|Researched by||Dr Heidi Tillin & Georgina Budd||Refereed by||This information is not refereed.|
|EUNIS 2008||A5.261||Abra alba and Nucula nitidosa in circalittoral muddy sand or slightly mixed sediment|
|JNCC 2015||SS.SSa.CMuSa.AalbNuc||Abra alba and Nucula nitidosa in circalittoral muddy sand or slightly mixed sediment|
|JNCC 2004||SS.SSa.CMuSa.AalbNuc||Abra alba and Nucula nitidosa in circalittoral muddy sand or slightly mixed sediment|
|1997 Biotope||SS.CMS._.AbrNucCor||Abra alba, Nucula nitida and Corbula gibba in circalittoral muddy sand or slightly mixed sediment|
Non-cohesive muddy sands or slightly shelly/gravelly muddy sand characterized by the bivalves Abra alba and Nucula nitidosa. Other important taxa include Nephtys spp., Chaetozone setosa and Spiophanes bombyx with Fabulina fabula also common in many areas. The echinoderms Ophiura albida and Asterias rubens may also be present. The epibiotic biotope EcorEns may overlap this biotope. This biotope is part of the Abra community defined by Thorson (1957) and the infralittoral etage described by Glemarec (1973) (JNCC, 2015).
No text entered.
The muddy sand / mixed substratum of the biotope offers little habitat complexity. However, structural diversity may be provided by either localized physiographic features created by the hydrodynamic regime or the biota. Some diversity within the substratum is provided by the burrows and burrowing activity of infauna. Most species living within the sediment are limited to the area above the anoxic layer, the depth of which will vary depending on sediment particle size and organic content. However, the presence of burrows allows a larger surface area of sediment to become oxygenated, and probably enhances the survival of a considerable variety of small species and to a greater depth (Pearson & Rosenberg, 1978).
Benthic communities in deeper water, where light is insufficient for primary production, depend almost entirely on an input of energy via sedimentation of organic matter (Wood, 1987). Organic matter may be derived from phytoplankton, zooplankton, bacteria and faecal pellets, the supply of which is one of the main factors affecting production in these communities. If the majority of the phytoplankton and organic material is utilized in the surface waters, productivity of the seabed community would consequently be low. Estimates of productivity are available for individual species in the biotope, but specific community information was not found. For example:
The bivalves which characterize the biotope typically have an 'r' type life-cycle strategy (from Krebs, 1978, after Pianka, 1970), characteristics of which are high fecundity and rapid development, that allow rapid exploitation of available habitat. For instance, high densities of newly settled Corbula gibba (30, 000-67, 000 m⊃2;) and Abra alba (16, 000-22, 000 m⊃2;) were found at locations in the Limfjord, Denmark by Jensen (1988) and growth of both species was very rapid. However, recruitment in bivalves is heavily influenced by larval and post-settlement mortality so that large population increases are offset. Larval mortality results from predation during larval pelagic stages, predation from suspension feeding macrofauna (including conspecific adults) prior to settlement, deposit feeders after settlement and from settlement in unsuitable habitats. Mortality of the juveniles of marine benthic invertebrates can exceed 30% in the first day, and several studies report 90% mortality (Gosselin & Qian, 1997). In addition to larval dispersal, dispersal of juveniles and adults occurs via burrowing (Bonsdorff, 1984; Guenther, 1991), floating (Sörlin, 1988), and also possibly by bedload transport. It is expected therefore that recruitment can occur from both local and distant populations.
Polychaete worms in the biotope also tend to be 'r' type life-cycle strategists, with a dispersive planktonic larval stage that follows release and fertilization of gametes. Recruitment of Lagis koreni via its pelagic larvae is typically erratic between years, but newly settled juveniles may number several thousand per square metre (e.g. Macer, 1967; Basimi & Grove, 1985). Nichols (1977) recorded an early and late summer recruitment of Lagis koreni in Kiel Bay, but with additional sporadic recruitment occurring throughout the year. However, off the North Wales coast, Nicolaidou (1983) observed only one recruitment event (in June). Colonization of any new or disturbed substrata may also occur by colonization of adults (Eagle, 1975; Rees et al., 1977). Nephtys hombergii matures between two and three years of age. In the Tyne Estuary spawning of Nephtys hombergii occurred in May and September, whilst in Southampton Water the species spawned throughout the year with peaks in July and November (Oyenekan, 1986). The pelagic life cycle of Nephtys hombergii lasts seven to eight weeks at the end of which larvae metamorphose into benthic juveniles.
Echinocardium cordatum demonstrates a 'K' type life-cycle strategy (Rees & Dare, 1993) and subtidal populations of Echinocardium cordatum are reported to reproduce sporadically, e.g. one population recruited in only three years over a ten year period (Buchanan, 1966). Although, the species is fecund (> 1,000,000 eggs) recruitment is infrequently successful. Recruitment success may depend on temperatures of the preceding winter in some areas (Beukema, 1985).
Diaz-Castaneda et al. (1989) experimentally investigated recolonization sequences of benthic associations over a period of one year, following defaunation of the sediment. Recovery of the Abra alba community was rapid, recruitment occurring from surrounding populations via the plankton. The abundance, total biomass and diversity of the community all increased until a maximum was reached after 20 to 24 weeks, according to the season. The community within the experimental containers matched that of the surrounding areas qualitatively but quantitatively within 4 to 8 months depending on the seasonal availability of recruits, food supply and faunal interactions. The experimental data suggested that Abra alba would colonize available sediments within the year following environmental perturbation. Summer settled recruits may grow very rapidly and spawn in the autumn, whilst autumn recruits experience delayed growth and may not reach maturity until the following spring/summer. In the worst instance, a breeding population of Abra alba may take up to two years to fully establish. Dittman et al. (1999) observed that Nephtys hombergii was included amongst the macrofauna that colonized experimentally disturbed tidal flats within two weeks of the disturbance that caused defaunation of the sediment. In addition to larval recruitment, recolonization by polychaete worms could also occur via adult migration. In contrast to the dominant characterizing bivalves and polychaetes, Echinocardium cordatum is a long lived species and takes a relatively long time to reach reproductive maturity. Echinocardium cordatum breeds for the first time when two to three years old and recruitment of is often sporadic with reports of recruiting in only three years over a ten year period for a subtidal population (Buchanan, 1966).
Thus it is likely that the dominant infaunal bivalve and polychaete community of the CMS.AbrNucCor biotope would recover rapidly from a disturbance and mature populations of important characterizing species be present within a year. However, other components of the community that take longer to attain maturity, e.g. Echinocardium cordatum, and in their absence the CMS.AbrNucCor biotope would be recognized but may be considered impoverished.
No text entered.
|Depth Range||10-20 m, 20-30 m, 30-50 m|
|Water clarity preferences||Field Unresearched|
|Limiting Nutrients||Field unresearched|
|Salinity||Full (30-40 psu)|
|Physiographic||Enclosed coast / Embayment|
|Substratum||Mixed, Muddy sand|
|Tidal||Moderately Strong 1 to 3 knots (0.5-1.5 m/sec.), Weak < 1 knot (<0.5 m/sec.)|
|Wave||Exposed, Moderately exposed|
|Other preferences||Muddy sand|
No text entered.
The initial project scoping exercise to update the existing MarLIN assessments with the MarESA approach grouped biotopes where sensitivities were likely to be similar. The biotopes SS.SSa.IMuSa.FfabMag; SS.SSa.IMuSa.SsubNhom; SS.SSa.CFiSa.EpusOborApri; SS.SSa.CFiSa.ApriBatPo and SS.SSa.CMuSa.AalbNuc represent a continuum along depth and sediment gradients and were grouped as their responses to pressures are probably similar, given the presence of similar species and functional groups. These biotopes have been reviewed as a group, although the resultant reviews and sensitivity assessments are presented separately for each biotope. The specific biotope assessments have been updated with information on characterizing species, where this was readily available, however, the assessments are relatively generic for this group.
This biotope may recover from impacts via in-situ repair of damaged individuals, migration of adults of mobile species such as the errant polychaetes Glycera lapidum and Nephtys cirrosa, amphipods and urchins. Adults may also be transported in the water column following washout from sediments.Storm events may lead to the displacement of large numbers of individuals. Most bivalves will be able to reposition within the sediment and some, such as Glycymeris glycymeris, are also able to move and to relocate following displacement and disturbance (Thomas, 1975). For immobile species or where depopulation has occurred over a large area, recovery will depend on recolonization by pelagic larvae.
A large number of species are recorded in the biotopes within the assessed group and there may be large natural variation in species abundance over the course of a year or between years (see Dauvin, 1985 for Timoclea ovata; Fahy et al., 2003 for Spisula solida; Sardá et al., 1999 multispecies). These variations may not alter the biotope classification where habitat parameters, such as sediment type, remain as described in the classification and many of the characteristic species groups are present. For many of the bivalve species studied, recruitment is sporadic and depends on a successful spat fall but recruitment by the characterizing polychaetes may be more reliable. However, due to the large number of pre- and post-recruitment factors such as food supply, predation, and competition, recruitment of venerid bivalves and other species is unpredictable (Olafsson et al., 1994).
The life history characteristics of the characterizing bivalves and polychaetes and other species were reviewed. Little information was found for Moerella spp. Morton (2009) noted that despite the wide global distribution of the characterizing venerid bivalve, Timoclea ovata, little was known about its anatomy or basic biology. This appears to be the case for many of the other characterizing venerid bivalves and much more information was available for the polychaete species that occur in this biotope. Two linked factors that may explain this are the greater research effort in soft sediments with higher mud contents where sampling is easier than in coarse sediments. Venerid bivalves are also considered to be under-represented in grab samples (JNCC, 2015), so less is known of their occurrence on ecological and impact gradients.
The venerid bivalves in the biotope reach sexual maturity within two years, spawn at least once a year and have a pelagic dispersal phase (Guillou & Sauriau, 1985; Dauvin, 1985). No information was found concerning number of gametes produced, but the number is likely to be high as with other bivalves exhibiting planktotrophic development (Olafsson et al., 1994). Recruitment in venerids is likely to be episodic, some species such as Chamelea gallina may be long-lived (11-20 years). The long lifespan & slow growth rate suggest that this group is likely to take several years, even if initial recolonization were to occur rapidly (MES, 2010). Dauvin (1985) reported that Timoclea ovata (studied as Venus ovata) recruitment occurred in July-August in the Bay of Morlaix. However, the population showed considerable pluriannual variations in recruitment, which suggests that recruitment is patchy and/or post settlement processes are highly variable.
The species that are present in the biotope can be broadly characterized as either opportunist species that rapidly colonize disturbed habitats and increase in abundance, or species that are larger and longer-lived and that may be more abundant in an established, mature assemblage.
Species with opportunistic life strategies (small size, rapid maturation and short-life span of 1-2 years with production of large numbers of small propagules), include the bivalve Spisula solida; and the polychaetes Spiophanes bombyx, Spio filicornis, and Chaetozone setosa; also cumaceans; barnacles Balanus crenatus; and the tube worm Spirobranhchus (formerly Pomatoceros) lamarckii. These are likely to recolonize disturbed areas first, although the actual pattern will depend on recovery of the habitat, season of occurrence and other factors. The recovery of bivalves that recruit episodically and the establishment of a representative age-structured population for larger, longer lived organisms may require longer than two years. In an area that had been subjected to intensive aggregate extraction for 30 years, abundances of juvenile and adults Nephtys cirrosa had greatly increased three years after extraction had stopped (Mouleaert & Hostens, 2007). An area of sand and gravel subject to chronic working for 25 years had not recovered after 6 years when compared to nearby reference sites unimpacted by operations (Boyd et al., 2005). The characterizing Moerella (now Tellina) spp. are a relatively long-lived genus (6-10 years; MES, 2008, 2010) and the number of eggs is likely to be fewer than genera that have planktotrophic larvae. Similarly, Chamelea sp. and Dosinia sp. are long lived (11-20 years and up to 20 years, respectively; MES, 2008, 2010). While recruitment may be rapid, restoration of the biomass by growth of the colonizing individuals is likely to take many years.
Other longer lived species that may represent a more developed and stable assemblage include the polychaete Owenia fusiformis which lives for 4 years and reproduces annually (Gentil et al., 1990). Nepthys species and Glycera spp. are also longer-lived. Glycera are monotelic having a single breeding period towards the end of their life but may recover through migration and may persist in disturbed sediments through their ability to burrow (Klawe & Dickie, 1957). Glycera spp. have a high potential rate of recolonization of sediments, but the relatively slow growth-rate and long lifespan suggests that recovery of biomass following initial recolonization by post-larvae is likely to take several years (MES, 2010). Following dredging of subtidal sands in summer and autumn to provide material for beach nourishment in the Bay of Blanes, (north-west Mediterranean sea, Spain) recovery was tracked by Sardá et al. (2000). Recolonization in the dredged habitats was rapid, with high densities of Spisula subtruncata and Owenia fusiformis in the spring following dredging, although most of these recruits did not survive summer. However, Glycera spp. and Protodorvillea kefersteini had not recovered within two years (Sardá et al., 2000).
A number of studies have tracked recovery of sand and coarse sand communities following disturbance from fisheries (Gilkinson et al., 2005) and aggregate extraction (Boyd et al., 2005). The available studies confirm the general trend that, following severe disturbance, habitats are recolonized rapidly by opportunistic species (Pearson & Rosenberg, 1978). Experimental deployment of hydraulic clam dredges on a sandy seabed on Banquereau, on the Scotian Shelf, eastern Canada showed that within 2 years of the impact, polychaetes and amphipods had increased in abundance after 1 year (Gilkinson et al., 2005). Two years after dredging, abundances of opportunistic species were generally elevated relative to pre-dredging levels while communities had become numerically dominated (50-70%) by Spiophanes bombyx (Gilkinson et al., 2005). Van Dalfsen et al. (2000) found that polychaetes recolonized a dredged area within 5-10 months (reference from Boyd et al., 2005), with biomass recovery predicted within 2-4 years. The polychaete and amphipods are therefore likely to recover more rapidly than the characterizing bivalves and the biotope classification may revert, during recovery, to a polychaete dominated biotope.
Sardá et al. (1999) tracked annual cycles within a Spisula community in Bay of Blanes (north west Mediterranean sea, Spain) for 4 years. Macroinfaunal abundance peaked in spring, decreased sharply throughout the summer, with low density in autumn and winter. The observed trends were related to a number of species, including many that characterize this biotope such as Owenia fusiformis; Glycera sp.; Protodorvillea kefersteini; Mediomastus fragilis; Spisula subtruncata; and Branchiostoma lanceolatum. The Spisula subtruncata populations were dominated by juveniles, with high abundances in spring followed by declines in summer, with very few survivors 3 months after recruitment. Inter-annual differences in recruitment of Owenia fusiformis were apparent and this species showed spring/summer increases. Mediomastus fragilis also had spring population peaks but more individuals persisted throughout the year. Protodorvillea kefersteini exhibited a similar pattern with spring recruitment and a population that persisted throughout the year.
Where impacts also alter the sedimentary habitat, recovery of the biotope will also depend on recovery of the habitat to the former condition to support the characteristic biological assemblage. Recovery of sediments will be site-specific and will be influenced by currents, wave action and sediment availability (Desprez, 2000). Except in areas of mobile sands, the process tends to be slow (Kenny & Rees, 1996; Desprez, 2000 and references therein). Boyd et al. (2005) found that in a site subject to long-term extraction (25 years), extraction scars were still visible after six years and sediment characteristics were still altered in comparison with reference areas, with ongoing effects on the biota.
Resilience assessment. Where resistance is ‘None’ or ‘Low’ and an element of habitat recovery is required, resilience is assessed as ‘Medium’ (2-10 years), based on evidence from aggregate recovery studies in similar habitats including Boyd et al. (2005). Where resistance of the characterizing species is ‘Low’ or ‘Medium’ and the habitat has not been altered, resilience is assessed as ‘High’ as, due to the number of characterizing species and variability in recruitment patterns, it is likely that the biotope would be considered representative and hence recovered after two years although some parameters such as species richness, abundance and biotopes may be altered . Recovery of the seabed from severe physical disturbances that alter sediment character may also take up to 10 years or longer (Le Bot et al., 2010), although extraction of gravel may result in more permanent changes and this will delay recovery.
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 recognisable as the initial habitat of interest. It should be noted that the recovery rates are only indicative of the recovery potential.
Davenport & Davenport (2005) demonstrated that the limits of thermal tolerance to high and low temperatures reflect distribution of intertidal macroinvertebrate species. Species that occur highest on the shore are more tolerant of a wider range of temperatures than species that occurred low on the shore or subtidally. As subtidal biotopes are less exposed to temperature fluctuations, the characterizing species may be less able to tolerate temperature fluctuations.
No direct evidence was found to support assessment of this pressure. Very few laboratory studies have been carried out on the characterizing species and the assessment relies on information on larvae in the plankton or monitoring of settlement and records of species distribution. Species from different areas may be acclimated to prevailing conditions and life histories may vary, e.g. Chamelea gallina longevity varies between populations (Gaspar et al., 2004) as does the longevity of Amphipholis squamata in different locations and habitats (Emson et al., 1989).
Kröncke et al. (1998) examined long-term changes in the macrofauna in the subtidal zone off Norderney, one of the East Frisian barrier islands. The analysis suggested that macrofauna were severely affected by cold winters whereas storms and hot summers have no impact on the benthos. A long-term increase in temperature might cause a shift in species composition. Long‐term analysis of the North Sea pelagic system has identified yearly variations in larval abundance of Echinodermata, Arthropoda, and Mollusca larvae that correlate with sea surface temperatures. Larvae of benthic echinoderms and decapod crustaceans increased after the mid‐1980s, coincident with a rise in North Sea sea surface temperature, whereas bivalve larvae underwent a reduction (Kirby et al., 2008). An increase in temperature may alter larval supply and in the long-term, and over large spatial scales, may result in changes in community composition.
Temperature cues influence the timing of gametogenesis and spawning in several species present in the biotope. Seasonal variations in reproductive cycle of Spisula solida were studied at a site off Vilamoura, southern Portugal. The onset of spawning took place in February when the seawater temperature began to increase and spawning ended in May. It is possible that Spisula solida does not spawn at a definite temperature, rather responding to the increase in seawater temperature (Gaspar & Monteiro, 1999). Many polychaete species including Mediomastus fragilis, Owenia fusiformis and Protodorvillea kefersteini also show spring/early summer recruitment (Sardá et al., 1999).
Sensitivity assessment. Little evidence was available to assess this pressure. Assemblages in fine sands and muddy sands contain many of the characterizing species occur in the Mediterranean (see resilience section Sardá et al., 1999; Sardá et al., 2000), where temperatures are higher than experienced in the UK. It is considered likely, therefore, that a chronic change in temperature at the pressure benchmark would be tolerated by species with a wide distribution or that some species or groups of species would be resistant. An acute change may exceed thermal tolerances or lead to spawning or other biological effects. These effects may be sub-lethal or result in the removal of only a proportion of less tolerant species. Biotope resistance is therefore assessed as ‘Medium’ and resilience is assessed as ‘High’. Biotope sensitivity is therefore assessed as ‘Low’.
Davenport & Davenport (2005) demonstrated that the limits of thermal tolerance to high and low temperatures reflect the distribution of intertidal macroinvertebrate species. Species that occur highest on the shore are more tolerant of a wider range of temperatures than species that occurred low on the shore or subtidally. As subtidal biotopes are less exposed to temperature fluctuations the characterizing species may be less able to tolerate temperature fluctuations.
Long‐term analysis of the North Sea pelagic system has identified yearly variations in larval abundance of Echinodermata, Arthropoda, and Mollusca larvae that correlate with sea surface temperatures. Larvae of benthic echinoderms and decapod crustaceans increased after the mid‐1980s, coincident with a rise in North Sea sea surface temperature, whereas bivalve larvae underwent a reduction (Kirby et al., 2008). A decrease in temperature may alter larval supply and in the long-term, and over large spatial scales, may result in changes in community composition.
Sensitivity assessment. Many of the characterizing species are found in more northern waters than the UK and may be adapted to colder temperatures. Plankton studies suggest that colder waters may favour bivalve larvae. An acute change may exceed thermal tolerances or lead to spawning or other biological effects. These effects may be sub-lethal or remove only a proportion of less tolerant species. Biotope resistance is therefore assessed as ‘Medium’ and resilience is assessed as ‘High’. Biotope sensitivity is therefore assessed as ‘Low’.
The assessed biotope occurs in full salinity (30-35 ppt) (JNCC, 2015). A change at the pressure benchmark therefore refers to a change to full salinity. No directly relevant evidence was found to assess this pressure. A study from the Canary Islands indicates that exposure to high salinity effluents (47- 50 psu) from desalination plants alter the structure of biological assemblages, reducing species richness and abundance (Riera et al., 2012). Bivalves and amphipods appear to be less tolerant of increased salinity than polychaetes and were largely absent at the point of discharge. Polychaetes, including species or genera that occur in this biotope, such as Spio filicornis, Glycera spp. and Lumbrineris sp., were present at the discharge point (Riera et al., 2012). The ophiuroid Amphipholis squamata has been recorded in areas of high salinity (52-55 ppt) in the Arabian Gulf (Price, 1982), indicating local acclimation may be possible.
Sensitivity assessment. High saline effluents alter the structure of biological assemblages. Polychaete species may be more tolerant than bivalves so that an increase in salinity may lead to a shift in community composition. Biotope resistance is therefore assessed as ‘Low’ and resilience as ‘Medium’, as bivalve recovery may depend on episodic recruitment. Biotope sensitivity is assessed as ‘Medium’.
Species tolerances to decreases in salinity are likely to vary but changes in sensitivity are likely to result in changes in species richness and abundance with some shift in species composition. Biotope resistance is therefore assessed as ‘Low’ and resilience as ‘Medium’, as bivalve recovery may depend on episodic recruitment. Biotope sensitivity is assessed as ‘Medium’.
This biotope is recorded in areas where tidal flow varies between moderately strong (0.5-1,5 m/s) and weak (>0.5 m/s) (JNCC, 2015). Sands are less cohesive than mud sediments and a change in water flow at the pressure benchmark may alter sediment transport patterns within the biotope. Hjulström (1939), concluded that fine sand (particle diameter of 0.3-0.6 mm) was easiest to erode and required a mean velocity of 0.2 m/s. Erosion and deposition of particles greater than 0.5 mm require a velocity > 0.2 m/s to alter the habitat. The topography of this habitat is shaped by currents and wave action that influence the formation of ripples in the sediment. Specific fauna may be associated with troughs and crests of these bedforms. may form following an increase in water flow, or disappear following a reduction in flow.
Sensitivity assessment. This biotope occurs in areas subject to moderately strong water flows and these are a key factor maintaining the clean sand habitat. Changes in water flow may alter the topography of the habitat and may cause some shifts in abundance. However, a change at the pressure benchmark (increase or decrease) is unlikely to affect biotopes that occur in mid-range flows and biotope resistance is therefore assessed as ‘High’ and resilience is assessed as ‘High’ so that the biotope is considered to be ‘Not sensitive’.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Changes in emergence are 'Not relevant' to this biotope which is restricted to fully subtidal habitats.
As this biotope occurs in sublittoral habitats, it is not directly exposed to the action of breaking waves. Associated polychaete species that burrow are protected within the sediment but the characterizing bivalves would be exposed to oscillatory water flows at the seabed. They and other associated species may be indirectly affected by changes in water movement where these impact the supply of food or larvae or other processes. No specific evidence was found to assess this pressure.
Sensitivity assessment. The range of wave exposures experienced by this biotpe and similar infralittoral and circalittoral biotopes is considered to indicate, by proxy, that the biotope would have ‘High’ resistance and by default ‘High’ resilience to a change in significant wave height at the pressure benchmark. The biotope is therefore classed as ‘Not sensitive’.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
The capacity of bivalves to accumulate heavy metals in their tissues, far in excess of environmental levels, is well known. Reactions to sub-lethal levels of heavy metal stressors include siphon retraction, valve closure, inhibition of byssal thread production, disruption of burrowing behaviour, inhibition of respiration, inhibition of filtration rate, inhibition of protein synthesis and suppressed growth (see review by Aberkali & Trueman, 1985). Stirling (1975) investigated the effect of exposure to copper on Tellina tenuis. The 96 hour LC50 for Cu was 1000 µg/l. Exposure to Cu concentrations of 250 µg/l and above inhibited burrowing behaviour and would presumably result in greater vulnerability to predators. Similarly, burial of the venerid bivalve, Venerupis senegalensis, was inhibited by copper spiked sediments, and at very high concentrations, clams closed up and did not bury at all (Kaschl & Carballeira, 1999). The copper 10 day LC50 for Venerupis senegalensis was found to be 88 µg/l in sandy sediments (Kaschl & Carballeira, 1999).
Echinoderms are also regarded as being intolerant of heavy metals (e.g. Bryan, 1984; Kinne, 1984) while polychaetes are tolerant (Bryan, 1984).
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Suchanek (1993) reviewed the effects of oil on bivalves. Generally, contact with oil causes an increase in energy expenditure and a decrease in feeding rate, resulting in less energy available for growth and reproduction. Sublethal concentrations of hydrocarbons also reduce byssal thread production (thus weakening attachment) and infaunal burrowing rates. Conan (1982) investigated the long term effects of the Amoco Cadiz oil spill at St Efflam beach in France. It was estimated that the delayed mortality effects on sand and mud biotas were 1.4 times as large as the immediate effects. Fabulina fabula (studied as Tellina fabula) started to disappear from the intertidal zone a few months after the spill and from then on was restricted to subtidal levels. In the following 2 years, recruitment of Fabulina fabula was very much reduced. The author commented that, in the long term, the biotas most severely affected by oil spills are low energy sandy and muddy shores, bays and estuaries. In such places, populations of species with long and short term life expectancies (e.g. Fabulina fabula, Echinocardium cordatum and Ampelisca sp.) either vanished or displayed long-term decline following the Amoco Cadiz oil spill. Polychaetes, however, including Nephtys hombergii, cirratulids and capitellids were largely unaffected.
Dauvin (1998) reported the effects of the Amoco Cadiz spill on the fine sand community in the Bay of Morlaix. Reductions in abundance, biomass and production of the community were very evident through the disappearance of the dominant populations of the amphipods Ampelisca sp. which are very intolerant of oil contamination. 2 weeks after the spill, the level of hydrocarbons in subtidal sediments reached 200 ppm (Dauvin, 1984; cited in Poggiale & Dauvin, 2001). This caused the disappearance of the Ampelisca populations, leaving behind a single species, Ampelisca sarsi, in very low densities. The sediment rapidly depolluted and in 1981 benthic recruitment occurred in normal conditions (Dauvin, 1998). However, the recovery of the Ampelisca populations took up to 15 years. This was probably due to the amphipods' low fecundity, lack of pelagic larvae and the absence of local unperturbed source populations (Poggiale & Dauvin, 2001).
Echinoderms also seem to be especially intolerant of the toxic effects of oil, probably because of the large amount of exposed epidermis (Suchanek, 1993). The high intolerance of Echinocardium cordatum to hydrocarbons was seen by the mass mortality of animals, down to about 20m depth, shortly after the Amoco Cadiz oil spill (Cabioch et al., 1978).
Dauvin (2000) The muddy fine sand Abra alba-Melinna palmata community from the Bay of Morlaix (western English Channel) was strongly polluted by hydrocarbons from the Amoco Cadiz oil spill in April 1978. Long-term changes of this community (1977-1996) showed that it was weakly affected by the spill. This was due to the low number and low abundance of sensitive species present on the community in normal conditions. Polychaetes, such as Chaetozone setosa dominated the community, supporting high levels of organic matter. Only two opportunistic polychaetes Mediomastus fragilis and Tharyx marioni increased in abundance just after the spill.
Glycera spp. were described by Hiscock et al. (2004; 2005, from Levell et al., 1989) as a very tolerant taxa, found in enhanced abundances in the transitional zone along hydrocarbon contamination gradients surrounding oil platforms.
The Braer oil spill in Shetland in 1993 provided an opportunity to identify species that increased or declined in abundance where oiling occurred. Severe weather conditions meant that oil was incorporated into sediments. Kingston et al. (1995) noted that the congener Lumbrineris gracilis (from Hiscock et al., 2004b) declined at oiled sites.
Diaz-Castaneda et al. (1989) looked at colonization of defaunated and polluted sediments in Dunkerque harbour. The sediment was polluted with both heavy metals and oil. Capitella capitata was generally the first polychaete to colonize the polluted sediment. Spio filicornis took between 7 weeks and 3 months to appear in the sediment suggesting it is tolerant of oil pollution.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Stirling (1975) investigated the effects of phenol, a non-persistent, semi-synthetic organic pollutant, on Tellina tenuis. Exposure to phenol produced a measurable effect on burrowing at all concentrations tested, i.e. 50 mg/l and stronger. Sub-lethal effects of exposure to phenol included delayed burrowing and valve adduction to exclude the pollutant from the mantle cavity. After exposure to 100 mg/l for 24 hours, the majority of animals were extended from their shells and unresponsive to tactile stimulation. Following replacement of the phenol solution with clean seawater, good recovery was exhibited after 2 days for animals exposed to 50 mg/l and some recovery occurred after 4 days for animals exposed to 100 mg/l.
Similarly, no evidence was found directly relating to the effects of synthetic chemicals on Magelona mirabilis. However, there is evidence from other polychaete species. Collier & Pinn (1998) investigated the effect on the benthos of ivermectin, a feed additive treatment for infestations of sea-lice on farmed salmonids. The polychaete Hediste diversicolor was particularly susceptible, exhibiting 100% mortality within 14 days when exposed to 8 mg/m² of ivermectin in a microcosm. Arenicola marina was also intolerant of ivermectin through the ingestion of contaminated sediment (Thain et al., 1998; cited in Collier & Pinn, 1998) and it was suggested that deposit feeding was an important route for exposure to toxins. Beaumont et al. (1989) investigated the effects of tri-butyl tin (TBT) on benthic organisms. At concentrations of 1-3 µg/l there was no significant effect on the abundance of Hediste diversicolor after 9 weeks in a microcosm. However, no juvenile polychaetes were retrieved from the substratum and hence there is some evidence that TBT had an effect on the larval and/or juvenile stages.
Detergents used to disperse oil from the Torrey Canyon oil spill caused mass mortalities of Echinocardium cordatum (Smith, 1968) and its intolerance to TBT is similar to that of other benthic organisms with LC50 values of 222ng Sn/l in pore water and 1594ng Sn/g dry weight of sediment (Stronkhorst et al., 1999). Gammaridean amphipods have also been reported to be intolerant of TBT with 10 day LC50 values of 1-48 ng/l (Meador et al., 1993).
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
Riedel et al. (2012) assessed the response of benthic macrofauna to hypoxia advancing to anoxia in the Mediterranean. The hypoxic and anoxic conditions were created for 3-4 days in a box that enclosed in-situ sediments. In general molluscs were more resistant than polychaetes, with 90% surviving hypoxia and anoxia, whereas only 10% of polychaetes survived. Exposed individual Timoclea ovata and Tellina serrata survived the experiment but the exposed Glycera spp. died. In general epifauna were more sensitive than infauna, mobile species more sensitive than sedentary species and predatory species more sensitive than suspension and deposit feeders. The test conditions did not lead to the production of hydrogen sulphide which may have reduced mortalities compared to some observations.
Further evidence of sensitivity was available for some of the polychaete species associated with this biotope. Rabalais et al. (2001) observed that hypoxic conditions in the north Coast of the Gulf of Mexico (oxygen concentrations from 1.5 to 1 mg/l (1 to 0.7 ml/l) led to the emergence of Lumbrineris sp. from the substrate these then lie motionless on the surface. Glycera alba was found to be able to tolerate periods of anoxia resulting from inputs of organic rich material from a wood pulp and paper mill in Loch Eil (Scotland) (Blackstock & Barnes, 1982). Nierman et al. (1990) reported changes in a fine sand community for the German Bight in an area with regular seasonal hypoxia. In 1983, oxygen levels were exceptionally low (<3 mg O2/l) in large areas and <1 mg O2/l in some areas. Species richness decreased by 30-50% and overall biomass fell. Owenia fusiformis were reduced in abundance significantly by the hypoxia Spiophanes bombyx was found in small numbers at some, but not all areas, during the period of hypoxia. Once oxygen levels returned to normal Spiophanes bombyx increased in abundance; the evidence suggests that at least some individuals would survive hypoxic conditions.
Sensitivity assessment. Riedel et al. (2012) provide evidence on general sensitivity trends. The characterizing bivalves are likely to survive hypoxia at the pressure benchmark although the polychaetes present, particularly the mobile predatory species such as Glycera and Nephtys may be less tolerant. As the biotope is characterized by bivalves and polychaetes, resistance is assessed as ‘Low’ and resilience as ‘High’ based on migration, water transport of adults and recolonization by pelagic larvae. Biotope sensitivity is assessed as ‘Low’.
This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations. The pressure benchmark is set at compliance with Water Framework Directive (WFD) criteria for good status, based on nitrogen concentration (UKTAG, 2014).
Bivalves, polychaetes and other invertebrate species are unlikely to be directly affected by changes in nutrient enrichment.
Sensitivity assessment. As this biotope is structured by the sediments and water flow rather than nutrient enrichment and is not characterized by macroalgae (although some may be present), the biotope is considered to have ‘High’ resistance to this pressure and ‘High’ resilience, (by default) and is assessed as ‘Not sensitive’.
At the pressure benchmark, organic inputs are likely to represent a food subsidy for the associated deposit feeding species and are unlikely to significantly affect the structure of the biological assemblage or impact the physical habitat. Biotope sensitivity is therefore assessed as ‘High’ and resilience as ‘High’ (by default), and the biotope is therefore considered to be ‘Not sensitive’.
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.
The biotope is characterized by the sedimentary habitat (JNCC, 2015), so a change to an artificial or rock substratum would alter the character of the biotope leading to reclassification and the loss of the sedimentary community including the characterizing bivalves, polychaetes and echinoderms that live buried within the sediment.
Sensitivity assessment. Based on the loss of the biotope, resistance is assessed as ‘None’, recovery is assessed as ‘Very Low’ (as the change at the pressure benchmark is permanent), and sensitivity is assessed as ‘High’.
This biotope is found in medium to very fine sand with some silt (JNCC, 2015). The change referred to at the pressure benchmark is a change in sediment classification (based on Long, 2006) rather than a change in the finer-scale original Folk categories (Folk, 1954). For sand sediments, resistance is assessed based on a change to either mixed sediments or mud and sandy muds.
Sediment type is a key factor structuring the biological assemblage present in the biotope. Surveys over sediment gradients and before-and-after impact studies from aggregate extraction sites where sediments have been altered indicate patterns in change. The biotope classification (JNCC, 2015) provides information on the sediment types where biotopes are found and indicates likely patterns in change if the sediment were to alter.
Differences in biotope assemblages in areas of different sediment type are likely to be driven by pre and post recruitment processes. Sediment selectivity by larvae will influence levels of settlement and distribution patterns. Snelgrove et al. (1999) demonstrated that Spisula solidissima, selected coarse sand over muddy sand, and capitellid polychaetes selected muddy sand over coarse sand, regardless of site. Both larvae selected sediments typical of adult habitats, however, some species were nonselective (Snelgrove et al., 1999) and presumably in unfavourable habitats post recruitment, mortality will result for species that occur in a restricted range of habitats. Some species may, however, be present in a range of sediments. Post-settlement migration and selectivity also occurred on small scales (Snelgrove et al., 1999).
Cooper et al. (2011) found that characterizing species from sand dominated sediments were equally likely to be found in gravel dominated sediments, and an increase in sediment coarseness may not result in loss of characterizing species but biotope classification may revert to the biotope SS.SCS.CCS.MedLumVen, which occurs in gravels (JNCC, 2015).
Desprez (2000) found that a change of habitat to fine sands from coarse sands and gravels (from deposition of screened sand following aggregate extraction) changed the biological communities present. Tellina pygmaea and Nephtys cirrosa dominated the fine sand community. Dominant species of coarse sands, Echinocyamus pusillus and Amphipholis squamata, were poorly represented and the characteristic species of gravels and shingles were absent (Desprez, 2000).
Sensitivity assessment. A change to finer, muddy and mixed sediments is likely to reduce the abundance of the characterizing Tellina spp., venerid bivalves and other bivalves such as Spisula solida, and favour polychaetes. Such changes would lead to biotope reclassification. Biotope resistance is therefore assessed as ‘Low’ (as some species may remain), resilience is Very low (the pressure is a permanent change) and sensitivity is assessed as High.
Most of the animals that occur in this biotope are shallowly buried and extraction of the sediment will remove the biological assemblage.
Sensitivity assessment. Resistance is assessed as ‘None’ as extraction of the sediment swill remove the characterizing and associated species present. Resilience is assessed as ‘Medium’ as some species may require longer than two years to re-establish (see resilience section) and sediments may need to recover (where exposed layers are different). Biotope sensitivity is therefore assessed as ‘Medium’.
Abrasion is likely to damage epifauna and flora and may damage a proportion of the characterizing species, biotope resistance is therefore assessed as ‘Medium’. Resilience is assessed as ‘High’ as opportunistic species are likely to recruit rapidly and some damaged characterizing species may recover or recolonize. Biotope sensitivity is assessed as ‘Low’.
The epifauna and infaunal assemblages of both stable and dynamic fine sands are susceptible to direct physical disturbance from towed demersal gears and dredges which penetrate and disturb the sediment e.g. Eleftheriou & Robertson, 1992; Kaiser et al., 1998; Robinson & Richardson, 1998; Schwinghamer et al., 1996; Freese et al., 1999; Prena et al,. 1999; Bergman & Van Santbrick 2000a,b; Tuck et al., 2000; Kenchington et al., 2001; Gilkinson et al., 2005). In general, fishing using towed gears results in the mortality of non-target organisms either through physical damage inflicted by the passage of the trawl or indirectly by disturbance, damage, exposure and subsequent predation. Beam trawling, for example, decreases the density of common echinoderms, polychaetes and molluscs (Bergman & Hup, 1992) and decreases the density and diversity of epifauna in stable sand habitats (Kaiser & Spencer, 1996).
Gilkinson et al. (1998) simulated the physical interaction of otter trawl doors with the seabed in a laboratory test tank using a full-scale otter trawl door model. Between 58% and 70% of the bivalves in the scour path that were originally buried were completely or partially exposed at the test bed surface. However, only two out of a total of 42 specimens showed major damage. The pressure wave associated with the otter door pushes small bivalves out of the way without damaging them. Where species can rapidly burrow and reposition (typically within species occurring in unstable habitats) before predation mortality rates will be relatively low. These experimental observations are supported by diver observations of fauna dislodged by a hydraulic dredge used to catch Ensis spp. Small bivalves were found in the trawl tracks that had been dislodged from the sediments, including the venerid bivalves Dosinia exoleta, Chamelea striatula and the hatchet shell Lucinoma borealis. These were usually intact (Hauton et al., 2003a) and could potentially reburrow.
Larger, fragile species are more likely to be damaged by sediment penetration and disturbance than smaller species (Tillin et al., 2006). Bergman & van Santbrink (2000a,b) suggested that the megafauna were amongst the species most vulnerable to direct mortality due to bottom trawling in sandy sediments. Stomach analysis of fish caught scavenging in the tracks of beam trawls found parts of Spatangus purpureus and Ensis spp. indicating that these had been damaged and exposed by the trawl (Kaiser & Spencer, 1994a). Capasso et al. (2010) compared benthic survey datasets from 1895 and 2007 for an area in the English Channel. Although methodological differences limit direct comparison, the datasets appear to show that large, fragile urchin species including Echinus esculentus, Spatangus purpureus and Psammechinus miliaris and larger bivalves had decreased in abundance. Small, mobile species such as amphipods and small errant and predatory polychaetes (Nephtys, Glycera, Lumbrineris) appeared to have increased (Capasso et al., 2010). The area is subject to beam trawling and scallop dredging and the observed species changes would correspond with predicted changes following physical disturbance. Two small species: Timoclea ovata and Echinocyamus pusillus had increased in abundance between the two periods.
Sensitivity assessment. The trawling studies and the comparative study by Capasso et al. (2010) suggest that the biological assemblage present in this biotope is characterized by species that are relatively tolerant of penetration and disturbance of the sediments. Either species are robust or buried within sediments or are adapted to habitats with frequent disturbance (natural or anthropogenic) and recover quickly. The results suggest that a reduction in physical disturbance may lead to the development of a community with larger, more fragile species including large bivalves. Biotope resistance is assessed as ‘Medium’ as some species will be displaced and may be predated or injured and killed. Biotope resilience is assessed as ‘High’ as most species will recover rapidly and the biotope is likely to still be classified as the same type following disturbance. Biotope sensitivity is therefore assessed as ‘Low’.
A change in turbidity at the pressure benchmark is assessed as an increase from intermediate 10-100 mg/l to medium (100-300 mg/l) and a change to clear (<10 mg/l). An increase or decrease in turbidity may affect primary production in the water column and indirectly alter the availability of phytoplankton food available to species in filter feeding mode. However, phytoplankton will also be transported from distant areas and so the effect of increased turbidity may be mitigated to some extent. According to Widdows et al. (1979), growth of filter-feeding bivalves may be impaired at suspended particulate matter (SPM) concentrations >250 mg/l.
Sensitivity assessment. No direct evidence was found to assess impacts on the characterizing and associated species. The characterizing, suspension feeding bivalves are not predicted to be sensitive to decreases in turbidity and may be exposed to, and tolerant of, short-term increases in turbidity following sediment mobilization by storms and other events. An increase in suspended solids, at the pressure benchmark may have negative impacts on growth and fecundity by reducing filter feeding efficiency and imposing costs on clearing. Biotope resistance is assessed as ‘Medium’ as there may be some shift in the structure of the biological assemblage and resilience is assessed as ‘High’ (following restoration of typical conditions). Biotope sensitivity is assessed as ‘Low’.
Addition of fine material will alter the character of this habitat by covering it with a layer of dissimilar sediment and will reduce suitability for the species associated with this feature. Recovery will depend on the rate of sediment mixing or removal of the overburden, either naturally or through human activities. Recovery to a recognisable form of the original biotope will not take place until this has happened. In areas where the local hydrodynamic conditions are unaffected, fine particles will be removed by wave action moderating the impact of this pressure. The rate of habitat restoration would be site-specific and would be influenced by the type of siltation and rate. Long-term or permanent addition of fine particles would lead to re-classification of this biotope type (see physical change pressures). The additions of silts to a Spisula solida bed in Waterford Harbour (Republic of Ireland) from earthworks further upstream, for example, reduced the extent of the bed (Fahy et al., 2003). No information was provided on the depth of any deposits.
Most bivalve species are capable of burrowing through sediment to feed, e.g. Abra alba are capable of upwardly migrating if lightly buried by additional sediment (Schafer, 1972). There may be an energetic cost expended by species to either re-establish burrow openings, to self-clean feeding apparatus or to move up through the sediment, though this is not likely to be significant. Most animals will be able to reburrow or move up through the sediment within hours or days. Bijkerk (1988, results cited from Essink, 1999) indicated that the maximal overburden through which small bivalves could migrate was 20 cm in sand for Donax and approximately 40 cm in mud for Tellina sp. and approximately 50 cm in sand. No further information was available on the rates of survivorship or the time taken to reach the surface. Little direct evidence was found to assess the impact of this pressure at the benchmark level. Powilleit et al. (2009) studied the response of the polychaete Nephtys hombergii to smothering. This species successfully migrated to the surface of 32-41 cm deposited sediment layer of till or sand/till mixture and restored contact with the overlying water. The high escape potential could partly be explained by the heterogeneous texture of the till and sand/till mixture with ‘voids’. While crawling upward to the new sediment surfaces burrowing velocities of up to 20 cm/day were recorded for Nephtys hombergii. Similarly, Bijkerk (1988, results cited from Essink 1999) indicated that the maximal overburden through which species could migrate was 60 cm through mud for Nephtys and 90 cm through sand. No further information was available on the rates of survivorship or the time taken to reach the surface.
Sensitivity assessment. Bivalves and polychaetes and other species are likely to be able to survive short periods under sediments and to reposition. However, as the pressure benchmark refers to fine material, this may be cohesive and species characteristic of sandy habitats may be less adapted to move through this than sands. Biotope resistance is assessed as 'Medium' as some mortality of characterizing and associated species may occur. Biotope resilience is assessed as 'High' and biotope sensitivity is assessed as 'Low'.
Bijkerk (1988, results cited from Essink, 1999) indicated that the maximal overburden through which small bivalves could migrate was 20 cm in sand for Donax and approximately 40 cm in mud for Tellina sp. and approximately 50 cm in sand. No further information was available on the rates of survivorship or the time taken to reach the surface.
Sensitivity assessment. The character of the overburden is an important factor determining the degree of vertical migration of buried bivalves. Individuals are more likely to escape from a covering similar to the sediments in which the species is found than a different type. Resistance is assessed as ‘Low’ as few individuals are likely to reposition. Resilience is assessed as ‘Medium’ and sensitivity is assessed as ‘Medium’.
|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)|
Not relevant. No information was found concerning the intolerance of the biotope or the characterizing species to noise. The siphons of bivalves and palps of polychaetes are likely to detect vibrations and are probably withdrawn as a predator avoidance mechanism. However, it is unlikely that the biotope will be affected by noise or vibrations caused by noise at the level of the benchmark.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
|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)|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Key characterizing species within this biotope are not cultivated or translocated. This pressure is therefore considered ‘Not relevant’ to this biotope group.
Few invasive non-indigenous species may be able to colonize mobile sands, due to the high-levels of sediment disturbance. However, two species may be of concern. The slipper limpet Crepidula fornicata may settle on stones in substrates and hard surfaces such as bivalve shells and can sometimes form dense carpets which can smother bivalves and alter the seabed, making the habitat unsuitable for larval settlement. Dense aggregations trap suspended silt, faeces and pseudofaeces altering the benthic habitat. Where slipper limpet stacks are abundant, few other bivalves can live amongst them (Fretter & Graham, 1981; Blanchard, 1997). Muddy and mixed sediments in wave sheltered areas are probably optimal but Crepidula fornicata has been recorded from a wide variety of habitats including clean sands and areas subject to moderately strong tidal streams (Blanchard, 1997; De Montaudouin & Sauriau, 1999). Bohn et al. (2015) report that in the Milford Haven Waterway (MHW) in south-west Wales, UK, subtidally, highest densities were found in areas of high gravel content (grain sizes 16-256 mm), suggesting that the availability of this substrata type is beneficial for its establishment.
The colonial ascidian Didemnum vexillum is present in the UK but appears to be restricted to artificial surfaces such as pontoons, this species may, however, have the potential to colonize and smother offshore gravel habitats. Valentine et al., (2007) describe how Didemnum sp. appear to have rapidly colonized gravel areas on the Georges Bank (US/Canada boundary). Colonies can coalesce to form large mats that may cover more than 50% of the seabed in parts. Areas of mobile sand bordered communities of Didemnum sp. and these, therefore, do not appear to be suitable habitats (Valentine et al., 2007).
Although not currently established in UK waters, the whelk Rapana venosa, may spread to UK habitats from Europe. Both Rapana venosa and the introduced oyster drill Urosalpinx cinerea both predate on bivalves and could therefore negatively affect the characterizing bivalve species.
Sensitivity assessment. The sediments characterizing this biotope are likely to too mobile or otherwise unsuitable for most of the recorded invasive non-indigenous species currently recorded in the UK. The slipper limpet may colonize this habitat resulting in habitat change and potentially classification to the biotope which is found in similar habitats SS.SMx.IMx.CreAsAn. Didemnum sp. and non-native predatory gastropods may also emerge as a threat to this biotope, although more mobile sands may exclude Didemnum. Based on Crepidula fornicata biotope resistance is assessed as ‘None’ and resilience as ‘Very low’ (as removal of established non-native is unlikely), so that biotope sensitivity is assessed as ‘High’.
No evidence was found for the characterizing polychaete species. Populations of bivalve species may be subject to a variety of diseases and parasites but evidence for the characterizing bivalves is limited. Berilli et al. (2000) conducted a parasitological survey of the bivalve Chamelea gallina in natural beds of the Adriatic Sea, where anomalous mortalities had been observed in 1997-1999. The occurrence of protozoans belonging to the families Porosporidae, Hemispeiridae and Trichodinidae was recorded. Porosporidae of the genus Nematopsis, present with 4 species, showed a prevalence of 100%. The results suggested that severe infections of protozoans of the genus Nematopsis could cause a not negligible respiratory sufferance, with a possible role in the decline of the natural banks of Chamelea gallina (Berilli et al., 2000).
Bacterial diseases are frequently found in molluscs during their larval stages, but seem to be relatively insignificant in populations of adult animals (Lόpez-Flores et al., 2004). This may be due to the primary defence mechanisms of molluscs, phagocytosis and encapsulation, which fight against small-sized pathogens, and whose resistance may be age related (Sinderman, 1990; Lόpez-Flores et al., 2004).
Individuals of Fabulina fabula from Boulogne-sur-Mer (studied as Angulus fabula) were infected with the trematode parasite Gymnophallus strigatus, causing erosion of the shell (Giard, 1897, cited in Kinne, 1983).
Sensitivity assessments. Pathogens may cause mortality and there may be a minor decline in species richness or abundance in the biotope. As there is no evidence for mass mortalities of characterizing species that would alter biotope classification biotope resistance is assessed as ‘Medium’. Biotope resilience is assessed as ‘High’ as changes may fall within natural population variability and a recognizable biotope is likely to be present after two years. Biotope sensitivity is therefore assessed as ‘Low’.
A number of the larger bivalve species that may be associated with this biotope group are targeted by commercial fishers in some parts of their range. These include Chamelea gallina (Ballarin et al., 2003); Spisula solida (Fahy et al., 2003; Joaquim et al., 2008); Glycymeris glycymeris and Paphia spp. (Savina & Pouvreau, 2004); Ensis spp., Donax spp. and Pharus spp. (Chícharo et al., 2002). In targeted areas, the populations of fished bivalves may be depleted, for example, fishing has led to declines in Spisula solida (Joaquim et al., 2008; Fahy et al., 2003).
Sensitivity assessment. In general dredges that are used to target bivalves are likely to be efficient at removing targeted species. Removal of commercially targeted bivalves may lead to biotope reclassification based on the dominance of polychaetes to a similar biotope Biotope resistance, based on the characterizing bivalves is assessed as ‘Low’. Undersized juveniles may be returned and can re-burrow but are likely to suffer from stress. Targeted removal of adult bivalves within the biotope may allow successful recruitment of juveniles where intra-specific competition for space and food and possibly consumption of larvae has prevented successful spatfall. Some species such as Glycymeris glycymeris are characteristic of habitats with low levels of competition and may benefit from removal of other species. Biotope resilience is assessed as ‘Medium’, as recruitment in many bivalve species is episodic and unpredictable. Biotope sensitivity is therefore assessed as ‘Medium’.
Species within the biotope are not functionally dependent on each other, although biological interactions will play a role in structuring the biological assemblage through predation and competition. Removal of adults may support recruitment of juvenile bivalves by reducing competition for space and consumption of larvae.
Removal of species would also reduce the ecological services provided by these species such as secondary production and nutrient cycling.
Sensitivity assessment. Species within the biotope are relatively sedentary or slow moving, although the infaunal position may protect some burrowing species from removal. Biotope resistance is therefore assessed as ‘Low’ and resilience as ‘High’, as the habitat is likely to be directly affected by removal and some species will recolonize rapidly, biotiope sensitivity is therefore assessed as Low'. Some variability in species recruitment, abundance and composition is natural and therefore a return to a recognizable biotope should occur within 2 years. Repeated chronic removal would, however, impact recovery.
|Habitats of Principal Importance||Subtidal sands and gravels|
|Habitats of Conservation Importance||Subtidal sands and gravels|
|UK Biodiversity Action Plan Priority||Subtidal sands and gravels|
Aberkali, H.B. & Trueman, E.R., 1985. Effects of environmental stress on marine bivalve molluscs. Advances in Marine Biology, 22, 101-198.
Ansell, A.D., Barnett, P.R.O., Bodoy, A. & Masse, H., 1980. Upper temperature tolerances of some European molluscs. 1. Tellina fabula and T. tenuis. Marine Biology, 58, 33-39.
Arndt, C. & Schiedek, D., 1997. Nephtys hombergii, a free living predator in marine sediments: energy production under environmental stress. Marine Biology, 129, 643-540.
Arntz, W.E. & Rumohr, H., 1986. Fluctuations of benthic macrofauna during succession and in an established community. Meeresforschung, 31, 97-114.
Arntz, W.E., 1980. Predation by demersal fish and its impact on the dynamics of macrobenthos. In Marine benthos dynamics, (eds. K.R. Tenore & B.C. Coull), pp. 121-149. Columbia: University of South Carolina Press.
Ballarin, L., Pampanin, D.M. & Marin, M.G., 2003. Mechanical disturbance affects haemocyte functionality in the Venus clam Chamelea gallina. Comparative Biochemistry and Physiology Part A: Molecular & Integrative Physiology, 136 (3), 631-640.
Barnes, R.S.K. & Hughes, R.N., 1992. An introduction to marine ecology. Oxford: Blackwell Scientific Publications.
Basimi, R.A. & Grove, D.J., 1985. Estimates of daily food intake by an inshore population of Pleuronectes platessa L. off eastern Anglesey, north Wales. Journal of Fish Biology, 27, 505-520.
Beaumont, A.R., Newman, P.B., Mills, D.K., Waldock, M.J., Miller, D. & Waite, M.E., 1989. Sandy-substrate microcosm studies on tributyl tin (TBT) toxicity to marine organisms. Scientia Marina, 53, 737-743.
Bergman, M.J.N. & Hup, M., 1992. Direct effects of beam trawling on macro-fauna in a sandy sediment in the southern North Sea. ICES Journal of Marine Science, 49, 5-11.
Bergman, M.J.N. & Van Santbrink, J.W., 2000b. Fishing mortality of populations of megafauna in sandy sediments. In The effects of fishing on non-target species and habitats (ed. M.J. Kaiser & S.J de Groot), 49-68. Oxford: Blackwell Science.
Bergman, M.J.N. & Van Santbrink, J.W., 2000a. Mortality in megafaunal benthic populations caused by trawl fisheries on the Dutch continental shelf in the North Sea in 1994. ICES Journal of Marine Science, 57 (5), 1321-1331.
Berrilli, F., Ceschia, G., De Liberato, C., Di Cave, D. & Orecchia, P., 2000. Parasitic infections of Chamelea gallina (Mollusca, Bivalvia) from commercially exploited banks of the Adriatic Sea. Bulletin of European Association of Fish Pathologists, 20 (5), 199-205.
Beukema, J.J., 1985. Growth and dynamics in populations of Echinocardium cordatum living in the North Sea off the Dutch corth coast. Netherlands Journal of Sea Research, 19, 129-134.
Bijkerk, R., 1988. Ontsnappen of begraven blijven: de effecten op bodemdieren van een verhoogde sedimentatie als gevolg van baggerwerkzaamheden: literatuuronderzoek: RDD, Aquatic ecosystems.
Blackstock, J. & Barnes, M., 1982. The Loch Eil project: biochemical composition of the polychaete, Glycera alba (Müller), from Loch Eil. Journal of Experimental Marine Biology and Ecology, 57 (1), 85-92.
Blanchard, M., 1997. Spread of the slipper limpet Crepidula fornicata (L.1758) in Europe. Current state and consequences. Scientia Marina, 61, Supplement 9, 109-118.
Bohn, K., Richardson, C.A. & Jenkins, S.R., 2015. The distribution of the invasive non-native gastropod Crepidula fornicata in the Milford Haven Waterway, its northernmost population along the west coast of Britain. Helgoland Marine Research, 69 (4), 313.
Bonsdorff, E., 1984. Establishment, growth and dynamics of a Macoma balthica (L.) population. Limnologica (Berlin), 15, 403-405.
Bosselmann, A., 1989. Larval plankton and recruitment of macrofauna in a subtidal area in the German Bight. In Reproduction, Genetics and Distributions of Marine Organisms (ed. J.S. Ryland & P.A. Tyler), pp. 43-54.
Boyd, S., Limpenny, D., Rees, H. & Cooper, K., 2005. The effects of marine sand and gravel extraction on the macrobenthos at a commercial dredging site (results 6 years post-dredging). ICES Journal of Marine Science: Journal du Conseil, 62 (2), 145-162.
Bradshaw, C., Veale, L.O., Hill, A.S. & Brand, A.R., 2000. The effects of scallop dredging on gravelly seabed communities. In: Effects of fishing on non-target species and habitats (ed. M.J. Kaiser & de S.J. Groot), pp. 83-104. Oxford: Blackwell Science.
Bradshaw, C., Veale, L.O., Hill, A.S. & Brand, A.R., 2002. The role of scallop-dredge disturbance in long-term changes in Irish Sea benthic communities: a re-analysis of an historical dataset. Journal of Sea Research, 47, 161-184.
Brey, T., 1990. Estimating productivity of macrobenthic invertebrates from biomass abd mean individual weight. Meeresforschung, 32,329-343.
Brousseau, D.J., 1978b. Population dynamics of the soft-shell clam Mya arenaria. Marine Biology, 50, 67-71.
Bryan, G.W. & Gibbs, P.E., 1991. Impact of low concentrations of tributyltin (TBT) on marine organisms: a review. In: Metal ecotoxicology: concepts and applications (ed. M.C. Newman & A.W. McIntosh), pp. 323-361. Boston: Lewis Publishers Inc.
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.
Buchanan, J.B. & Moore, J.B., 1986. A broad review of variability and persistence in the Northumberland benthic fauna - 1971-85. Journal of the Marine Biological Association of the United Kingdom, 66, 641-657.
Buchanan, J.B., 1966. The biology of Echinocardium cordatum (Echinodermata: Spatangoidea) from different habitats. Journal of the Marine Biological Association of the United Kingdom, 46, 97-114.
Cabioch, L., Dauvin, J.C. & Gentil, F., 1978. Preliminary observations on pollution of the sea bed and disturbance of sub-littoral communities in northern Brittany by oil from the Amoco Cadiz. Marine Pollution Bulletin, 9, 303-307.
Capasso, E., Jenkins, S., Frost, M. & Hinz, H., 2010. Investigation of benthic community change over a century-wide scale in the western English Channel. Journal of the Marine Biological Association of the United Kingdom, 90 (06), 1161-1172.
Carter, C.G., Grove, D.J. & Carter, D.M., 1991. Trophic resource partitioning between two co-existing flatfish species off the North coast of Anglesey, North Wales. Netherlands Journal of Sea Research, 27, 325-335.
Chícharo, L., Chícharo, M., Gaspar, M., Regala, J. & Alves, F., 2002. Reburial time and indirect mortality of Spisula solida clams caused by dredging. Fisheries Research, 59, 247-257.
Collier, L.M. & Pinn, E.H., 1998. An assessment of the acute impact of the sea lice treatment Ivermectin on a benthic community. Journal of Experimental Marine Biology and Ecology, 230, 131-147.
Conan, G., 1982. The long-term effects of the Amoco Cadiz oil spill. Philosophical Transactions of the Royal Society of London B, 297, 323-333.
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.
Cooper, K., Ware, S., Vanstaen, K. & Barry, J., 2011. Gravel seeding - A suitable technique for restoring the seabed following marine aggregate dredging? Estuarine, Coastal and Shelf Science, 91 (1), 121-132.
Costa, M.J. & Elliot, M., 1991. Fish usage and feeding in two industrialised estuaries - the Tagus, Portugal and the Forth, Scotland. In Estuaries and Coasts: Spatial and Temporal Intercomparisons (ed. B. Knights & A.J. Phillips), pp. 289-297. Denmark: Olsen & Olsen.
Cotter, A.J.R., Walker, P., Coates, P., Cook, W. & Dare, P.J., 1997. Trial of a tractor dredger for cockles in Burry Inlet, South Wales. ICES Journal of Marine Science, 54, 72-83.
Crisp, D.J. (ed.), 1964. The effects of the severe winter of 1962-63 on marine life in Britain. Journal of Animal Ecology, 33, 165-210.
Dame, R.F.D., 1996. Ecology of Marine Bivalves: an Ecosystem Approach. New York: CRC Press Inc. [Marine Science Series.]
Dauvin, J-C. & Gentil, F., 1989. Long-term changes in populations of subtidal bivalves (Abra alba and Abra prismatica) from the Bay of Morlaix (Western English Channel). Marine Biology, 103, 63-73.
Dauvin, J.-C., 1986. Biologie, dynamique et production d'une population d' Abra alba (Wood) (mollusque-bivalve) de la baie de Morlaix (Manche occidentiale). Journal of Experimental Marine Biology and Ecology, 97, 151-180.
Dauvin, J.C., 1985. Dynamics and production of a population of Venus ovata (Pennant) (Mollusca-Bivalvia) of Morlaix Bay (western English Channel). Journal of Experimental Marine Biology and Ecology, 91, 109-123.
Dauvin, J.C., 1998. The fine sand Abra alba community of the Bay of Morlaix twenty years after the Amoco Cadiz oil spill. Marine Pollution Bulletin, 36, 669-676.
Dauvin, J.C., 2000. The muddy fine sand Abra alba - Melinna palmata community of the Bay of Morlaix twenty years after the Amoco Cadiz oil spill. Marine Pollution Bulletin, 40, 528-536.
Davenport, J. & Davenport, J.L., 2005. Effects of shore height, wave exposure and geographical distance on thermal niche width of intertidal fauna. Marine Ecology Progress Series, 292, 41-50.
Davies, C.E. & Moss, D., 1998. European Union Nature Information System (EUNIS) Habitat Classification. Report to European Topic Centre on Nature Conservation from the Institute of Terrestrial Ecology, Monks Wood, Cambridgeshire. [Final draft with further revisions to marine habitats.], Brussels: European Environment Agency.
de Groot, S.J. & Apeldoorn, J., 1971. Some experiments on the influence of the beam trawl on the bottom fauna. International Council for the Exploration of the Sea (CM Papers and Reports) CM 1971/B:2, 5 pp. (mimeo).
De Montaudouin, X. & Sauriau, P.G., 1999. The proliferating Gastropoda Crepidula fornicata may stimulate macrozoobenthic diversity. Journal of the Marine Biological Association of the United Kingdom, 79, 1069-1077.
Desprez, M., 2000. Physical and biological impact of marine aggregate extraction along the French coast of the Eastern English Channel: short- and long-term post-dredging restoration. ICES Journal of Marine Science, 57 (5), 1428-1438.
Diaz-Castaneda, V., Richard, A. & Frontier, S., 1989. Preliminary results on colonization, recovery and succession in a polluted areas of the southern North Sea (Dunkerque's Harbour, France). Scientia Marina, 53, 705-716.
Dittmann, S., Günther, C-P. & Schleier, U., 1999. Recolonization of tidal flats after disturbance. In The Wadden Sea ecosystem: stability, properties and mechanisms (ed. S. Dittmann), pp.175-192. Berlin: Springer-Verlag.
Drake, D.E. & Cacchione, D.A., 1985. Seasonal variation in sediment transport on the Russian River shelf, California. Continental Shelf Research, 4, 495-514.
Duineveld, G.C.A. & Jenness, M.I., 1984. Differences in growth rates of the sea urchin Echinocardium cordatum as estimated by the parameters of the von Bertalanffy equation applied to skeletal rings. Marine Ecology Progress Series, 19, 64-72.
Dyer, K.R., 1998. Estuaries - a Physical Introduction. John Wiley & Son, Chichester.
Eagle, R.A., 1975. Natural fluctuations in a soft bottom benthic community. Journal of the Marine Biological Association of the United Kingdom, 55, 865-878.
Eleftheriou, A. & Robertson, M.R., 1992. The effects of experimental scallop dredging on the fauna and physical environment of a shallow sandy community. Netherlands Journal of Sea Research, 30, 289-299.
Elliott, M., 1994. The analysis of macrobenthic community data. Marine Pollution Bulletin, 28, 62-64.
Emerson, C.W. & Grant, J., 1991. The control of soft-shell clam (Mya arenaria) recruitment on intertidal sandflats by bedload sediment transport. Limnology and Oceanography, 36, 1288-1300.
Emson, R.H., Jones, M. & Whitfield, P., 1989. Habitat and latitude differences in reproductive pattern and life-history in the cosmopolitan brittle-star Amphipholis squamata (Echinodermata). In: Ryland, J.S., Tyler, P.A. (Eds.), Reproduction, Genetics and Distributions of Marine Organisms, pp. 75-81. Olsen & Olsen, Fredensborg.
Eno, N.C., 1991. Marine Conservation Handbook. English Nature, Peterborough.
Essink, K., 1999. Ecological effects of dumping of dredged sediments; options for management. Journal of Coastal Conservation, 5, 69-80.
Fahy, E., Carroll, J. & O'Toole, M., 2003. A preliminary account of fisheries for the surf clam Spisula solida (L) (Mactracea) in Ireland [On-line] http://www.marine.ie, 2004-03-16
Fiege, D., Licher, F. & Mackie, A.S.Y., 2000. A partial review of the European Magelonidae (Annelida : Polychaeta) Magelona mirabilis redefined and M. johnstoni sp. nov. distinguished. Journal of the Marine Biological Association of the United Kingdom, 80, 215-234.
Fish, J.D. & Fish, S., 1996. A student's guide to the seashore. Cambridge: Cambridge University Press.
Folk, R.L., 1954. The distinction between grain size and mineral composition in sedimentary-rock nomenclature. 62, The Journal of Geology, 344-359.
Ford, E., 1923. Animal communities of the level sea-bottom in the water adjacent to Plymouth. Journal of the Marine Biological Association of the United Kingdom, 13, 164-224.
Franzen, N.C.M., 1995. Shear wave detection by Macoma balthica.
Freese, L., Auster, P.J., Heifetz, J. & Wing, B.L., 1999. Effects of trawling on seafloor habitat and associated invertebrate taxa in the Gulf of Alaska. Marine Ecology Progress Series, 182, 119-126.
Fretter, V. & Graham, A., 1981. The Prosobranch Molluscs of Britain and Denmark. Part 6. olluscs of Britain and Denmark. part 6. Journal of Molluscan Studies, Supplement 9, 309-313.
Gaspar, M.B. & Monteiro, C.C., 1999. Gametogenesis and spawning in the subtidal white clam Spisula solida, in relation to temperature. Journal of the Marine Biological Association of the United Kingdom, 79, 753-755.
Gaspar, M.B., Pereira, A.M., Vasconcelos, P. & Monteiro, C.C., 2004. Age and growth of Chamelea gallina from the Algarve coast (southern Portugal): influence of seawater temperature and gametogenic cycle on growth rate. Journal of Molluscan Studies, 70 (4), 371-377.
Gentil, F., Dauvin, J.C. & Menard, F., 1990. Reproductive biology of the polychaete Owenia fusiformis Delle Chiaje in the Bay of Seine (eastern English Channel). Journal of Experimental Marine Biology and Ecology, 142, 13-23.
Gilkinson, K., Paulin, M., Hurley, S. & Schwinghamer, P., 1998. Impacts of trawl door scouring on infaunal bivalves: results of a physical trawl door model/dense sand interaction. Journal of Experimental Marine Biology and Ecology, 224 (2), 291-312.
Gilkinson, K.D., Gordon, D.C., MacIsaac, K.G., McKeown, D.L., Kenchington, E.L., Bourbonnais, C. & Vass, W.P., 2005. Immediate impacts and recovery trajectories of macrofaunal communities following hydraulic clam dredging on Banquereau, eastern Canada. ICES Journal of Marine Science: Journal du Conseil, 62 (5), 925-947.
Glémarec, M., 1973. The benthic communities of the European North Atlantic continental shelf. Oceanography and Marine Biology: an Annual Review, 11, 263-289.
Gosselin, L.A. & Qian, P., 1997. Juvenile mortality in benthic marine invertebrates. Marine Ecology Progress Series, 146, 265-282.
Graham, M., 1955. Effects of trawling on animals on the sea bed. Deep-Sea Research, 3 (Suppl.), 1-6.
Grant, J. & Thorpe, B., 1991. Effects of suspended sediment on growth, respiration, and excretion of the soft shelled clam (Mya arenaria). Canadian Journal of Fisheries and Aquatic Sciences, 48, 1285-1292.
Guenther, C.P., 1991. Settlement of Macoma balthica on an intertidal sandflat in the Wadden Sea. Marine Ecology Progress Series, 76, 73-79.
Guillou, J. & Sauriau, F.G., 1985. Some observations on the biology and ecology of a Venus striatula population in the Bay of Douarnenez, Brittany. Journal of the Marine Biological Association of the United Kingdom, 65, 889-900.
Hall, S.J., 1994. Physical disturbance and marine benthic communities: life in unconsolidated sediments. Oceanography and Marine Biology: an Annual Review, 32, 179-239.
Hauton, C., Hall-Spencer, J.M. & Moore, P.G., 2003. An experimental study of the ecological impacts of hydraulic bivalve dredging on maerl. ICES Journal of Marine Science, 60, 381-392.
Hayward, P., Nelson-Smith, T. & Shields, C. 1996. Collins pocket guide. Sea shore of Britain and northern Europe. London: HarperCollins.
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., Langmead, O. & Warwick, R., 2004. Identification of seabed indicator species from time-series and other studies to support implementation of the EU Habitats and Water Framework Directives. Report to the Joint Nature Conservation Committee and the Environment Agency from the Marine Biological Association. Marine Biological Association of the UK, Plymouth. JNCC Contract F90-01-705. 109 pp.
Hiscock, K., Langmead, O., Warwick, R. & Smith, A., 2005a. Identification of seabed indicator species to support implementation of the EU Habitats and Water Framework Directives. Report to the Joint Nature Conservation Committee and the Environment Agency The Marine Biological Association, Plymouth, 77 pp.
Hjulström, F., 1939. Transportation of detritus by moving water: Part 1. Transportation. Recent Marine Sediments, a Symposium (ed. P.D. Trask), pp. 5-31. Dover Publications, Inc.
Houghton, R.G., Williams, T. & Blacker, R.W., 1971. Some effects of double beam trawling. International Council for the Exploration of the Sea CM 1971/B:5, 12 pp. (mimeo)., International Council for the Exploration of the Sea CM 1971/B:5, 12 pp. (mimeo).
Hughes, D.J., 1998b. Subtidal brittlestar beds. An overview of dynamics and sensitivity characteristics for conservation management of marine SACs. Natura 2000 report prepared for Scottish Association of Marine Science (SAMS) for the UK Marine SACs Project., Scottish Association for Marine Science. (UK Marine SACs Project, Vol. 3). Available from: http://www.ukmarinesac.org.uk/pdfs/britstar.pdf
Hunt, J.D., 1925. The food of the bottom fauna of the Plymouth fishing grounds. Journal of the Marine Biological Association of the United Kingdom, 13, 560-599.
Hylland, K., Sköld, M., Gunnarsson, J.S. & Skei, J., 1996. Interactions between eutrophication and contaminants. IV. Effects on sediment-dwelling organisms. Marine Pollution Bulletin, 33, 90-99.
Jensen, J.N., 1988. Recruitment, growth and mortality of juvenile Corbula gibba and Abra alba in the Limfjord, Denmark. The Baltic Sea environment: history, eutrophication, recruitment and toxicology. Kieler Meeresforschungen (Sonderheft), 6, 357-365.
JNCC, 2015. The Marine Habitat Classification for Britain and Ireland Version 15.03. JNCC: JNCC. 2015(20/05/2015). jncc.defra.gov.uk/MarineHabitatClassification
JNCC (Joint Nature Conservation Committee), 1999. Marine Environment Resource Mapping And Information Database (MERMAID): Marine Nature Conservation Review Survey Database. [on-line] http://www.jncc.gov.uk/mermaid
Joaquim, S., Gaspar, M.B., Matias, D., Ben-Hamadou, R. & Arnold, W.S., 2008. Rebuilding viable spawner patches of the overfished Spisula solida (Mollusca: Bivalvia): a preliminary contribution to fishery sustainability. ICES Journal of Marine Science: Journal du Conseil, 65 (1), 60-64.
Jones, M.L., 1968. On the morphology, feeding and behaviour of Magelona sp. Biological Bulletin of the Marine Laboratory, Woods Hole, 134, 272-297.
Jones, N.S., 1950. Marine bottom communities. Biological Reviews, 25, 283-313.
Jones, N.S., 1952. The bottom fauna and the food of flatfish off the Cumberland Coast. Journal of Animal Ecology, 20, 132-144.
Jorgensen, B.B., 1980. Seasonal oxygen depletion in the bottom waters of a Danish fjord and its effect on the benthic community. Oikos, 32, 68-76.
Künitzer, A., Basford, D., Craeymeersch, J.A., Dewarumez, J.M., Derjes, J., Duinevald, G.C.A., Eleftheriou, A., Heip, C, Herman, P., Kingston, P., Neirmann, U., Rachor, E., Rumohr, H. & Wilde, P.A.J. de, 1992. The benthic infauna of the North Sea: species distribution and assemblages. ICES Journal of Marine Science, 49, 127-143.
Kaiser, M.J. & Spencer, B.E., 1995. Survival of by-catch from a beam trawl. Marine Ecology Progress Series, 126, 31-38.
Kaiser, M.J. & Spencer, B.E., 1996. The effects of beam-trawl disturbance on infaunal communities in different habitats. Journal of Animal Ecology, 65, 348-358.
Kaiser, M.J., & Spencer, B.E., 1994a. A preliminary assessment of the immediate effects of beam trawling on a benthic community in the Irish Sea. In Environmental impact of bottom gears on benthic fauna in relation to natural resources management and protection of the North Sea. (ed. S.J. de Groot & H.J. Lindeboom). NIOZ-Rapport, 11, 87-94.
Kaschl, A. & Carballeira, A., 1999. Behavioural responses of Venerupis decussata (Linnaeus, 1758) and Venerupis pullastra (Montagu, 1803) to copper spiked marine sediments. Boletin. Instituto Espanol de Oceanografia, 15, 383-394.
Kenchington, E.L.R., Prena, J., Gilkinson, K.D., Gordon, D.C., Macisaac, K., Bourbonnais, C.; Schwinghamer, P.J., Rowell, T.W., McKeown, D.L. & Vass, W.P., 2001. Effects of experimental otter trawling on the macrofauna of a sandy bottom ecosystem on the Grand Banks of Newfoundland. Canadian Journal of Fisheries and Aquatic Sciences, 58, 1043-1057.
Kenny, A.J. & Rees, H.L., 1996. The effects of marine gravel extraction on the macrobenthos: results 2 years post-dredging. Marine Pollution Bulletin, 32 (8-9), 615-622.
Kingston, P.F., Dixon, I.M.T., Hamilton, S. & Moore, D.C., 1995. The impact of the Braer oil spill on the macrobenthic infauna of the sediments off the Shetland Islands. Marine Pollution Bulletin, 30 (7), 445-459.
Kinne, O. (ed.), 1983. Diseases of marine animals, volume II. Introduction, Bivalvia to Scaphopoda. Biologische Anstalt Helgoland.
Kinne, O. (ed.), 1984. Marine Ecology: A Comprehensive, Integrated Treatise on Life in Oceans and Coastal Waters.Vol. V. Ocean Management Part 3: Pollution and Protection of the Seas - Radioactive Materials, Heavy Metals and Oil. Chichester: John Wiley & Sons.
Kirby, R.R., Beaugrand, G. & Lindley, J.A., 2008. Climate-induced effects on the meroplankton and the benthic-pelagic ecology of the North Sea. Limnology and Oceanography, 53 (5), 1805.
Kiørboe, T. & Møhlenberg, F., 1981. Particle selection in suspension-feeding bivalves. Marine Ecology Progress Series, 5, 291-296.
Klawe, W.L. & Dickie, L.M., 1957. Biology of the bloodworm, Glycera dibranchiata Ehlers, and its relation to the bloodworm fishery of the Maritime Provinces. Bulletin of Fisheries Research Board of Canada, 115, 1-37.
Kröncke, I., Dippner, J., Heyen, H. & Zeiss, B., 1998. Long-term changes in macrofaunal communities off Norderney (East Frisia, Germany) in relation to climate variability. Marine Ecology Progress Series, 167, 25-36.
Krebs, C.J., 1978. Ecology: the experimental analysis of distribution and abundance (2nd edition). New York: Harper & Row, 678pp.
Lawrence, J.M., 1996. Mass mortality of echinoderms from abiotic factors. In Echinoderm Studies Vol. 5 (ed. M. Jangoux & J.M. Lawrence), pp. 103-137. Rotterdam: A.A. Balkema.
Levell, D., Rostron, D. & Dixon, I.M.T., 1989. Sediment macrobenthic communities from oil ports to offshore oilfields. In Ecological Impacts of the Oil Industry, Ed. B. Dicks. Chicester: John Wiley & Sons Ltd.
Lin, J. & Hines, A.H., 1994. Effects of suspended food availability on the feeding mode and burial depth of the Baltic clam, Macoma balthica. Oikos, 69, 28-36.
Lockwood, S.J., 1980. The daily food intake of O-group plaice (Pleuronectes platessa L.) under natural conditions. Journal du Conseil Permanent International pour l'Exploration de la Mer, 39, 154-159.
Long, D., 2006. BGS detailed explanation of seabed sediment modified Folk classification. Available from: http://www.emodnet-seabedhabitats.eu/PDF/GMHM3_Detailed_explanation_of_seabed_sediment_classification.pdf
Lopez, G.R. & Levinton, J.S., 1987. Ecology of deposit-feeding animals in marine sediments. Quarterly Review of Biology, 62, 235-260.
Lopez-Flores I., De la Herran, R., Garrido-Ramos, M.A., Navas, J.I., Ruiz-Rejon, C. & Ruiz-Rejon, M., 2004. The molecular diagnosis of Marteilia refringens and differentiation between Marteilia strains infecting oysters and mussels based on the rDNA IGS sequence. Parasitology, 19 (4), 411-419.
Lopez-Jamar, E., Francesch, O., Dorrio, A.V. & Parra, S., 1995. Long term variation of the infaunal benthos of La Coruna Bay (NW Spain): results from a 12-year study (1982-1993). Scientia Marina, 59(suppl. 1), 49-61.
Møhlenberg, F. & Kiørboe, T., 1983. Burrowing and avoidance behaviour in marine organisms exposed to pesticide contaminated sediments. Marine Pollution Bulletin, 14, 57-60.
Macer, C.T., 1967. The food web in Red Wharf Bay (North Wales) with particular reference to young plaice (Pleuronectes platessa). Helgolander Wissenschaftliche Meeresuntersuchungen, 15, 560-573.
Mackie, A.S.Y., James, J.W.C., Rees, E.I.S., Darbyshire, T., Philpott, S.L., Mortimer, K., Jenkins, G.O. & Morando, A., 2006. Biomôr 4. The outer Bristol Channel marine habitat study. , Cardiff: National Museum Wales.
Mackie, A.S.Y., Oliver, P.G. & Rees, E.I.S., 1995. Benthic biodiversity in the southern Irish Sea. Studies in Marine Biodiversity and Systematics from the National Museum of Wales. BIOMOR Reports, no. 1.
Maurer, D., Keck, R.T., Tinsman, J.C., Leatham, W.A., Wethe, C., Lord, C. & Church, T.M., 1986. Vertical migration and mortality of marine benthos in dredged material: a synthesis. Internationale Revue der Gesamten Hydrobiologie, 71, 49-63.
McDermott, J.J., 1984. The feeding biology of Nipponnemertes pulcher (Johnston) (Hoplonemertea), with some ecological implications. Ophelia, 23, 1-21.
McIntyre, A.D., 1958. The ecology of Scottish inshore fishing grounds. 1. The bottom fauna of east coast grounds. Marine Research, 1, 1-24.
Meador, J.P., Varanasi, U. & Krone, C.A., 1993. Differential sensitivity of marine infaunal amphipods to tributyltin. Marine Biology, 116, 231-239.
Marine Ecological Surveys Limited (MES), 2008. Marine Macrofauna Genus Trait Handbook. Marine Ecological Surveys Limited: Bath.
MES, 2010. Marine Macrofauna Genus Trait Handbook. Marine Ecological Surveys Limited. http://www.genustraithandbook.org.uk/
Mills, E.L., 1967. The biology of an ampeliscid amphipod crustacean sibling species pair. Journal of the Fisheries Research Board of Canada, 24, 305-355.
Molander, A.R., 1962. Studies on the fauna in the fjords of Bohsula with reference to the distribution of different associations. Arkiv fur Zoologi, Series 2, 15, 1-64.
Morton, B., 2009. Aspects of the biology and functional morphology of Timoclea ovata (Bivalvia: Veneroidea: Venerinae) in the Azores, Portugal, and a comparison with Chione elevata (Chioninae). Açoreana, 6, 105-119.
Moulaert, I. & Hostens, K., 2007. Post-extraction evolution of a macrobenthic community on the intensively extracted Kwintebank site in the Belgian part of the North Sea. CM Documents-ICES, (A:12).
Navarro, J.M. & Widdows, J., 1997. Feeding physiology of Cerastoderma edule in response to a wide range of seston concentrations. Marine Ecology Progress Series, 152, 175-186.
Nichols, F.H., 1977. Dynamics and production of Pectinaria koreni (Malmgren) in Kiel Bay, West Germany. In Biology of benthic organisms, (eds. B.F. Keegan, P. O'Ceidigh & P.J.S. Boaden), pp. 453-463.
Nicolaidou, A., 1983. Life history and productivity of Pectinaria koreni Malmgren (Polychaeta). Estuarine, Coastal and Shelf Science, 17, 31-43.
Niermann, U., 1996. Fluctuation and mass occurrence of Phoronis muelleri (Phoronidea) in the south-eastern North Sea during 1983-1988. Senckenbergiana Maritima, 28, 65-79.
Niermann, U., 1997. Macrobenthos of the south-eastern North Sea during 1983-1988. Berichte der Biologischen Anstalt Helgoland, 13, 144pp.
Niermann, U., Bauerfeind, E., Hickel, W. & Westernhagen, H.V., 1990. The recovery of benthos following the impact of low oxygen content in the German Bight. Netherlands Journal of Sea Research, 25, 215-226.
Nilsson, H.C. & Rosenberg, R., 1994. Hypoxic response of two marine benthic communities. Marine Ecology Progress Series, 115, 209-217.
Olafsson, E.B. & Persson, L.E., 1986. The interaction between Nereis diversicolor (Muller) and Corophium volutator (Pallas) as a structuring force in a shallow brackish sediment. Journal of Experimental Marine Biology and Ecology, 103, 103-117.
Olafsson, E.B., Peterson, C.H. & Ambrose, W.G. Jr., 1994. Does recruitment limitation structure populations and communities of macro-invertebrates in marine soft sediments: the relative significance of pre- and post-settlement processes. Oceanography and Marine Biology: an Annual Review, 32, 65-109
Oyenekan, J.A., 1986. Population dynamics and secondary production of Nephtys hombergii (Polychaeta: Nephtyidae). Marine Biology, 93, 217-223.
Pearson, T.H. & Rosenberg, R., 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine environment. Oceanography and Marine Biology: an Annual Review, 16, 229-311.
Petersen, C.G.J., 1918. The sea bottom and its production of fish food. A survey of the work done in connection with valuation of the Denmark waters from 1883-1917. Report of the Danish Biological Station, 25, 1-62.
Pianka, E.R., 1970. On r- and K- selection. American Naturalist, 104, 592-597.
Picton, B.E. & Costello, M.J., 1998. BioMar biotope viewer: a guide to marine habitats, fauna and flora of Britain and Ireland. [CD-ROM] Environmental Sciences Unit, Trinity College, Dublin., http://www.itsligo.ie/biomar/
Poggiale, J.C. & Dauvin, J.C., 2001. Long term dynamics of three benthic Ampelisca (Crustacea - Amphipoda) populations from the Bay of Morlaix (western English Channel) related to their disappearance after the Amoco Cadiz oil spill. Marine Ecology Progress Series, 214, 201-209.
Powilleit, M., Graf, G., Kleine, J., Riethmuller, R., Stockmann, K., Wetzel, M.A. & Koop, J.H.E., 2009. Experiments on the survival of six brackish macro-invertebrates from the Baltic Sea after dredged spoil coverage and its implications for the field. Journal of Marine Systems, 75 (3-4), 441-451.
Price, H., 1982. An analysis of factors determining seasonal variation in the byssal attachment strength of Mytilus edulis. Journal of the Marine Biological Association of the United Kingdom, 62 (01), 147-155
Rabalais, N.N., Harper, D.E. & Turner, R.E., 2001. Responses of nekton and demersal and benthic fauna to decreasing oxygen concentrations. In: Coastal Hypoxia Consequences for Living Resources and Ecosystems, (Edited by: Rabalais, N. N. and Turner, R. E.), Coastal and Estuarine Studies 58, American Geophysical Union, pp. 115–128. Washington D.C.
Rainer, S.F., 1985. Population dynamics and production of the bivalve Abra alba and implications for fisheries production. Marine Biology, 85, 253-262.
Rauck, G., 1988. What influence have bottom trawls on the seafloor and bottom fauna? Informationen fur die Fischwirtschaft, Hamberg, 35, 104-106.
Rees, E.I.S. & Walker, A.J.M., 1983. Annual and spatial variation in the Abra community in Liverpool Bay. Oceanologica Acta, Special issue (0399-1784), 165-169.
Rees, E.I.S., Nicholaidou, A. & Laskaridou, P., 1977. The effects of storms on the dynamics of shallow water benthic associations. In Proceedings of the 11th European Symposium on Marine Biology, Galway, Ireland, October 5-11, 1976. Biology of Benthic Organisms, (ed. B.F. Keegan, P. O'Ceidigh & P.J.S. Boaden), pp. 465-474.
Rees, H.L. & Dare, P.J., 1993. Sources of mortality and associated life-cycle traits of selected benthic species: a review. MAFF Fisheries Research Data Report, no. 33., Lowestoft: MAFF Directorate of Fisheries Research.
Rhoads, D.C. & Young, D.K., 1970. The influence of deposit-feeding organisms on sediment stability and community trophic structure. Journal of Marine Research, 28, 150-178.
Riedel, B., Zuschin, M. & Stachowitsch, M., 2012. Tolerance of benthic macrofauna to hypoxia and anoxia in shallow coastal seas: a realistic scenario. Marine Ecology Progress Series, 458, 39-52.
Riera, R., Tuya, F., Ramos, E., Rodríguez, M. & Monterroso, Ó., 2012. Variability of macrofaunal assemblages on the surroundings of a brine disposal. Desalination, 291, 94-100.
Robinson, R.F. & Richardson, C.A., 1998. The direct and indirect effects of suction dredging on a razor clam (Ensis arcuatus) population. ICES Journal of Marine Science, 55, 970-977.
Rosenberg, R. & Loo, L., 1988. Marine eutrophication induced oxygen deficiency: effects on soft bottom fauna, western Sweden. Ophelia, 29, 213-225.
Rowden, A.A., Jones, M.B. & Morris, A.W., 1998. The role of Callianassa subterranea (Montagu) (Thalassinidea) in sediment resuspension in the North Sea. Continental Shelf Research, 18, 1365-1380.
Rygg, B., 1985. Effect of sediment copper on benthic fauna. Marine Ecology Progress Series, 25, 83-89.
Salzwedel, H., 1979. Reproduction, growth, mortality and variations in abundance and biomass of Tellina fabula (Bivalvia) in the German Bight in 1975/1976. Veroffentlichungen des Instituts fur Meeresforschung in Bremerhaven, 18, 111-202.
Salzwedel, H., Rachor, E. & Gerdes, D., 1985. Benthic macrofauna communities in the German Bight. Verifflithungen des Institut fur Meeresforschung in Bremerhaven, 20, 199-267.
Sanders, H.L., 1978. Florida oil spill impact on the Buzzards Bay benthic fauna: West Falmouth. Journal of the Fisheries Board of Canada, 35, 717-730.
Sardá, R., Pinedo, S. & Martin, D., 1999. Seasonal dynamics of macroinfaunal key species inhabiting shallow soft-bottoms in the Bay of Blanes (NW Mediterranean). Publications Elsevier: Paris.
Sardá, R., Pinedo, S., Gremare, A. & Taboada, S., 2000. Changes in the dynamics of shallow sandy-bottom assemblages due to sand extraction in the Catalan Western Mediterranean Sea. ICES Journal of Marine Science, 57 (5), 1446-1453.
Savina, M. & Pouvreau, S., 2004. A comparative ecophysiological study of two infaunal filter-feeding bivalves: Paphia rhomboıdes and Glycymeris glycymeris. Aquaculture, 239 (1), 289-306.
Schäfer, H., 1972. Ecology and palaeoecology of marine environments, 568 pp. Chicago: University of Chicago Press.
Sinderman, C.J., 1990. Principle diseases of marine fish and shellfish, 2nd edition, Volume 2. Diseases of marine shellfish. Academic Press, 521 pp.
Smith, J.E. (ed.), 1968. 'Torrey Canyon'. Pollution and marine life. Cambridge: Cambridge University Press.
Snelgrove, P.V., Grassle, J.P., Grassle, J.F., Petrecca, R.F. & Ma, H., 1999. In situ habitat selection by settling larvae of marine soft‐sediment invertebrates. Limnology and Oceanography, 44 (5), 1341-1347.
Soemodinoto, A., Oey, B.L. & Ibkar-Kramadibrata, H., 1995. Effect of salinity decline on macrozoobenthos community of Cibeurum River estuary, Java, Indonesia. Indian Journal of Marine Sciences, 24, 62-67.
Sörlin, T., 1988. Floating behaviour in the tellinid bivalve Macoma balthica (L.). Oecologia, 77, 273-277.
Stamouli, M. & Papadopoulou, C., 1990. Trivalent Cr-51 bioaccumulation study in two mollusc species. Thalassographica. Athens, 13 suppl. 1, 49-52.
Stenton-Dozey, J.M.E. & Brown, A.C., 1994. Short term changes in the energy balance of Venerupis corrugatus (Bivalvia) in relation to tidal availability of natural suspended particles. Marine Ecology Progress Series, 103, 57-64.
Stickle, W.B. & Diehl, W.J., 1987. Effects of salinity on echinoderms. In Echinoderm Studies, Vol. 2 (ed. M. Jangoux & J.M. Lawrence), pp. 235-285. A.A. Balkema: Rotterdam.
Stirling, E.A., 1975. Some effects of pollutants on the behaviour of the bivalve Tellina tenuis. Marine Pollution Bulletin, 6, 122-124.
Strasser, M., 1999. Mya arenaria - an ancient invader of the North Sea coast. Helgoländer Meeresuntersuchungen, 52, 309-324.
Stronkhorst, J., Hattum van, B. & Bowmer, T., 1999. Bioaccumulation and toxicity of tributyltin to a burrowing heart urchin and an amphipod in spiked, silty marine sediments. Environmental Toxicology and Chemistry, 18, 2343-2351.
Suchanek, T.H., 1993. Oil impacts on marine invertebrate populations and communities. American Zoologist, 33, 510-523.
Tait, R.V. & Dipper, R.A., 1998. Elements of Marine Ecology. Reed Elsevier.
Thomas, R., 1975. Functional morphology, ecology, and evolutionary conservatism in the Glycymerididae (Bivalvia). Palaeontology, 18 (2), 217-254.
Thomsen, L. & Altenbach, A.V., 1993. Vertical and areal distribution of foraminiferal abundance and biomass in microhabitats around inhabited tubes of marine echiurids. Marine Micropaleontology, 20, 303-309.
Thorson, G., 1957. Bottom communities (sublittoral or shallow shelf). Memoirs of the Geological Society of America, 67, 461-534.
Thouzeau, G., Jean, F. & Del Amo, Y., 1996. Sedimenting phytoplankton as a major food source for suspension-feeding queen scallops (Aequipecten opercularis L.) off Roscoff (western English Channel) ? Journal of Shellfish Research, 15, 504-505.
Thrush, S.F., 1986. Community structure on the floor of a sea-lough: are large epibenthic predators important? Journal of Experimental Marine Biology and Ecology, 104, 171-183.
Tillin, H.M., Hiddink, J.G., Jennings, S. & Kaiser, M.J., 2006. Chronic bottom trawling alters the functional composition of benthic invertebrate communities on a sea-basin scale. Marine Ecology Progress Series, 318, 31-45.
Tuck, I.D., Atkinson, R.J.A. & Chapman, C.J., 2000. Population biology of the Norway lobster Nephrops norvegicus (L.) in the Firth of Clyde, Scotland II: Fecundity and size at onset of sexual maturity. ICES Journal of Marine Science, 57, 1227-1239.
Tyler, P.A., 1977b. Sublittoral community structure of Oxwich Bay, South Wales, in relation to sedimentological, physical oceanographical and biological parameters. In Biology of Benthic Organisms: 11th European Symposium on Marine Biology, Galway, October 1976 (ed. B.F. Keegan, P.O. Ceidigh and P.J.S. Boaden), 559-566. Pergamon Press, Oxford.
UKTAG, 2014. UK Technical Advisory Group on the Water Framework Directive [online]. Available from: http://www.wfduk.org
Valentine, P.C., Carman, M.R., Blackwood, D.S. & Heffron, E.J., 2007. Ecological observations on the colonial ascidian Didemnum sp. in a New England tide pool habitat. Journal of Experimental Marine Biology and Ecology, 342 (1), 109-121.
Van Dalfsen, J.A., Essink, K., Toxvig Madsen, H., Birklund, J., Romero, J. & Manzanera, M., 2000. Differential response of macrozoobenthos to marine sand extraction in the North Sea and the Western Mediterranean. ICES Journal of Marine Science, 57 (5), 1439-1445.
Warwick, R.M. & Davis, J.R., 1977. The distribution of sublittoral macrofauna communities in the Bristol Channel in relation to the substrate. Estuarine and Coastal Marine Science, 5, 267-288.
Warwick, R.M. & George, C.L., 1980. Annual macro-fauna production in an Abra community. In Industrialised embayments and their environmental problems: a case study of Swansea Bay (ed. M.B. Collins et al.), pp. 517-538. Oxford: Pergamon Press.
Warwick, R.M. & Uncles, R.J., 1980. Distribution of benthic macrofauna associations in the Bristol Channel in relation to tidal stress. Marine Biology Progress Series, 3, 97-103.
Warwick, R.M., George, C.L. & Davies, J.R., 1978. Annual macrofauna production in a Venus community. Estuarine and Coastal Marine Science, 7, 215-241.
Weigelt, M. & Rumohr, H., 1986. Effects of wide range oxygen depletion on benthic fauna and demersal fish in Kiel Bay. Meeresforschung, 31, 124-136.
Widdicombe, S. & Austen, M.C., 1999. Mesocosm investigation into the effects of bioturbation on the diversity and structure of a subtidal macrobenthic community. Marine Ecology Progress Series, 189, 181-193.
Widdows, J., Bayne, B.L., Livingstone, D.R., Newell, R.I.E. & Donkin, P., 1979. Physiological and biochemical responses of bivalve molluscs to exposure to air. Comparative Biochemistry and Physiology, 62A, 301-308.
Wilson, J.G., 1978. Upper temperature tolerances of Tellina tenuis and T. fabula. Marine Biology, 45, 123-128.
Wood, E. (ed.), 1988. Sea Life of Britain and Ireland. Marine Conservation Society. IMMEL Publishing, London
Wood, E.M., 1987. Subtidal Ecology. London: Edward Arnold.
Ziegelmeier, E., 1978. Macrobenthos investigations in the eastern part of the German Bight from 1950 to 1974. Rapports et Proces-verbaux des Reunions. Commission Internationale pour l'Exploration Scientifique de la Mer Mediterranee. Paris, 172, 432-444.
This review can be cited as:
Last Updated: 01/07/2016