Biodiversity & Conservation

LR.FLR.Rkp.SwSed

Explanation of sensitivity and recoverability


Physical Factors

Substratum Loss
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Loss of the substratum would involve loss of all the species within the rockpool and hence loss of the biotope. Break up of the rocky substratum (e.g. by a grounded vessel) and or infill of the rockpool would constitute loss of available substratum and hence the habitat. Infilling of the rockpool by permanent material (e.g. by cement) or occlusion by revetment material would constitute a permanent loss of the rockpool and biotope. However, in other instances the species could recolonize the remaining pool and recoverability is likely to be high (see additional information below).
Smothering
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Seapy & Littler (1982) reported a decrease in macroalgal cover from 47.3 to 37.5% on a Californian rocky shore due to sediment deposition on the mid to lower shore following rain and flooding. Corallina sp. and Pelvetia sp. were the most affected macroalgal species, while associated red algae were only slightly affected by the resultant scour. Macroinvertebrates declined in cover from 15.8% to 6.5% particularly barnacle species. Daly & Mathieson (1977) examined intertidal zonation on a shore affected by sand scour, and noted that fucoids were reduced to small or young plants, while sand tolerant species such as Ahnfeltia plicata dominated on areas affected by sediment. Smothering by 5 cm of sediment (see benchmark) is likely to increase scour and be detrimental to macroalgae, especially Corallina officinalis and fucoids, and the more fleshy red algae. While red algae such as Chondrus crispus and Ceramium spp. are large enough not to be smothered completely by 5 cm of sediment, the resultant scour is likely to damage fronds but, in particular, remove juveniles, sporelings and other propagules. In addition, the rockpool environment is likely to be more vulnerable to smothering as sediment is likely to accumulate in, and be retained by the rockpool itself, effectively increasing the depth of the sediment layer in the pool. In wave exposed conditions the sediment may be removed but in sheltered areas it is likely to be retained for longer than indicated by the benchmark. In deep pools, the macroalgae and associated invertebrates are likely to reduce in depth penetration into the pool while sediment tolerant algae increase. In shallow pools, the depth may be further reduced and the macroalgae restricted to sand tolerant species alone. Overall, while the biotope will remain, smothering is likely to reduce the diversity of the pool, exclude grazing littorinids, and smother small epifaunal species such as sponges, bryozoans and ascidians, although large anemones may survive (e.g. Urticina felina). Therefore, an intolerance of intermediate has been recorded. Recoverability is likely to be high (see additional information below). However, in extremely high suspended sediment loads, as found in estuaries rockpools may become completely filled with fine sediment, so that only infaunal species survive.
Increase in suspended sediment
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An increase in suspended sediment could potentially result in increased turbidity (see below), smothering, especially on sheltered shores (see above), and increased scour. The characterizing sediment tolerant red algae are unlikely to be adversely affected by an increase in suspended sediment. However, other macroalgae, and the community they support, are likely to be adversely affected, as shown above (Daly & Mathieson, 1977; Seapy & Littler, 1982). On wave sheltered shores, sediment may accumulate in low to mid shore pools, which again will favour sand tolerant species and infauna. Overall, the biotope is likely to remain but the species diversity decrease (for example see Daly & Mathieson, 1977). Therefore, an intolerance of intermediate has been recorded, although recovery is potentially high (see additional information below). However, in extreme situations deposition of fine sediments may result in smothering of the rockpool (see above).
Decrease in suspended sediment
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The presence of sediment in the rockpool environment, a defining feature of this biotope, suggest occasional or regular sediment supply. Therefore, a decrease in suspended sediment may result in erosion of the sediment from the bottom of pools. Erosion is likely to be greatest in shallow pools or moderately wave exposed shores. The internal topography of the pools will affect the rate of erosion i.e. presence of depressions and crevices may retain sediment longer. A decrease in suspended sediment loads may reduce food availability of suspension feeders within the biotope, however, they would probably feed on plankton within the pools. However, erosion or removal of sediment from the pools would be detrimental for an infauna, and especially sand tolerant algae which are likely to be out-competed by other red algae. Therefore, the character of the pool may change, becoming more like £LR.FK£. Hence an intolerance of high has been recorded since the biotope may be lost in the long term, although it should be noted that a healthy but different community is likely to remain, potentially of higher species diversity.
Desiccation
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Rockpools are natural refuges from desiccation but may be drained due to slow seepage or due to 'bucketing' by shore users, resulting in a decrease in the water level and hence desiccation exposure. Many members of the biotope are common on the emergent rock surface (e.g. fucoids, red algae, littorinids) and therefore, exhibit relative tolerance of desiccation. However, the presence of the rockpool allows species to occur in niches higher on the shore than they would otherwise. Low shore, sublittoral fringe or sublittoral species within the pool would be particularly intolerant of desiccation, e.g. Furcellaria lumbricalis and low shore algae. However, such drainage is likely to be short-lived, and the water level return to normal levels after the next high tide. Therefore, an increase in desiccation at the benchmark level, an increase equivalent to a rise in shore height, is likely to result in a decrease in species richness, although the biotope itself is likely to remain and an intolerance of intermediate has been recorded. Recoverability is likely to be high (see additional information).
Increase in emergence regime
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An increase in emergence is likely to significantly affect physico-chemical environment of the rockpool and its resident community. An increase in emergence will increase the time that the pool is exposed to fluctuating air temperatures, wind, rain and sunlight, all of which will affect the and temperature, salinity regime within the pool. Lower shore pools will come to resemble mid shore pool communities, with a reduction in sublittoral species and species sensitive to extremes of temperature, for example the laminarians (see individual reviews). For example, the upper limit of Bifurcaria bifurcata within rockpools in Roscoff, France was shown to be limited by the summer temperatures where the surface pool water temperatures exceeded 20 °C (Kooistra et al., 1989). Mid shore examples of this biotope are likely to be worst affected. High shore pools tend to support communities of temperature tolerant or opportunistic algae, especially green algae such as Ulva spp., and temperature and salinity tolerant species as harpacticoid copepods, ostracods, and small gastropods (for example see £LR.G£). This biotope would be lost from mid shore areas as a result of an increase in emergence at the benchmark level. Therefore, an intolerance of high has been recorded, although recoverability is potentially high.
Decrease in emergence regime
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A decrease in emergence will reduce the time the pool spends exposed to the air and cut off from the sea. Therefore, the range of temperatures and oxygen levels characteristic of rockpool environments is likely to decrease. Hence the pool communities will come to resemble low shore pools. Low shore pools are characterized by higher abundance of large macroalgae, such as Halidrys siliquosa, Cystoseira sp. and laminarians, especially deep pools, and a larger diversity of red algae and macrofauna. However, the presence of sediment within the pools will still favour the sand tolerant algae. Therefore, although the community is likely to increase in diversity the biotope is likely to remain. Therefore, an intolerance of low has been recorded to reflect changes in community structure.
Increase in water flow rate
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Water flow rate in this biotope is typically only that of the ebb and flood tide speed, which hardly affects intertidal habitats and is far exceeded by the strength of wave action. A change in water flow rate is therefore considered not relevant.
Decrease in water flow rate
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Water flow rate in this biotope is typically only that of the ebb and flood tide speed, which hardly affects intertidal habitats and is far exceeded by the strength of wave action. A change in water flow rate is therefore considered not relevant.
Increase in temperature
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Rockpools experience considerable variation in temperature on a daily and seasonal basis. The range and extremes of temperature change increasing with shore height but also dependent on shading, aspect, topography and depth of the pool (Pyefinch, 1943; Ganning, 1971; Daniel & Boyden, 1975; Goss-Custard et al., 1979; Morris & Taylor, 1983; Huggett & Griffiths, 1986; Metaxas & Scheibling, 1993). For example, reported temperature ranges for mid to low shore pools include annual maxima and minima of 1-25 °C and 2-22 °C (Morris & Taylor, 1983), a diurnal range of 24 °C (day) and 13 °C (night) for a mid shore pool (Daniel & Boyden, 1975), and surface water temperature ranges of 14-19.25 °C and 15.5-20.75 °C in mid shore pools (Pyefinch, 1943). Temperature stratification within pools may result in higher surface temperatures and lower deep water temperatures in sunlight (Daniel & Boyden, 1977) or be reversed due to wind cooling, night or in winter (Naylor & Slinn, 1958; Ganning, 1971; Morris & Taylor, 1983). The temperature range will limit the distribution of sensitive species within the pools, especially normally sublittoral species, e.g. laminarians (see individual reviews). For example, the upper limit of Bifurcaria bifurcata within rockpools in Roscoff, France was shown to be limited by the summer temperatures where the surface pool water temperatures exceeded 20 °C (Kooistra et al., 1989). Therefore, an increase in ambient temperatures is likely to reduce the abundance or vertical extent of sensitive species within the biotope, especially in shallow examples of the biotope. However, the range and extremes of temperature routinely experienced by the biotope are greater than the benchmark level and an intolerance of low has been recorded to represent a potential decrease in species diversity.
Decrease in temperature
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Rockpools experience considerable variation in temperature on a daily and seasonal basis. The range and extremes of temperature change increasing with shore height but also dependent on shading, aspect, topography and depth of the pool (Pyefinch, 1943; Ganning, 1971; Daniel & Boyden, 1975; Goss-Custard et al., 1979; Morris & Taylor, 1983; Huggett & Griffiths, 1986; Metaxas & Scheibling, 1993). For example, reported temperature ranges for mid to low shore pools include annual maxima and minima of 1-25 °C and 2-22 °C (Morris & Taylor, 1983), a diurnal range of 24 °C (day) and 13 °C (night) for a mid shore pool (Daniel & Boyden, 1975), and surface water temperature ranges of 14-19.25 °C and 15.5-20.75 °C in mid shore pools (Pyefinch, 1943). Temperature stratification within pools may result in higher surface temperatures and lower deep water temperatures in sunlight (Daniel & Boyden, 1977) or be reversed due to wind cooling, or in winter (Naylor & Slinn, 1958; Ganning, 1971; Morris & Taylor, 1983). Morris & Taylor (1983) reported that the surface of an upper shore was seen to freeze one winter night, although that this was a rare event. Freezing is likely to be rare in mid or low shore pools. The pool is likely to represent a buffer from the extreme cold and frosts experienced by fauna and flora on the emergent rock surface. In addition, few macroalgae were damaged as a result of the severe winter of 1962/63 (Crisp, 1964). Overall, the range of temperatures routinely experienced by mid to low shore rock pools is greater than the benchmark level. Therefore, tolerant has been recorded.
Increase in turbidity
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An increase in turbidity will reduce the availability of light for macroalgal growth. Macroalgae near the surface who's fronds float on the water surface will probably be unaffected. However, the depth within the pool that large macroalgae can penetrate will be reduced, in favour of shade tolerant red algae. Red algae attract fewer grazers and support fewer mesoherbivores (e.g. amphipods) and meiofauna, so that faunal diversity will decrease. The macroalgal abundance is likely to decrease but the biotope will still be recognizable. Therefore, an intolerance of low has been recorded to represent loss of diversity.
Decrease in turbidity
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A decrease in turbidity will allow fucoids and laminarians to grow to greater depths within the pool, in competition with red algae, except at the sediment/ rock surface interface. Overall, fucoid abundance is likely to increase at the expense of some red algae, depending on the depth of the pool, but the biotope will remain. Therefore, tolerant has been reported.
Increase in wave exposure
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This biotope has been recorded from wave exposed to wave sheltered conditions. The effect of increased wave exposure is probably dependant on the depth of the pool. Sediment is unlikely to remain in shallow pools in wave exposed conditions and the biotope is likely to be replaced by coralline pool (£LR.Cor£). In deep pools, increased wave exposure is likely to result in increased scour, resulting in bare rock at the bottom of the pool, especially where cobbles and pebbles are present. Fine sediments and the infauna they support are likely to be lost. However if the pool is deep enough, the upper levels of the pool are likely to continue to support macroalgae, especially laminarians and erect corallines. For example, an increase in wave exposure from moderately exposed to very exposed is likely to remove the sediment from all but the deepest pools, so that the biotope may come to resemble £LR.FK£. Overall, the biotope is likely to change and an intolerance of high has been recorded. Recoverability is likely to be high (see additional information below).
Decrease in wave exposure
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This biotope has been recorded from wave exposed to wave sheltered conditions. A decrease in wave exposure from moderately exposed to sheltered is likely to encourage the deposition of sediments and favour sand tolerant red algal species. Therefore, sand tolerant species are likely to increase in abundance. However, a decrease in wave exposure from sheltered to very sheltered is likely to result in smothering of the biotope (see above), and in extremely wave sheltered environments rockpools may fill with sediment and only infauna survive. Overall, an intolerance of intermediate has been recorded at the benchmark level.
Noise
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Few organisms within the biotope are likely to respond to noise or vibration at the benchmark level. Fish may attempt to leave the biotope at high tide but would otherwise be trapped at low tide. Overall, little if any effect on the biotope is expected.
Visual Presence
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Mobile invertebrates and fish are able to react to shading, usually darting to cover in order to avoid a potential predator. However, their visual acuity is low, and they are unlikely to be adversely affected by visual presence.
Abrasion & physical disturbance
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Abrasion by an anchor or mooring may remove some fronds of the large macroalgae, foliose red algae and coralline turf, although most species would grow back from their remaining holdfasts. However, trampling and netting for shrimps or fish may be more damaging. Deep pools and the species they contain are protected by their depth but both small and large shallow pools are probably more vulnerable.

No studies of the effects of trampling or netting on rockpools were found but studies of the effects on emergent algal communities are probably indicative. For example, moderate (50 steps per 0.09 sq. metre) or more trampling on intertidal articulated coralline algal turf in New Zealand reduced turf height by up to 50%, and the weight of sand trapped within the 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 fucoid algae and coralline turf species on the New Zealand rocky shore. Low trampling intensity (10 tramples) reduced fucoid cover by 25%, while high intensity (200 tramples) reduced fucoid cover by over 90%, although over 97% cover returned within 21 months after spring trampling; autumn treatments took longer to recover due to the delay in recruitment. Coralline bases were seen to peel from the rocks (Schiel & Taylor, 1999) due to increased desiccation caused by loss of the algal canopy. Brosnan & Cumrie (1994) demonstrated that foliose species (e.g. fucoids and Mastocarpus papillatus) were the most susceptible to trampling disturbance, while turf forming species were more resistant. Barnacles were also crushed and removed. However, the algae and barnacles recovered in the year following the trampling (Brosnan & Cumrie, 1994). Boalch et al. (1974) and Boalch & Jephson (1981) noted a reduction in fucoid cover (especially of Ascophyllum nodosum) at Wembury, Devon, when compared with the same transects surveyed 43 years previously. They suggested that the reduction in fucoid cover was due to the large number of visitors and school groups received by the site.

Dethier (1984) noted that low shore rockpools on the coast of Washington State, suffered physical disturbance from storms (wave action and wave driven logs) in winter months. The frequency of disturbance ranged from one every 2-5 years, while recovery of dominant to species to its original level ranged from 3 month to over 2 years. As a result, she estimated that ca 20-50% of the populations of dominant pools species were in a state of recovery in her study area.

Rockpools form natural mesocosms and so attract considerable attention from the general public, educational events and scientists alike. In addition to trampling within shallow pools and the vicinity of deeper pools, turning of rocks within the pool is likely to disturb underboulder communities (e.g. see MLR.Fser.Fser.Bo). Overall, a proportion of the macroalgal community, and the invertebrates it supports are likely to be removed, depending on trampling intensity, and an intolerance of intermediate has been recorded. Recoverability is likely to be high (see additional information below) once physical disturbance has stopped. However, it should be noted that ongoing trampling is likely to result in a long term reduction in the diversity of affected pools.
Displacement
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The majority of the epiphytic fauna, such as the isopods, amphipods and harpacticoid copepods are highly mobile are unlikely to be adversely affected by displacement. Similarly, gastropods are likely to survive and migrate back to suitable feeding areas. But the dominant macroalgae and sessile epifauna (e.g. barnacles and tubeworms) are permanently attached to the substratum and if removed will be lost. Loss of the red algal species especially will result in loss of the biotope overall. If macroalgal holdfasts and bases are also removed then recovery will be prolonged but still relatively rapid.

Chemical Factors

Synthetic compound contamination
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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. They also suggested that red algae were effective indicators of detergent damage since they undergo colour changes when exposed to relatively low concentration of detergent. Smith (1968) reported that red algae such as Ahnfeltia plicata, Chondrus crispus, Furcellaria fastigiata, Mastocarpus stellatus, Polyides rotundus and Osmundea pinnatifida were amongst the algae least affected by detergents, whereas other species, including Ceramium spp., Cryptopleura ramosa, Cladophora rupestris, Lomentaria articulata and Ulva lactuca were either killed or unhealthy, although the effects were worst higher on the shore, which had received the most detergents. Laboratory studies of the effects of oil and dispersants on several red algal species concluded that they were all sensitive to oil/dispersant mixtures, with little difference between adults, sporelings, diploid or haploid life stages (Grandy, 1984; cited in Holt et al., 1995). Smith (1968) reported that oil and detergent dispersants from the Torrey Canyon spill affected high water specimens of Corallina officinalis more than low shore specimens and some specimens were protected in deep pools. In areas of heavy detergent spraying, however, Corallina officinalis was killed, and was affected down to 6m depth at one site, presumably due to wave action and mixing (Smith, 1968). However, regrowth of fronds had begun within 2 months after spraying ceased (Smith, 1968).

Gastropods and amphipods were found to be amongst the most sensitive species to detergents and oils. For example, limpets are extremely intolerant of aromatic solvent based dispersants used in oil spill clean-up. During the clean-up response to the Torrey Canyon oil spill nearly all the limpets were killed in areas close to dispersant spraying. Viscous oil will not be readily drawn in under the edge of the shell by ciliary currents in the mantle cavity, whereas detergent, alone or diluted in sea water, would creep in much more readily and be liable to kill the limpet (Smith, 1968). A concentration of 5ppm killed half the limpets tested in 24 hours (Southward & Southward, 1978; Hawkins & Southward, 1992). Toxicity experiments with gastropods demonstrated that 10 ppm of BP1002 was enough to cause the animals to close and stop climbing (Smith, 1968). Smith (1968) noted that over a 100 ppm of BP1002 was required to kill the majority of Nucella lapillus in experiments, while different concentrations of BP1002 killed the majority of the following: Littorina littorea (100 ppm); Calliostoma zizyphinum (10 ppm); Aplysia punctata (50 ppm), and Patella vulgata (5 ppm) (see individual reviews).

Smith (1968) also noted that after detergent treatment only the beadlet anemone Actinia equina and tufts of Bifurcaria sp., Corallina sp., and other algae were present in a rockpool. The pool had previously supported a community of anemones, gastropods, Corallina, Lithophyllum, Enteromorpha, crabs, prawns and fish.

Cole et al. (1999) suggested that herbicides were, not surprisingly, very toxic to algae and macrophytes. Hoare & Hiscock (1974) noted that all red algae except Phyllophora sp. were excluded from near to 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. 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). For example, Lindane is likely to bioaccumulate significantly and is considered to be highly toxic to fish (Cole et al., 1999). Ebere & Akintonwa (1992) conducted experiments on the toxicity of various pesticides to Gobius spp. They found Lindane and Diazinon to be very toxic, with 96 hr LC50's of 0.25 µg/l and 0.04 µg/l respectively. TBT is generally very toxic to algae and fish. However, toxicity of TBT is highly variable with 96-hr LC50 ranging from 1.5 to 36 µg/l, with larval stages being more intolerant than adults (Cole et al., 1999). PCBs are highly persistent in the water column and sediments, have the potential to bioaccumulate significantly and can be very toxic to marine invertebrates. However their toxicity to fish was not clear (Cole et al., 1999). The pesticide ivermectin is very toxic to crustaceans, and has been found to be toxic towards some benthic infauna such as Arenicola marina (Cole et al., 1999).

Overall, the evidence suggests that, on balance, the characterizing red algae are probably very intolerant to synthetic chemicals, while resident gastropods, crustaceans and fish vary in their sensitivity. Loss of grazing invertebrates will significantly affect community structure. Therefore, biotope intolerance is assessed as high. Rockpools might be expected to accumulate chemical contaminants, depending on the rate of flushing, so that mid shore pools may be more vulnerable than low shore examples of the biotope. Recoverability is probably high (see additional information below).

Heavy metal contamination
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Bryan (1984) suggested that the general order for heavy metal toxicity in seaweeds is: organic Hg > inorganic Hg > Cu > Ag > Zn > Cd >Pb. Cole et al. (1999) reported that Hg was very toxic to macrophytes. The sub-lethal effects of Hg (organic and inorganic) on the sporelings of an intertidal red algae, Plumaria elegans, were reported by Boney (1971). 100% growth inhibition was caused by 1 ppm Hg. Burdin & Bird (1994) reported that both gametophyte and tetrasporophyte forms of Chondrus crispus accumulated Cu, Cd, Ni, Zn, Mn and Pb when immersed in 0.5 mg/l solutions for 24 hours. No effects were reported however, and no relationship was detected between hydrocolloid characteristics and heavy metal accumulation. It is generally accepted that adult fucoids are relatively tolerant of heavy metal pollution (Holt et al., 1997).

Bryan (1984) suggested that adult gastropod molluscs were relatively tolerant of heavy metal pollution. Cole et al. (1999) suggested that Pb, Zn, Ni and As were very toxic to algae, while Cd was very toxic to Crustacea (amphipods, isopods, shrimp, mysids and crabs), and Hg, Cd, Pb, Cr, Zn, Cu, Ni, and As were very toxic to fish. Bryan (1984) reported sublethal effects of heavy metals in crustaceans at low (ppb) levels. In laboratory investigations Hong & Reish (1987) observed 96hr LC50 (the concentration which produces 50% mortality) of between 0.19 and 1.83 mg/l in the water column for several species of amphipod.

Cd, Hg, Pb, Zn and Cu are highly persistent, have the potential to bioaccumulate significantly and are all considered to be very toxic to fish (Cole et al., 1999). Mueller (1979) found that in Pomatoschistus sp., very low concentrations of Cd, Cu and Pb (0.5 g/l Cd2+; 5 g/l Cu2+; 20 g/l Pb2+) brought about changes in activity and an obstruction to the gill epithelia by mucus. This may also be true for other goby species. Inorganic Hg concentrations as low as 30 µg/l (96-h LC5) are considered to be toxic to fish, whereas organic Hg concentrations are more toxic to marine organisms (WHO, 1989, 1991). Oertzen et al. (1988) found that the toxicity of the organic Hg complex exceeded that of HgCl2 by a factor of 30 for the goby Pomatoschistus microps.

Heavy metal contamination could potentially persist in deep rock pools due to depth and /or the presence of sediments onto which the heavy metals could adsorb. The intolerance of crustaceans and fish to heavy metal contaminants suggests that amphipod and isopod grazers and fish grazers and predators would be lost, allowing rapid growth of opportunistic algae such as Ulva spp. However, the presence of sediment will offset the loss of grazers, so that sand tolerant algae are likely to dominate shallow pools and other macroalgae are likely to be little affected. Therefore, an intolerance of intermediate has been recorded to represent the loss of species richness. Recoverability is likely to be high (see additional information below).
Hydrocarbon contamination
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Hydrocarbon contamination, e.g. from spills of fresh crude oil or petroleum products, may cause significant loss of component species in the biotope, through impacts on individual species viability or mortality, and resultant effects on the structure of the community. Rockpools are potentially vulnerable habitats, depending on depth, flushing rate and tidal height. Rockpool organisms may be protected, since oil will float on the pool surface. However, rockpool organisms will be exposed to the water soluble fraction of fresh oils, and a surface film of oil will prevent gaseous exchange and may reduce or exclude light. If exposed to oil the resident sediment is likely to adsorb oil and release it slowly, causing chronic long-term contamination and potentially prolonged recovery. The effects of oil contamination on marine organisms were reviewed by Suchanek (1983) and are summarized below.
  • Holt et al. 1995 reported that oil spills in the USA and from the 'Torrey Canyon' had little effect on kelp forest. Similarly, surveys of subtidal communities at a number sites between 1-22.5m below chart datum, including Laminaria hyperborea communities, showed no noticeable impacts of the Sea Empress oil spill and clean up (Rostron & Bunker, 1997)
  • Fucus vesiculosus shows limited intolerance to oil. After the Amoco Cadiz oil spill Fucus vesiculosus suffered very little (Floc'h & Diouris, 1980). Indeed, Fucus vesiculosus may increase significantly in abundance on a shore where grazing gastropods have been killed by oil, although very heavy fouling could reduce light available for photosynthesis and in Norway a heavy oil spill reduced fucoid cover.
  • Littoral barnacles (e.g. Semibalanus balanoides) have a high resistance to oil (Holt et al., 1995) but may suffer some mortality due to the smothering effects of thick oil (Smith, 1968).
  • Gastropods (e.g. Littorina littorea and Patella vulgata) and especially amphipods have been shown to be particularly intolerant of hydrocarbon and oil contamination (see Suchanek, 1993).
  • The abundance of littorinids decreased after the Esso Bernica oil spill in Sullom Voe in December 1978 (Moore et al., 1995). The abundance of Patella sp., Littorina saxatilis, Littorina littorea and Littorina neglecta and Littorina obtusata were reduced but had returned to pre-spill levels by May 1979. In heavily impacted sites, subjected to clean-up, where communities were destroyed in the process, Littorina saxatilis recovered an abundance similar to pre-spill levels within ca 1 year, while Littorina littorea took ca 7 years to recover prior abundance (Moore et al., 1995).
  • Widdows et al. (1981) found Littorina littorea surviving in a rockpool, exposed to chronic hydrocarbon contamination due to the presence of oil from the Esso Bernica oil spill.
  • The anemones Actinia and Anthopleura were reported to survive in waters with severe oil pollution (Smith, 1968; Suchanek, 1993).
  • Echinoderms are thought to be especially sensitive to oil (Suchanek, 1993). In a survey of rock pool at West Angle Bay, Pembrokeshire, Crump & Emson (1997) noted that limpets, crustaceans (amphipods and Palaemon) and the echinoderms Amphipholis squamata and rare Asterina phylactica were adversely affected. However, the majority of adult Asterina gibbosa survived. The macrofauna, except Asterina phylactica, had recovered its diversity and abundance within 12 weeks of the spill (Crump & Emson, 1997).
  • Laboratory studies of the effects of oil and dispersants on several red algae species (Grandy 1984 cited in Holt et al. 1995) concluded that they were all intolerant of oil/ dispersant mixtures, with little differences between adults, sporelings, diploid or haploid life stages. O'Brien & Dixon (1976) suggested that red algae were the most sensitive group of algae to oil or dispersant contamination. However, Smith (1968) noted that ed algae such as Ahnfeltia plicata, Chondrus crispus, Furcellaria fastigiata, Mastocarpus stellatus, Polyides rotundus and Osmundea pinnatifida were amongst the algae least affected by detergents, whereas other species, including Ceramium spp., Cryptopleura ramosa, Cladophora rupestris, Lomentaria articulata and Ulva lactuca were either killed or unhealthy, although the effects were worst higher on the shore, which had received the most detergents.
  • The lugworm Arenicola marina was driven to the surface by high concentrations of fresh no. 2 fuel oil, and by the presence no. 2 fuel oil in the water column, resulting in death within 3 days (Prouse & Gordon, 1976; Suchanek, 1993).
  • Cole et al. (1999) suggested a moderate to high toxicity of oils and petrochemicals for fish. Bowling et al. (1983) found that anthracene, a Polyaromatic hydrocarbon (PAH) had a photo-induced toxicity to the bluegill sunfish. They reported that when exposed to sunlight anthracene was at least 400 times more toxic than when no sunlight was present. According to Ankley et al. (1997) only a subset of PAH's are phototoxic (fluranthene, anthracene, pyrene etc.). Effects of these compounds are destruction of gill epithelia, erosion of skin layers, hypoxia and asphyxiation (Bowling et al., 1983). In PAH contaminated areas, fish have been observed to develop tumours (GESAMP, 1993). Oil spills were reported to have low acute toxicity to adult fish (GESAMP, 1993), probably since adults can avoid contaminated areas, but that fish kills may occur after exposure to emulsified oil in shallow waters, e.g. after the Braer oil spill (GESAMP, 1993). However, in the rockpool environment, fish are unlikely to be able to avoid the water soluble fractions, and may suffer chronic or acute toxicity depending on the oil type and fish species concerned.
  • Loss of grazing gastropods and mesoherbivores after oil spills results in marked increases in the abundance of ephemeral green algae (e.g. Ulva spp.) and fucoids (Southward & Southward, 1978; Hawkins & Southward, 1992; Raffaelli & Hawkins, 1999).
Overall, red algae, gastropods, amphipods and other crustaceans, echinoderms and fish within the rockpool community are likely to be adversely affected. However, fucoids and some of the characterizing red algae (e.g. Ahnfeltia plicata and Furcellaria fastigiata) are likely to survive and the biotope is likely to remain, although with a greatly reduced species richness. Therefore, an intolerance of intermediate has been recorded. The loss of grazers will allow increased growth of ephemeral greens and fucoids, although shade tolerant and sand tolerant red algae should still prevail, as the depth that fucoids penetrate will depend on self-shading and/or the depth of the sediment layer. However, the extent of damage may be exaggerated by the clean-up techniques employed e.g. detergents (see synthetic chemicals above) or high pressure water sprays. High water pressure sprays are likely to denude the rock surface of most life.

On wave exposed rocky coasts oil will be removed relatively quickly. Recovery of rocky shore populations was intensively studied after the Torrey Canyon oil spill in March 1967. Loss of grazers results in an initial flush of ephemeral green then fucoid algae, followed by recruitment by grazers including limpets, which free space for barnacle colonization. On shores that were not subject to clean-up procedures, the community recovered within ca 3 years. However, on shores treated with dispersants recovery took 5-8 years but was estimated to take up to 15 years on the worst affected shores (Southward & Southward, 1978; Hawkins & Southward, 1992; Raffaelli & Hawkins, 1999). Therefore, the community may take longer to recover, especially if oil is retained within pool bound sediments. Hence, a recoverability of moderate has been recorded (see additional information below).

Radionuclide contamination
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Insufficient information
Changes in nutrient levels
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Little information on the nutrient regime of rockpools was found. Rockpools are cut off from the sea for periods of time, depending on their shore height, and hence nutrients could potentially become limiting (e.g. nitrogen and phosphorous) within the period of emersion. Similarly, pools could also become eutrophic due to the presence of washed up seaweeds and bird droppings and in some cases due to sewage effluents.

Increased nutrient may increase growth in fast growing species, especially green algae (e.g. Ulva spp., Cladophora spp., and Chaetomorpha spp.) and some browns (e.g. Ectocarpus spp.) (Fletcher, 1996) to the detriment of slower growing and perennial species of macroalgae. Red algae such as Gracilaria sp. , Gracilariopsis sp., Corallina sp., Ceramium spp., Gelidium sp., Bangia sp. and in a few instances Furcellaria lumbricalis and Phycodrys rubens were reported to increase in abundance in eutrophicated waters (Fletcher, 1996).

Fucus vesiculosus was observed to grow in the vicinity of a sewage outfall (Holt et al., 1997) and is probably not sensitive directly. However, one of the most noticeable changes associated with eutrophication is the decline in abundance of fucoids (e.g. Fucus spp., Ascophyllum nodulosum, and Cystoseira spp.), possibly due to increased competition with opportunistic tolerant green algae, and associated effects of eutrophication such as suspended sediment levels (Fletcher, 1996).

Eutrophication can potentially increase oxygen consumption leading to deoxygenation. However, the rockpool environment normally experience considerable variation in oxygen levels. Overall, the macroalgal community is likely to change, favouring ephemeral green and brown algae, red and coralline algae. However, the sand tolerant species characteristic of the biotope will probably remain. Therefore, an intolerance of intermediate has been recorded to reflect the change in species composition.
Increase in salinity
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High air temperatures cause surface evaporation of water from pools, so that salinity steadily increases, especially in pools not flooded by the tide for several days. However, Daniel & Boyden (1975) and Morris & Taylor (1983) reported little variability in salinity over one tidal cycle, and Ganning (1971) suggested that changes in salinity were of limited importance. Morris & Taylor (1983) reported an annual maximum salinity of 36.5 ppt in the pools studied on the west coast of Scotland. Goss-Custard et al. (1979) recorded salinities of 34.8 and 35.05 ppt in mid-shore pools. Therefore, the biotope is probably tolerant of small increases in salinity and an intolerance of low has been recorded. High shore pools exhibit greater variation and higher extremes of salinity (Pyefinch, 1943; Ganning, 1971) and different communities but mid to low shore pools are unlikely to experience such extremes unless the emergence regime is increased (see above) or they are exposed to hypersaline effluents. Therefore, an intolerance of low has been recorded, with a very high recoverability.
Decrease in salinity
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During periods of emersion, high rainfall will reduce pool salinity or create a surface layer of brackish/nearly fresh water for a period. The extremes of salinity experienced will depend on the depth of the pool, shore height and flushing rate, and season. For example, Morris & Taylor (1983) stated that a low salinity layer of 2-10 mm was normal but after one storm the low salinity layer increased in depth, eventually resulting in a homogeneous pool of brackish water. Morris & Taylor (1983) reported an annual salinity range in mid to low shore pools of 26-36.5 ppt. Therefore, decreases in salinity equivalent of a reduction from full to reduced (see benchmark) are likely to be a regular occurrence in rockpool communities. Hence, tolerant has been recorded.
Changes in oxygenation
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During emergence, rockpools are closed systems and gaseous exchange occurs over the air/water interface. In shallow pools the volume to surface area ratio is likely to be high, whereas in deep pools the ratio is likely to be low. In addition, the oxygen concentration is dependant on the community present. During the day, photosynthesis uses up CO2 and produces O2, in excess of respiration. However, at night respiration by flora and fauna deplete oxygen levels. As a result rockpool environments exhibit marked variation in oxygen levels. In summer, rockpools are likely to be supersaturated with oxygen during the day (Pyefinch, 1943). For example, the greatest range of oxygen saturation of 101.7% occurred in a seaweed dominated, sediment floored pool, which reach over 190%saturation on some days (Pyefinch, 1943). Daniel & Boyden (1975) noted that a mid shore, seaweed dominated pool reached 194% saturation (ca 15 mg O2/l) but that oxygenation was also marked in shaded pools. A pool with dense fauna exhibited a maximum saturation of 210% (Pyefinch, 1943). During photosynthesis, algae absorb carbon dioxide and as concentrations fall, the pH rises. Morris & Taylor (1983) recorded pH values >9 in rockpools on the Isle of Cumbrae. At night, oxygen levels may fall below 100% saturation and pH will decrease as CO2 levels increase. Morris & Taylor (1983) noted an annual maximum of oxygen concentration of 400-422 mm Hg (ca 23.4-24.7 mg/l) and an annual minimum of 18-38 mm Hg (ca 1-2.2 mg/l) in mid shore pools (containing Furcellaria). Daniel & Boyden (1975) reported oxygen depletion at night, with mid to low shore pools reduced to 8-44% saturation. They noted that the crab Carcinus maenas leaves the pools at night, and that other species with the ability to air-breathe could also do so, e.g. limpets, littorinids, and the shanny Lipophrys pholis. They also observed that shrimps gathered at the edge of high shore pools at night, presumably to take advantage of the better oxygenated surface layer (Daniel & Boyden, 1975).

The range of extremes in oxygen concentration were greater in summer than in winter. On immersion, the rockpool community was exposed to potentially large, sudden fluctuations in oxygen concentrations depending on season and time of day (Morris & Taylor, 1983). Therefore, rockpools communities are probably exposed to variations equivalent to or greater than the benchmark level on a regular basis and tolerant has been recorded.

Biological Factors

Introduction of microbial pathogens/parasites
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Several coralline and non-coralline species are epiphytic on Corallina officinalis. Irvine & Chamberlain (1994) cite tissue destruction caused by Titanoderma corallinae. However, no information on pathogenic organisms in the British Isles 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). Intertidal gastropods often act a secondary hosts for trematode parasites of sea birds. For example, Nucella lapillus may be infected by cercaria larvae of the trematode Parorchis acanthus. Infestation causes castration and continued growth (Feare, 1970b; Kinne, 1980; Crothers, 1985). Overall, a wide variety of pathogens may affect members of the community but no information on associated mortality was found.
Introduction of non-native species
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Several non-native species may occur in this biotope. Codium fragile subsp. tomentosoides and Codium fragile subsp. atlanticum were introduced from mainland Europe and Japan respectively and may out-compete the native Codium tomentosum (Eno et al., 1997). The non-native harpoon weed Asparagopsis armata was first recorded in Ireland in Galway Bay in 1939 and Britain in 1949 at Lundy in the Bristol Channel, and may come to dominate rockpools (Keith Hiscock, pers. comm.), although its effect on other species is not known.

Sargassum muticum is a non-native macroalgae spreading around the coasts of Britain and Europe (see Eno et al., 1997) and is often found in low to mid shore rockpools in the intertidal in areas it has colonized. Although, no studies on its effects on rockpool species were found, studies of its effect on shallow sublittoral macroalgae suggest that it can out-compete fucoids and kelps. For example, Stæhr et al. (2000) reported that an increase in the abundance of Sargassum muticum in the Limfjorden (Denmark) from 1990 to 1997 was accompanied by a decrease in the abundance of thick, slow growing macroalgae such as Saccharina latissima (studied as Laminaria saccharina), Codium fragile, Halidrys siliquosa, Fucus vesiculosus, and Fucus serratus, together with other algae such as Ceramium virgatum (as rubrum) and Dictyota dichotoma. In Sargassum muticum removal experiments on the coast of Washington State, Britton-Simmonds (2004) concluded that Sargassum muticum reduced the abundance of native canopy algae (especially kelps) by 75% and native understorey algae by 50% probably as a result of shading. However, Viejo (1999) noted that mobile epifauna (e.g. amphipods, isopods) successfully colonized Sargassum muticum which provided additional habitat. Overall, Sargassum muticum can successfully invade rockpools, and would probably out-compete resident fucoids and kelp species, and some red algae. The presence of sediment may still favour sand tolerant red algae, which may be little effected. In addition, mesoherbivores will probably adapt to the new substratum offered by Sargassum muticum since they feed primarily on epiphytes. Therefore, the biotope is likely to remain but with a reduced species richness due to the loss of some species of macroalgae, and an intolerance of intermediate has been recorded. Recovery is potentially high but assumes removal of Sargassum muticum which is unlikely. Hence, a recoverability of 'none' has been recorded since the biotope is likely to change, although a viable community will remain.

Extraction
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Several of the characterizing red algae species are subject to harvesting. Ahnfeltia plicata is one of the world's principal commercial agarophytes. It is harvested mainly on the Russian coast of the White Sea as a source of high quality, low sulphate agar (Chapman & Chapman, 1980). In Britain and Ireland, however, Ahnfeltia plicata does not occur in sufficient quantities to harvest on a commercial scale (Dickinson, 1963). Chondrus crispus is extracted commercially in Ireland, but the harvest has declined since its peak in the early 1960s (Pybus, 1977). Mathieson & Burns (1975) described the recovery of Chondrus crispus following experimental drag raking (see review) and concluded that control levels of biomass and population structure are probably re-established after 18 months of regrowth. Commercial utilization of Furcellaria lumbricalis is based on the gelling properties of its extracted structural polysaccharide, furcellaran (Bird et al., 1991). Extraction of Furcellaria lumbricalis was reviewed by Guiry & Blunden (1991). Plinski & Florczyk (1984) noted that over-exploitation of Furcellaria lumbricalis resulted in severe depletion of stocks. However, no commercial harvest as yet occurs in Britain or Ireland. Overall, while rockpools in areas subject to commercial harvesting may be directly affected, most examples of the biotope are unlikely to be affected by commercial harvesting in the UK. However, due to the relative small size of the community, even small scale hand collecting may have a significant effect.

Hand collection may reduce the population of Littorina littorea within rockpools and hence reduce grazing pressure, resulting in an increase in macroalgal cover, especially of opportunistic green algae and epiphytes.

An intolerance of intermediate has been recorded to represent the loss of a proportion of the macroalgae and the invertebrate community it supports. However, recovery is likely to be rapid since holdfasts and sporelings are likely to remain. Although, Furcellaria lumbricalis will recovery slowly, it is only one of the sand tolerant algae characteristic of this biotope. The littorinid will probably recover quickly by migration and recruitment.

Additional information icon Additional information

Recoverability
Red algae produce non motile spores, dependant on the hydrography and most recruitment is likely to occur within about 10 m of the parent plants (Norton, 1992). Therefore, within a rock pool or a pool surrounded by macroalgae, recruitment is likely to be good. However, recruitment from remote populations is likely to be more protracted and sporadic.

The life history characteristics of Ahnfeltia plicata suggest that the species is likely to recover within 5 years if local populations exist (see MarLIN review). Recovery of a population of Chondrus crispus following a perturbation is likely to be largely dependent on whether holdfasts remain, from which new thalli can regenerate (Holt et al., 1995). Following experimental harvesting by drag raking in New Hampshire, USA, populations recovered to 1/3 of their original biomass after 6 months and totally recovered after 12 months (Mathieson & Burns, 1975). Raking is designed to remove the large fronds but leave the small upright shoots and holdfasts. The authors suggested that control levels of biomass and reproductive capacity are probably re-established after 18 months of regrowth. It was noted however, that time to recovery was much extended if harvesting occurred in the winter, rather than the spring or summer (Mathieson & Burns, 1975). Minchinton et al. (1997) documented the recovery of Chondrus crispus after a rocky shore in Nova Scotia, Canada, was totally denuded by an ice scouring event. Initial recolonization was dominated by diatoms and ephemeral macroalgae, followed by fucoids and then perennial red seaweeds. After 2 years, Chondrus crispus had re-established approximately 50% cover on the lower shore and after 5 years it was the dominant macroalga at this height, with approximately 100% cover. The authors pointed out that although Chondrus crispus was a poor colonizer, it was the best competitor.

Kain (1975) examined recolonization of cleared concrete blocks in a subtidal kelp forest. Red algae colonized blocks within 26 weeks in the shallow subtidal (0.8m) and 33 weeks at 4.4m. After about 2.5 years, Laminaria hyperborea standing crop, together with an understorey of red algae, was similar to that of virgin forest. Red algae were present throughout the succession increasing from 0.04 to 1.5 percent of the biomass within the first 4 years. Colonizing species varied with time of year, for example blocks cleared in August 1969 were colonized by primarily Saccharina latissima (studied as Laminaria saccharina) and subsequent colonization by Laminaria hyperborea and other laminarians was faster than blocks colonized by Saccorhiza polyschides; within 1 year the block was occupied by laminarians and red algae only. Succession was similar at 4.4m, and Laminaria hyperborea dominated within about 3 years. Blocks cleared in August 1969 at 4.4m were not colonized by Saccorhiza polyschides but were dominated by red algae after 41 weeks, e.g. Cryptopleura ramosa. Kain (1975) cleared one group of blocks at two monthly intervals and noted that brown algae were dominant colonists in spring, green algae (solely %) in summer and red algae were most important in autumn and winter. Overall, red algae are likely to be able to recolonize and recover abundance with a year in some instances and probably within 5 years. Similarly, laminarians could potentially colonize low shore rockpools within 3-4 years, depending on grazing and competition for space. Fucoids (e.g. Fucus serratus) are highly fecund, reproduce throughout the years, are widespread and could potentially recovery quickly. For example, after the Torrey Canyon, oil spill fucoids attained maximum cover within 1-3 years (Southward & Southward, 1978; Hawkins & Southward, 1992; Raffaelli & Hawkins, 1999).

Furcellaria lumbricalis is an exception. Although highly fecund (Austin, 1960a), the species grows very slowly compared to other red algae (Bird et al., 1991) and takes a long time to reach maturity, typically 5 years (Austin, 1960b). Christensen (1971; cited in Bird et al., 1991) noted that following harvesting of Furcellaria lumbricalis forma aegagropila in the Baltic Sea, harvestable biomass had not been regained 5 years after the suspension of harvesting. In view of its slow growth, time to maturity and limited dispersal, recovery of Furcellaria lumbricalis is likely to take between 5 and 10 years to recover in situations where intolerance to a factor is high. Where a portion of the population remains for vegetative regrowth, recovery is likely to occur within 5 years.

Gastropods and other mobile grazers (e.g. amphipods, isopods) are likely to be attracted by developing microalgae and macroalgae and could return quickly by either migration or larval recruitment. Epifaunal species vary in their recruitment rates. Sebens (1985, 1986) reported that rapid colonizers such as encrusting corallines, encrusting bryozoans, amphipods and tubeworms recolonized cleared rock surfaces within 1-4 months. Ascidians such as Aplidium spp. achieved significant cover in less than a year, and, together with Halichondria panicea, reached pre-clearance levels of cover after 2 years. Anemones colonized within 4 years (Sebens, 1986) and would probably take longer to reach pre-clearance levels. The anemone Urticina felina has poor powers of recoverability due to poor dispersal (Sole-Cava et al., 1994 for the similar Tealia crassicornis) and slow growth (Chia & Spaulding, 1972), though populations should recover within 5 years.

Overall, members of the rockpool community could potentially recolonize with a year and a recognizable biotope return within 5 years. However, rockpool recruitment is reported to be sporadic and variable (Metaxas & Scheibling, 1993). Therefore, while a recognizable biotope will return the exact community may differ from that present prior to perturbation.

This review can be cited as follows:

Tyler-Walters, H. 2005. Seaweeds in sediment (sand or gravel)-floored eulittoral rockpools. Marine Life Information Network: Biology and Sensitivity Key Information Sub-programme [on-line]. Plymouth: Marine Biological Association of the United Kingdom. [cited 26/11/2014]. Available from: <http://www.marlin.ac.uk/habitatbenchmarks.php?habitatid=326&code=2004>