|Researched by||Dr Harvey Tyler-Walters||Refereed by||This information is not refereed.|
Deep rockpools in the mid to lower eulittoral zone often contain a community characterized by Fucus serratus and Laminaria digitata. Other large brown algae, including Saccharina latissima, Himanthalia elongata and Halidrys siliquosa, may also occur. The rock surface is usually covered by encrusting coralline algae. A wide variety of filamentous and foliose algae, which are typical of lower shore and shallow sublittoral zones (e.g. Palmaria palmata, Chondrus crispus, Ceramium spp., Membranoptera alata and Gastroclonium ovatum) occur beneath the brown algal canopy. Algal-free vertical and overhanging faces often support the sponge Halichondria panicea and anemones Actinia equina. The abundance of grazing molluscs varies considerably. In some, large numbers of littorinids and limpets are probably responsible for the limited variety of red seaweeds present. In other pools, fewer grazers may result in an abundance of these algae. Where boulders occur in these pools they provide a greater variety of micro-habitats which support a variety of fauna. Mobile crustaceans (Pagurus bernhardus and Carcinus maenas), brittlestars (Ophiothrix fragilis and Amphipholis squamata), encrusting bryozoans and ascidians are typically found beneath and between boulders. (Information taken from the Marine Biotope Classification for Britain and Ireland, Version 97.06: Connor et al., 1997a, b).
Factors such as pool depth, surface area, volume, orientation to sunlight, shading, internal topography, sediment content and type, together with wave exposure, shore height, and hence flushing rate, and the presence of absence of freshwater runoff, results in large spatial variation in community structure, even between adjacent pools at the same shore height (Ganning, 1971; Metaxas & Scheibling, 1993). Individual rockpools and the communities that occupy them are highly variable.
|Depth Range||Mid shore, Lower shore|
|Water clarity preferences|
|Limiting Nutrients||No information found|
|Salinity preferences||Full (30-40 psu)|
|Physiographic preferences||Enclosed coast / Embayment, Open coast|
|Biological zone preferences||Lower eulittoral, Eulittoral|
|Tidal strength preferences|
|Wave exposure preferences||Exposed, Moderately exposed, Sheltered|
|Other preferences||Deep rockpools|
This MarLIN sensitivity assessment has been superseded by the MarESA approach to sensitivity assessment. MarLIN assessments used an approach that has now been modified to reflect the most recent conservation imperatives and terminology and are due to be updated by 2016/17.
Palmaria palmata and Ceramium virgatum have been chosen to represent the sensitivity of characteristic foliose and filamentous red algae, and Corallina officinalis to represent corallines. Littorinids, limpets and amphipods have been shown to be important grazers in rockpool environments (see 'ecological relationships') that affect community structure and development. Their sensitivities are represented by Littorina littorea and Patella vulgata. The sensitivity of amphipods is treated as a functional group, although reference has been made to relevant species reviews e.g. Hyale prevostii.
|Community Importance||Species name||Common Name|
|Important structural||Ceramium virgatum||A red seaweed|
|Important structural||Chondrus crispus||Carrageen|
|Important structural||Corallina officinalis||Coral weed|
|Important characterizing||Fucus serratus||Toothed wrack|
|Important characterizing||Halidrys siliquosa||Sea oak|
|Important characterizing||Laminaria digitata||Oarweed|
|Important functional||Littorina littorea||Common periwinkle|
|Important structural||Palmaria palmata||Dulse|
|Important functional||Patella vulgata||Common limpet|
|Important characterizing||Saccharina latissima||Sugar kelp|
|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).|
|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 laminarians and 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. Overall, smothering is likely to reduce the macroalgal diversity of the pool, exclude grazing littorinids, and smother small epifaunal species such as sponges, bryozoans, small anemones and ascidians, although large anemones may survive (e.g. Urticina felina). Where sediment is retained the sediment tolerant algae may come to dominate and the biotope will resemble A1.413. Therefore, an intolerance of high has been recorded. Recoverability is likely to be moderate (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.|
|An increase in suspended sediment could potentially result in increased turbidity (see below), smothering, especially on sheltered shores (see above), and increased scour. Fucoids, kelps and 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 will favour sand tolerant species and infauna. Overall, macroalgae are likely to be damaged but 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).|
|Tolerant||Not relevant||No change||Low|
|A decrease in suspended sediment could reduce the turbidity (see below) and potentially reduce the food availability for suspension feeders, due to a reduction in organic particulates. However, suspension feeders will continue to feed on available plankton and detritus and be little affected. Similarly, the resident macroalgae are unlikely to be adversely affected by reduced sediment loads, except that scour is reduced. Therefore, tolerant has been recorded.|
|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).|
|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 temperature and 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 bifurcaria 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 LR.FK 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 such as harpacticoid copepods, ostracods, and small gastropods (for example see A1.421). 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 be recorded and recoverability is probably moderate (see additional information below).|
|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 mid shore 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 and a larger diversity of red algae and macrofauna. Low shore pools will probably be colonized by an increasing number of sublittoral species. 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.|
|Not relevant||Not relevant||Not relevant||Not relevant||Moderate|
|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.|
|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.|
|Low||Very high||Very Low||Minor decline||Low|
|Rockpools experience 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 bifurcaria 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 shallower 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.|
|Rockpools experience 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. Nevertheless the severe winter of 1962/63 resulted in a wide variety of mortalities in the intertidal and shallow subtidal (Crisp, 1964a). For example, few macroalgae were damaged but specimens of Cystoseira spp. in the south and south west were smaller than usual. However, the anemone Anemonia viridis was missing from shallow pools and that only a single specimen of Cereus pedunculatus was found in an area of usual abundance, while many dead specimens of both species were found in south Wales. Similarly, many dead porcelain crabs (Porcellana spp.) were found. Patella vulgata exhibited increasing mortality with shore height and hence emersion (Crisp, 1964a), and several species of gastropod exhibited mortality. Although southern, lusitanian, species were worst affected, mortalities of individual species varied with location. However, rockpools, especially deep pools and low shore pools are likely to represent a buffer from the extreme cold and frosts experienced by fauna and flora on the emergent rock surface. Overall, the range of temperatures routinely experienced by mid to low shore rock pools is greater than the benchmark level. However, the severe winter of 1962/63 suggests that some sensitive species, particularly limpets and gastropods, and anemones near the surface of deep pools may be affected. The loss of grazers may benefit the macroalgal community, resulting in increased growth of fucoids and green algae. Therefore, an intolerance of intermediate has been recorded to represent the loss of species diversity and changes in community structure, especially in mid shore examples of the biotope. Recoverability is probably high (see additional information below).|
|An increase in turbidity due to suspended sediment, dissolved organics or phytoplankton blooms will reduce the depth that light can penetrate the pool and hence the depth within the pool that different groups of algae can grow, particularly kelps. For example, in the silt-laden waters around Helgoland, Germany the depth limit for Laminaria digitata growth may be reduced to between 0 m and 1.5 m (Birkett et al., 1998b). Increased turbidity around a sewage treatment plant was thought to be responsible for the absence of Laminaria digitata plants in the Firth of Forth (Read et al., 1983). In Narragansett Bay, Rhode Island growth rates of Laminaria digitata fell during a summer bloom of microalgae that dramatically reduced down welling irradiance. Quality of light is also important with blue light necessary for gametogenesis and development of gametophytes in laminarians. Dissolved organic materials (yellow substance or gelbstoff) absorbs blue light strongly, therefore changes in riverine input or other land based runoff are likely to influence kelp density and distribution. Light levels often determine the maximum depth for survival of Saccharina latissima (studied as Laminaria saccharina) at a particular site (Lüning & Dring, 1975; Gerard, 1988) therefore an increase in turbidity may lead to the mortality of some plants towards the deeper end of their depth range, although Gerard (1988) reported that Saccharina latissima populations may adapt to low or variable light conditions. Moss & Sheader (1973) demonstrated that the growth of Halidrys siliquosa germlings was dependent on light intensity but that germlings could survive total darkness for 120 days (see general biology). Fucus serratus can normally photosynthesize when emersed so that increased turbidity on emergent rocks is unlikely to be detrimental although growth rates are likely to be reduced. Overall, an increase in turbidity of the water will reduce the depth within the pool that macroalgae can grow, so that kelps and to a lesser extent the fucoids are likely to be limited to the upper margin of the pool. However, shade tolerant red algae may benefit and dominate the deeper parts of the pool. An increase in turbidity at the benchmark level may result in loss of laminarians from deep pools, especially Laminaria digitata, but fucoids and hence the biotope will probably remain at the surface. Therefore, an intolerance of intermediate has been recorded to represent the potential loss of kelp species, although recoverability is likely to be high.|
|Low||Very high||Moderate||No change||Low|
|A decrease in turbidity will increase light penetration, and hence the growth of all macroalgae, especially kelps species, and increase the depth at which red algae or fucoids may grow, possibly increasing competition for space between the algae themselves and other space occupiers such as sponges and ascidians. However, the effects are likely to depend on the size of the pool. In smaller pools, increased growth of kelps and fucoids is likely to result in self-shading, so that the net effect is likely to be minimal. Therefore, an intolerance of low has been recorded.|
|This rockpool biotope occurs in wave sheltered to wave exposed habitats. The rockpool provides a degree of shelter from wave action, especially deep pools, allowing more fragile sublittoral algae to survive. However, an increase in wave exposure from, for example moderately exposed to very exposed is likely change the community. Fucoid abundance in characteristic of wave sheltered conditions, and on more wave exposed shores shallow rockpools are dominated by Corallina officinalis (see A1.411). Therefore, an increase in wave exposure at the benchmark level is likely to reduce the abundance or remove fucoids from the margin of the pool, in favour of corallines. Laminaria digitata is likely to be replaced by Alaria esculenta, which tolerates strong water movement. Lewis (1964) noted that Halidrys siliquosa, and Cystoseira spp. were restricted to deep mid shore pools with increasing wave exposure. Similarly, the increased turbulence within the pool itself will favour species that prefer strong water movement, such as the passive suspension feeders e.g. hydroids (e.g. Tubularia larynx) and anemones (e.g. Metridium dianthus) and other epifauna, together with more wave exposure tolerant red algae, e.g. Porphyra sp., Plocamium sp. and Gigartina sp.. However, with increasing wave exposure the biotope is likely to change, and may come to resemble Corallina officinalis rockpool biotopes, depending on the relative abundance of Bifurcaria bifurcata (in the south west) and Cystoseira spp (see LR.Cor). Therefore, the biotope is likely to be lost, and although replaced by another healthy community, an intolerance of high has been recorded. Recoverability is likely to be moderate (see additional information below).|
|This rockpool biotope occurs in wave sheltered to wave exposed habitats. A decrease in wave exposure from e.g. sheltered to very sheltered, or extremely sheltered is likely to adversely affect the biotope. The resultant lack of water movement is likely to result in increased suspended sediment and siltation of the rockpool, smothering and filing the rockpool. Fucoids will survive on the margins of the pool and emergent rock, however laminarians, and epifauna are likely to be lost and only sediment tolerant red algae survive within the pool. The biotope may come to resemble LR.SwSed, or in worst case situations become silted up, so that only infauna survive. Therefore, an intolerance of high has been recorded, with moderate recoverability (see additional information below).|
|Tolerant||Not relevant||Not sensitive||No change||High|
|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.|
|Tolerant||Not relevant||Not sensitive||No change||High|
|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 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 may be more damaging. Deep pools are protected by their depth but shallower pools or the shallower margins of larger pools are probably more vulnerable.
No studies of the effects of trampling on rockpools were found but studies of the effects on emergent algal communities are probably indicative. For example, moderate (50 steps per 0.09 square 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). Similarly, 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.Rockpools form natural mesocosms and so attract considerable attention from the general public, educational events and scientists alike. In addition to trampling within shallower pools and the margins of deeper pools, turning of rocks within the pool is likely to disturb underboulder communities (e.g. see A1.2142). 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 trampling has stopped. However, it should be noted that ongoing trampling is likely to result in a long term reduction in the diversity of the margins of the affected pools.
|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 fucoids and kelps especially will result in loss of the biotope overall. If macroalgal holdfasts and bases are also removed then recovery will be prolonged.|
|The different groups of organisms within the biotope are likely to vary in their response to synthetic chemical pollution. Key examples are summarized below.
|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). The effect of heavy metals on the growth rate of adult Fucus serratus plants has been studied by Strömgren (1979b;1980a, b). Copper significantly reduces the growth rate of vegetative apices at 25 µg/l over 10 days (Strömgren, 1979b). Zinc, lead, cadmium & mercury significantly reduce growth rate at 1400 µg/l, 810 µg/l, 450 µg/l and 5 µg/l respectively (Strömgren, 1980a, b).
Zinc was found to inhibit growth in Laminaria digitata at a concentration of 100 µg/L and at 515 µg/L growth had almost completely ceased (Bryan, 1969). Axelsson & Axelsson (1987) investigated the effect of exposure to mercury (Hg), lead (Pb) and nickel (Ni) for 24 hours by measuring ion leakage to indicate plasma membrane damage. Inorganic and organic Hg concentrations of 1 mg/L resulted in the loss of ions equivalent to ion loss in seaweed that had been boiled for 5 minutes. Laminaria digitata was unaffected when subjected to Pb and Ni at concentrations up to 10 mg/L. Their results also indicated that the species is intolerant of the tin compounds butyl-Sn and phenyl-Sn. Sporophytes of Saccharina latissima (studied as Laminaria saccharina) have a low intolerance to heavy metals but the early life stages are more intolerant (Thompson & Burrows, 1984). Growth of sporophytes was significantly inhibited at 50 µg Cu /l, 1000 µg Zn/l and 50 µg Hg/l. Zoospores were found to be more intolerant and significant reductions in survival rates were observed at 25 µg Cu/l, 1000 µg Zn/l and 5 µg/l (Thompson & Burrows, 1984).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 96 hr LC50 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.The intolerance of crustaceans to heavy metal contaminants suggests that amphipod and isopod grazers would be lost, allowing rapid growth of opportunistic algae such as Ulva spp. In addition, the characterizing laminarians and their propagules may be adversely affected, and the growth rates of fucoids reduced. Therefore, an intolerance of intermediate has been recorded to represent a decrease in species diversity, although a recognizable biotope is likely to remain. Recoverability is likely to be high (see additional information below).
|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 (1993) and are summarized below.|
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 limpet, which free space for barnacle colonization. On shores that were not subject to clean up procedures, the community recovered within ca 3 years, however, in 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 in oil is retained within pool bound sediments or as a coating of tar. Hence, a recoverability of moderate has been recorded (see additional information below).
|No information||Not relevant||No information||Not relevant||Not relevant|
|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 sewage effluent. The effluent from rotting seaweeds on the strandline can severely impact upper shore pools (e.g. at Wembury, Devon) although lower shore pools are unlikely to be affected in LR.FK. However, eutrophication only likely to be a problem in high shore pools cut off from the sea for days at a time.
Increased nutrient may increase growth in fast growing species (e.g. Ulva spp.) to the detriment of slower growing species of macroalgae. However, Fucus vesiculosus was observed to grow in the vicinity of a sewage outfall (Holt et al., 1997) and is probably not sensitive.Eutrophication can potentially increase oxygen consumption leading to deoxygenation. However, the rockpool environment normally experience considerable variation in oxygen levels. Overall, an intolerance of intermediate has been recorded.
|Low||Very high||Very Low||No change||Low|
|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.|
|Tolerant||Not relevant||No change||Low|
|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. Mid shore examples of this biotope may lack more sensitive species, such as Laminaria digitata and some sublittoral species. Nevertheless, decreases in salinity equivalent of a reduction from full to reduced (see benchmark) are likely to be a regular occurrence in rockpool communities, and the biotope is unlikely to be adversely affected. Hence, tolerant has been recorded.|
|Tolerant||Not relevant||Not sensitive||No change||Low|
|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. 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). Goss-Custard et al. (1979) noted that oxygen saturation levels decreased with depth in deep mid shore pools, while Morris & Taylor (1983) noted that oxygen saturation varied with depth and proximity to algae, especially green algae such as Cladophora spp.
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.
|Low||Very high||Very Low||No change||Low|
|Laminarians are susceptible to brown spot disease, caused by the brown alga Streblonema aecidioides. Infected algae show symptoms of Streblonema disease, i.e. alterations of the blade and stipe ranging from dark spots to heavy deformations and completely crippled thalli (Peters & Scaffelke, 1996). The occurrence of hyperplasia or gall growths, seen as dark spots, on Laminaria digitata is well known and may be associated with the presence of endophytic brown filamentous algae. Ectocarpus deformans, for example, was considered the cause of galls in Laminaria digitata by Apt (1988). In Helgoland, Ellertsdottir and Peters (1997) found 86% of Laminaria digitata thalli infected with endophytic brown algae and all those that exhibited weak to moderate but visible thallus alterations such as dark spots on the lamina or small warts on the stipe were infected. 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. Therefore, an intolerance of low has been recorded.|
|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 nodulosum (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. 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 resemble the sub-biotope A1.4121. Therefore, 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.|
|Several of the characterizing red algae species are subject to harvesting. 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 MarLIN review) and concluded that control levels of biomass and population structure are probably re-established after 18 months of regrowth. Palmaria palmata is used as a vegetable substitute or animal fodder although harvesting on a commercial scale only takes place in Ireland and France (Guiry & Blunden, 1991). Littorina littorea is also subject to harvesting in the UK and limpets in France. Hand collection may reduce the population of Littorina littorea within rockpools and hence reduce grazing pressure which may actually benefit the algal component of the biotope, especially opportunistic green algae and epiphytes.
Overall, while rockpools in areas subject to commercial algal harvesting may be directly affected, most examples of the biotope are unlikely to be affected by commercial harvesting in the UK. In deep pools characterized by this biotope, only the margins of the pool are likely to be affected. However, due to the relative small size of the community, even small scale hand collecting may have a significant effect. Therefore, an intolerance of intermediate has been recorded to represent the loss of a proportion of the macroalgae and the invertebrate community it supports, and loss of some littorinids. However, recovery is likely to be rapid since holdfasts and sporelings are likely to remain and the littorinids will probably recover quickly by migration and recruitment.
|Low||Very high||Very Low||No change||Low|
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.
Fucoids (e.g. Fucus serratus and Fucus vesiculosus) recruit readily to cleared areas, especially in the absence of grazers (Holt et al., 1997). However, fucoid propagules tend to settle near to the parent plants, due to turbulent deposition by water flow. Within monospecific stands recruitment of conspecifics is most likely, and community recovery is likely to be rapid. 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). However, in cleared areas, recruitment is likely to be rapid but recovery of the original community structure is likely to take some years (Holt et al., 1997). For example, after the Torrey Canyon oil spill, although maximum cover of fucoids occurred within 1-3 years, the abundance of barnacles increased in 1-7 years, limpet number were still reduced after 6-8 years and species richness was regained in 2 to >10 years (Southward & Southward, 1978; Hawkins & Southward, 1992; Raffaelli & Hawkins, 1999).
Sousa et al. (1981) reported that experimental removal of sea urchins significantly increased recruitment in long-lived brown algae. In experimental plots cleared of algae and sea urchins in December, Halidrys dioica colonized the plots, in small numbers, within 3-4 months. Plots cleared in August received few , if any recruits, suggesting that recolonization was dependant on zygote availability and therefore the season. Wernberg et al. (2001) suggested that the lack of long range dispersal success in Halidrys siliquosa was responsible for its regional distribution in the north east Atlantic
Corallina officinalis probably has good recruitment and settled on artificial substrata within 1 week of their placement in the intertidal during summer in New England (Harlin & Lindbergh, 1977). New fronds of Corallina officinalis appeared on sterilized plots within six months and 10% cover was reached with 12 months (Littler & Kauker 1984). Bamber & Irving (1993) reported that new plants grew back in scraped transects within 12 months, although the resistant crustose bases were probably not removed. Similarly, in experimental plots, up to 15% cover of Corallina officinalis fronds returned within 3 months after removal of fronds and all other epiflora/fauna but not the crustose bases (Littler & Kauker, 1984). Although new crustose bases may recruit and develop quickly the formation of new fronds from these bases and recovery of original cover may take longer, and it is suggested that the population is likely to recover within a few 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). While a recognizable biotope will return the exact community may differ from that present prior to perturbation. In addition, although the biotope is likely to be recognizable within less than 5 years, if the community was completely destroyed by perturbation, it may take longer for a typically diverse community to become established, especially the biotopes supported anemones and the rarer red algal species.
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Last Updated: 16/12/2015