stationary Office.
Mainwaring, K., Tillin, H. & Tyler-Walters, H., 2014. Assessing the sensitivity of blue mussel beds to pressures associated with human activities. Joint Nature Conservation Committee, JNCC Report No. 506., Peterborough, 96 pp.
Marchan, S., Davies, M.S., Fleming, S. & Jones, H.D., 1999. Effects of copper and zinc on the heart rate of the limpet Patella vulgata (L.) Comparative Biochemistry and Physiology, 123A, 89-93.
Marshall, D.J. & McQuaid, C.D., 1989. The influence of respiratory responses on the tolerance to sand inundation of the limpets Patella granularis L.(Prosobranchia) and Siphonaria capensis Q. et G.(Pulmonata). Journal of Experimental Marine Biology and Ecology, 128 (3), 191-201.
Martel, A. & Chia, F.S., 1991b. Drifting and dispersal of small bivalves and gastropods with direct development. Journal of Experimental Marine Biology and Ecology, 150, 131-147.
May, V., 1985. Observations on algal floras close to two sewerage outlets. Cunninghamia, 1, 385-394.
McGrath, D., 1992. Recruitment and growth of the Blue-rayed limpet, Helcion pellucidum in south east Ireland. Journal of Molluscan Studies, 58, 425-431.
Miossec, L., Le Deuff, R.-M. & Goulletquer, P., 2009. Alien species alert: Crassostrea gigas (Pacific oyster). ICES Cooperative Research Report, 299
Moss, B., Mercer, S., & Sheader, A., 1973. Factors Affecting the Distribution of Himanthalia elongata (L.) S.F. Gray on the North-east Coast of England. Estuarine and Coastal Marine Science, 1, 233-243.
Moylan, E., 1997. Gonad conditioning and wild stock enhancement of the purple sea urchin Paracentrotus lividus on the west coast of Ireland. Bulletin-Aquaculture Association of Canada, 38-41.
Newell, R.C., 1979. Biology of intertidal animals. Faversham: Marine Ecological Surveys Ltd.
Norton, T.A., 1992. Dispersal by macroalgae. British Phycological Journal, 27, 293-301.
O'Brien, P.J. & Dixon, P.S., 1976. Effects of oils and oil components on algae: a review. British Phycological Journal, 11, 115-142.
Palacı́n, C., Giribet, G., Carner, S., Dantart, L. & Turon, X., 1998. Low densities of sea urchins influence the structure of algal assemblages in the western Mediterranean. Journal of Sea Research, 39 (3), 281-290.
Pastor. R., 1971. Distribución del erizo de mar, Paracentrotus lividus (Lmk), en la Ría de Vigo. Publ Técn Dir Gen Pesca Marít, Spain, 9, 255-270.
Pedrotti, M.L., 1993. Spatial and temporal distribution and recruitment of echinoderm larvae in the Ligurian Sea. Journal of the Marine Biological Association of the United Kingdom, 73, 513-530.
Perkol-Finkel, S. & Airoldi, L., 2010. Loss and recovery potential of marine habitats: an experimental study of factors maintaining resilience in subtidal algal forests at the Adriatic Sea. PLoS One, 5 (5), e10791.
Pinn, E.H. & Rodgers, M., 2005. The influence of visitors on intertidal biodiversity. Journal of the Marine Biological Association of the United Kingdom, 85 (02), 263-268.
Povey, A. & Keough, M.J., 1991. Effects of trampling on plant and animal populations on rocky shores. Oikos, 61: 355-368.
Raffaelli, D. & Hawkins, S., 1999. Intertidal Ecology 2nd edn.. London: Kluwer Academic Publishers.
Reed, R.H. & Russell, G., 1979. Adaptation to salinity stress in populations of Enteromorpha intestinalis (L.) Link. Estuarine and Coastal Marine Science, 8, 251-258.
Riisgård, H.U., Bondo Christensen, P., Olesen, N.J., Petersen, J.K, Moller, M.M. & Anderson, P., 1993. Biological structure in a shallow cove (Kertinge Nor, Denmark) - control by benthic nutrient fluxes and suspension-feeding ascidians and jellyfish. Ophelia, 41, 329-344.
Robles, C., 1982. Disturbance and predation in an assemblage of herbivorous Diptera and algae on rocky shores. Oecologia, 54 (1), 23-31.
Sala, E., Boudouresque, C. & Harmelin-Vivien, M., 1998. Fishing, trophic cascades, and the structure of algal assemblages: evaluation of an old but untested paradigm. Oikos, 82, 425-439.
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.
Schiel, D.R. & Taylor, D.I., 1999. Effects of trampling on a rocky intertidal algal assemblage in southern New Zealand. Journal of Experimental Marine Biology and Ecology, 235, 213-235.
Schmidt, A.L. & Scheibling, R.E., 2007. Effects of native and invasive macroalgal canopies on composition and abundance of mobile benthic macrofauna and turf-forming algae. Journal of Experimental Marine Biology and Ecology, 341 (1), 110-130.
Seapy , R.R. & Littler, M.M., 1982. Population and Species Diversity Fluctuations in a Rocky Intertidal Community Relative to Severe Aerial Exposure and Sediment Burial. Marine Biology, 71, 87-96.
Seed, R. & Suchanek, T.H., 1992. Population and community ecology of Mytilus. In The mussel Mytilus: ecology, physiology, genetics and culture, (ed. E.M. Gosling), pp. 87-169. Amsterdam: Elsevier Science Publ. [Developments in Aquaculture and Fisheries Science, no. 25.]
Sewell, J. 2011c. Wireweed, Sargassum muticum. Great Britain Non-native Species Secretariat. [cited 16/06/2015]. Available from: <http://www.nonnativespecies.org
Shanks, A.L. & Wright, W.G., 1986. Adding teeth to wave action- the destructive effects of wave-bourne rocks on intertidal organisms. Oecologia, 69 (3), 420-428.
Smith, J.E. (ed.), 1968. 'Torrey Canyon'. Pollution and marine life. Cambridge: Cambridge University Press.
Sousa, W.P., 1979b. Experimental investigations of disturbance and ecological succession in a rocky intertidal algal community. Ecological Monographs, 49, 227-254.
Steneck, R.S., 1986. The ecology of coralline algal crusts: convergent patterns and adaptive strategies. Annual Review of Ecology and Systematics, 17, 273-303.
Stewart, J.G., 1989. Establishment, persistence and dominance of Corallina (Rhodophyta) in algal turf. Journal of Phycology, 25 (3), 436-446.
Storey, K.B., Lant, B., Anozie, O.O. & Storey, J.M., 2013. Metabolic mechanisms for anoxia tolerance and freezing survival in the intertidal gastropod, Littorina littorea. Comparative Biochemistry and Physiology Part A: Molecular & Integrative Physiology, 165 (4), 448-459.
Suchanek, T.H., 1993. Oil impacts on marine invertebrate populations and communities. American Zoologist, 33, 510-523.
Svåsand, T., Crosetti, D., García-Vázquez, E. & Verspoor, E., 2007. Genetic impact of aquaculture activities on native populations. Genimpact final scientific report (EU contract n. RICA-CT-2005-022802).
Sweet, N.S. 2011j. Green sea-fingers (tomentosoides), Codium fragile subsp. tomentosoides. Great Britain Non-native Species Secretariat. [cited 16/06/2015]. Available from: <http://www.nonnativespecies.org
Sweet, N.S. 2011g. Devil’s Tongue Weed, Grateloupia turuturu . Great Britain Non-native Species Secretariat. [cited 16/06/2015]. Available from:
Sweet, N.S. & Sewell, J. 2014. Asian shore crab, Hemigrapsus sanguineus. Great Britain Non-native Species Secretariat. [cited 16/06/2015]. Available from: <http://www.nonnativespecies.org
Taskin, E., Ozturk, M. & Kurt, O., 2007. Antibacterial activities of some marine algae from the Aegean Sea (Turkey). African Journal of Biotechnology, 6 (24), 2746-2751.
Trowbridge, C.D., Little, C., Dlouhy-Massengale, B., Stirling, P. & Pilling, G.M., 2013. Changes in brown seaweed distributions in Lough Hyne, SW Ireland: a long-term perspective. Botanica Marina, 56 (4), 323-338.
Tyler-Walters, H. & Arnold, C., 2008. Sensitivity of Intertidal Benthic Habitats to Impacts Caused by Access to Fishing Grounds. Report to Cyngor Cefn Gwlad Cymru / Countryside Council for Wales from the Marine Life Information Network (MarLIN) [Contract no. FC 73-03-327], Marine Biological Association of the UK, Plymouth, pp.
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.
Vadas, R.L., Keser, M. & Rusanowski, P.C., 1976. Influence of thermal loading on the ecology of intertidal algae. In Thermal Ecology II, (eds. G.W. Esch & R.W. McFarlane), ERDA Symposium Series (Conf-750425, NTIS), Augusta, GA, pp. 202-212.
Vermaat J.E. & Sand-Jensen, K., 1987. Survival, metabolism and growth of Ulva lactuca under winter conditions: a laboratory study of bottlenecks in the life cycle. Marine Biology, 95 (1), 55-61.
Watson, D.C. & Norton, T.A., 1985. Dietary preferences of the common periwinkle, Littorina littorea (L.). Journal of Experimental Marine Biology and Ecology, 88, 193-211.
Wiedemann, T., 1994. Oekologische Untersuchungen in Gezeitentuempeln des Helgolaender Nord-Ost Felswatts. , Diploma thesis, University of Kiel, Germany.
Williams, G.A. & Seed, R., 1992. Interactions between macrofaunal epiphytes and their host algae. In Plant-animal interactions in the marine benthos (ed. D.M John, S.J. Hawkins & J.H. Price), pp. 189-211. Oxford: Clarendon Press. [Systematics Association Special Volume, no. 46.]
Wood, C. 2015. Watersipora subatra. Great Britain Non-native Species Secretariat. [On-line][cited 16/06/2015]. Available from:
This review can be cited as:
Last Updated: 04/01/2016