|Researched by||Dr Heidi Tillin & Will Rayment||Refereed by||This information is not refereed.|
Low-lying rock surrounded by mobile sand and often subject to burying by the sand, with a turf of resilient red seaweeds Chondrus crispus, Polyides rotunda and Ahnfeltia plicata typically protruding through the sand on the upper surfaces of the rock. Other scour-tolerant seaweeds include Rhodomela confervoides, Phyllophora pseudoceranoides, Phyllophora crispa, Furcellaria lumbricalis, Gracilaria gracilis, Ceramium rubrum, Plocamium cartilagineum, Heterosiphonia plumosa, Cryptopleura ramosa and Dilsea carnosa. Coralline crusts typically cover the rock, while scattered individuals of the brown seaweeds Halidrys siliquosa, Cladostephus spongiosus, Dictyota dichotoma and Saccharina latissima can be present. The large anthozoan Urticina felina can occur in this biotope but there are few other conspicuous animals (Connor et al., 2004: JNCC).
The dominant algal species in the biotope are perennial and therefore present throughout the year. However, they do exhibit seasonality in terms of growth and reproduction. For example, maximum growth of Furcellaria lumbricalis occurs in March/April (Austin, 1960b) and release of carpospores and tetraspores occurs in December/January (Bird et al., 1991). Reproductive bodies are present on the gametophytes of Ahnfeltia plicata between July and January and mature carposporophytes occur between October and July (Maggs & Pueschel, 1989). The annual algal species, for example the filamentous greens, are likely to proliferate in spring and summer in conjunction with increased irradiance and temperatures, and then die back in autumn and winter.
Recruitment processes and recolonization by macroalgae are very dependent on time of year as spores are only available for limited periods. The advantage of being fertile through the winter, as in the case of Ahnfeltia plicata, Furcellaria lumbricalis and Chondrus crispus, is the availability of substrata for colonization as other annual species die back (Kain, 1975). Dickinson (1963) reported that Chondrus crispus was fertile in the UK from autumn to spring, but that the exact timing varied according to local environmental conditions. Similarly, Pybus (1977) reported that although carposporic plants were present throughout the year in Galway Bay, Ireland, maximum reproduction occurred in the winter and estimated that settling of spores occurred between January and May.
Storms and increased wave action are more likely to occur in the winter months and may cause physical damage to the community. Austin (1960b) reported damage to Furcellaria lumbricalis plants during storms and Sharp et al. (1993) reported that plants may be cast ashore by increased wave action. Dudgeon & Johnson (1992) noted wave induced disturbance of intertidal Chondrus crispus on shores of the Gulf of Maine, USA, during winter. 25-30% of cover of large Chondrus crispus thalli was lost in one winter. Physical disruption of the algal turf is likely to promote diversity as spaces become available for colonization.
Habitat complexity is provided by the mixed substratum of bedrock, cobbles, pebbles and mobile sand. It is this complexity which determines the species of algae which characterize the biotope. Only species tolerant of sand cover and sand scour, e.g. Ahnfeltia plicata, Polyides rotunda and Furcellaria lumbricalis, are able to persist in the community.
The dense algal turf provides shelter for a variety of fauna and sites for attachment of both epifauna (e.g. ascidians, bryozoans and hydroids) and epiphytes (Lewis, 1964).
Primary production by the slow growing, perennial red algae which dominate the biotope is probably low. Wallentinus (1978) measured in situ primary production by macroalgae in the northern Baltic Sea. Productivity of Furcellaria lumbricalis was 0.36-0.54 mg C/g dry wt/hour. The comparative figure for Cladophora glomerata, a filamentous green alga was 1.47-11.38 mg C/g dry wt/hour. These figures suggest that the contribution made by the perennial algal turf to macroalgal production in the biotope is likely to be very small. Fast growing, ephemeral, annual species with rapid turnover probably account for the majority of macroalgal primary production. However, the contribution to primary production of all macroalgae in the biotope is likely to be small in comparison with the phytoplankton. Jansson & Kautsky (1976), for example, recorded annual macroscopic plant production of hard bottoms in the Baltic shallow subtidal to be approximately 4% of the total primary production, suggesting that phytoplankton are by far the most important carbon fixers. Additionally, they noted that fast growing species with rapid turnover, for example the filamentous brown algae, contributed approximately one third of macroalgal production and that there was a relatively small contribution made by the slow growing perennials.
Vadas et al. (1992) reviewed recruitment and mortality of early post settlement stages of benthic algae. They identified 6 intrinsic and 17 extrinsic factors affecting recruitment and mortality. They concluded that grazing, canopy and turf effects were the most important but that desiccation and water movement may be as important for the early stages. The review indicated that recruitment is highly variable and episodic and that mortality of algae at this period is high. Chance events during the early post settlement stages are therefore likely to play a large part in survival.
As with all red algae, the spores of Ahnfeltia plicata, Chondrus crispus, Furcellaria lumbricalis and Polyides rotunda are non-flagellate and therefore dispersal is a wholly passive process (Fletcher & Callow, 1992). In general, due to the difficulties of re-entering the benthic boundary layer, it is likely that successful colonization is achieved under conditions of limited dispersal and/or minimum water current activity. Norton (1992) reported that although spores may travel long distances (e.g. Ulva sp. 35 km, Phycodrys rubens 5 km), the reach of the furthest propagule does not equal useful dispersal range, and most successful recruitment occurs within 10 m of the parent plants. It is expected, therefore, that recruitment of Ahnfeltia plicata, Chondrus crispus, Furcellaria lumbricalis, Polyides rotunda and the majority of other macroalgae in the biotope would occur from local populations and that establishment and recovery of isolated populations would be patchy and sporadic. Scrosati et al. (1994) commented that viability of spores of Chondrus crispus was low (<30%) and suggested that reproduction by spores probably does not contribute much to maintenance of the intertidal population in Nova Scotia, compared to vegetative growth of gametophytes.
As and when bare substratum becomes available for colonization, for instance following storm events, it is expected that algal recruitment and succession would follow a predictable sequence (Hawkins & Harkin, 1985). Initial colonizers on bare rock are often epiphytic species, suggesting that it is competition from canopy forming algae that usually restricts them to their epiphytic habit (Hawkins & Harkin, 1985). Gradually, the original canopy or turf forming species, in this case Ahnfeltia plicata, Furcellaria lumbricalis, Polyides rotunda and Chondrus crispus, then become established. These findings suggest that interactions between macrophytes are often more important than grazing in structuring algal communities (Hawkins & Harkin, 1985).
The anemone, Urticina felina, disperses via a large pelagic larvae (Chia & Spaulding, 1972) or may be able to brood its offspring until they are well developed (Spaulding, 1974). Either way the species has poor dispersive powers (Sole-Cava et al., 1994) and therefore is most likely to recruit from local populations.
Maturity of the community is likely to be limited by the time it takes the climax, perennial algae to settle, grow and reach reproductive viability. 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. Furcellaria lumbricalis grows even more slowly than Chondrus crispus (Bird et al., 1979) and may take 5 years to reach fertility (Austin, 1960b).
It is expected that, although the species which characterize the biotope would probably establish themselves after 2-3 years, a climax reproductive community may not be achieved until 5 years or more.
|Depth Range||5-10 m|
|Water clarity preferences|
|Limiting Nutrients||Nitrogen (nitrates), Phosphorus (phosphates)|
|Salinity preferences||Full (30-40 psu)|
|Physiographic preferences||Open coast|
|Biological zone preferences||Infralittoral|
|Substratum/habitat preferences||Bedrock, Cobbles, Pebbles, Sand|
|Tidal strength preferences||Moderately Strong 1 to 3 knots (0.5-1.5 m/sec.), Weak < 1 knot (<0.5 m/sec.)|
|Wave exposure preferences||Exposed, Moderately exposed|
The biotope description and characterizing species is taken from Connor et al. (2004). This biotope occurs on low-lying rock surrounded by mobile sand and often subject to burying by the sand, with a turf of resilient red seaweeds Chondrus crispus, Polyides rotunda and Ahnfeltia plicata typically protruding through the sand on the upper surfaces of the rock. The sensitivity assessments are based on these named characterizing species. Other scour-tolerant seaweeds that may be present include Rhodomela confervoides, Phyllophora pseudoceranoides, Phyllophora crispa, Furcellaria lumbricalis, Gracilaria gracilis, Ceramium rubrum, Plocamium cartilagineum, Heterosiphonia plumosa, Cryptopleura ramosa and Dilsea carnosa. The sensitivity of theses is considered generally. Coralline crusts typically cover the rock, while scattered individuals of the brown seaweeds Halidrys siliquosa, Cladostephus spongiosus, Dictyota dichotoma and Laminaria saccharina can be present. Evidence is presented for the sensitivity of these species and the large anthozoan Urticina felina where available.
This biotope occurs in areas subject to chronic disturbance from scour and sediment mobility, these factors maintain the biotope by preventing the growth of larger and more-long lived species that are typical of more stable biotopes occurring in the infralittoral. As an early successional biotope subject to frequent perturbation, species present will either be adapted to resist the stressors acting on the biotope or to recover rapidly. As the substratum is relatively mobile, it is likely that the scattered kelps and other larger brown algae, such as Halidrys siliqosa that occur in this biotope are usually present as smaller, seasonal recruits that are removed during periods of disturbance and regrow annually. These larger species do not define the biotope and recovery of this element is assessed as ‘High’ at all levels of impact (resistance is High, Medium, Low or None).
The turf forming red algae is the key group characterizing this biotope. Depending on the level of impact, recovery of the turf may occur through repair and regrowth of damaged fronds, regrowth from crustose bases or via recolonization of rock surfaces where all the plant material is removed. Although there are few case studies following recovery some general trends are apparent. All the red algae (Rhodophyta) exhibit distinct morphological stages over the reproductive life history. This phenomenon is known as heterotrichy or heteromorphy and describes cases where the algal thallus consists of two parts; a prostrate creeping system exhibiting apical growth and functioning as a holdfast. The thalli can regrow from these crusts where they remain supporting recovery of the biotope (Mathieson & Burns, 1975; Dudgeon & Johnson, 1992). The basal crusts are perennial, tough, resistant stages that prevent other species from occupying the rock surface and allow rapid regeneration and where these remain they provide a significant recovery mechanism. The bases can spread laterally across rock and for species such the characterizing Chondrus crispus they coalesce over time and can form an extensive crust on rock (Taylor et al., 1981). Some species exhibit annual growth and die back patterns. In Phyllophora pseudoceranoides for example the plant is perennial but blades are lost and regrown each year (Molenaar & Breeman, 1994). Similarly Plocamium cartilagineum, loses blades in winter in wave exposed conditions while the crustose bases survive and spread laterally (Kain, 1987). Some temporal variation in abundance and biomass is therefore normal within this biotope. Resistant crustose bases enable the turf of red algae and the crustose corallines to withstand and recover from physical disturbance and scour while preventing the establishment of other species.
Of the characterizing species Chondrus crispus is the most studied red algae within this biotope due to its commercial value. Growth patterns vary seasonally with the highest biomass usually in late Spring or Summer and lowest in Winter (Fernández & Menéndez, 1991). Pybus (1977) estimated that Chondrus crispus from Galway Bay, Ireland, reached maturity approximately 2 years after the initiation of the basal disc, at which stage, the fronds were approximately 12 cm in length. The fronds of Chondrus crispus typically have a life of 2-3 years (Taylor, cited in Pringle & Mathieson, 1986) but may live up to 6 years in sheltered waters (Harvey & McLachlan, 1973). Dickinson (1963) reported that Chondrus crispus was fertile in the UK from autumn to spring, but that the exact timings varied according to local environment. Similarly, Pybus (1977) reported that in Galway Bay, Ireland, maximum reproduction occurred in the winter and estimated that settling of spores occurred between January and May. In Nova Scotia, and most likely other areas where conditions are not optimal for reproduction and settlement, reproduction by spores probably does not contribute much to maintenance of population of Chondrus crispus compared to vegetative growth of gametophytes (Scrosati et al., 1994). Most of the evidence for recovery of Chondrus crispus is based on intertidal experiments that simulate the effects of different harvesting mechanisms and intensities (Macfarlane, 1952; Mathieson & Burns, 1975). Macfarlane (1952) in a series of experiments identified that where Chondrus crispus was removed by cutting of fronds or thorough raking (leaving the crusts undamaged) the turf had recovered and there were no notable differences bwetween the experimental areas and control sites. However, where the crusts were removed by scraping or damaged the experimental plots were still recovering nearly two years after the treatment. Following experimental harvesting by drag raking (hodfasts and small blades undamaged) 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). The authors suggested that control levels of biomass and reproductive capacity are probably reestablished after 18 months of regrowth (where crusts are not removed). 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. Similarly MacFarlane (1952) reports that the particularly harsh winter of 1947/48 destroyed Chondrus beds near Pubnico. By the next summer, the annual brown alga Chordaria had colonized the area, and by the summer of 1950, Fucus had taken over as the dominant successional stage. Chondrus did not noticeably start to grow back in the area until summer 1951, four years later. Pringle and Semple (1980) estimated it would take about four years for a bare patch in a Chondrus bed to fill in with harvestable plants and five to ten years for Chondrus to re-establish in barren areas. The applicability of recovery patterns and rates from intertidal biotopes to this scoured, subtidal biotope is not clear
No information was found concerning the longevity of Ahnfeltia plicata. However, it is a slow maturing perennial (Dickinson, 1963) and the thallus survives several years without considerable losses (Lüning, 1990). Maggs & Pueschel (1989) reported that mature gametophytes in Nova Scotia varied in size from 3-21 cm, and that 14 months after germination, gametophyte fronds had reached up to 5 cm in length. No definitive information was found concerning age at maturity. However, Maggs & Pueschel (1989) made observations of Ahnfeltia plicata from Nova Scotia. Tetrasporophyte crusts matured and released tetraspores after 15 months. The associated red algae Furcellaria lumbricalis also grows very slowly compared to some other red algae (Bird et al., 1979) and takes a long time to reach maturity. For example, Austin (1960b) reported that in Wales, Furcellaria lumbricalis typically takes 5 years to attain fertility.
The spores of red algae are non-motile (Norton, 1992) and therefore entirely reliant on the hydrographic regime for dispersal. Norton (1992) reviewed dispersal by macroalgae and concluded that dispersal potential is highly variable, recruitment usually occurs on a much more local scale, typically within 10 m of the parent plant. Hence, it is expected that the red algal turf would normally rely on recruit from local individuals and that recovery of populations via spore settlement, where adults are removed, would be protracted.
Coralline crust is a generic term that in UK biotopes refers to nongeniculate (crustose) species from the family Corallinacea that could include Lithophyllum incrustans , Lithothamnion spp. and Phymatolithon spp. Due to the lack of evidence the assessments are generic, although species specific information is presented where available. A number of papers by Edyvean & Ford (1984a & b; 1986; 1987) describe aspects of reproduction and growth of encrusting coralline, Lithophyllum incrustans. Studies by Edyvean & Forde (1987) in populations of Lithophyllum incrustans in Pembroke south-west Wales suggest that reproduction occurs on average early in the third year. Reproduction may be sexual or asexual. Populations release spores throughout the year but abundance varies seasonally, with the populations studied in Cullercoats Bay, and Lannacombe Bay (North East and South West England, respectively) producing less spores in the summer. Spore release is initiated by changes in temperature or salinity (see relevant pressure information) at low tide so that spore dispersal is restricted to within the tide pool enhancing local recruitment. Witihn subtidal biotopes this is not possible and recruitment success may be altered (although this may be compensated by avoidance of desiccation). Spore survival is extremely low with only a tiny proportion of spores eventually recruiting to the adult population (Edyvean & Ford, 1986). The spores are released from structures on the surface called conceptacles, these are formed annually and subsequently buried by the new layer of growth. Plants can be aged by counting the number of layers of conceptacles. Edyvean & Ford (1984a) found that the age structure of populations sampled from Orkney (Scotland) Berwick (northern England) and Devon (England) were similar, mortality seemed highest in younger year classes with surviving individuals after the age of 10 years appear relatively long-lived (up to 30 years). In St Mary’s Northumberland, the population was dominated by the age 6-7 year classes (Edyvean & Ford, 1984a). Growth rates were highest in young plants measured at Pembroke (south-west Wales) with an approximate increase in diameter of plants of 24 mm in year class 0 and 155 mm in year 1 and slowing towards an annual average horizontal growth rate of 3mm/year (Edyvean & Ford, 1987).
Some repair of damaged encrusting coralline occurs through vegetative growth. Chamberlain (1996) observed that although Lithophyllum incrustans was quickly affected by oil during the Sea Empress spill, recovery occurred within about a year. The oil was found to have destroyed about one third of the thallus thickness but regeneration occurred from thallus filaments below the damaged area. Recolonization by propagules is also an important recovery mechanism, Airoldi (2000) observed that encrusting coralline algae recruited rapidly on to experimentally cleared subtidal rock surfaces in the Mediterranean Sea, reaching up to 68% cover in 2 months.
Little information is available for life-history and reproductive strategies to inform a recovery assessment for the large anemone Urticina felina. Some damage to individuals can be repaired, for example removal of tentacles by clipping does not alter behaviour and the tentacle regenerates within a few days (Mercier et al., 2011). Recovery is likely to be slow in populations where nearby individuals do not exist as the species broods juvenile stages and does not have a pelagic dispersal phase. Dispersal ability was considered to be poor in the similar species Urticina eques (Solé-Cava et al. 1994). The large size, slow growth rate and evidence from aquarium populations suggest that Urticina felina is long lived (Hiscock, pers comm.). Adults can detach from the substratum and relocate but locomotive ability is very limited. Impacts that remove large proportions of the population over a wide area will effectively reduce the availability of colonists. However, the species colonized ex-HMS Scylla in the fourth year of the vessel being on the seabed (Sköld et al., 2001).
Resilience assessment. Growth, maturity and longevity vary between the turf forming red algae species. Although some general trends are apparent. Recovery rates, for example, will be greatly influenced by whether the crust stages remain from which the thalli can regrow of the characterizing red algae. Biotope resilience is assessed as ‘High’ where resistance is ‘High’. Where resistance is assessed as ‘Medium’ (loss of <25% of individuals or cover) and the bases remain then recovery is assessed as ‘High’ based on regrowth from crusts and remaining plants. Where resistance is assessed as 'Medium, ‘Low’ or ‘None’, and a high proportion of bases are lost, then recovery will depend on either vegetative regrowth from remaining bases or the supply of propagules from neighbouring populations. Dispersal is limited and propagule supply will be influenced site-specific factors, particularly local water transport. Where resistance is assessed as ‘Low’ or ‘None’ then resilience is assessed as ‘Medium’ (2-10 years) for the red algal turf, encrusting corallines and Urticina felina. Biotope composition may be altered in favour of species with better dispersal ability and higher growth rates but some variation is natural.
NB: The resilience and the ability to recover from human induced pressures is a combination of the environmental conditions of the site, the frequency (repeated disturbances versus a one-off event) and the intensity of the disturbance. Recovery of impacted populations will always be mediated by stochastic events and processes acting over different scales including, but not limited to, local habitat conditions, further impacts and processes such as larval-supply and recruitment between populations. Full recovery is defined as the return to the state of the habitat that existed prior to impact. This does not necessarily mean that every component species has returned to its prior condition, abundance or extent but that the relevant functional components are present and the habitat is structurally and functionally recognizable as the initial habitat of interest. It should be noted that the recovery rates are only indicative of the recovery potential.
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
The key characterizing red algal species found in this biotope have broad geographic distributions and are found in warmer waters than those around the UK. As this biotope is subtidal it is protected from exposure to air so that the thermal regime is more stable and desiccation is not a factor. Examples of distribution and thermal tolerances tested in laboratory experiments are provided as evidence to support assessments. Populations can acclimate to prevailing conditions which can alter tolerance thresholds and care should therefore be used when interpreting reported tolerances.
The key characterizing species Chondrus crispus has a wide distribution, it is found extensively throughout Europe and North America records also recorded under a number of synonyms from Africa and Asia (Guiry & Guiry, 2015). Spore germination in Chondrus crispus appears to be temperature dependent with spores discharged at temperatures of 5oC failing to germinate although in laboratory culture at 10oC spores were viable all year round (Bhattacharya, 1985). In New Hampshire, USA, Chondrus crispus grows abundantly in waters with an annual variation in surface temperature from -1 to +19°C (Mathieson & Burns, 1975). The optimum temperature for growth has been reported as 10-15°C (Fortes & Lüning, 1980), 15°C (Bird et al., 1979), 15-17°C (Tasende & Fraga, 1999) and 20°C (Simpson & Shacklock, 1979). Above the optimum temperature, growth rate is reported to decline (Bird et al., 1979; Simpson & Shacklock, 1979). Compared to Chondrus crispus plants grown at 5°C, plants grown at 20°C had higher growth rates in terms of length, biomass, surface area, dichotomy and branch production. The differences resulted in growth of morphologically more complex thalli at higher temperatures with more efficient nutrient exchange and light harvesting (Kuebler & Dudgeon, 1996). Chondrus crispus plants acclimated to growth at 20°C (vs. 5°C) had higher levels of chlorophyll a and phycobilins, resulting in higher rates of light limited photosynthesis for a given photon flux density (Kuebler & Davison, 1995). Plants grown at 20°C were able to maintain constant rates of light saturated photosynthesis at 30°C for 9 hours. In contrast, in plants acclimated to 5°C, light saturated photosynthetic rates declined rapidly following exposure to 30°C (Kuebler & Davison, 1993). Prince & Kingsbury (1973) reported cessation of growth in Chondrus crispus cultures at 26°C, first mortality of spores at 21.1°C and total mortality of spores at 35-40°C, even if exposed for just 1 minute.
The key characterizing species Ahnfeltia plicata, has a very wide geographic range, occurring from northern Russia to Portugal. The species is therefore likely to be tolerant of higher temperatures than it experiences in Britain and Ireland. Lüning & Freshwater (1988) incubated Ahnfeltia plicata from British Columbia at a range of temperatures for 1 week and tested their survivability by ability to photosynthesize at the end of the incubation period. The species survived from the coldest temperature tested (-1.5°C) to 28 °C. Total mortality occurred at 30 °C. Lüning & Freshwater (1988) suggested that Ahnfeltia plicata was therefore amongst the group of most eurythermal heat tolerant algae. Haglund et al. (1987) incubated Ahnfeltia plicata from the subtidal in Sweden at a range of temperatures and measured photosynthetic rate. There were no significant results, but photosynthetic rate appeared to be optimal at 15°C and decreased either side of this temperature.The key characterizing species Polyides rotunda, is found throughout the North Atlantic Ocean and related populations occur in the North Pacific. Growth and survival was tested over a temperature range of -5 to 30 oC. Polyides rotunda tolerated temperatures from -5 to 27 oC, grew well from 5 to 25 oC, and had a broad optimal range of 10-25 oC. This species tolerated 3 months in darkness at 0 oC (Novaczek & Breeman, 1990).
The associated species Furcellaria lumbricalis has a wide geographic range, occurring in Europe from northern Norway to the Bay of Biscay. Novaczek & Breeman (1990) recorded that specimens of Furcellaria lumbricalis grew well in the laboratory from 0-25°C with optimal growth between 10 and 15°C. Growth ceased at 25°C and 100% mortality resulted after 3 months exposure to 27°C. Similarly, Bird et al. (1979) recorded optimum growth at 15°C and cessation of growth at 25°C with associated necrosis of apical segments. Samples of Phyllophora pseudoceranoides from Nova Scotia, Iceland, Roscoff and Helgoland grew from 3 oC to 25 oC and survived from -2 oC to 27 oC but not 30 oC. Luning (1984) found that Phyllophora pseudoceranoides collected from Helgoland died after 1 week exposure to 33 oC (Molenaar & Breeman,1994). Edyvean & Ford (1984b) suggest that populations of Lithophyllum incrustans are affected by temperature changes and salinity and that temperature and salinity ‘shocks’ induce spawning but no information on thresholds was provided (Edyvean & Ford, 1984b). Populations of Lithophyllum incrustans were less stable in tide pools with a smaller volume of water that were more exposed to temperature and salinity changes due to lower buffering capacity. Sexual plants (or the spores that give rise to them) were suggested to be more susceptible than asexual plants to extremes of local environmental variables (temperature, salinity etc.) as they occur with greater frequency at sites where temperature and salinity were more stable (Edyvean & Forde, 1984b). Lithophyllum incrustans in the UK is close to the northern edge of its range and is likely to tolerate increased temperatures.
Sensitivity assessment. The geographic distribution and laboratory experiments indicate that the key characterizing species and many of the associated red algal turf species would tolerate either an acute or chronic increase in temperature at the pressure benchmark, although some sub-lethal decreases in photosynthesis may occur where temperatures exceed the optimal. Gamete release may also depend on lower winter temperatures and therefore non-lethal effects on reproduction may occur where optimal temperatures are exceeded. As these effects do not result in mortality, resistance is assessed as ‘High’ and recovery as ‘High’ (by default) so that the biotope is considered to be ‘Not sensitive’ to short-term (not grater than a year) changes at the pressure benchmark..
The key characterizing red algal species found in this biotope have broad geographic distributions and are found in colder waters than those around the UK. As this biotope is subtidal it is protected from exposure to air so that the thermal regime is more stable. Examples of distribution and thermal tolerances tested in laboratory experiments are provided as evidence to support assessments. Populations can acclimate to prevailing conditions which can alter tolerance thresholds and care should therefore be used when interpreting reported tolerances.
The key characterizing species, Chondrus crispus has a broad geographical distribution (Guiry & Guiry, 2015) and throughout the range experience wide variation in temperatures (although local populations may be acclimated to the prevailing thermal regime). In New Hampshire, USA, Chondrus crispus grows abundantly in waters with an annual variation in surface temperature from -1 to +19 °C (Mathieson & Burns, 1975). The photosynthetic rate of Chondrus crispus recovered after 3hrs at -20 °C but not after 6 hrs (Dudgeon et al. (1989). Frond bleaching and declines in photosynthesis and growth also occur in long-term experimental exposure to periodic freezing in Chondrus crispus (Dudgeon et al., 1990). Plants from Maine, USA, were frozen at -5 °C for 3 hours a day for 30 days. Photosynthesis was reduced to 55 % of control values, growth rates were reduced and fronds were eventually bleached and fragmented resulting in biomass losses. Additionally, fronds of Chondrus crispus which were frozen daily had higher photosynthetic rates following subsequent freezing events than unfrozen controls, indicating that the species is able to acclimate to freezing conditions (Dudgeon et al., 1990). Spore germination in Chondrus crispus appears to be temperature dependent with spores discharged at temperatures of 5 oC failing to germinate although in laboratory culture at 10 oC spores were viable all year round (Bhattacharya, 1985). Acute or chronic changes in temperature below 5 oC may therefore reduce reproductive success although reproduction and vegetative growth in warmer months should compensate for any reduction in output.
Ahnfeltia plicata has a very wide geographic range, occurring from northern Russia to Portugal. The species is therefore likely to be tolerant of lower temperatures than it experiences in Britain and Ireland. Lüning & Freshwater (1988) incubated Ahnfeltia plicata from British Columbia at a range of temperatures for 1 week and tested their survivability by ability to photosynthesize at the end of the incubation period. The species survived from the coldest temperature tested (-1.5 °C) to 28 °C. Haglund et al. (1987) incubated Ahnfeltia plicata from the subtidal in Sweden at a range of temperatures and measured photosynthetic rate. There were no significant results, but photosynthetic rate appeared to be optimal at 15 °C and decreased either side of this temperature.
Growth and survival of the key characterizing species, Polyides rotunda was tested over a temperature range of -5 to 30 oC. Polyides rotunda tolerated temperatures from -5 to 27 oC, grew well from 5 to 25 oC, and had a broad optimal range of 10-25 oC (Novaczek & Breeman, 1990). This species tolerated 3 months in darkness at 0 oC (Novaczek & Breeman, 1990).
Edyvean & Forde (1984b) suggest that populations of Lithophyllum incrustans are affected by temperature changes and salinity and that temperature and salinity ‘shocks’ induce spawning but no information on thresholds was provided (Edyean & Ford, 1984b).
Furcellaria lumbricalis has a wide geographic range, occurring in Europe from northern Norway to the Bay of Biscay. Novaczek & Breeman (1990) recorded that specimens of Furcellaria lumbricalis grew well in the laboratory from 0-25 °C with optimal growth between 10 and 15 °C. Growth ceased at 25 °C and 100 % mortality resulted after 3 months exposure to 27 °C. Similarly, Bird (1979) recorded optimum growth at 15 °C.
Sensitivity assessment. The key characterizing species occur over a wide geographical range and can tolerate temperatures below 0 oC. Reduced temperatures may result in suboptimal growth and may affect reproduction. Biotope resistance is assessed as ‘High’ as these effects do not result in mortality and resilience is assessed as ‘High’, so that the biotope is not considered to be sensitive to this pressure.
This biotope occurs in full salinity (Connor et al., 2004) a change above the pressure benchmark is assessed as a change to above 40 ppt. Species within this biotope such as Chondrus crispus that are also found in intertidal biotopes including rock pools are likely to have some form of physiological adaptations to increases in salinity as these habitats encounter more variation in salinity. However, local populations are likely to be acclimated to the prevailing conditions even over small spatial distances and gradients and caution should be used in extrapolating sensitivities.
More evidence was found to assess Chondrus crispus than the other turf forming red algae. Chondrus crispus is found in a range of salinities across its range and has been reported from sites with yearly salinity range 0-10 psu and 10-35 psu (Lindgren & Åberg 1996) and sites from an average of 26-32 psu. However, at different salinities the ratio between the abundance of the tetrasporophyte phase and the gametophyte alters (Guidone & Grace, 2010). Mathieson & Burns (1971) recorded maximum photosynthesis of Chondrus crispus in culture at 24 psu, but rates were comparable at 8, 16 and 32 psu. Photosynthesis continued up to 60 psu. Bird et al. (1979) recorded growth of Canadian Chondrus crispus in culture between 10 and 50 psu, with a maximum at 30 psu. Chondrus crispus would therefore appear to be euryhaline and tolerant of a range of salinities.
Haglund et al. (1987) studied photosynthetic rate of Ahnfeltia plicata from the subtidal in Sweden and found that, at constant temperature, rate increased up to the maximum salinity tested (33 psu).
Furcellaria lumbricalis is a euryhaline species which occurs in a wide range of salinity conditions down to 6-8 psu (Bird et al., 1991). In the Kattegat and the Gulf of St Lawrence, it is reported to compete well with other species at salinities ranging from 25-32 psu (see review by Bird et al., 1991). Growth experiments in the laboratory revealed that optimum growth occurred at 20 psu, the species grew well at 10 psu and 30 psu, but that growth declined above 30 psu to negligible levels at 50 psu (Bird et al., 1979). It is expected that an increase in salinity may cause reduced growth and fecundity, but that mortality is unlikely.
Edyvean & Ford (1984b) suggest that populations of the crustose coralline Lithophyllum incrustans are affected by temperature changes and salinity and that temperature and salinity ‘shocks’ induce spawning but no information on thresholds was provided (Edyvean & Ford, 1984b). Populations of Lithophyllum incrustans were less stable in rockpools with a smaller volume of water that were more exposed to temperature and salinity changes due to lower buffering capacity. Sexual plants (or the spores that give rise to them) were suggested to be more susceptible than asexual plants to extremes of local environmental variables (temperature, salinity etc.) as they occur with greater frequency at sites where temperature and salinity were more stable (Edyvean & Ford, 1984b).
Sensitivity assessment. No specific evidence was found for the salinity tolerance of the key characterizing species Polyides rotunda. Based on the reported salinity tolerances of Chondrus crispus and other red algae species, it is considered that an increase at the pressure benchmark may lead to some changes in the composition of the red algal turf that characterizes this biotope and reduce species richness and abundance, although some more tolerant species may persist. Resistance (of the biotope) is therefore assessed as ‘Low’ and resilience as ‘Medium’ and biotope sensitivity is assessed as ‘Medium’.
As this biotope is present in full salinity (30-35 ppt, Connor et al., 2004), the assessed change at the pressure benchmark is a reduction in salinity to a variable regime (18-35 ppt) or reduced regime (18-30 ppt). Populations can acclimate to local conditions and caution should be used in extrapolating observations from areas where populations are likely to he adapted to the prevailing conditons.A comparative study of salinity tolerances of macroalgae collected from North Zealand in the South Kattegat (Denmark) where salinity is 16 psu. Showed that species generally had a high tolerance (maintained more than half of photosynthetic capacity) to short-term exposure (4 days) to salinities lower than 3.7. However, tolerances varied between species with Ahnfeltia plicata, Phyllophora pseudoceranoides and Chondrus crispus exhibiting greater tolerance than Rhodomela confervoides (Larsen & Sand-Jensen, 2006). The result illustrates that responses to this pressure will vary between species and that a change at the pressure benchmark is likely to alter the composition of the red algal turf that characterizes the biotope and may alter the biomass and density of more tolerant species.
Chondrus crispus is found in a range of salinities across its range and has been reported from sites with yearly salinity range 0-10 psu and 10-35 psu (Lindgren & Åberg 1996) and sites from an average of 26-32 psu. However, at different salinities the ratio between the abundance of the tetrasporophyte phase and the gametophyte alters (Guidone & Grace, 2010). Mathieson & Burns (1971) recorded maximum photosynthesis of Chondrus crispus in culture at 24 psu, but rates were comparable at 8, 16 and 32 psu. Photosynthesis continued up to 60 psu. Bird et al. (1979) recorded growth of Canadian Chondrus crispus in culture between 10 and 50 psu, with a maximum at 30 psu. Chondrus crispus would therefore appear to be euryhaline and tolerant of a range of salinities.
Ahnfeltia plicata occurs over a very wide range of salinities. The species penetrates almost to the innermost part of Hardanger Fjord in Norway where it experiences very low salinity values and large salinity fluctuations due to the influence of snowmelt in spring (Jorde & Klavestad, 1963). Ahnfeltia plicata penetrates further than the euryhaline species, Polyides rotunda, and probably has a similar salinity tolerance to Furcellaria lumbricalis, which is limited only by the 4 psu isohaline (see review by Bird et al., 1991). Haglund et al. (1987) studied photosynthetic rate of Ahnfeltia plicata from the subtidal in Sweden and found that, at constant temperature, photosynthesis was positively correlated with salinity between 15 and 33 psu. It is likely therefore that the benchmark decrease in salinity would not result in mortality, but photosynthesis would not be optimal and so growth and reproduction may be compromised.
Furcellaria lumbricalis occurs in the lowest category on the salinity scale (Connor et al., 1997a) and therefore probably relatively tolerant of decreases in salinity. The species forms extensive populations in the main basin of the Baltic Sea where salinity is 6-8 psu in the upper 60-70 m and its extension into the Gulfs of Bothnia and Finland is limited by the 4 psu isohaline (see review by Bird et al., 1991).
Rhodomela confervoides shows local acclimation to prevailing salinity regimes in the brackish Baltic Sea, with populations surviving at 11 ppt or 2.5 ppt depending on the local salinity where collected (Rietema, 1995).
Edyvean & Ford (1984b) suggest that populations of the crustose coralline Lithophyllum incrustans are affected by temperature changes and salinity and that temperature and salinity ‘shocks’ induce spawning but no information on thresholds was provided (Edyvean & Ford, 1984b). Populations of Lithophyllum incrustans were less stable in tide pools with a smaller volume of water that were more exposed to temperature and salinity changes due to lower buffering capacity. Sexual plants (or the spores that give rise to them) were suggested to be more susceptible than asexual plants to extremes of local environmental variables (temperature, salinity etc.) as they occur with greater frequency at sites where temperature and salinity were more stable (Edyvean & Ford, 1984b).
The anemone Urticina felina occurs in estuaries e.g. the Thames estuary at Mucking and the River Blackwater estuary (Davis, 1967). Braber and Borghouts (1977) found that Urticina (as Tealia) felina penetrated to about the 11ppt Chlorinity isohaline (corresponding to about 20psu based on conversion rates) at mid tide during average water discharge in the Westerschelde estuary suggesting that it would be tolerant of reduced salinity conditions. Intertidal and rock pool individuals will also be subject to variations in salinity because of precipitation on the shore; albeit for short periods on the lower shore. Therefore, the species seems to have a high tolerance to reduction in salinity but may have to retract tentacles and suffer reduced opportunity to feed.
Sensitivity assessment. Two of the key defining species,Chondrus crispus and Ahnfeltia plicata are euryhaline and occur over a range of salinities. Other species associated with the biotope such as the red algae Furcellaria lumbricalis and Urticina felina are likely to tolerate a reduction in alinity at the pressure benchmark. Some changes in lgal composition may occur as a response to a decrease in salinity but may not significantly alter it from the biotope description. Resistance (of the biotope) is therefore assessed as ‘Medium’ and resilience as ‘Medium’ (as bases may not remain to enhance recovery), and biotope sensitivity is judged to be 'Medium'.
This biotope occurs across a range of flow speeds, from moderately strong (0.5-1.5 m/s) to areas where water flow is negligible (Connor et al., 2004). As water velocity increases foliose macroalgae can flex and reconfigure to reduce the size of the alga when aligned with the direction of flow, this minimises drag and hence the risk of dislodgement (Boller & Carrington, 2007). These characteristics allow these species to persist in areas that experience a range of flow speeds. Biogenic habitat structures, including the fronds of algae, reduce the effects of water flows on individuals by slowing and disrupting flow. Boller and Carrington (2006) found that the canopy created by a turf of Chondrus crispus reduced drag forces on individual plants by 15-65%. The compact, turf forming growth of the algal species will therefore reduce water flow and the risk of displacement through turbulence and friction.
The characterizing red algal turf is unlikely to be affected by changes in water flow alone as these can endure high current speeds. For example, Chondrus crispus occurs at intertidal sites in Maine, USA experiencing peak Autumn flow speeds as measured by current meters of 9.2 m/s and 5.8 m/s.
The anemone Urticina felina favours areas with strong tidal currents (Holme & Wilson, 1985; Migné & Davoult, 1997a) although it is also found in calmer and sheltered areas as well as deep water. Records from the MNCR database were used as a proxy indicator of the resistance to water flow changes by this species by Tillin & Tyler-Walters (2014). The records indicate the water flow categories for biotopes characterized by Urticina felina range from very strong to very weak (negligible) (negligible to >3m/s) suggesting that a change in the maximum water flow experienced by mid-range populations for the periods of peak spring tide flow would not have negative effects (Tillin & Tyler-Walters, 2014).
Scour is a key factor structuring this biotope, changes in the flow may increase or decrease sediment transport and associated scour. Reductions in flow may lead to increased deposition of silts and alter the sediment character. An increase in water flow at the pressure benchmark may re-suspend and remove sand particles which are less cohesive than mud particles. The level of impact will depend on site specific hydrodynamic and sediment conditions. Some periodic movement of sediments and changes in coverage is part of the natural temporal variation and periodic disturbances from storms may be more important than water flow in maintaining the character of the biotope, particularly in sheltered areas.
Sensitivity assessment. As the biotope can occur in a range of flow speeds, resistance of the biotope to changes in water flow that do not alter the substrata is assessed as ‘High’ and resilience as ‘High’ (by default) so that the biotope is assessed as ‘Not sensitive’.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Changes in emergence are not relevant to this biotope (group) which is restricted to fully subtidal habitats. Many of the species found in this biotope such as Ahnfeltia plicata, Chondrus crispus, Rhodomela confervoides occur intertidally but a change in emergence would alter the character of the habitat and lead to biotope reclassification.
This biotope is recorded from locations that are judged to range from extremely exposed or exposed (Connor et al., 2004). The degree of wave exposure influences wave height, as in more exposed areas with a longer fetch waves would be predicted to be higher. As this biotope occurs across a range of exposures, this was therefore considered to indicate, by proxy, that biotopes in the middle of the wave exposure range would tolerate either an increase or decrease in significant wave height at the pressure benchmark. As water movement increases foliose macroalgae can flex and reconfigure to reduce the size of the alga when aligned with the direction of flow, this minimises drag and hence the risk of dislodgement (Boller & Carrington, 2007). These characteristics allow these species to persist in areas that experience a range of flow speeds resulting from wave action. The crustose corallines associated with this biotope have a flat growth form and are unlikely to be dislodged by increased wave action.
Records from the MNCR database were used as a proxy indicator of the resistance to wave height changes by Urticina felina. The latest version of the JNCC National Biodiversity Database was used as the source of the MNCR data. The records indicate the wave exposure categories for biotopes characterized by members of this ecological group as extremely sheltered; very sheltered; sheltered; moderately exposed; exposed; very exposed (Tillin & Tyler-Walters, 2014).
Sensitivity assessment. The biotope is found across a range of wave exposures, mid-range populations are considered to have 'High' resistance to a change in significant wave height at the pressure benchmark. Resilience is assessed as ‘High’, by default, and the biotope is considered ‘Not sensitive’.
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed but evidence is presented where available.
Contamination at levels greater than the pressure benchmark may adversely impact the biotope. Little information was found concerning the intolerance of Chondrus crispus to heavy metals. Burdin & Bird (1994) reported that both gametophyte and tetrasporophyte forms 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. 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.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed but evidence is presented where available
Contamination at levels greater than the pressure benchmark may adversely impact the biotope. The long-term effects on Chondrus crispus of continuous doses of the water accommodated fraction (WAF) of diesel oil were determined in experimental mesocosms (Bokn et al., 1993). Mean hydrocarbon concentrations tested were 30.1 µg/l and 129.4 µg/l. After 2 years, there were no demonstrable differences in the abundance patterns of Chondrus crispus. Kaas (1980) (cited in Holt et al., 1995) reported that the reproduction of adult Chondrus crispus plants on the French coast was normal following the Amoco Cadiz oil spill. However, it was suggested that the development of young stages to adult plants was slow, with biomass still reduced 2 years after the event. O'Brien & Dixon (1976) and Grandy (1984) (cited in Holt et al., 1995) comment on the high intolerance of red algae to oil/dispersant mixtures, but it is unclear which factor is responsible for the intolerance.
Crump et al. (1999) described "dramatic and extensive bleaching" of 'Lithothamnia' following the Sea Empress oil spill. Observations following the Don Marika oil spill (K. Hiscock, pers. comm.) were of rockpools with completely bleached coralline algae. However, Chamberlain (1996) observed that although Lithophyllum incrustans was affected in a short period of time by oil during the Sea Empress spill, recovery occurred within about a year. The oil was found to have destroyed about one third of the thallus thickness but regeneration occurred from thallus filaments below the damaged area.
The examples stated refer to intertidal areas, where the contaminants consist of lighter fractions and float on water subtidal habitats will be less exposed.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed but evidence is presented where available.
Contamination at levels greater than the pressure benchmark may adversely impact the biotope. No evidence was found specifically relating to the intolerance of Chondrus crispus to synthetic chemicals. However, inferences may be drawn from the sensitivities of red algal species generally. 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 report that red algae are effective indicators of detergent damage since they undergo colour changes when exposed to relatively low concentration of detergent. Smith (1968) reported that 10 ppm of the detergent BP 1002 killed the majority of specimens in 24hrs in toxicity tests, although Chondrus crispus was amongst the algal species least affected by the detergent used to clean up the Torrey Canyon oil spill. 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). Cole et al. (1999) suggested that herbicides, such as simazine and atrazine, were very toxic to macrophytes. The evidence suggests that in general red algae are very sensitive to synthetic chemicals. Intolerance of Chondrus crispus is therefore recorded as high.
Cole et al. (1999) suggested that herbicides were (not surprisingly) very toxic to algae and macrophytes. Hoare & Hiscock (1974) noted that with the exception of Phyllophora species, all red algae including encrusting coralline forms, were excluded from the vicinity of an acidified halogenated effluent discharge in Amlwch Bay, Anglesey and that intertidal populations of Corallina officinalis occurred in significant amounts only 600m east of the effluent. Chamberlain (1996) observed that although Lithophyllum incrustans was quickly affected by oil during the Sea Empress spill, recovery occurred within about a year. The oil was found to have destroyed about one third of the thallus thickness but regeneration occurred from thallus filaments below the damaged area.
|No evidence (NEv)||Not relevant (NR)||No evidence (NEv)|
No evidence was found to assess this pressure at the benchmark. Algae bioaccumulate radionuclides (with extent depending on the radionuclide and the algae species). A study in France found that Chondrus crispus was capable of absorbing a large number of artificial radioactive elements and that this had consequences considering the exploitation of this species as a harvestable resource (Cosson et al., 1984). However, no information was found concerning the actual effects of radionuclide on Chondrus crispus.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
This pressure is Not assessed.
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
The effects of reduced oxygenation on algae are not well studied. Plants require oxygen for respiration, but this may be provided by production of oxygen during periods of photosynthesis. Lack of oxygen may impair both respiration and photosynthesis (see review by Vidaver, 1972). A study of the effects of anoxia on a red algae found within this biotope, Delesseria sanguinea, revealed that specimens died after 24 hours at 15°C but that some survived at 5°C (Hammer, 1972). No evidence is available to make an intolerance assessment for the key characterizing species.
Although Chondrus crispus and Polyides rotunda and other characterizing species may be out-competed by faster growing or ephemeral green and red algal species where nutrient enrichment is extensive or prolonged (Johansson et al., 1998) , the levels of scour and sand abrasion are likely to limit the growth of competing species as these tend to be thin and less robust.
Short-term experiments with low levels of nutrient enrichment showed little impacts on red algal turfs. Atalah & Crowe (2010) added nutrients to rockpoolsoccupied by a range of algae including encrusting corallines, turfs of Mastocarpus stellatus, Chondrus crispus and Corallina officinalis and green and red filamentous algae. Nitrogen and phosphorous enhancement was via the addition of fertilisers, as either 40 g/litre or 20 g/litre. The treatments were applied for seven months and experimental conditions were maintained every two weeks. The experimental treatments do not directly relate to the pressure benchmark or biotope type but indicate some general trends in sensitivity. Nutrients had no significant effect on the cover of crustose coralline algae or the cover of red turfing algae. However, the cover of green filamentous algae increased only where grazers were removed (Atalah & Crowe, 2010).
Over geological timescales periods of increased nutrient availability have experienced increases in the distribution of crustose coralline species at the expense of corals (Littler & Littler, 2013), suggesting that this group have some tolerance for enhanced nutrient levels. Overall. Littler & Littler (2013) suggest that corallines as a group can tolerate both low and elevated levels of nutrients. The crusting coralline Lithophyllum incrustans were present at sites dominated by Ulva spp. in the Mediterranean exposed to high levels of nutrient enrichment from domestic sewage (Arévalo et al., 2007).
Sensitivity assessment. The pressure benchmark is relatively protective and may represent a reduced level of nutrient enrichment in previously polluted areas. Due to the tolerance of high levels of nutrient input demonstrated generally by red algal turfs, resistance to this pressure is assessed as ‘High’ and resilience as ‘High’ so that the biotope is assessed as ‘Not sensitive’.
Where the biotope occurs in tide swept or wave exposed areas (Connor et al., 2004) water movements will disperse organic matter reducing the level of exposure. the crusting coralline Lithophyllum incrustans were present at sites dominated by Ulva spp. in the Mediterranean exposed to high levels of organic pollution from domestic sewage (Arévalo et al., 2007). As turf forming algae, including the red algal turf trap large amounts of sediment the turf itself is not considered sensitive to sedimentation. The turfs probably host a variety of associated species and deposit feeders amongst these would be able to consume inputs of organic matter.
In a recent review, assigning species to groups based on tolerances to organic pollution, the anemone Urticina felina was assigned to AMBI Group II described as 'species indifferent to enrichment, always present in low densities with non-significant variations with time, from initial state, to slight unbalance' (Gittenberger & van Loon, 2011).
Sensitivity assessment. Based on resistance to sedimentation, exposure to wave action, and the dominance of red algal turfs in areas subject to sewage inputs, resistance is assessed as ‘High’ and resilience as ‘High’ (by default). The biotope is therefore considered to be ‘Not sensitive’ to this pressure at the benchmark.
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
All marine habitats and benthic species are considered to have a resistance of ‘None’ to this pressure and to be unable to recover from a permanent loss of habitat (resilience is ‘Very Low’). Sensitivity within the direct spatial footprint of this pressure is therefore ‘High’. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.
This biotope is characterized by a hard rock substratum that is overlain or periodically exposed to, a layer of coarse sand. Removal of the bedrock would remove the attachment surface for the red algal turf that characterizes this biotope and the absence of sand may allow colonisation ofless scour-tolerant species, significantly altering the character of the biotope.
Artificial hard substratum may also differ in character from natural hard substratum, so that replacement of natural surfaces with artificial may lead to changes in the biotope through changes in species composition, richness and diversity (Green et al., 2012; Firth et al., 2014) or the presence of non-native species (Bulleri & Airoldi, 2005). The key characterizing species, Chondrus crispus readily colonised artificial settlement plates and by the end of the experiment was the dominant species on plates (Harlin & Lindbergh, 1977). Chondrus crispus were however significantly more abundant on the substratum with the largest particles (1-2 mm) and only a few individuals were found on the smooth surface, demonstrating that artificial smooth surfaces may not provide an optimal habitat.
In the absence of hard substratum crustose corralines can propagate as free-living rhodolith nodules and can form extensive subtidal habitats (Littler & Little, 2013). However, these biogenic reefs are not analogous to this habitat type.
Sensitivity assessment. Based on the loss of suitable habitat, biotope resistance is assessed as ‘None’ and recovery is assessed as ‘Very Low’ as the change at the pressure benchmark is permanent. Sensitivity is therefore ‘High’.
Generally this pressure is considered to be 'Not relevant' to biotopes occurring on bedrock. However, as this biotope is often overlain by sands or exposed to shifting sands this pressure is assessed. The sand covering and scour is an important factor supporting development and maintenance of this biotope. Removal of sands may allow species with less scour tolerance to colonize altering the character of the biotope. Siltation by finer sediments may have less impact but may lead to subtle changes and changes such as anoxia at the bedrock/sediment interface may lead to removal or damage of holdfasts and bases. A change to coarser gravels, pebbles and cobbles would increase the degree of abrasion where these are mobile and this may also remove the red algal turf.
Sensitivity assessment. A change in the character of the overlying sediment may alter the character of the biotope. Resistance is, therefore, assessed as 'Low' and resilience is Very low (the pressure is a permanent change), so that the biotope is considered to have 'High' sensitivity to this pressure.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
The species characterizing this biotope are epifauna or epiflora occurring on rock and would be sensitive to the removal of the habitat. However, extraction of rock substratum is considered unlikely and this pressure is considered to be ‘Not relevant’ to hard substratum habitats.
The red algal turf that characterizes this biotope has some resistance to abrasion as these species persist where scour rates are high from resuspension and transport of sand particles. Little empirical evidence was found to support this asessment.
Chondrus crispus is flexible (Dixon & Irvine, 1977) and would be expected to be relatively resistant to physical abrasion. Indeed, Worm & Chapman (1998) suggested that Chondrus crispus was highly resistant to intense physical and herbivore induced disturbance, ensuring competitive dominance on the lower shore. Chondrus crispus is capable of regenerating from its holdfasts (e.g. Dudgeon & Johnson, 1992) and so no mortality is expected.
Taylor (1970) (cited in Sharp et al., 1993) stated that clumps of fronds of Furcellaria lumbricalis were easily removed from the substratum by drag-raking, but only where the plant had a sufficient number of dichotomies (usually more than 3) to snag in the rake.
Dethier (1994) experimentally manipulated surface abrasion on a range of encrusting algae including Lithophyllum impressum. Crusts were brushed with either a nylon or steel brush for 1 minute a month for 24 months. Unbrushed controls grew by approximately 50% where the cover of nylon brushed crusts and steel brushed crusts decreased by approximately 25% and 40% respectively (interpreted from figures in Dethier, 1994). In laboratory tests on chips of Lithophyllum impressum brushing with a steel brush for 1 minute once a week for 3 weeks, resulted in no cover loss of two samples while a third ‘thinned and declined’ (Dethier, 1994). Mechanical abrasion from scuba divers was also reported to impact encrusting corallines, with cover of Lithophyllum stictaeforme greater in areas where diving was forbidden than visited areas (abundance, 6.36 vs 1.4; it is presumed this refers to proportion of cover, although this is not clear from the text, Guarinieri et al., 2012).
Sensitivity assessment. The impact of surface abrasion will depend on the footprint, duration and magnitude of the pressure. Based on evidence from the step experiments and the relative robustness of the red algal turf and encrusting corallines, resistance, to a single abrasion event is assessed as ‘Medium’ (loss of <25% cover/abundance) and recovery as ‘High’ (as bases are likely to remain), so that sensitivity is assessed as ‘Low’. Resistance and resilience will be lower (and hence sensitivity greater) to abrasion events that exert a greater crushing force and remove the bases.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
The species characterizing this biotope group are epifauna or epiflora occurring on rock which is resistant to subsurface penetration. The assessment for abrasion at the surface only is therefore considered to equally represent sensitivity to this pressure.
As this biotope occurs where sands cover the rock and scour is a significant factor this biotope is potentially sensitive to changes in suspended solids. Siltation and abrasion, which may be associated with changes in suspended solids are assessed separately.
The red algal species that characterize this biotope are scour tolerant and occur in turbid waters and in general algal turfs replace fucoids and kelps in areas where turbidity and sedimentation increase (Airoldi, 2003). Many of the red algal species, including the key characterizing Chondrus crispus, occur beneath canopies of larger macroalgae and are tolerant of low light levels (Gantt, 1990). Ahnfeltia plicata is not likely to be affected directly by an increase in suspended sediment. In general, subtidal red algae are able to exist at relatively low light levels (Ahnfeltia plicata typically occurs as an understory alga beneath Laminaria sp. (Lüning, 1990) and so is presumably well adapted to growth in low light conditions. Furcellaria lumbricalis often occurs in relatively turbid waters. Laboratory experiments by Bird et al. (1979) revealed that Furcellaria lumbricalis was growth saturated at very low light levels (ca 20µE/m²/s) compared to other algae such as Chondrus crispus (50-60µE/m²/s). They suggest that this may be an explanation why Furcellaria lumbricalis is able to proliferate in relatively deep and turbid waters. Similarly, in their review, Bird et al. (1979) comment that in all studies, saturation and inhibition radiances were low for Furcellaria lumbricalis compared to other macroalgae indicating good competitive ability in the attenuated light of deeper or more turbid waters. Increases in turbidity may provide the species with a competitive advantage over other macroalgae.
Crustose coralline alga are thought to be fairly resistant of sediment scour (Littler & Kauker, 1984) and are tolerant of very low light levels. Encrusting corallines can occur in deeper water than other algae where light penetration is limited. Samples of Lithophyllum impressum suspended from a raft and shaded (50-75% light reduction) continued to grow over two years (Dethier, 1994).
A decrease in suspended solids and a reduction in scour may favour species more typical of areas of lower turbidity and lower levels of scour. This may result in an increase in kelp or Alaria esculenta (where the sand layer characterizing this biotope is removed).
Sensitivity assessment. Changes in suspended solids at the pressure benchmark may lead to changes in the biotope assemblage, however, no empirical evidence was found to assess this pressure. Increases in suspended solids may lead to greater scour damaging plants and reduce light availability impacting growth. Decreased suspended solids may increase the suitability of the biotope for species less tolerant of scour and turbidity that can outcompete the characterizing species. The assessment is based on the characteristic sands remaining in the biotope covering the rocks and limiting the establishment of other competing species, In this scenario the biotope will still be subject to episodes of re-suspension and water transport. As the species present are tolerant of scour and turbidity, biotope resistance is assessed as ‘Medium’ and resilience as ‘High’ (as crustose bases are expected to remain to support recovery) sensitivity is therefore assessed as ‘Low’.
This biotope is characterized by a turf of red algae occurring on rock that is overlain with coarse sediments. These sediments are likely to move and therefore the turf and associated species will be subject to periodic burial. In areas exposed to wave action or strong water flows deposited sediments ar e likely to be moved rapidly mitigating the impact. The key characterizing species Chondrus crispus, Polyides rotunda and Ahnfeltia plicata are erect and grow to over 20 cm, therefore mature plants are unlikely to be affected by smothering with 5 cm of sediment. However, recently settled propagules, regenerating holdfasts and small developing plants would be buried by 5 cm of sediment and be unable to photosynthesize.
In a review of the effects of sedimentation on rocky coast assemblages, Airoldi (2003) outlined the evidence for the sensitivity of coralline algae to sedimentation. The reported results are contradictory with some authors suggesting that coralline algae are negatively affected by sediments while others report that encrusting corallines are often abundant or even dominant in a variety of sediment impacted habitats (Airoldi, 2003 and references therein). Crustose corallines have been reported to survive under a turf of filamentous algae and sediment for 58 days (the duration of experiment) in the Galapagos (species not identified, Kendrick, 1991). The crustose coralline Hydrolithon reinboldii, has also been reported to survive deposition of silty sediments on subtidal reefs off Hawaii (Littler, 1973).
Communities dominated by Urticina felina were described on tide swept seabed, exposed to high levels of suspended sediment, sediment scour and to periodic smothering by thin layers of sand, up to ca 5cm in the central English Channel (Home & Wilson, 1985). Urticina felina is abundant in the sediment-scoured, silty rock communities CR.HCR.XFa.FluCoAs and CR.MCR.EcCr.UrtScr (Connor et al. 2004). Laboratory experiments have shown that another anemone Sagartiogeton laceratus is able to survive under sediments for 16 days and to be capable of re-emerging under shallow (2cm) burial (Last et al., 2011). The percentage mortality increased with both depth and increasingly finer sediment fraction. Bijkerk (1988, results cited from Essink (1999) indicated that the maximal overburden through which the anemone Sagartia elegans could migrate was <10cm in sand. No further information was available on the rates of survivorship or the time taken to reach the surface.
Sensitivity assessment. Based on the growth form of the characterizing red algae and the presence of these algae and Urticina felina in biotopes subject to sand covering (including the assessed biotope one), resistance to this pressure, at the benchmark, is assessed as 'High', resilience is assessed as 'High' (by default) and the biotope is considered to be 'Not sensitive'.
The available evidence for siltation pressures is outlined for the ‘light’ deposition pressure. At the pressure benchmark ‘heavy deposition’ represents a considerable thickness of deposit. Complete burial of algal turf and encrusting corallines and associated animals would occur. Removal of the sediments by wave action and tidal currents would result in considerable scour. The effect of this pressure will be mediated by the length of exposure to the deposit.
Sensitivity assessment. Resistance is assessed as ‘Medium-Low’ as the impact on the characterizing and associated species could be significant but may be mitigated by rapid removal. Resilience is assessed as ‘High’ based on vegetative re-growth from the scour-tolerant surviving bases of the characterizing species. Sensitivity is therefore assessed as 'Low'.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Coralline crusts are shade tolerant algae, often occurring under a macralgal canopy that reduces light penetration. These species can acclimate to different levels of light intensity and quality and encrusting corallines can occur in deeper water than other algae where light penetration is limited. Samples of Lithophyllum impressum suspended from a raft and shaded (50-75% light reduction) continued to grow over two years (Dethier, 1994). In areas of higher light levels, the fronds and bases may be lighter in colour due to bleaching (Colhart & Johansen, 1973). Other red algae in the biotope are flexible with regard to light levels and can also acclimate to different light levels. Canopy removal experiments in a rocky sub tidal habitat in Nova Scotia, Canada by Schmidt & Scheibling (2007) did not find a shift in understorey macraoalgal turfs (dominated by Corallina officinalis, Chondrus crispus and Mastocarpus stellatus) to more light-adapted species over 18 months.
Ahnfeltia plicata is not likely to be affected directly by an increase in suspended sediment. In general, subtidal red algae are able to exist at relatively low light levels (Gantt, 1990). Ahnfeltia plicata typically occurs as an understory alga beneath Laminaria sp. (Lüning, 1990) and so is presumably well adapted to growth in low light conditions. An increase in turbidity would reduce the amount of light reaching the understory. Over the course of a year, this may result in mortality of the Ahnfeltia plicata individuals at the limit of their depth range. Intolerance is therefore assessed as intermediate. Recoverability is recorded as high (see additional information below).
A decrease in turbidity would result in greater light availability for Ahnfeltia plicata. Haglund et al. (1987) reported no inhibition of photosynthesis up to 500 µE/m²/s and suggested that Ahnfeltia plicata had a high potential for growth provided no other factors were limiting. Ahnfeltia plicata is therefore assessed as being tolerant to decrease in turbidity, with the potential to benefit from the factor.
Two species of cold-temperate algae from the North Atlantic Ocean, Polyides rotunda and Furcellaria lumbricalis, were tested for growth and survival over a temperature range of -5 to 30-degrees-C. In comparisons of eastern and western isolates, both Furcellaria lumbricalis, a North Atlantic endemic, and Polyides rotunda, a species having related populations in the North Pacific, were quite homogeneous. Furcellaria lumbricalis tolerated -5 to 25-degrees-C and grew well from 0 to 25-degrees-C, with optimal growth at 10-15-degrees-C. P. rotundus tolerated -5 to 27-degrees-C, grew well from 5 to 25-degrees-C, and had a broad optimal range of 10-25-degrees-C. Both species tolerated 3 months in darkness at 0-degrees-C (Novaczek & Breeman, 1990).
Some populations of Plocamium cartilagineum were found to be light-limited in the Isle of Man and removal of the Laminaria canopy enhanced growth (Kain, 1987). However, in shallow water in summer, light levels are high enough to inhibit growth and it can only inhabit shallow waters where shaded by Laminarians (Kain, 1987 and references therein). A maximum irradiance of 0.5 mmol/m2/s was inhibitory for Plocamium cartilagineum (Kain, 1987)
The responses of the red algae to shading Plocamium cartilagineum and exposure to light vary, some such as are found in the deeper subtidal and are more tolerant of shade than species such as Chondrus crispus that are found in the intertidal .
Furcellaria lumbricalis often occurs in relatively turbid waters. Laboratory experiments by Bird et al. (1979) revealed that Furcellaria lumbricalis was growth saturated at very low light levels (ca 20µE/m²/s) compared to other algae such as Chondrus crispus (50-60µE/m²/s) and Fucus serratus (100µE/m²/s). They suggest that this may be an explanation why Furcellaria lumbricalis is able to proliferate in relatively deep and turbid waters. Similarly, in their review, Bird et al. (1999) comment that in all studies, saturation and inhibition radiances were low for Furcellaria lumbricalis compared to other macroalgae indicating good competitive ability in the attenuated light of deeper or more turbid waters. In light of its tolerance of turbid conditions it is expected that the majority of the Furcellaria lumbricalis population would be unaffected by increases in turbidity, indeed, such changes may even provide the species with a competitive advantage over other macroalgae.
Sensitivity assessment. As the key structuring and characterizing species colonize a broad range of light environments from intertidal to deeper sub tidal and shaded understorey habitats the biotope is considered to have ‘High’ resistance and, by default, ‘High’ resilience and therefore is ‘Not sensitive’ to this pressure.
Barriers that reduce the degree of tidal excursion may alter larval supply to suitable habitats from source populations. Conversely the presence of barriers may enhance local population supply by preventing the loss of larvae from enclosed habitats. Barriers and changes in tidal excursion are not considered relevant to the characterizing red algal species dispersal is limited by the rapid rate of settlement and vegetative growth from bases rather than reliance on recruitment from outside of populations. Resistance to this pressure is assessed as 'High' and resilience as 'High' by default. This biotope is therefore considered to be 'Not sensitive'.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Not relevant to seabed habitats. NB. Collision by grounding vessels is addressed under ‘surface abrasion.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
|Use / to open/close text displayed||Resistance||Resilience||Sensitivity|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
No information was found on current production of Chondrus crispus or other turf forming red seaweeds in the UK and it is understood that wild harvesting rather than cultivation is the method of production for these. No evidence was found for the effects of gene flow between cultivated species and wild populations. Although cultivation of different genotypes may lead to gene flow between wild and cultivated populations the limited dispersal may reduce exposure. Some negative effects may arise from hybridisation between very geographically separated populations but there is no evidence to suggest that gene flow between different UK haplotypes would lead to negative effects. This pressure is therefore considered ‘Not relevant’ to this biotope group.
The high levels of scour in this biotope will limit establishment of all but scour resistant invasive non-indigenous species (INIS) from this biotope and no direct evidence was found for effects of INIS on this biotope. A number of invasive red algae have been recorded in the UK, from reported habitat preferences Bonnemaisonia hamifera does not appear to be present in scoured environments although the harpoon weed, Asparagopsis armata is found in sandy pools (Guiry & Guiry, 2015). In North America Grateloupia turuturu is a major competitor of Chondrus crispus, although Grateloupia turuturu is present in the UK, this large foliose species may not be able to colonize this scoured biotope due to the effects of drag and abrasion.
Sensitivity assessment. As sand scouring of this biotope limits establishment of all but robust species, resistance to INIS is assessed as ‘High’ and resilience as ‘High’ (by default) so that the biotope is considered to be ‘Not sensitive’.
Extracts of some red algae show antimicrobial and antifungal activity limiting disease, Dilsea carnosa extracts, for example, limit colony extension in Microsporum canis and Trichophyton varrucosum (Tariq, 1991), providing some protection against pathogens. Some examples of infection by microbial pathogens have been found in the literature. Craigie & Correa (1996), for example, described 'green spot' disease in the key characterizing species Chondrus crispus, caused by the interaction of several biotic agents including fungi, bacteria, algal endophytes and grazers, and resulting in tissue necrosis. Correa & McLachlan (1992) infected Chondrus crispus with the green algal endophytes Acrochaete operculata and Acrochaete heteroclada. Infections resulted in detrimental effects, including slower growth, reduced carrageenan yield, reduced generation capacity and tissue damage. Stanley (1992) described the fungus Lautita danica being parasitic on cystocarpic Chondrus crispus and Molina (1986) was the first to report Petersenia pollagaster, a fungal invasive pathogen of cultivated Chondrus crispus. At usual levels of infestation in wild populations these are not considered to lead to high levels of mortality. Barton (1901) noted that Furcellaria lumbricalis may become infested with nematode worms and reacts by gall formation.
Diseased encrusting corallines were first observed in the tropics in the early 1990’s when the bacterial pathogen Coralline Lethal Orange Disease (CLOD) was discovered (Littler & Littler, 1995). All species of articulated and crustose species tested to date are easily infected by CLOD and it has been increasing in occurrence at sites where first observed and spreading through the tropics. Another bacterial pathogen causing a similar CLOD disease has been observed with a greater distribution and a black fungal pathogen first discovered in American Samoa has been dispersing (Littler & Littler, 1998). An unknown pathogen has also been reported to lead to white ‘target-shaped’ marks on corallines, again in the tropics (Littler et al., 2007). No evidence was found that these are impacting temperate coralline habitats.
Sensitivity assessment. Based on the lack of evidence for major pathogens and significant mortalities of other turf-forming macroalgae this biotope is considered to have ‘High’ resistance and hence ‘High’ resilience and is classed as ‘Not sensitive’ at the pressure benchmark.
Direct, physical impacts from harvesting are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of target species on this biotope. Incidental removal of the key characterizing species and associated species would alter the character of the biotope. Intertidal populations of Chondrus crispus are harvested commercially in Scotland and Ireland, the stipe is removed but the base is left intact to allow the algae to re-grow. As a key characterizing and structuring species extensive removal of Chondrus crispus would alter the character of the biotope. The effect of harvesting Chondrus crispus has been best studied in Canada. Prior to 1980, the seaweed beds of Prince Edward Island were dominated by Chondrus crispus and the species was heavily exploited, following this period a marked increase in abundance of the red seaweed, Furcellaria lumbricalis, which was avoided by the commercial harvest, and an associated decline in abundance of Chondrus crispus (Sharp et al., 1993). The authors suggested that harvesting has brought about the shift in community structure. Sharp et al. (1986) reported that the first drag rake harvest of the season, on a Nova Scotian Chondrus crispus bed, removed 11% of the fronds and 40% of the biomass. Efficiency declined as the harvesting season progressed. Chopin et al. (1988) noted that non-drag raked beds of Chondrus crispus in the Gulf of St Lawrence showed greater year round carposporangial reproductive capacity than a drag raked bed. In the short-term, therefore, harvesting of Chondrus crispus may remove biomass and impair reproductive capacity, while in the long-term, it has the potential to alter community structure and change the dominant species. The other key characterizing red algal species are not commercially targeted but have been investigated as alternative sources of carrageenan and agar (Mathieson et al., 1984) and in the future may be subject to harvesting.
The removal of kelps or brown algae from this biotope would reduce shading and is not considered to negatively affect this biotope (for assessment of removal of the key characterizing species as by-catch, see the removal of non-target species pressure).
Sensitivity assessment. The species that are harvested, or potentially harvested, in this biotope are all attached and relatively conspicuous. A single event of targeted harvesting could therefore efficiently remove individuals and resistance is assessed as ‘Low’. This assessment is supported by evidence from Sharp et al., (1993) on the proportion of biomass of Chondrus crispus removed commercially. Resilience of the turf forming red seaweeds is assessed as ‘Medium’ (based on evidence for recovery from harvesting where some damage occurred to the bases although see caveats in the resilience section) and biotope sensitivity is assessed as ‘Medium'. This assessment refers to a single collection event, long-term harvesting over wide spatial scales will lead to greater impacts, with lower resistance and longer recovery times.
Incidental removal of the key characterizing species and associated species would alter the character of the biotope. The biotope is characterized by the turf of red algae. The loss of this due to incidental removal as by-catch would therefore alter the character of the habitat and result in the loss of species richness. The ecological services such as primary production and the habitat structure provided by these species would also be lost.
Sensitivity assessment. Removal of a large percentage of the characterising species resulting in bare rock would alter the character of the biotope, species richness and ecosystem function. Resistance is therefore assessed as ‘Low’ and recovery as ‘Medium’, so that sensitivity is assessed as 'Medium'.
Airoldi, L., 2003. The effects of sedimentation on rocky coast assemblages. Oceanography and Marine Biology: An Annual Review, 41,161-236
Airoldi, L., 2000. Responses of algae with different life histories to temporal and spatial variability of disturbance in subtidal reefs. Marine Ecology Progress Series, 195 (8), 81-92.
Arévalo, R., Pinedo, S. & Ballesteros, E., 2007. Changes in the composition and structure of Mediterranean rocky-shore communities following a gradient of nutrient enrichment: descriptive study and test of proposed methods to assess water quality regarding macroalgae. Marine Pollution Bulletin, 55 (1), 104-113.
Atalah, J. & Crowe, T.P., 2010. Combined effects of nutrient enrichment, sedimentation and grazer loss on rock pool assemblages. Journal of Experimental Marine Biology and Ecology, 388 (1), 51-57.
Austin, A.P., 1960a. Life history and reproduction of Furcellaria fastigiata (L.) Lamouroux. Annals of Botany, New Series, 24, 257-274.
Austin, A.P., 1960b. Observations on the growth, fruiting and longevity of Furcellaria fastigiata (L.) Lamouroux. Hydrobiologia, 15, 193-207.
Barnes, R.S.K. & Hughes, R.N., 1992. An introduction to marine ecology. Oxford: Blackwell Scientific Publications.
Barton, E.S., 1901. On certain galls in Furcellaria and Chondrus. Journal of Botany, 39, 49-51.
Bellgrove, A., McKenzie, P.F., McKenzie, J.L. & Sfiligoj, B.J., 2010. Restoration of the habitat-forming fucoid alga Hormosira banksii at effluent-affected sites: competitive exclusion by coralline turfs. Marine Ecology Progress Series, 419, 47-56.
Bhattacharya, D., 1985. The demography of fronds of Chondrus crispus Stackhouse. Journal of Experimental Marine Biology and Ecology, 91, 217-231.
Bijkerk, R., 1988. Ontsnappen of begraven blijven: de effecten op bodemdieren van een verhoogde sedimentatie als gevolg van baggerwerkzaamheden: literatuuronderzoek: RDD, Aquatic ecosystems.
Bird, C.J., Saunders, G.W. & McLachlan, J., 1991. Biology of Furcellaria lumbricalis (Hudson) Lamouroux (Rhodophyta: Gigartinales), a commercial carrageenophyte. Journal of Applied Phycology, 3, 61-82.
Bird, N.L., Chen, L.C.-M. & McLachlan, J., 1979. Effects of temperature, light and salinity of growth in culture of Chondrus crispus, Furcellaria lumbricalis, Gracilaria tikvahiae (Gigartinales, Rhodophyta), and Fucus serratus (Fucales, Phaeophyta). Botanica Marina, 22, 521-527.
Birkett, D.A., Maggs, C.A., Dring, M.J. & Boaden, P.J.S., 1998b. Infralittoral reef biotopes with kelp species: an overview of dynamic and sensitivity characteristics for conservation management of marine SACs. Natura 2000 report prepared by Scottish Association of Marine Science (SAMS) for the UK Marine SACs Project., Scottish Association for Marine Science. (UK Marine SACs Project, vol V.). Available from: http://www.ukmarinesac.org.uk/publications.htm
Bokn, T.L., Moy, F.E. & Murray, S.N., 1993. Long-term effects of the water-accommodated fraction (WAF) of diesel oil on rocky shore populations maintained in experimental mesocosms. Botanica Marina, 36, 313-319.
Boller, M.L. & Carrington, E., 2007. Interspecific comparison of hydrodynamic performance and structural properties among intertidal macroalgae. Journal of Experimental Biology, 210 (11), 1874-1884.
Boney, A.D., 1971. Sub-lethal effects of mercury on marine algae. Marine Pollution Bulletin, 2, 69-71.
Braber, L. & Borghouts, C.H., 1977. Distribution and ecology of Anthozoa in the estuarine region of the rivers Rhine, Meuse and Scheldt. Hydrobiologia, 52, 15-21.
Bryan, G.W., 1984. Pollution due to heavy metals and their compounds. In Marine Ecology: A Comprehensive, Integrated Treatise on Life in the Oceans and Coastal Waters, vol. 5. Ocean Management, part 3, (ed. O. Kinne), pp.1289-1431. New York: John Wiley & Sons.
Bulleri, F. & Airoldi, L., 2005. Artificial marine structures facilitate the spread of a non‐indigenous green alga, Codium fragile ssp. tomentosoides, in the north Adriatic Sea. Journal of Applied Ecology, 42 (6), 1063-1072.
Burdin, K.S. & Bird, K.T., 1994. Heavy metal accumulation by carrageenan and agar producing algae. Botanica Marina, 37, 467-470.
Chamberlain, Y.M., 1996. Lithophylloid Corallinaceae (Rhodophycota) of the genera Lithophyllum and Titausderma from southern Africa. Phycologia, 35, 204-221.
Chapman, V.J. & Chapman, D.J., 1980. Seaweeds and their uses. Chapman & Hall.
Chia, F.S. & Spaulding, J.G., 1972. Development and juvenile growth of the sea anemone Tealia crassicornis. Biological Bulletin, Marine Biological Laboratory, Woods Hole, 142, 206-218.
Chopin, T. & Wagey, B.T., 1999. Factorial study of the effects of phosphorus and nitrogen enrichments on nutrient and carrageenan content in Chondrus crispus (Rhododphyceae) and on residual nutrient concentration in seawater. Botanica Marina, 42, 23-31.
Chopin, T., Pringle, J.D. & Semple, R.E., 1988. Reproductive capacity of dragraked and non-dragraked Irish moss (Chondrus crispus Stackhouse) beds in the southern Gulf of St Lawrence. Canadian Journal of Fisheries and Aquatic Sciences, 45, 758-766.
Cole, S., Codling, I.D., Parr, W., Zabel, T., 1999. Guidelines for managing water quality impacts within UK European marine sites [On-line]. UK Marine SACs Project. [Cited 26/01/16]. Available from: http://www.ukmarinesac.org.uk/pdfs/water_quality.pdf
Cole, S., Codling, I.D., Parr, W. & Zabel, T., 1999. Guidelines for managing water quality impacts within UK European Marine sites. Natura 2000 report prepared for the UK Marine SACs Project. 441 pp., Swindon: Water Research Council on behalf of EN, SNH, CCW, JNCC, SAMS and EHS. [UK Marine SACs Project.], http://www.ukmarinesac.org.uk/
Colhart, B.J., & Johanssen, H.W., 1973. Growth rates of Corallina officinalis (Rhodophyta) at different temperatures. Marine Biology, 18, 46-49.
Connor, D.W., Allen, J.H., Golding, N., Howell, K.L., Lieberknecht, L.M., Northen, K.O. & Reker, J.B., 2004. The Marine Habitat Classification for Britain and Ireland. Version 04.05. Joint Nature Conservation Committee, Peterborough. www.jncc.gov.uk/MarineHabitatClassification.
Connor, D.W., Dalkin, M.J., Hill, T.O., Holt, R.H.F. & Sanderson, W.G., 1997a. Marine biotope classification for Britain and Ireland. Vol. 2. Sublittoral biotopes. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06., Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06.
Correa, J.A. & McLachlan, J.L., 1992. Endophytic algae of Chondrus crispus (Rhodophyta). 4. Effects on the host following infections by Acrochaete operculata and A. heteroclada (Chlorophyta). Marine Ecology Progress Series, 81, 73-87.
Cosson, J., Lepy, M.C., Patry, M.C. & Saur, H., 1984. Etude sur les radioelements emetteurs presents dans les algues des cotes du Calvados (France) pendant les annees, 1980 - 1982. Botanica marina, 27, 301-308.
Craigie, J.S. & Correa, J.A., 1996. Etiology of infectious diseases in cultivated Chondrus crispus (Gigartinales, Rhodophyta). Hydrobiologia, 326-327, 97-104.
Crump, R.G., Morley, H.S., & Williams, A.D., 1999. West Angle Bay, a case study. Littoral monitoring of permanent quadrats before and after the Sea Empress oil spill. Field Studies, 9, 497-511.
Daly, M.A. & Mathieson, A.C., 1977. The effects of sand movement on intertidal seaweeds and selected invertebrates at Bound Rock, New Hampshire, USA. Marine Biology, 43, 45-55.
Davies, C.E. & Moss, D., 1998. European Union Nature Information System (EUNIS) Habitat Classification. Report to European Topic Centre on Nature Conservation from the Institute of Terrestrial Ecology, Monks Wood, Cambridgeshire. [Final draft with further revisions to marine habitats.], Brussels: European Environment Agency.
Davis, D.S., 1967. The marine fauna of the Blackwater Estuary and adjacent waters. Essex Naturalist, 32, 1-60.
Dethier, M.N., 1994. The ecology of intertidal algal crusts: variation within a functional group. Journal of Experimental Marine Biology and Ecology, 177 (1), 37-71.
Dickinson, C.I., 1963. British seaweeds. London & Frome: Butler & Tanner Ltd.
Dixon, P.S. & Irvine, L.M., 1977. Seaweeds of the British Isles. Volume 1 Rhodophyta. Part 1 Introduction, Nemaliales, Gigartinales. London: British Museum (Natural History) London.
Dudgeon, S., Davison, I. & Vadas, R., 1989. Effect of freezing on photosynthesis of intertidal macroalgae: relative tolerance of Chondrus crispus and Mastocarpus stellatus (Rhodophyta). Marine Biology, 101 (1), 107-114.
Dudgeon, S.R. & Johnson, A.S., 1992. Thick vs. thin: thallus morphology and tissue mechanics influence differential drag and dislodgement of two co-dominant seaweeds . Journal of Experimental Marine Biology and Ecology, 165, 23-43.
Dudgeon, S.R., Davison, I.R. & Vadas, R.L., 1990. Freezing tolerance in the intertidal red algae Chondrus crispus and Mastocarpus stellatus: relative importance of acclimation and adaptation. Marine Biology, 106, 427-436.
Dudgeon, S.R., Kuebler, J.E., Vadas, R.L. & Davison, I.R., 1995. Physiological responses to environmental variation in intertidal red algae: does thallus morphology matter ? Marine Ecology Progress Series, 117, 193-206.
Edyvean, R.G.J. & Ford, H., 1987. Growth rates of Lithophyllum incrustans (Corallinales, Rhodophyta) from south west Wales. British Phycological Journal, 22 (2), 139-146.
Edyvean, R.G.J. & Ford, H., 1984a. Population biology of the crustose red alga Lithophyllum incrustans Phil. 2. A comparison of populations from three areas of Britain. Biological Journal of the Linnean Society, 23 (4), 353-363.
Edyvean, R.G.J. & Ford, H., 1984b. Population biology of the crustose red alga Lithophyllum incrustans Phil. 3. The effects of local environmental variables. Biological Journal of the Linnean Society, 23, 365-374.
Edyvean, R.G.J. & Ford, H., 1986. Population structure of Lithophyllum incrustans (Philippi) (Corallinales Rhodophyta) from south-west Wales. Field Studies, 6, 397-405.
Essink, K., 1999. Ecological effects of dumping of dredged sediments; options for management. Journal of Coastal Conservation, 5, 69-80.
Fernandez, C. & Menendez, M.P., 1991. Ecology of Chondrus crispus on the northern coast of Spain. 2. Reproduction. Botanica Marina, 34, 303-310.
Firth, L., Thompson, R., Bohn, K., Abbiati, M., Airoldi, L., Bouma, T., Bozzeda, F., Ceccherelli, V., Colangelo, M. & Evans, A., 2014. Between a rock and a hard place: Environmental and engineering considerations when designing coastal defence structures. Coastal Engineering, 87, 122-135.
Fletcher, R.L. & Callow, M.E., 1992. The settlement, attachment and establishment of marine algal spores. British Phycological Journal, 27, 303-329.
Fortes, M.D. & Lüning, K., 1980. Growth rates of North Sea macroalgae in relation to temperature, irradiance and photoperiod. Helgolander Meeresuntersuchungen, 34, 15-29.
Gantt, E., 1990. Pigmentation and photoacclimation. In Biology of the Red Algae (ed. K.M. Cole and R.G. Sheath), 203-219. Cambridge University Press.
Gittenberger, A. & Van Loon, W.M.G.M., 2011. Common Marine Macrozoobenthos Species in the Netherlands, their
Grandy, N., 1984. The effects of oil and dispersants on subtidal red algae. Ph.D. Thesis. University of Liverpool.
Green, D., Chapman, M. & Blockley, D., 2012. Ecological consequences of the type of rock used in the construction of artificial boulder-fields. Ecological Engineering, 46, 1-10.
Guarnieri, G., Terlizzi, A., Bevilacqua, S. & Fraschetti, S., 2012. Increasing heterogeneity of sensitive assemblages as a consequence of human impact in submarine caves. Marine Biology, 159 (5), 1155-1164.
Guidone, M. & Grace, S., 2010. The ratio of gametophytes to tetrasporophytes of intertidal Chondrus crispus (Gigartinaceae) across a salinity gradient. Rhodora, 112 (949), 80-84.
Guiry, M.D. & Blunden, G., 1991. Seaweed Resources in Europe: Uses and Potential. Chicester: John Wiley & Sons.
Guiry, M.D. & Guiry, G.M. 2015. AlgaeBase [Online], National University of Ireland, Galway [cited 30/6/2015]. Available from: http://www.algaebase.org/
Gutierrez, L.M. & Fernandez, C., 1992. Water motion and morphology in Chondrus crispus (Rhodophyta). Journal of Phycology, 28, 156-162.
Haglund, K., Axelsson, L. & Pedersen, M., 1987. Photosynthesis and respiration in the alga Ahnfeltia plicata in a flow through system. Marine Biology, 96, 409-412.
Harlin, M.M., & Lindbergh, J.M., 1977. Selection of substrata by seaweed: optimal surface relief. Marine Biology, 40, 33-40.
Harvey, M.J. & McLachlan, J., 1973. Chondrus crispus. Proceedings of the Transactions of the Nova Scotian Institute of Science, 27 (Suppl.1), 1-155.
Hawkins, S.J. & Harkin, E., 1985. Preliminary canopy removal experiments in algal dominated communities low on the shore and in the shallow subtidal on the Isle of Man. Botanica Marina, 28, 223-30.
Hiscock, K., 1983. Water movement. In Sublittoral ecology. The ecology of shallow sublittoral benthos (ed. R. Earll & D.G. Erwin), pp. 58-96. Oxford: Clarendon Press.
Hiscock, K., ed. 1998. Marine Nature Conservation Review. Benthic marine ecosystems of Great Britain and the north-east Atlantic. Peterborough, Joint Nature Conservation Committee.
Hoare, R. & Hiscock, K., 1974. An ecological survey of the rocky coast adjacent to the effluent of a bromine extraction plant. Estuarine and Coastal Marine Science, 2 (4), 329-348.
Holme, N.A. & Wilson, J.B., 1985. Faunas associated with longitudinal furrows and sand ribbons in a tide-swept area in the English Channel. Journal of the Marine Biological Association of the United Kingdom, 65, 1051-1072.
Holt, T.J., Jones, D.R., Hawkins, S.J. & Hartnoll, R.G., 1995. The sensitivity of marine communities to man induced change - a scoping report. Countryside Council for Wales, Bangor, Contract Science Report, no. 65.
Jansson, M. & Kautsky, M., 1976. Quantitative survey of hard bottom communities in a Baltic archipelago. In Proceedings of the 11th European Symposium on Marine Biology, Galway, 5-11 October, 1976. Biology of Benthic Organisms (ed. B.F. Keegan, P.O. Ceidigh & Boaden, P.J.S.), pp. 359-366.
Johansson ,G., Eriksson, B.K., Pedersen, M. & Snoeijs, P., 1998. Long term changes of macroalgal vegetation in the Skagerrak area. Hydrobiologia, 385, 121-138.
Jorde, I. & Klavestad, N., 1963. The natural history of the Hardangerfjord. 4. The benthonic algal vegetation. Sarsia, 9, 1-99.
Juanes, J.A. & McLachlan, J.L., 1992. Productivity of Chondrus crispus Stackhouse (Rhodophyta, Gigartinales) in sublittoral Prince Edward Island. 2. Influence of temperature and nitrogen reserves. Botanica Marina, 35, 399-405.
Kaas, R., 1980. Les consequences de l'echouement de l'Amoco Cadiz" sur les peuplements algaux exploitables. Revue des Travaux de l'Institut des Pêches Maritimes, 44 (2), 157-194.
Kain, J.M., 1975a. Algal recolonization of some cleared subtidal areas. Journal of Ecology, 63, 739-765.
Kain, J.M., 1987. Photoperiod and temperature as triggers in the seasonality of Delesseria sanguinea. Helgolander Meeresuntersuchungen, 41, 355-370.
Kain, J.M., & Norton, T.A., 1990. Marine Ecology. In Biology of the Red Algae, (ed. K.M. Cole & Sheath, R.G.). Cambridge: Cambridge University Press.
Kautsky, N., Kautsky, H., Kautsky, U. & Waern, M., 1986. Decreased depth penetration of Fucus vesiculosus (L.) since the 1940s indicates eutrophication of the Baltic Sea. Marine Ecology Progress Series, 28, 1-8.
Kendrick, G.A., 1991. Recruitment of coralline crusts and filamentous turf algae in the Galapagos archipelago: effect of simulated scour, erosion and accretion. Journal of Experimental Marine Biology and Ecology, 147 (1), 47-63
Kuebler, J.E. & Davison, I.R., 1993. High temperature tolerance of photosynthesis in the red alga Chondrus crispus. Marine Biology, 117, 327-335.
Kuebler, J.E. & Davison, I.R., 1995. Thermal acclimation of light use characteristics of Chondrus crispus (Rhodophyta). Journal of Mycology, 30, 189-196.
Kuebler, J.E. & Dudgeon, S.R., 1996. Temperature dependent change in the complexity of form of Chondrus crispus fronds. Journal of Experimental Marine Biology and Ecology, 207, 15-24.
Larsen, A. & Sand-Jensen, K., 2006. Salt tolerance and distribution of estuarine benthic macroalgae in the Kattegat-Baltic Sea area. Phycologia, 45 (1), 13-23.
Last, K.S., Hendrick V. J, Beveridge C. M & Davies A. J, 2011. Measuring the effects of suspended particulate matter and smothering on the behaviour, growth and survival of key species found in areas associated with aggregate dredging. Report for the Marine Aggregate Levy Sustainability Fund,
Lewis, J.R., 1964. The Ecology of Rocky Shores. London: English Universities Press.
Lindgren, A. & Åberg, P., 1996. Proportion of life cycle stages of Chondrus crispus and its population structure: a comparison between a marine and an estuarine environment. Botanica Marina, 39 (1-6), 263-268.
Littler, M. & Littler, D., 1998. An undescribed fungal pathogen of reef-forming crustose corraline algae discovered in American Samoa. Coral Reefs, 17 (2), 144-144.
Littler, M. & Littler, D.S. 2013. The nature of crustose coralline algae and their interactions on reefs. Smithsonian Contributions to the Marine Sciences, 39, 199-212
Littler, M.M., 1973. The population and community structure of Hawaiian fringing-reef crustose Corallinaceae (Rhodophyta, Cryptonemiales). Journal of Experimental Marine Biology and Ecology, 11 (2), 103-120.
Littler, M.M. & Littler, D.S., 1995. Impact of CLOD pathogen on Pacific coral reefs. Science, 267, 1356-1356.
Littler, M.M., & Kauker, B.J., 1984. Heterotrichy and survival strategies in the red alga Corallina officinalis L. Botanica Marina, 27, 37-44.
Littler, M.M., Littler, D.S. & Brooks, B.L. 2007. Target phenomena on south Pacific reefs: strip harvesting by prudent pathogens? Reef Encounter, 34, 23-24
Lüning, K., 1990. Seaweeds: their environment, biogeography, and ecophysiology: John Wiley & Sons.
Lüning, K. & Freshwater, W., 1988. Temperature tolerance of northeast Pacific marine algae. Journal of Phycology, 24, 310-315.
Lüning, K., 1984. Temperature tolerance and biogeography of seaweeds: the marine algal flora of Helgoland (North Sea) as an example. Helgolander Meeresuntersuchungen, 38, 305-317.
MacFarlane, C.I., 1952. A survey of certain seaweeds of commercial importance in southwest Nova Scotia. Canadian Journal of Botany, 30, 78-97.
Maggs, C.A. & Pueschel, C.M., 1989. Morphology and development of Ahnfeltia plicata (Rhodophyta) : proposal of Ahnfeltiales ord. nov. Journal of Phycology, 25, 333-351.
Mathieson, A., Emerich Penniman, C. & Tveter-Gallagher, E., 1984. Phycocolloid ecology of underutilized economic red algae. Hydrobiologia, 116 (1), 542-546.
Mathieson, A.C. & Burns, R.L., 1971. Ecological studies of economic red algae. 1. Photosynthesis and respiration of Chondrus crispus (Stackhouse) and Gigartina stellata (Stackhouse) Batters. Journal of Experimental Marine Biology and Ecology, 7, 197-206.
Mathieson, A.C. & Burns, R.L., 1975. Ecological studies of economic red algae. 5. Growth and reproduction of natural and harvested populations of Chondrus crispus Stackhouse in New Hampshire. Journal of Experimental Marine Biology and Ecology, 17, 137-156.
Mercier, A., Sun, Z. & Hamel, J.-F., 2011. Internal brooding favours pre-metamorphic chimerism in a non-colonial cnidarian, the sea anemone Urticina felina. Proceedings of the Royal Society of London B: Biological Sciences, 282, 1-6.
Minchinton, T.E., Schiebling, R.E. & Hunt, H.L., 1997. Recovery of an intertidal assemblage following a rare occurrence of scouring by sea ice in Nova Scotia, Canada. Botanica Marina, 40, 139-148.
Molenaar, F.J. & Breeman, A.M., 1994. Ecotypic variation in Phyllophora pseudoceranoides (Rhodophyta) ensures winter reproduction throughout its geographic range. Journal of Phycology, 30 (3), 392-402.
Molenaar, F.J. & Breeman, A.M., 1997. Latitudinal trends in the growth and reproductive seasonality of Delesseria sanguinea, Membranoptera alata, and Phycodrys rubens (Rhodophyta). Journal of Phycology, 33, 330-343.
Molina, F.I., 1986. Petersenia pollagaster (Oomycetes): an invasive fungal pathogen of Chondrus crispus (Rhodophyceae). In The Biology of Marine Fungi (ed. S.T. Moss), 165-175.
Norton, T.A., 1992. Dispersal by macroalgae. British Phycological Journal, 27, 293-301.
Novaczek, I. & Breeman, A.M., 1990. Thermal ecotypes of amphi-Atlantic algae. 2. Cold-temperate species (Furcellaria lumbricalis and Polyides rotundus). Helgolander Meeresuntersuchungen, 44, 475-485.
O'Brien, P.J. & Dixon, P.S., 1976. Effects of oils and oil components on algae: a review. British Phycological Journal, 11, 115-142.
Plinski, M. & Florczyk, I., 1984. Changes in the phytobenthos resulting from the eutrophication of Puck Bay. Limnologica, 15, 325-327.
Prince, J.S. & Kingsbury, J.M., 1973. The ecology of Chondrus crispus at Plymouth, Massachusetts. 3. Effect of elevated temperature on growth and survival. Biology Bulletin, 145, 580-588.
Pringle, J., & Semple, R. 1980. The benthic algal biomass, commercial harvesting, and Chondrus growth and colonization off southwestern Nova Scotia. In Proceedings of the workshop on the relationship between sea urchin grazing and commercial plant / animal harvesting. Edited by J. Pringle, G. Sharp, and J. Caddy. Canadian Technical Report of Fisheries and Aquatic Sciences 954. pp. 144-169.
Pringle, J.D. & Mathieson, A.C., 1986. Chondrus crispus Stackhouse. Case Studies of Seven Commercial Seaweed Resources, 281, 49-122, FAO Fisheries Technical Paper.
Pybus, C., 1977. The ecology of Chondrus crispus and Gigartina stellata (Rhodophyta) in Galway Bay. Journal of the Marine Biological Association of the United Kingdom, 57, 609-628.
Rietema, H., 1995. Ecoclinal variation in Rhodomela confervoides along a salinity gradient in the North Sea and Baltic Sea. Botanica Marina, 38 (1-6), 475-480.
Schwenke, H., 1971. Water movement: 2. Plants. In Marine Ecology. Volume 1. Environmental Factors (2), 705-820 (ed. O. Kinne). Wiley-Interscience, London.
Scrosati, R., Garbary, D.J. & McLachlan, J., 1994. Reproductive ecology of Chondrus crispus (Rhodophyta, Gigartinales) from Nova Scotia, Canada. Botanica Marina, 37, 293-300.
Sharp, G.J., Tetu, C., Semple, R. & Jones, D., 1993. Recent changes in the seaweed community of western Prince Edward Island: implications for the seaweed industry. Hydrobiologia, 260-261, 291-296.
Sharp, G.J., Tremblay, D.M. & Roddick, D.L., 1986. Vulnerability of the southwestern Nova Scotia Chondrus crispus resource to handraking. Botanica Marina, 29, 449-453.
Simpson, F.J. & Shacklock, P.F., 1979. The cultivation of Chondrus crispus. Effect of temperature on growth and carageenan production. Botanica Marina, 22, 295-298.
Sköld, M., Josefson, A.B. & Loo, L.-O., 2001. Sigmoidal growth in the brittlestar Amphiura filiformis (Echinodermata: Ophiuroidea). Marine Biology, 139, 519-526.
Smith, J.E. (ed.), 1968. 'Torrey Canyon'. Pollution and marine life. Cambridge: Cambridge University Press.
Solé-Cava, A.M., Thorpe, J.P. & Todd, C.D., 1994. High genetic similarity between geographically distant populations in a sea anemone with low dispersal capabilities. Journal of the Marine Biological Association of the United Kingdom, 74, 895-902.
Stanley, S.J., 1992. Observations on the seasonal occurrence of marine endophytic and parasitic fungi. Canadian Journal of Botany, 70, 2089-2096.
Tariq, V.N., 1991. Antifungal activity in crude extracts of marine red algae. Mycological Research, 95 (12), 1433-1435.
Tasende, M.G. & Fraga, M.I., 1999. The growth of Chondrus crispus Stackhouse (Rhodophyta, Gigartinaceae) in laboratory culture. Ophelia, 51, 203-213.
Taylor, A. R. A., Chen, L. C. M., Smith, B. D., & Staples, L. S. 1981. Chondrus holdfasts in natural populations and in culture. In Proceedings of the International Seaweed Symposium 8, 140-145.
Tillin, H. & Tyler-Walters, H., 2014. Assessing the sensitivity of subtidal sedimentary habitats to pressures associated with marine activities. Phase 2 Report – Literature review and sensitivity assessments for ecological groups for circalittoral and offshore Level 5 biotopes. JNCC Report No. 512B, 260 pp. Available from: www.marlin.ac.uk/publications
Vadas, R.L., Johnson, S. & Norton, T.A., 1992. Recruitment and mortality of early post-settlement stages of benthic algae. British Phycological Journal, 27, 331-351.
Vidaver, W., 1972. Dissolved gases - plants. In Marine Ecology. Volume 1. Environmental factors (3), (ed. O. Kinne), 1471-1490. Wiley-Interscience, London.
Vogt, H. & Schramm, W., 1991. Conspicuous decline of Fucus in Kiel Bay (Western Baltic): what are the causes ? Marine Ecology Progress Series, 69, 189-194.
Wallentinus, I., 1978. Productivity studies on Baltic macroalgae. Botanica Marina, 21, 365-380.
Worm, B. & Chapman, A.R.O., 1998. Relative effects of elevated grazing pressure and competition from a red algal turf on two post settlement stages of Fucus evanescens. Journal of Experimental Marine Biology and Ecology, 220, 247-268.
This review can be cited as:
Last Updated: 07/12/2015