The Marine Life Information Network

Temperature increase (local)

Benchmark. A 5°C increase in temperature for one month, or 2°C for one year. Further detail

Evidence

This biotope occurs intertidally and is therefore likely to be relatively tolerant of rapid changes in temperature as experienced during cyclical periods of immersion and emersion. Examples of distribution and thermal tolerances tested in laboratory experiments are provided as evidence to support the sensitivity assessment. In general, populations can acclimate to prevailing conditions which can alter tolerance thresholds and care should, therefore, be used when interpreting reported tolerances.

The key characterizing species Cerastoderma edule is found from Norway to Mauritania (Honkoop et al., 2008) and through the Baltic, Mediterranean and Black Sea (Longshaw & Malham, 2013). The eastern border of distribution is the Murmansk coast of the Barents sea (Genelt-Yanovskiy et al., 2010). The species is therefore likely to be exposed to warmer and colder water sand air temperatures than experienced in the UK over its geographic range.

Kristensen (1958) reported that Cerastoderma edule from the Dutch Wadden Sea have an upper temperature tolerance of 31°C for 24 hrs, but that spat (3-6 mm) were more tolerant. All cockles died after 6 min at 36°C. Ansell et al. (1981) reported an upper median lethal temperature of 35°C after 24hrs (29°C after 96 hrs exposure), and Wilson (1981) reported an upper lethal temperature of 42.5°C. These temperatures are likely to exceed the pressure benchmark.

Wilson (1981) noted that Cerastoderma edule had limited ability to acclimate and Smaal et al. (1997) stated that Cerastoderma edule is unable to acclimate to low temperatures. However, Newell & Bayne (1980) stated that Cerastoderma edule was able to acclimate to a temperature change of 10°C and regulate its metabolic rate in response to rising spring temperatures. Temperature tolerance in the above studies was dependant on the environmental temperature, i.e. specimens collected in summer or areas of higher average temperature tolerated higher temperatures than specimens collected in winter and/or at lower average temperatures.

Kingston (1974) reared artificially fertilized Cerastoderma edule (as Cardium edule) in the laboratory in the temperature range 10-20oC (fertilization did not occur at 5oC). Larval growth was ‘poor’ at 10oC, optimal between 15 and 20oC and most larvae grew poorly and died before metamorphosing at 30oC. No larval growth occurred at 35 °C and all larvae held at this temperature were dead within 4 days of the start of the experiment.  Honkoop and Van Der Meer (1998) found that winter temperatures influenced egg production by Cerastoderma edule, individuals kept in warmer waters produced smaller eggs.

Wilson (1993) concluded that Cerastoderma edule was probably tolerant of a long-term temperature rise of 2°C associated with climate change. Warmer temperatures during winter result in increased metabolic rate and hence depletion of energy reserves in a time of low food availability and may contribute to post winter mortality of adult cockles (Wilson & Elkaim, 1991). Therefore, the tolerance of Cerastoderma edule to temperature change will be dependent on season, an acute, short-term temperature rise in summer or decrease in winter may be detrimental.

Rapid increases in temperature during the spawning season may initiate spawning (Ducrotoy et al. 1991). High shore populations are likely to be more vulnerable to extremes of temperatures due to their longer emergence time (see emergence). However, Wilson (1981) showed that Cerastoderma edule had the highest upper lethal temperature of the species he studied, presumably due to acclimation from its close contact with the sediment surface. The upper lethal temperature of 42.8°C is unlikely to occur on mudflats except in extremely hot summers.

Changes in temperature may also lead to indirect ecological consequences. Experiments demonstrated that predation on Cerastoderma edule by shore crabs (Carcinus maenas) increases as temperature increases (Sanchez-Salazar et al., 1987a) Experiments were run at 6.0, 9.5, and 15.5 ^oC, representing the annual range of sea surface temperatures within the Menai Strait (north Wales).  (Sanchez-Salazar et al., 1987a). Mild winters that enhance predator survival are likely to result in increased predation of spat the following spring (Bukema & Dekker, 2005),

The cirratulid Aphelochaeta marioni (studied as Tharyx marioni) has been recorded from the Baltic to the Indian Ocean and so it probably has some degree of adaptation or tolerance to a range of temperatures (Hartmann-Schroder, 1974 and Rogall, 1977, cited in Farke, 1979).  However, acute rises in temperature may have a more deleterious effect. George (1964a) reported that a rapid rise or fall in temperature of 3 °C was sufficient to induce spawning in 25% of mature Cirriformia tentaculata. If this occurred at a time of year that was not suitable for larval survival then larval mortality could be high.

The upper lethal limits for Cirriformia tentaculata from the Hamble were reported to be of 32 °C and 29 °C for 5-6 day old and adult Cirriformia tentaculata respectively (George, 1964b). The upper temperature tolerance (that killed half of the test organisms after 96 hours) of the oligochaete Tubificoides benedii (studied as Peloscolex benedeni) was reported to be 28.5 °C (Diaz, 1980). However, temperatures of this magnitude are unlikely to be experienced by this intertidal biotope.

Cirriformia tentaculata is reported to be near its northern limit in the British Isles (George, 1968) and an increase in temperature may lead to the extension of its upper distribution range. An increase in temperature could also serve to decrease the length of time spent in the larval phase and so reduce the risk of predation. The rate of larval growth in Cirriformia tentaculata was found to be twice as fast at 20 °C than at 8 °C.

Capitella capitata were a dominant species in mud sediments receiving effluents that were typically 8-12^oC warmer than the receiving waters (Bamber & Spencer, 1984) and are considered to be tolerant to this pressure at the benchmark. Eteone longa and Pygospio elegans were summer visitors to the same effluent exposed habitats and these three species are considered tolerant of acute and chronic increases in temperature.

Sensitivity assessment. Typical surface water temperatures around the UK coast vary, seasonally from 4-19^oC (Huthnance, 2010). The characterizing and associated species are considered likely to be tolerant of acute and chronic increases in temperature at the pressure benchmark.  Cerastoderma edule has a wide geographic range and as experiments suggest that individuals can survive sudden increases in temperature. As an intertidal species, with some populations occurring above mid-shore, Cerastoderma edule experiences rapid fluctuations in temperature over the tidal cycle. The lack of evidence for mass mortalities in very hot summers (compared with reports for low winter temperatures suggest that this species is likely to tolerate a chronic increase at the pressure benchmark (2^oC for one year). An acute increase in temperature for one month may lead to changes in reproductive success and predation, particularly on spat and juveniles. Adults may, however survive. Biotope resistance is therefore assessed as ‘High’ and residence is ‘High’ (by default), the biotope is therefore considered to be ‘Not sensitive’.

Temperature decrease (local)

Benchmark. A 5°C decrease in temperature for one month, or 2°C for one year. Further detail

Evidence

This biotope occurs intertidally and is therefore likely to be relatively tolerant of rapid changes in temperature as experienced during cyclical periods of immersion and emersion. Examples of distribution and thermal tolerances tested in laboratory experiments are provided as evidence to support the sensitivity assessment. In general, populations can acclimate to prevailing conditions which can alter tolerance thresholds and care should, therefore, be used when interpreting reported tolerances.

The key characterizing species Cerastoderma edule is found from Norway to Mauritania (Honkoop et al., 2008) and through the Baltic, Mediterranean and Black Sea (Longshaw & Malham, 2013). The eastern border of distribution is the Murmansk coast of the Barents sea (Genelt-Yanovskiy et al., 2010).  Populations at the Barents sea may experience annual water temperatures from 3-8^oC but are exposed to air temperatures of -30oC. These populations are present between mid to low shore and although acclimated to lower temperatures are present in low densities compared to more central parts of the range (Genelt-Yanovskiy et al., 2010). The low densities may be due to thermal tolerances or restrictions on feeding and growth or other factors.

High mortalities of cockle populations due to severe winters have been reported by many authors (Kristensen, 1958; Hancock & Urquhart, 1964; Beukema, 1979, 1985, 1990; Ducrotoy et al., 1991, Strasser et al., 2001). Kristensen (1957) showed a direct influence of temperatures below about -2 ° C on cockle survival (cited from Beukema, 1979). Kristensen (1958) reported that the sediment froze to a depth of 10 cm and 15 cm, resulting in death of cockles in areas of the Wadden Sea in the severe winter of 1954. Hancock & Urquhart (1964) report almost 100% mortality of cockles in Llanrhidian Sands, Burry Inlet and high mortalities of cockles in other areas around the UK after the winter of 1962/63.  However, enhanced recruitment of bivalves, including Cerastoderma edule and Macoma balthica has been observed in  European estuaries after colder winters while densities following milder winters are lower (Beukema 1991, Walker & Dare, 1993, Young et al., 1996). The factors indirectly responsible for this pattern may be changes in reproductive success (Honkoop & Van Der Meer, 1998), changes in the spring phytoplankton bloom, predation (Beukema & Dekker, 2005), removal of larvae by off-shore currents and removal of adults (enhancing recruitment via reduced ingestion of larvae (André et al.,1993) and reduced competition between adults and juveniles).

Aphelochaeta marioni (studied as Tharyx marioni) has been recorded from the Baltic to the Indian Ocean and so it probably has some degree of adaptation or tolerance to a range of temperatures (Hartmann-Schroder, 1974 and Rogall, 1977, cited in Farke, 1979). Short periods of severe frost in November 1973 were not reported to have affected the population of Aphelochaeta marioni (studied as Tharyx marioni) in the German Bight (Farke, 1979).

Acute falls in temperature may have a more deleterious effect. George (1964a) reported that a rapid rise or fall in temperature of 3 °C was sufficient to induce spawning in 25% of mature Cirriformia tentaculata. If this occurred at a time of year that was not suitable for larval survival then larval mortality could be high. However, George (1964b) noted that although in Southampton the incoming tide incurred a drop of 6 °C in five minutes, such rapid changes in temperature had no significant effect on the mortality of either juvenile of adult Cirriformia tentaculata in the laboratory. The larvae of this species grow twice as slow at 8 °C than they do at 20 °C (George, 1964). Any increase in the length of time spent in the larval phase will increase the risk of predation. In adults, field data suggests that growth ceases at 6 °C (George, 1964). On the Hamble, lower lethal limits of -6 °C (by extrapolation) and 2 °C have been reported for 5-6 day old and adult Cirriformia tentaculata respectively (George, 1964b). These are temperatures that can reasonably be expected in winter in this intertidal biotope and so some mortality is likely. Furthermore, Cirriformia tentaculata is reported to be near its northern limit in the British Isles (George, 1968) and a long-term chronic decrease in temperature could serve to exclude this species from the northern extent of its distribution. George (1968) reported several major changes and a major reduction in the distribution range of Cirriformia tentaculata following the severe winter of 1962/3. In temperature tolerance experiments, no Cirriformia tentaculata survived even a brief exposure to -2 °C or 96 hours at 0 °C.

The cirratulid Cirratulus cirratus was found to be tolerant to lower temperatures and it is possible that this species will become more prevalent in this biotope if the temperature falls. George (1968) reported that the ciliary feeding mechanisms of Cirriformia tentaculata became so inefficient at low temperatures that, over long periods, the animal may die of starvation. George (1968) also mentioned that the animal does not withdraw its branchiae in cold weather. Due to their delicate nature, the branchiae may subsequently freeze on the surface. In such a case, the animal would be living under anaerobic conditions and so emerges from the burrow to enable them to respire through their body surface. This emergence would increase both risk of predation and of freezing.

Sensitivity assessment. Typical surface water temperatures around the UK coast vary, seasonally from 4-19^oC (Huthnance, 2010). The biotope is considered to tolerate a chronic change at the pressure benchmark (2^oC decrease in temperature for a year). An acute reduction in temperature may be tolerated by adults and spat outside of winter (although acclimation to warmer temperatures means that impacts on spawning and growth may occur). An acute reduction in temperature during winter may exceed thermal tolerances, biotope resistance (based on Cerastoderma edule evidence) is therefore assessed as ‘Low’ and resistance is assessed as ‘Medium’. Biotope sensitivity is therefore judged to be ‘Medium’, this precautionary assessment is presented in the table.

Salinity increase (local)

Benchmark. A increase in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

This biotope is reported to occur in variable (18-35 ppt) salinity (JNCC, 2015). A change at the pressure benchmark is considered to represent an increase to full salinity.

Boyden & Russell (1972) stated that Cerastoderma edule prefers salinities between 15 and 35 psu. Russell & Peterson (1973) reported lower median salinity limits of 12.5 psu and upper median salinity limits of 38.5 psu. Rygg (1970) noted that Cerastoderma edule did not survive 23 days exposure to <10 psu or at 60 psu, although they did survive at 46 psu. Rygg (1970) also demonstrated that salinity tolerance was temperature dependant (after 3 days, 100% survival at 33 psu and 35-38°C, but 50% mortality occurred at 20 psu and 37°C and 100% mortality at 13 psu and 37°C). Wilson (1984) noted that Cerastoderma edule remained open during 1 hour exposure to salinities between 13.3 and 59.3 psu. It should be noted that the tolerances reported above depend on the duration of the experiment.

Kingston (1974) found that Cerastoderma edule larvae grew optimally at 30 and 35 psu, and grew well at 40 psu but the growth increment declined at 45 psu and larvae did not metamorphose. He noted that Cerastoderma edule larvae survived between 20 -50 psu, but died after 11 days at 55 psu or 10 days at 10 psu.

Populations of Aphelochaeta marioni inhabit the open coast where seawater is at full salinity. They are clearly capable of thriving in fully saline conditions and hence probably relatively tolerant of increases in salinity. No information was found concerning the reaction to hypersaline conditions (>40psu). Farke (1979) studied the effects of changing salinity on Aphelochaeta marioni (studied as Tharyx marioni) in a microsystem in the laboratory. Over several weeks, the salinity in the microsystem was increased from 25-40 psu and no adverse reaction was noted. However, when individuals were removed from the sediment and displaced to a new habitat, they only dug into their new substratum if the salinities in the two habitats were similar. If the salinities differed by 3-5 psu, the worms carried out random digging movements, failed to penetrate the sediment and died at the substratum surface after a few hours. This would suggest that Aphelochaeta marioni can tolerate salinity changes when living infaunally but is far more intolerant when removed from its habitat.

Sensitivity assessment.  Little evidence was found to assess this pressure at the benchmark. Although species within the biotope are likely to tolerate short-term increases in salinity in sediment surface layers between tidal cycles a longer change is likely to exceed salinity tolerances of adult Cerastoderma edule and larvae. Biotope resistance is assessed as ‘Low’ as the results of Rygg (1970) suggest some adults may survive and acclimate. Biotope resilience (following a return to suitable habitat conditions) is assessed as ‘Medium’ and sensitivity is assessed as ‘Medium’.

Salinity decrease (local)

Benchmark. A decrease in one MNCR salinity category above the usual range of the biotope or habitat. Further detail

Evidence

This biotope is reported to occur in full (30-35 ppt) salinity (JNCC, 2015). As the characterizing species Cerastoderma edule is found in biotopes in both full and variable salinity habitats, the biotope is considered ‘Not sensitive’ to a decrease in salinity from full to reduced or variable.

The available studies indicate that Cerastoderma edule larvae and adults show a wide tolerance range of salinity for both adults and larvae, in accordance with the intertidal distribution.  Kristensen (1958), however, reported the death of young spat (1-2 mm) in the Dutch Wadden Sea due to heavy rain, whereas the adults were able to dig deeper into the sediment, reducing exposure.

Boyden & Russell (1972) stated that Cerastoderma edule prefers salinities between 15 and 35 psu. Russell & Peterson (1973) reported lower median salinity limits of 12.5 psu and upper median salinity limits of 38.5 psu. Rygg (1970) noted that Cerastoderma edule did not survive 23 days exposure to <10 psu. Rygg (1970) also demonstrated that salinity tolerance was temperature dependant (after 3 days, 100% survival at 33 psu and 35-38°C, but 50% mortality occurred at 20 psu and 37°C and 100% mortality at 13 psu and 37°C). Wilson (1984) noted that Cerastoderma edule remained open during 1 hour exposure to salinities between 13.3 and 59.3 psu. It should be noted that the tolerances reported above depend on the duration of the experiment.

Russell (1969) found that the optimum salinity for the survival of an adult cockle varies with the mean environmental salinity and suggested that the different salinity tolerance, demonstrated for various populations of Cerastodema edule are not inherent interspecific differences, but a result of localized environmental acclimation. It is possible that larvae settling in regions of low salinity could have developed elsewhere, under more favourable conditions, and have become gradually acclimatized to the low salinity conditions; alternatively, the larvae produced by parents from a low salinity environment might be adapted to lower salinities than those produced by populations from higher salinity (Russell, 1969).

In the Severn Estuary, Aphelochaeta marioni (studied as Tharyx marioni) characterized the faunal assemblage of very poorly oxygenated, poorly sorted mud with relatively high interstitial salinity (Broom et al., 1991).  Aphelochaeta marioni can tolerate lower salinities range. Wolff (1973) recorded Aphelochaeta marioni (studied as Tharyx marioni) from brackish inland waters in the Netherlands with a salinity of 16 psu, but not in areas permanently exposed to lower salinities. Farke (1979) reported that the species also penetrated into areas exposed to salinities of 4 psu during short periods at low tide when the freshwater discharge from rivers was high.

Sensitivity assessment.  The available evidence and distribution in estuaries indicates that adult Cerastoderma edule and other characterizing species may survive a reduction in salinity to reduced or variable and populations may become locally acclimated to reduced salinities. Biotope resistance is therefore assessed as 'HIgh' and resilience as 'High', so that the biotope is assessed as 'Not sensitive'.

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s to 0.2 m/s for more than one year. Further detail

Evidence

The biotope is associated with weak  tidal streams (<0.5 m/s) and a change at the pressure benchmark (increase or decrease may affect the sediments and biological assemblage. Sepending on site specific ydrodynamics an increase in water flow rate at the benchmark level may remove fine silts and organic matter. Although Cerastoderma edule and Abra nitida are found in coarser sediments (muddy sands) the cirratulid Aphelochaeta marioni, prefers a habitat with a high silt content (Gibbs, 1969).  George (1964b) found that particle size was negatively correlated with the density of Cirriformia tentaculata in Hamble Spit, Southampton. However, he suggested that this was probably as much to do with availability of organic matter, it being generally lower in the areas with higher grain sizes. There was a positive correlation between the amount of organic matter and abundance.

Cockles are dependent on water flow to deliver suspended food particles and are abundant in biotopes that experience stronger tidal streams than the biotope considered.  Decreasing water flow rate in sheltered habitats may increase siltation and favour muddy substrates that are less suitable for Cerastoderma edule. Boyden and Russell (1972) suggested that lack of tidal flow may exclude Cerastoderma edule possibly due to reduced food availability as suggested by Brock (1979). According to regression models developed by Ysebaert et al. (2002), Cerastoderma edule occurs in environments subject to flow velocities of up to 0.8 m/s, having a maximum predicted probability of occurrence at flow velocities around 0.35 m/s. Experimental studies of water velocity and clearance rate in Cerastoderma edule have produced a range of results which may be due to genetic or phenotypic differences in test populations (Widdows & Navarro, 2007). Wildish & Miyares (1990) recorded a reduction in flume experiments found that feeding efficiency was greatest at 0.15 m/s and gradually declined to 0.45 m/s, there was no significant difference in feeding rate between current velocities of  0.05 and 0.35 m/s(Widdows & Navarro, 2007). As this biotope occurs in sheltered areas and is characterized by muddy mixed sediments, water flows are already likely to be low and a further decrease may reduce habitat suitability for Cerastoderma edule.

Increasing water flow may remove adult Cerastoderma edule from the sediment surface and carry them to unfavourable substratum or deep water, where they may be lost from the population. Coffen-Smout & Rees (1999) reported that cockles could be distributed by flood and ebb tides, but especially flood tides (by rolling around the surface) up to 0.45 m on neap tides or between 94 m and 164 m on spring tides. Newly settled spat and juveniles (<4.8mm) are capable of bysso-pelagic dispersal. Therefore, water flow rates probably affect the distribution and dispersal of juveniles and adults but these changes are unlikely at the pressure benchmark.

Additionally, an increase in water flow that reduced the deposition of particulate matter at the sediment surface would  reduce food availability for all deposit feeders. . At the pressure benchmark. increased water flow may enhance food supply to the suspension feeding Cerastoderma edule, an increase in habitat suitability for this species may lead to changes in the biological assemblage as cockles can exclude other species by occupying more space and disturbing sediments leading to organic mater re-suspension and reduction in habitat suitability for deposit feeders.

Sensitivity assessment. The cirratulids are considered to be tolerant of a reduction in water flow at the pressure benchmark, as these are found in accreting environments with high mud content and are deposit feeders, benefitting from ncreased depoition of organic matter. Increased siltation will reduce habitat suitability for ephemeral alagae and the barnacle Elminius modestus if small stones and other suitable attachment surfaces become covered with silts.  The characterizing Cerastoderma edule is considered to have some resistance to this pressure as they are found within a range of flow speeds and can feed at a range of flow speeds. However decreased flow rates (at the pressure benchmark) in sheltered habitats may reduce food supply and enhance sediment deposition leading to replacement by deposit feeders. At the pressure benchmark some biotopes may be affected by an increase or decrease in water flow through effects on sediment and organicmatter and delivery of suspended food; biotope resistance is assessed as 'Medium' and resilience is assessed as 'High'

Emergence regime changes

Benchmark.  1) A change in the time covered or not covered by the sea for a period of ≥1 year or 2) an increase in relative sea level or decrease in high water level for ≥1 year. Further detail

Evidence

This biotope is found on the lower and mid shore (JNCC, 2015)  and the associated fauna are likely to be tolerant of a certain degree of desiccation. In addition, the majority of important characterizing species are infaunal and are therefore protected from major changes in aerial exposure.

Boyden (1972) reported that Cerastoderma edule survived 42.9% water loss. However, increased desiccation, equivalent to raising the biotope from mid to high water, is likely to reduce the abundance of this some associated fauna, especially those that don't build tubes such as the polychaetes Exogone naidina and Sphaerosyllis taylori. If the branchiae of the cirratulid Cirriformia tentaculata are exposed they will either be withdrawn into the burrow of the worm or clump together and stop functioning properly (Dales & Warren, 1980). If the branchiae of the cirratulid Cirriformia tentaculata are exposed they will either be withdrawn into the burrow of the worm or clump together and stop functioning properly (Dales & Warren, 1980). Aphelochaeta marioni, another cirratulid, can only feed when immersed and therefore will experience reduced feeding opportunities, reducing growth.

Cockle dominated biotopes can range from mean high water springs (Sanchez-Salazar et al., 1987b) to the sublittoral (JNCC, 2015).  Shore height influences a number of factors that affect cockle condition and survival and the size and age structure of populations can vary significantly with shore height (Sanchez-Salazar et al., 1987b), hence this biotope is likely to be sensitive to changes in emergence (both increase and decrease). 

At lower shore levels predation by shore crabs and fish may structure populations by removing smaller individuals and may set the lower distribution limit, while at higher shore levels predation by oyster-catchers targets larger size classes (Sanchez-Salazar et al., 1987b).  Higher shore populations are exposed to air temperatures for longer that may be warmer and colder than seawater creating thermal shocks, higher shore individuals  also have less time to feed, resulting in reduced growth (Jensen, 1992) body condition which may increase susceptibility to parasites and other factors (Wegeberg & Jensen, 2003). Dense populations lower on the shore may also deplete the available suspended food, reducing the supply to higher shore populations (Peterson & Black, 1987, 1991; Kamermans 1993).

DeMontadouin and Bachelet (1996) manipulated low and high population densities (160-2000 adults/m² and tidal elevation (low and mid-water levels of Cerastoderma edule at Arcachom bay, SW France to test the influence of adult densities and emersion time on growth, settlement and survival. Growth rates were affected by tidal height with higher growth rates at low water levels.

Sensitivity assessment. A change in emersion (particularly increase) is likely to alter the habitat suitability for Cerastoderma edule and associated species, resulting in changes in growth rates, predation and  assemblage structure, biotope resistance is therefore assessed as ‘Low’ and resilience is assessed as ’Medium’ following a return to previous habitat condition

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year. Further detail

Evidence

This biotope occurs in very sheltered habitats (JNCC, 2015) and is considered be tolerant of a decrease in wave exposure at the pressure benchmark.  

Increases in wave exposure greater than the pressure benchmark are likely to have marked effects on the sediment dynamics of the shore. Species on the sediment surface including cockles and tube building polychaetes are likely to be washed away and may end up in unfavourable habitats. Infauna may also be dislodged if the top layers centimetres of sediment are removed. This will render the worms more susceptible to predation. Rough seas in March 1960 were found to wash away young Cirriformia tentaculata from the top surface layers of mud at Hamble Spit, Southampton (George, 1964b). Polychaetes living further down in the sediment may be saved from dislodgement but the biotope per se will be lost. Increased exposure could also result in increased grain size or erosion of the sediment, while decreased exposure will lead to increased siltation and reduced grain size (muddy sediment). In both cases the sediment may become unsuitable for Cerastoderma edule populations resulting in a reduction of the extent or abundance of the population (see physical change pressures). Increased wave action during storms may also remove adult cockles from the sediment surface which may be subsequently lost from the population. These changes are not assessed in this section as they are considered to be the result of changes in wave action that exceeds the pressure benchmark.

Sensitivity assessment. At the pressure benchmark the biotope is considered to have 'High' resistance and 'High' resiliencee (by default) to changes (increase or decrease) in wave action at the pressure benchmark.

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed but evidence is presented where available.

The species present in the biotope may, however, be sensitive to increases in contaminants greater than the pressure benchmark. Studies of Cerastoderma edule populations from polluted and un-contaminated sites in Southampton Water showed that tissue heavy metal concentrations were lower in summer than winter/spring, tissue heavy metal concentrations decreased with size of the cockle, and that cockles in sediments contaminated with metals and hydrocarbons had lower life expectancies, growth rates and body condition index (Savari et al. 1991(a), (b)).

Bryan (1984) suggested that many polychaetes were resistant to heavy metals and evidence from the work of Bryan & Gibbs (1983) in the metal polluted Fal estuary supports this view. Bivalves, on the other hand, including Cerastoderma edule displayed a much lower tolerance and were found to be the most obvious absentees from the polluted Restronguet Creek area of the Fal (Bryan & Gibbs, 1983). The following information is taken from Bryan & Gibbs (1983).

•Aphelochaeta marioni (studied as Tharyx marioni) was found to contain exceptionally high concentrations of arsenic (> 2000 µg / gram dry body weight) without obvious adverse effects.

•Pygospio elegans appear to have adapted to the high concentrations of copper and zinc in Restronguet Creek and the larvae are subjected to widely fluctuating conditions of salinity and relatively high metal concentrations.

•Increased tolerance of copper was found in the amphipod Corophium volutator in the creek.

• Adult Cerastoderma edule were found to be more tolerant to metal toxicity than the juvenile or larval stages which appear unable to withstand the high concentrations of copper and zinc. However, transplantation of Cerastoderma edule into Restronguet Creek (highly polluted by heavy metals) resulted in 10-15% mortality within 63 days but 100% within about four months. The toxic body-burden of copper to Cerastoderma edule was found to be ca. 250 µg / g with zinc being less toxic. Bryan & Gibbs (1983) stated that Cerastoderma edule takes up heavy metals mainly from solution rather than from sediment and that it was excluded from Restronguet Creek by the high levels of Cu and Zn.

A 2-year microcosm experiment was undertaken to investigate the impact of Cu on the benthic fauna of the lower Tyne Estuary (UK) by Hall and Frid (1995). During a 1-year simulated contamination period, 1 mg l−1 Cu was supplied at 2-weekly 30% water changes, at the end of which the sediment concentrations of Cu in contaminated microcosms reached 411 μg g−1. Toxicity effects reduced populations of the four dominant taxa, including Capitella capitata. When Cu dosage was ceased and clean water supplied, sediment Cu concentrations fell by 50% in less than 4 days, but faunal recovery took up to 1 year, with the pattern varying between taxa. Since the Cu leach rate was so rapid it is concluded that after remediation, contaminated sediments show rapid improvements in chemical concentrations, but faunal recovery may be delayed with experiments in microcosms showing faunal recovery taking up to a year.

Hydrocarbon & PAH contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed but evidence is presented where available.

The species present in the biotope may, however, be sensitive to increases in contaminants greater than the pressure benchmark The impacts of hydrocarbon contamination on sediment shores were well documented by Rostron (1998) following the Sea Empress oil spill in Milford Haven, Wales and the information in the following section is taken mainly from that report including the references therein.

•High mortalities of the cockles Cerastoderma edule were reported. In Angle Bay, the presence of juveniles lower down the shore shortly after the spill enabled the reestablishment of adult populations on the middle shore within about six months.

•Additional species recorded at Sandy Haven in the summer following the spill included the polychaetes Pygospio elegans and Capitella capitata and the oligochaete Tubificoides benedii. The abundance of Capitella capitata increased dramatically at one site.

•At one station in Sandy Haven, the amphipod Corophium volutator disappeared completely following the Sea Empress oil spill. Indeed, Chasse & Morvan (1978, cited in Rostron 1998) calculated that only 10% of these amphipods survived the Amoco Cadiz oil spill.

•At Angle Bay, a sheltered bay with mixed sandy mud and mud, the cirratulid Chaetozone gibber showed a dramatic increase after the spill. However, this success was short lived and the numbers had fallen significantly by the following year.

Cirratulids appeared to Suchanek (1993) to be mostly immune to oil spills because their feeding tentacles are protected by a heavy secretion of mucus. This immunity is supported by observations of Aphelochaeta marioni following the Amoco Cadiz oil spill in March, 1978 (Dauvin, 1982, 2000). Prior to the spill, Aphelochaeta marioni (studied as Tharyx marioni) was present in very low numbers in the Bay of Morlaix, western English Channel. Following the spill, the level of hydrocarbons in the sediment increased from 10 mg/kg dry sediment to 1443 mg/kg dry sediment 6 months afterwards. In the same period, Aphelochaeta marioni increased in abundance to a mean of 76 individuals per m², which placed it among the top five dominant species in the faunal assemblage. It was suggested that the population explosion occurred due to the increased food availability because of accumulation of organic matter resulting from high mortality of browsers. Six years later, abundance of Aphelochaeta marioni began to fall away again, accompanied by gradual decontamination of the sediments.

Synthetic compound contamination

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed but evidence is presented where available.

The species present in the biotope may, however, be sensitive to increases in contaminants greater than the pressure benchmark. The close association of benthic invertebrates with contaminated sediments may cause some sub-lethal effects and in the long-term could interfere with reproductive potential (Rostron, 1998). Cerastoderma edule is known to accumulate PCBs (see MarLIN review) but no specific information concerning the actual effects that contamination with synthetic chemicals has on this species or on other characterizing species within the biotope was found.

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 Cirratulus cirratus (family Cirratulidae) 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 of the polychaetes.

Radionuclide contamination

Benchmark. An increase in 10µGy/h above background levels. Further detail

Evidence

No evidence.

Introduction of other substances

Benchmark. Exposure of marine species or habitat to one or more relevant contaminants via uncontrolled releases or incidental spills. Further detail

Evidence

This pressure is Not assessed.

De-oxygenation

Benchmark. Exposure to dissolved oxygen concentration of less than or equal to 2 mg/l for one week (a change from WFD poor status to bad status). Further detail

Evidence

A number of animals have behavioural strategies to survive periodic events of reduced dissolved oxygen. These include avoidance by mobile species such as crustaceans, shell closure and reduced metabolic rate in bivalve molluscs and either decreased burrowing depth or emergence from burrows for sediment dwelling crustaceans, molluscs and annelids. The sensitivity to reduced oxygen and recovery from episodes of hypoxia and anoxia varies between the characterizing and associated species of this biotope. The available evidence suggests that Cerastoderma edule is more sensitive to this pressure than polychaete species.

Rosenberg et al. (1991) reported 100% mortality of Cerastoderma edule exposed to 0.5-1.0ml/l oxygen for 43 days and 98% mortality after 32 days. Cerastoderma edule migrated to the surface of the sediment in response to decreased oxygen concentrations. Theede et al. (1969) reported 50% mortality after 4.25 days at 1.5 mg/l oxygen. Theede et al. (1969) also noted that Cerastoderma edule only survived 4 days exposure to 0.0-6.1 cm³/l of hydrogen sulphide, which is associated with anoxic conditions. This suggests that Cerastoderma edule could survive short periods of anoxia but it is likely that continued exposure to 2 mg/l oxygen for a week would be lethal.

Fifty percent (LT50) of cockles in anoxic seawater died after 3.5 days (Babarro & de Zwaan, 2001) The anoxic survival time of Cerastoderma edule from two different ecosystems and differing anoxia tolerances was studied in static (closed) and flow-through systems. The antibiotics chloramphenicol, penicillin and polymyxin were added, and molybdate ( a specific inhibitor of the process of sulfate reduction). Median mortality times were 2.7 and 2.9 days for Cerastoderma for static and flow-through incubations, respectively. The addition of chloramphenicol increased strongly survival time in both systems with corresponding values of 6.4 and 6.5 days for Cerastoderma. Overall the results indicate that proliferation of anaerobic pathogenic bacteria, associated with the bivalves, is a main cause of death besides the lack of oxygen. Bacterial damage is probably caused by injury of the tissues of the clams and not by the release of noxious compounds to the medium (de Zwaan et al. 2002).

Connor et al. (1997b) described sediments in which the cirratulid Aphelochaeta marioni is commonly found as usually having a "black anoxic layer close to the sediment surface". Broom et al. (1991) considered Aphelochaeta marioni (studied as Tharyx marioni) to be characteristic of faunal assemblage of very poorly oxygenated mud in the Severn Estuary. They found that it dominated sediments where the redox potential at 4 cm sediment depth was 56 mV and, therefore, concluded that the species was tolerant of very low oxygen tensions. Thierman et al. (1996) studied the distribution of Aphelochaeta marioni in relation to hydrogen sulphide concentrations. The species was found to be abundant at low sulphide concentrations (less than 50 µM) but only occasional at concentrations from 75-125 µM. They concluded that Aphelochaeta marioni does not display a massively adverse reaction to sulphidic conditions and is able to tolerate a low amount of sulphide. The evidence suggests that Aphelochaeta marioni is capable of tolerating hypoxia but it is difficult to determine to what degree.

The cirratulid Cirriformia tentaculata is reported to have several metabolic adaptations to the hypoxic conditions to which it is periodically subjected (Dales & Warren, 1980; Bestwick et al., 1989). The sediment around their burrows is often hydrogen-sulphide rich and therefore a sink for oxygen (Bestwick et al., 1989). The adaptations are, firstly, the filamentous branchiae of the worm, that are spread out over the surface of the substratum, are very thin and oxygen uptake can continue during tidal emersion providing the branchiae are covered by a film of water (Bestwick et al., 1989). If the branchiae are exposed they may be withdrawn into the burrow at which point the gaseous exchange occurring across the branchial epithelium starts to fall. Secondly, the haemoglobin has an extremely high affinity for oxygen and as the internal oxygen pressure falls, oxygen is released from the haemoglobin store (Dales & Warren, 1980). At an external oxygen pressure of 0.88 mg/l, oxygen uptake stops and the species cannot tolerate anoxia for more than three days (Dales & Warren, 1980).

The oligochaete Tubificoides benedii also inhabits sulfide rich environments and has a high capacity to tolerate anoxic conditions (Nubilier et al., 1997; Giere et al., 1999). Tubificoides benedii is often buried up to 10 cm deep and so has no contact with the surface but has a highly specialized adaptive physiology that allows it to maintain some oxygen consumption even at 2% (approximately 0.18 mg/l) oxygen saturation of the surrounding environment on the Isle of Sylt. The critical oxygen saturation for Capitella capitata is about 7.5 mg/l (Gamenick, 1996, cited in Giere et al., 1999). It has been suggested that tolerance to anoxia may be influenced by temperature. Tubificoides benedii (studied as Peloscolex benedeni) was found to be less tolerant to anoxia as temperature increased (Diaz, 1980). At 20 °C, it took almost 60 hours for half the worms to be killed but at 30 °C it took less than 18 hours.

Dense Capitella capitata populations are frequently located in areas with greatly elevated organic content, even though eutrophic sediments are often anoxic and highly sulfidic (Tenore 1977; Warren 1977; Tenore & Chesney 1985; Bridges et al. 1994). The polychaetes Capitella capitata, Pygospio elegans and Scoloplos armiger have all been reported to recolonize habitats following periods of anoxia and hypoxia. Following a period of anoxia in the Bay of Somme (north France) that removed cockles, Pygospio elegans increased rapidly but then decreased as cockle abundance recovered (Desprez et al., 1992; Rybarczyk et al.,1996).

Sensitivity assessment. Decreased oxygen levels, could lead to an alteration in sediment chemistry, including the production of hydrogen sulphides that would alter habitat conditions and is likely to lead to mortality of Cerastoderma edule although other species present are likely to be more tolerant. Based on Theede et al., 1996, the sensitivity of the biotope (based on Cerastoderma edule) is ‘Medium’ as periodic emmersion would reoxygenate sediments and exposure is likely to be short-term, limiting mortality. Resilience is assessed as ‘high’ and sensitivity is assessed as ‘Low’.

Nutrient enrichment

Benchmark. Compliance with WFD criteria for good status. Further detail

Evidence

Nutrient enrichment may result in increased primary productivity that could increase the amount of food available to both suspension feeders and deposit feeders. However, nutrient enrichment often culminates in eutrophic episodes which usually lead to increased light attenuation (see turbidity) and reduced oxygen concentration (see oxygenation); and increased algal growth often culminates in mats of algae covering the sediment surface.

Desprez et al. (1992) implicated a eutrophication-induced plankton bloom as the cause behind the decline of Cerastoderma edule populations in the Bay of Somme, France. Prior to the event in 1982, densities were several 1000 / m² but by 1982 this had fallen to just a few hundred individuals / m². By 1987, the cockle population had returned. Rosenberg & Loo (1988) suggested that the mass mortalities of Cerastoderma edule observed in Laholm Bay, western Sweden during the 1980s were correlated with increased nutrient levels, and the associated decrease in oxygen levels during this period. However, no direct causal link was established. Some authors have reported a decline in the abundance of Cerastoderma edule under algal mats (Raffaelli et al., 1998).

Raman & Ganapati (1983) studied the distribution of Aphelochaeta marioni (studied as Tharyx marioni) in relation to a sewage outfall in Visakhaptnam Harbour, Bay of Bengal. Aphelochaeta marioni was found to be dominant in the 'semi-healthy zone' characterized by high dissolved oxygen (median 7.2 mg/l), low biological oxygen demand (9.6 mg/l) and low nutrients (nitrate 0.02 mg/l, phosphate 0.88 mg/l). Aphelochaeta marioni was not found in high numbers in the polluted zone close to the sewage outfall, characterized by low dissolved oxygen (median 6.0 mg/l), high biological oxygen demand (14-60 mg/l) and high nutrients (nitrate 0.042-0.105 mg/l, phosphate 2.35-3.76 mg/l). This would suggest that Aphelochaeta marioni is intolerant of eutrophication.

The oligochaete Tubificoides benedii can be found living in abundance under algal mats (Nubilier et al., 1997). It is opportunistic and responds to organic pollution by increasing the size of the population (Diaz, 1977, cited in Diaz, 1980). Oligochaetes often become the dominant benthic fauna under algal mats (Raffaelli et al., 1998). Estuarine oligochaetes tend to become more abundant in areas where pollution or other physical factors result in a reduced habitat diversity and stressful conditions, concomitant with a decrease in polychaetes species (Diaz, 1980). Barnett (1983) found the maximum abundance of the oligochaete Tubificoides benedii at a site which received a significant input of both industrial and domestic effluent, including raw sewage, in the Humber estuary.

Capitella capitata is often associated with areas of high nutrient enrichment and is generally considered to be tolerant of increased nutrient load.

Sensitivity assessment. As Cerastoderma edule and other characterizing species in the biotope are not primary producers they are not considered directly sensitive to an increase or decrease in plant nutrients in the water column. Phytoplankton and algal detritus may be utilised as food by Cerastoderma edule and deposit feeders but supply is not considered to be affected at the pressure benchmark level. The biotope is therefore considered to be ‘Not Sensitive’ to this pressure. Resistance is therefore assessed as ‘High’ and resilience as ‘High’ (by default).

Organic enrichment

Benchmark. A deposit of 100 gC/m²/yr. Further detail

Evidence

Benthic responses to organic enrichment have been described by Pearson & Rosenberg (1978) and Gray (1981).  Moderate enrichment increases food supply enhancing productivity and abundance.

Cerastoderma edule has been categorised through expert judgement and literature review as AMBI Group III- species tolerant to excess organic matter enrichment. These species may occur under normal conditions, but their populations are stimulated by organic enrichment (slight unbalance situations). They are surface deposit feeding species (Borja et al., 2000, validated by Gittenberger & van Loon, 2011).

Organic enrichment beneath oyster cultivation trestles and mussel cultivation sites and fish cages has led to community replacement/dominance by cirratulid, capitellid and spionid polychaetes, in mudflats, that characterize disturbed areas enriched in organic matter (Pearson & Rosenberg 1978, Samuelson 2001, Bouchet & Saurier 2008). The associated cirratulids within the biotope are. therefore, likely to benefit from organic enrichment at the pressure benchmark.

Sensitivity assessment.  Areas with significant mud contents are likely to be rich in organic matter and low oxygen penetration coupled with high levels of bacterial activity means sediments are anoxic a short distance below the surface. Given their adaptation to these habitat conditions the characterizing Cerastoderma edule and other characterizing species are not considered sensitive to organic enrichment. The deposit feeding cirratulids are likely to benefit from the additional food source The biotope is considered ‘Not sensitive’ to this pressure based on ‘High’ resistance and ‘High’ recovery (by default). Gross organic pollution (greater than the pressure benchmark would be likely to lead to detrimental effects depending on the level of the pressure).

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site. Further detail

Evidence

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. 

Physical change (to another seabed type)

Benchmark. Permanent change from sedimentary or soft rock substrata to hard rock or artificial substrata or vice-versa. Further detail

Evidence

This biotope is found on muddy sands (Connor et al., 2004). A change to natural or artificial hard substratum would remove this sedimentary biotope and the species. If pockets of fine sediment accumulate in pockets within the substrata then these areas may be re-colonised by species associated with this biotope but these pockets of sediment would not be equivalent to the biotope. Recovery will depend on the re-instatement of suitable habitat.

Sensitivity assessment. Based on the loss of suitable habitat, biotope resistance to this pressure is assessed as ‘None’. Resilience is assessed as ‘Very low’ as the pressure benchmark refers to a permanent change. Biotope sensitivity is therefore ‘High’.

Physical change (to another sediment type)

Benchmark. Permanent change in one Folk class (based on UK SeaMap simplified classification). Further detail

Evidence

The benchmark for this pressure refers to a change in one Folk class.  The pressure benchmark originally developed by Tillin et al., (2010) used the modified Folk triangle developed by Long (2006) which simplified sediment types into four categories: mud and sandy mud, sand and muddy sand, mixed sediments and coarse sediments.  The change referred to is therefore a change in sediment classification rather than a change in the finer-scale original Folk categories (Folk, 1954).  The change in one Folk class is considered to relate to a change in classification to adjacent categories in the modified Folk triangle (Long, 2006). As this biotope is characterized by mud and gravelly mud (JNCC, 2015), the change at the pressure benchmark refers to a potential change to coarse sediments and sands.

The particle size of sediments and correlated physical and chemical factors (such as, organic matter content and hydrodynamic regime), is a key determinant of the structure of benthic invertebrate assemblages (Van Hoey et al., 2004; Yates et al., 1993). A change to coarse sediments would result in loss of characterizing and associated species, resulting in biotope reclassification..

A study in the intertidal of the Dutch Wadden Sea showed that suction-dredging for cockles (Cerastoderma edule) led to a significant long-term reduction in settlement and stocks of the target bivalve species (Piersma et al., 2001). Analysis of sediment characteristics before and after dredging showed an increase in median grain size and a reduction of silt content, and that these changes were most pronounced in the area dredged for cockles. Sediment characteristics only returned to pre-impact conditions 8-11 years after the suction dredging. The authors concluded that suction dredging of Cerastoderma edule had long lasting effects on the recruitment of bivalves (particularly the target species Cerastoderma edule, but also Macoma balthica) in sandy parts of the Wadden Sea basin.

Sensitivity assessment. The character of the habitat is largely determined by the sediment type, changes to this would lead to habitat re-classification. A change to coarse or sand sediment without a high proportion of mud would be unsuitable for the characterizing species and would lead to the development of a different habitat type.   Changes in sediment characteristics can lead to changes in community structure. An increase in coarse sediments would lead to the development of a community typical of mixed sediments, clean sands and/or gravels depending on the degree of change.  In general an increase to very coarse sediments may favour some amphipod species rather than Cerastoderma edule, and the associated species. This change would alter the character of the biotope present leading to re-classification.  Biotope resistance is assessed as ‘None’, as a change at the pressure benchmark would result in loss of the habitat. Biotope recovery is assessed as ‘Very low’ as the change at the pressure benchmark is considered to be permanent.  Sensitivity is therefore assessed as ‘High

Habitat structure changes - removal of substratum (extraction)

Benchmark. The extraction of substratum to 30 cm (where substratum includes sediments and soft rock but excludes hard bedrock). Further detail

Evidence

Sedimentary communities are likely to be highly intolerant of substratum removal, which will lead to partial or complete defaunation, expose underlying sediment which may be anoxic and/or of a different character or bedrock and lead to changes in the topography of the area (Dernie et al., 2003). Recovery by infilling will depend on local factors including the mobility of sediments, sediment supply, hydrodynamics and the spatial scale of the area affected (Van Hoey et al. 2008). The extraction of sediment to 30 cm (the pressure benchmark) would remove the characterizing species and associated species present.

Sensitivity assessment. Extraction of 30 cm of sediment will remove the characterizing biological component of the biotope and sediments. The resistance of the habitat to extraction is assessed as ‘None’ as sediment is removed: the depth of remaining sediments and their character will be site-specific. Recovery will depend on local factors including hydrodynamics, sediment supply and sediment mobility and the spatial scale affected. Resilience is assessed as ‘Medium’ as sediment infilling may be rapid in intertidal areas, however, recruitment of the characterizing Cerastoderma edule is episodic (see resilience section). Biotope sensitivity is therefore assessed as 'Medium'. If sediments do not return to the previous condition, larval recolonization may be inhibited (see physical change pressures).

Abrasion / disturbance of the surface of the substratum or seabed

Benchmark. Damage to surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

Sediment cohesion provides some sediment stabilisation to resist erosion following surface disturbance. Species associated with muddy sands/sandy muds are infaunal and hence have some protection against surface disturbance, although siphons and other body parts extended to the surface for respiration and/or feeding may be damaged. Cerastoderma edule has short siphons and requires contact with the surface for respiration and feeding and may be damaged by abrasion at the surface, Abra nitida however may be more deeply buried.  Surface compaction can collapse burrows and reduce the pore space between particles, decreasing penetrability and reducing stability and oxygen content (Sheehan, 2007).  Trampling (3 times a week for 1 month) associated with bait digging reduced the abundance and diversity of infauna (Sheehan, 2007; intertidal muds and sands). However, Cooke et al. (2002) found that trampling associated with bait digging had little effect on infaunal species composition (intertidal muddy sands).

Rossi et al. (2007) conducted experimental trampling on a mudflat (5 people, 3-5 hours, twice a month between March and September). Mobile fauna were not affected; however, the abundance of adult Cerastoderma edule was sharply reduced, probably due to the trampling directly killing or burying the animals, resulting in asphyxia. However, no effect was observed on small (<12 mm) individuals of Cerastoderma edule. The authors suggested that this was because the experiment was conducted in the reproductive season for these species and hence there were juveniles present in the water column to replace individuals displaced by trampling. The lack of observed effect was therefore due to continuous recruitment and replacement of impacted individuals.

Sensitivity assessment. Abrasion at the surface is likely to damage a proportion of the population of shallow buried bivalves (Cerastoderma edule and soft-bodied species that live on or very close to the surface (examples). The level of damage and mortality will depend on the force exerted. Biotope resistance is assessed as ‘Medium’ and resilience is assessed as ‘High’ so that biotope sensitivity is therefore assessed as ‘Low’

Penetration or disturbance of the substratum subsurface

Benchmark. Damage to sub-surface features (e.g. species and physical structures within the habitat). Further detail

Evidence

The characterizing species Cerastoderma edule, the cirratulids Aphelochaeta marioni, Chaetozone gibber and Cirriformia tentaculata all live buried in the top few centimetres of sediment and are therefore likely to be damaged by physical disturbance that penetrates the upper few centimetres of the sediment. Birds and fish could be attracted to the site of disturbance to feed on exposed and damaged individuals.

In the Burry Inlet, Wales, intertidal tractor towed cockle harvesting in muddy sand reduced the abundance of Cerastoderma edule by ca 34%. Populations of Cerastoderma edule had not recovered their original abundance after 174 days (Ferns et al., 2000). Hall and Harding (1997) studied the effect of suction and tractor dredging for cockles on non-target benthic fauna in the Solway Firth, Scotland where sediments contained 60-90% silt/clay in the more sheltered areas. The results showed that suction dredging resulted in significantly lower mean species numbers (by up to 30%) and mean numbers of individuals (up to 50%) and in the abundance of 3 of the 5 dominant species. The faunal structure of the dredged plots recovered (i.e. approached that of the undisturbed control plots) by 56 days. The results of the tractor dredge experiments showed fewer effects than the suction dredging (no significant effect on the number of species or individuals). The authors concluded that mechanical harvesting methods imposed high levels of mortality on non-target benthic fauna but that the recovery of disturbed sites was rapid and that the overall effects on populations were low. Although the results suggested that tractor dredging had less impact than suction dredging the authors proposed this may have been due to differences in the timing of the experiments (May-July suction dredging; July-September tractor dredging). They concluded that although significant mortality of Cerastoderma edule and other infauna occurred, recovery was rapid and the overall effects on populations were low. Hall & Harding (1997) found that abundance had returned to control levels within about 56 days and Moore (1991) also suggested that recovery was rapid. Rostron (1995) carried out experimental dredging of sandflats with a mechanical cockle dredge. Two distinct sites were sampled; Site A: poorly sorted fine sand with small pools and Arenicola marina casts with some algal growth, and Site B: well sorted fairly coarse sand, surface sediment well drained and rippled as a result of wave activity. At both sites, Cerastoderma edule reduced after dredging but recovery was rapid at Site B (no difference between control and experimental plots after 14 days), whilst at Site A, a significant reduction in numbers compared with the control was still apparent up to six months post-dredging.

A number of studies have found that the abundance of the polychaete Pygospio elegans is reduced by simulated cockle dredging (Hall & Harding, 1998; Moore, 1990; Ferns et al., 2000; Rostron, 1995).  Ferns et al. (2000) found that tractor-towed cockle harvesting, removed 83% of Pygospio elegans (initial density 1850/ m²).  In muddy sand habitats, Pygospio elegans had not recovered to its original abundance after 174 days (Ferns et al., 2000). Rostron (1995) also found that Pygospio elegans had not recovered to pre-dredging numbers after six months. Conversely, Hall & Harding (1998) found that the abundance of Pygospio elegans increased significantly over 56 days following suction dredging.  Pygospio elegans inhabits a fragile tube that projects above the sediment surface and is probably more vulnerable to physical disturbance and abrasion than other, more deeply buried, infaunal species. Other species may recover more rapidly Capitella capitata had almost trebled its abundance within the 56 days in a clean sandy area (Ferns et al., 2000).  Following experimental beam trawl disturbance in an area that had previously been closed to fishing populations of Melinna palmata increased by 41% (Tuck et al., 1998). The area was repeatedly disturbed over an 18-month period and recovery was tracked for a further 18 months. 

With respect to displacement, cockles are capable of burrowing rapidly into the substratum and >50% burrowed into the substratum within 1 hour in experimental trials (Coffen-Smout & Rees, 1999), although this rate was inhibited by prior disturbance. Brock (1979) reported that 80% began to burrow within 60 min and 50% had successfully burrowed into sediment within 60 min. He also noted that young cockles could burrow quickly, and were nearly buried within 5 min. Hand-raking for cockles was shown not to influence the re-burial rate of cockles in Strangford Lough, Northern Ireland (McLaughlin et al., 2007).

Sensitivity assessment. The available evidence indicates that small patches of physical disturbance are likely to be in-filled by adult cockle movement, large patches will recover through larval recruitment, which again is subject to many factors, and may be improved by the removal of adult cockles. Biotope resistance is assessed as ‘Low’ based on loss of characterizing species Cerastoderma edule and associated species. Resilience is assessed as ‘Medium’ to take account of recruitment variability and return of normal age structure. Sensitivity is therefore assessed as ‘Medium’

Changes in suspended solids (water clarity)

Benchmark. A change in one rank on the WFD (Water Framework Directive) scale e.g. from clear to intermediate for one year. Further detail

Evidence

An increase in the amount of suspended sediment could potentially increase the amount of food available to deposit feeders, the major trophic group within this biotope. However, this would only occur if the proportion of organic material within the suspended sediment increased. With regard to suspension feeders, increasing total particulate concentrations have been shown to decrease clearance rates and increase pseudofaeces production in Cerastoderma edule (Navarro et al., 1992; Navarro & Widdows, 1997). Furthermore, due to the sheltered nature of the habitat, siltation is likely. The increase in suspended sediment is likely to increase the proportion of mud, to the detriment of Cerastoderma edule.

A decrease in suspended sediment is likely to reduce the amount of available food for both suspension feeders and deposit feeders although at the benchmark level this is unlikely to cause mortality. Navarro & Widdows (1997) suggested that Cerastoderma edule was able to compensate for decrease in particulate quality (i.e. proportion of organic to inorganic seston) between 1.6 to 300 mg/l. Over the benchmark period the associated fauna may experience reduction in growth. On resumption of normal levels of suspended sediment. An increase in turbidity will mean that primary production in the water column may suffer from increased light attenuation.

Sensitivity assessment. A decrease in turbidity and hence increased light penetration may result in increased phytoplankton production and hence increased food availability for suspension feeders, including Cerastoderma edule. Therefore, reduced turbidity may be beneficial. In areas of high suspended sediment, a decrease may result in improved condition and recruitment due to a reduction in the clogging of filtration apparatus of suspension feeders and an increase in the relative proportion of organic particulates. However, a decrease in suspended organic particles in some areas may reduce food availability for deposit feeders resulting in lower growth or reduced energy for reproduction. Where increased turbidity results from organic particles then subsequent deposition may enhance food supply for deposit feeders within the biotope such as the cirratulids and oligochaetes. Alternatively, if turbidity results from an increase in suspended inorganic particles then energetic costs may be imposed on these species as sorting and feeding becomes less efficient reducing growth rates and reproductive success. Lethal effects are considered unlikely given the occurrence of Cerastoderma edule and other associated species in estuaries where turbidity is frequently high from suspended organic and inorganic matter. Resistance and resilience are therefore assessed as 'High' and the biotope is conisdered to be 'Not sensitive'.

Smothering and siltation rate changes (light)

Benchmark. ‘Light’ deposition of up to 5 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

Cerastoderma edule has short siphons and needs to keep in contact with the surface of the sediment. Bait digging (for large polychaete worms) disturbs the sediment down to a depth of 30cm (Jackson & James, 1979). Intensification of bait digging on intertidal flats on the north Norfolk coast for lugworm (Arenicola marina) and rag worm (Hediste diversicolor) was associated with declines in the Cerastoderma edule populations (Jackson & James, 1979). Experimental simulation of bait digging (sediment dug over to a depth of 30cm with a garden fork) led to high mortalities of cockles in dug areas rather than undug areas (48% mortality in 9 days to a maximum of 85% after 11 days) probably due to smothering (Jackson & James, 1979). The observation was texted experimentally in the laboratory by burying 3 size-classes of cockles at 0, 5 or 10cm depth in a mix of oxidated and deeper anoxic sands (mixed as a ratio of 3:1) was used as the sediment. Movements were recorded after 24, 48 and 72 hours. When buried to 5 cm depth most cockles returned to the surface but few were able to reposition to the surface if buried at 10cm depth. None had died after 72 hours. Additional burial experiments under 10 cm of sediment assessed movement and survival after 3, 6 and 9 days in two sediment types (a mix of surface mud and sand in two ratios 9:1 and 1:9).  Movement towards the surface was slower in the predominantly muddy sediment and all cockles died between 3 and 6 days. Substantial mortality resulted in the predominantly sandy mixture although some cockles were able to move towards the surface and survive for 9 days.

Cerastoderma edule have been categorised through expert and literature review as AMBI sedimentation Group II – species sensitive to high sedimentation. They prefer to live in areas with some sedimentation, but don’t easily recover from strong fluctuations in sedimentation (Gittenberger & van Loon 2011).

The associated species Pygospio elegans is limited by high sedimentation rates (Nugues et al., 1996) and the species does not appear to be well adapted to oyster culture areas where there are high rates of accumulation of faeces and pseudo faeces (Sornin et al., 1983; Deslous-Paoli et al., 1992; Mitchell, 2006 and Bouchet & Sauriau 2008).  Pygospio elegans is known to decline in areas following re-deposition of very fine particulate matter (Rhoads & Young, 1971; Brenchley, 1981). Experimental relaying of mussels on intertidal fine sands led to the absence of Pygospio elegans compared to adjacent control plots. The increase in fine sediment fraction from increased sediment deposition and biodeposition alongside possible organic enrichment and decline in sediment oxygen levels was thought to account for this (Ragnarsson & Rafaelli, 1999).

Mobile and/or burrowing species (including molluscs and polychaetes such as Hydrobia ulvae, Eteone longa and Scoloplos armiger) are generally considered to be able to reposition following periodic siltation events or low levels of chronic siltation. Field experiments where 10 cm of sediment were placed on intertidal sediments to investigate the effects of the beneficial use of dredged materials found that the abundance of H. ulvae had returned to ambient levels within 1 week (Bolam et al. 2004). However, survival depends on several factors. The snail can only burrow up through certain sorts of sediment. If the silt content of the smothering sediment is high and the water content low then it is unlikely that the surface will be regained from 5 cm down. Looser sediment with high water and low silt content can be negotiated quite rapidly. The surface is generally regained within a day. If the surface cannot be regained then Hydrobia ulvae can survive burial for quite extended periods although this is highly temperature dependent. Temperatures of 20^oC result in all individuals dying after 10 days. Survival is much better at lower temperatures. It is thought that oxygen stress is the cause of mortality (Jackson, 2000).

Melinna palmata lives in a mucous-lined tube covered in sediment that projects obliquely above the sediment (Fauchald & Jumars, 1979). In general, mucus tube feeders and labial palp deposit feeders were most intolerant to burial (Maurer et al., 1986). Smothering may result in this tube being broken which may result in the displacement or mortality of some individuals. It is not known whether other important characterizing fauna including the oligochaetes Tubificoides benedii, Tubificoides pseudogaster and the polychaete Pygospio elegans would be adversely affected by smothering but their mobility may enable them to dig back up through the sediment to the surface.

Sensitivity assessment. Biotope resistance based on the characterizing Cerastoderma edule is assessed as having ‘Medium’ resistance to siltation, (as many would be able to survive and re-emerge from a 5 cm depth of sediment). Resilience is assessed as ‘High’ based on adult migration and repopulation by larvae. Many of the associated species are also likely to reposition although Pygospio elegans may be more sensitive.

Smothering and siltation rate changes (heavy)

Benchmark. ‘Heavy’ deposition of up to 30 cm of fine material added to the seabed in a single discrete event. Further detail

Evidence

Cerastoderma edule has short siphons and needs to keep in contact with the surface of the sediment. Bait digging (for large polychaete worms) disturbs the sediment down to a depth of 30cm (Jackson & James, 1979) and leads to repositioning within sediment or burial from piled sediments. Intensification of bait digging on intertidal flats on the north Norfolk coast for lugworm (Arenicola marina) and ragworm (Hediste diversicolor) was associated with declines in the Cerastoderma edule populations (Jackson & James, 1979). Experimental simulation of bait digging (sediment dug over to a depth of 30cm with a garden fork) led to high mortalities of cockles in dug areas rather than undug areas (48% mortality in 9 days to a maximum of 85% after 11 days) probably due to smothering (Jackson & James, 1979). The observation was texted experimentally in the laboratory by burying 3 size-classes of cockles at 0, 5 or 10cm depth in a mix of oxidated and deeper anoxic sands (mixed as a ratio of 3:1) was used as the sediment. Movements were recorded after 24, 48 and 72 hours. When buried to 5 cm depth most cockles returned to the surface but few were able to reposition to the surface if buried at 10cm depth. None had died after 72 hours. Additional burial experiments under 10 cm of sediment assessed movement and survival after 3, 6 and 9 days in two sediment types (a mix of surface mud and sand in two ratios 9:1 and 1:9).  The movement towards the surface was slower in the predominantly muddy sediment and all cockles died between 3 and 6 days. Substantial mortality resulted in the predominantly sandy mixture although some cockles were able to move towards the surface and survive for 9 days.

Field experiments where 10 cm of sediment were placed on intertidal sediments to investigate the effects of the beneficial use of dredged materials found that the abundance of Hydrobia ulvae had returned to ambient levels within 1 week (Bolam et al., 2004). It is not clear whether this species could reposition after 30 cm of sediment was deposited on the sediment.

Sensitivity assessment. The addition of 30 cm of sediment would prevent Cerastoderma edule and Abra nitida from extending siphons to the surface. It is unlikely that these species could emerge from this depth of sediment although some individuals may survive and sediment may be rapidly removed by tide and wave action. It is likely however that there would be considerable mortality of the characterizing Cerastoderma edule and biotope sensitivity is based on this species. Resistance is assessed as ‘Low’ and resilience as ‘Medium’ (based on episodic recruitment). Sensitivity is therefore assessed as ‘Medium’. The sensitivity of the associated species is unclear. Although some polychaetes may be able to reposition following sedimentation at the pressure benchmark this will depend on the characteristics of the overburden and sedentary species such as Pygospio elegans are likely to suffer high levels of mortality.

Litter

Benchmark. The introduction of man-made objects able to cause physical harm (surface, water column, seafloor or strandline). Further detail

Evidence

Not assessed. No evidence was found for ingestion of microplastics by Cerastoderma edule. Polychaete worm fecal casts analyzed by Mathalon & Hill (2014) had microplastic fiber concentrations resembling those found in low tide sediments. This is an indication that polychaete deposit feeders are indiscriminately feeding on microplastics, and appear to be excreting most if not all the microplastics they consume. However, polychaetes may still be affected by contaminants that are absorbed in microplastics upon ingestion. Wright et al., (2013) found that deposit-feeding marine worms maintained in sediments spiked with microscopic unplasticised polyvinylchloride (UPVC) at concentrations overlapping those in the environment had significantly depleted energy reserves by up to 50%. The effect was suggested to result from a combination of reduced feeding activity, longer gut residence times of ingested material and inflammation (Wright et al., 2013)

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT. Further detail

Evidence

No evidence,

Underwater noise changes

Benchmark. MSFD indicator levels (SEL or peak SPL) exceeded for 20% of days in a calendar year. Further detail

Evidence

Cerastoderma edule can probably detect the vibration caused by predators and will withdraw its siphons. However, little information was found concerning the effect of noise or vibration on cockle populations. The polychaetes and other worms are unlikely to have the ability to detect noise and other associated fauna are also unlikely to be adversely affected. This pressure is therefore considered 'Not relevant'.

Introduction of light or shading

Benchmark. A change in incident light via anthropogenic means. Further detail

Evidence

Cerastoderma edule carries about 60 eyes on the sensory siphonal tentacles (Barber & Wright 1968). The basic response to light in bivalves is defensive with responses including digging, closing of valves and siphonal withdrawal (Morton, 2008). Changes in light levels may also simulate other behavioural responses and emergence of cockles in response to darkness has been observed (Richardson et al., 1993). Light levels may act as cues for reproduction (although temperature also regulates reproduction for many species) supporting synchronised spawning for species with external fertilisation.  No evidence was found to suggest that light levels are an important reproductive cue for characterizing and associated species.

Aphelochaeta marioni is only active at night and Farke (1979) noted their intolerance to visual disturbance in a microsystem in the laboratory. In order to observe feeding and breeding in the microsystem, the animals had to be gradually acclimated to lamp light. Even then, additional disturbance, such as an electronic flash, caused the retraction of palps and cirri and cessation of all activity for some minutes. Visual disturbance, in the form of direct illumination during the species' active period at night, may therefore result in loss of feeding opportunities, which may compromise growth and reproduction.

Sensitivity assessment. Light penetration into sediments is limited to the surface layers and permanently buried infauna are unlikely to be affected by changes in light levels. The characterizing Cerastoderma edule and other species present can perceive light but the effects of changes in light level and shading or the duration of light and darkness are not clear. As it is considered unlikely that changes in light levels would have significant effects on the key and associated species (where cirratulids acclimate to the changed light level), biotope resistance is assessed as ‘High’ and resilience as ‘High’ (by default) and the biotope is considered to be ‘Not sensitive’.

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion. Further detail

Evidence

No direct evidence was found to assess this pressure. As the larvae of Cerastoderma edule are planktonic and are transported by water movements, barriers that reduce the degree of tidal excursion may alter larval supply to suitable habitats from source populations. However the presence of barriers may enhance local population supply by preventing the loss of larvae from enclosed habitats.  Species that do not have a pelagic larval stage such as cirratulids or those that alternate between pelagic and benthic dispersal stages such as Pygospio elegans and Capitella capitata, are less likely to be impacted by this pressure. As both these key characterizing species are widely distributed and have larvae capable of long distance transport and long residence times in the water column, resistance to this pressure is assessed as 'High' and resilience as 'High' by default. This biotope is therefore considered to be 'Not sensitive'.

Death or injury by collision

Benchmark. Injury or mortality from collisions of biota with both static or moving structures due to 0.1% of tidal volume on an average tide, passing through an artificial structure. Further detail

Evidence

Not relevant.

Visual disturbance

Benchmark. The daily duration of transient visual cues exceeds 10% of the period of site occupancy by the feature. Further detail

Evidence

Not relevant.

Genetic modification & translocation of indigenous species

Benchmark. Translocation of indigenous species or the introduction of genetically modified or genetically different populations of indigenous species that may result in changes in the genetic structure of local populations, hybridization, or change in community structure. Further detail

Evidence

The key characterizing and associated species within this biotope are not cultivated or translocated. This pressure is therefore considered ‘Not relevant’ to this biotope group.

Due to long distance transport of pelagic larvae populations of the key characterizing species, Cerastoderma edule, may be interconnected and populations are not genetically isolated, with populations such as those at Pembroke showing mixing between British and Irish populations (Coscia et al., 2013). It should be noted that where local hydrodynamics prevent larval transport some genetically isolated populations may occur, as in the Burry Inlet, south Wales (Coscia et al., 2013).

Introduction or spread of invasive non-indigenous species

Benchmark. The introduction of one or more invasive non-indigenous species (INIS). Further detail

Evidence

The OSPAR (2009) background document identifies the threat to mudflats (considered to apply to muddy sand sediments) from INIS as follows: ‘Coastal and estuarine areas are among the most biologically invaded systems in the world, especially by molluscs such as the slipper limpet Crepidula fornicata and the Pacific oyster Magallana gigas. The two species have not only attained considerable biomasses from Scandinavian to Mediterranean countries but have also generated ecological consequences such as alterations of benthic habitats and communities, or food chain changes (OSPAR, 2009).

The Pacific oyster, Magallana (syn. Crassostrea) gigas, is native to warm temperate regions from the northwest Pacific to Japan and northeast Asia, including Cape Mariya (Russia) to Hong Kong (China) (Carrasco & Baron, 2010; GBNNSS, 2011, 2012). It is a fast-growing and tolerant species that has become a successful invader in the coastal waters of all continents, aside from Antarctica (Wrange et al., 2010; Carrasco & Baron, 2010; Padilla, 2010). Magallana gigas is recognised as a beneficial and important species in aquaculture worldwide (Padilla, 2010). It was initially introduced for aquaculture in Europe and the UK in the 1960s due to a decline in the Portuguese oyster (Crassostrea angulata) and the European flat oyster (Ostrea edulis) (Spencer et al., 1994; GBNNSS, 2011, 2012; Humphreys et al., 2014 cited in Alves et al., 2021; Hansen et al., 2023).

Since introduction, the species has invaded and established self-sustaining natural populations throughout Europe from the North Sea, Wadden Sea and Scandinavian coastlines to the Atlantic coastlines of Spain and Portugal, as well as the Mediterranean and Adriatic Sea (Wrange et al., 2010; GBNNSS, 2011, 2012; Ezgeta-Balic et al., 2019; Spagnolo et al., 2019; Bergstrom et al., 2021; Hansen et al., 2023). In the UK, the species predominantly occurs around the southern and western coastlines (OBIS, 2024; NBN, 2024). Shipping activity has also been associated with the introduction of Magallana gigas in the northeastern Adriatic Sea, where it was not introduced for aquaculture (Ezgeta-Balic et al., 2019). It was also suggested that some Magallana gigas populations were established in southwest England from France possibly via fouling on ships (GBNNSS, 2011, 2012; Padilla, 2010; Ezgeta-Balic et al., 2019).

Magallana gigas has a high fecundity, a long-lived pelagic larval phase (2 to 4 weeks) and can produce up to 200 million eggs during spawning (Herbert et al., 2012, 2016; Alves et al., 2021; Wood et al., 2021; Hansen et al., 2023). Hence, as a broadcast spawner, it has a high dispersal potential of more than 1000 km (Padilla, 2010; Wood et al., 2021). Larval mortality can be as large as 99%, as larvae are sensitive to environmental conditions (Alves et al., 2021). However, adults are long-lived so populations can survive with infrequent recruitment (Padilla, 2010). Larval dispersal and mass spawning events have facilitated the settlement and establishment of Pacific oysters, as seen in the Oosterschelde estuary, Netherlands (Hansen et al., 2023). It has been suggested that the spread of the Pacific oyster in Scandinavia is due to northward larval drift on tidal and wind-driven currents (Hansen et al., 2023). Wood et al. (2021) suggested that larval dispersal of the Pacific oyster from populations within and outside the UK was possible via unaided (passive) transport by currents, but that aquaculture and offshore structures (e.g. windfarms) increased the risk of the invasive species spreading and the geographical extent of spread.

Magallana gigas is an ecosystem engineer and can dramatically change habitat structure when it invades. Once successfully settled, groups of Pacific oysters may form dense aggregations, potentially forming a reef, which in some regions can reach densities of 700 individuals m² (Herbert et al., 2012, 2016). Once, the density of live or dead Pacific oysters reaches or exceeds 200 ind./m² little of the underlying substratum remains visible (Herbert et al., 2016). These reefs can stabilize the sediment surface locally (Troost, 2010). When such reefs are formed or, particularly when the species colonizes soft sediments such as mud or sand, it can change and affect local communities, by creating hard substrata for mobile species, which might not otherwise be present before the invasion (Padilla, 2010). However, Hansen et al. (2023) suggested that no immediate ecosystem risk is observed where the Pacific oyster occurs sporadically.

Magallana gigas requires hard substrata for successful settlement and establishment, including littoral rock, bedrock, chalk, bare boulders, cobbles and pebbles and shells (Kochmann, 2012; Kochmann et al., 2013; Mckinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). It also prefers mudflats with mixed sediment composed of shingle and sand, attaching to whatever hard substrata are available within otherwise unsuitable fine muddy sediment (Spencer et al., 1994; Mckinstry & Jensen, 2013; Tillin et al., 2020).

Magallana gigas has been reported from estuaries growing on intertidal mudflats, sandflats, and other soft sediments (Padilla, 2010; Herbert et al., 2016; Cabral et al., 2020). The settlement of spat on hard substrata within sediments has been observed in the estuaries of the River Dart, Exe, Fal, Fowey, Tamar, Teign, and Yealm in Devon and Cornwall, the Menai Straits, Wales and large estuaries of Lough Swilly, Lough Foyle and the Shannon in Ireland, and the Tagus Estuary in Portugal (Spencer et al., 1994; Kochmann, 2012; Kochmann et al., 2013; Cabral et al., 2020). In Lough Swilly, Lough Foyle and the Shannon, the Pacific oyster was often associated with intertidal mud or sandflats (Kochmann et al., 2013). In contrast, the Pacific oysters were absent from sandflat areas in Poole Harbour (Mckinstry & Jensen, 2013).

Although shorelines comprised of mainly mud were suggested to be unsuitable for spat settlement (Spencer et al., 1994), the presence of smaller hard substrata, such as shells or pebbles, can enable larvae to settle (Tillin et al., 2020). For example, in the River Teign estuary, Pacific oyster settlement was observed on shell-covered ground mainly attached to mussel shells, and occasionally attached to cockles, stones and common periwinkle (Littorina littorea) shells on a mud flat in the estuarine intertidal zone otherwise mainly comprised of sand and mud (Spencer et al., 1994). In addition, the Blue Lagoon on the north shore of Poole Harbour had the highest abundance of oysters on mud mixed with shingle and shell (Mckinstry & Jensen, 2013). Outside of the Blue Lagoon, oysters were also recorded on mixed substrata composed of mud, gravel, and shell (McKinstry & Jensen, 2013). Tillin et al. (2020) concluded that while successful invasions occurred on mudflats, Magallana gigas prefers mixed substrata. Fine mud sediments without hard substrata (such as small stones, gravel, and shell) are unlikely to be suitable (Tillin et al., 2020). The speed of Magallana gigas reef formation on soft substrata seems to be dependent on the amount of hard substrata present, developing quicker once there is a sufficient amount (Troost, 2010). Bergstrom et al. (2021) reported that the presence of Magallana gigas was partially dependent on increasing gravel content up to 15% but remained stable with increasing percentages (measured up to 80%).

In the Wadden Sea, the distribution of Magallana gigas on soft sediment shores can overlap with native bivalve species such as Cerastoderma edule, Macoma balthica and Scrobicularia plana (Troost, 2010; Herbert et al., 2012, 2016). However, these native species are likely to occur at higher shore elevations compared to the lower shore habitats preferred by the Pacific oyster (Troost, 2010; Herbert et al., 2012, 2016). For example, in the Wadden Sea, greater densities of Cerastoderma edule and Macoma balthica were found above the level of Magallana gigas reef development, with initial colonization of the oyster occurring on former mussel bed and cockle shell ridges (Herbert et al., 2012). Troost (2010) suggested that competition for space between Cerastoderma edule and Magallana gigas was likely to occur at locations unsuitable for the cockles. In addition, Pacific oysters have been observed settling on native cockle shell debris within Zostera noltei beds in South Devon and have been found on the lower shore in a cockle-dominated habitat (LS.LSa.MuSa.CerPo) in the Tamar-Tavy estuary (Morgan et al., 2021). 

The oyster reefs, in the Wadden Sea and Brittany, on littoral muddy and sandy habitats formed predominantly at lower tidal levels from Mean Low Water levels to the shallow subtidal (Troost, 2010; Herbert et al., 2012, 2016). Pacific oyster spatfall was recorded in the estuarine intertidal zone on areas with hard substrata of stone and shell, particularly between the low water of spring tides and high water of neap tides, such as in the Menai Strait (Spencer et al., 1994).

At high densities the Pacific oyster reef smothers sediment, provides hard substrata in an otherwise sedimentary environment with additional niches for colonization by other species that require hard substratum (e.g. barnacles), and changes surface roughness and local hydrography (Troost, 2010; Herbert et al., 2012, 2016; Tillin et al., 2020). Lejart & Hily (2011) found the surface available for epibenthic species in the Bay of Brest, increased 4-fold when oysters were present on mud, for every 1 m² of colonized substrata the oyster reef added 3.87 m² of surface area on mud sediment. An increase in available settlement substrata, free of epibiota, could be the reason oyster reefs see an increase in macrofaunal abundance. This can change the community composition and habitat structure in reefs on soft mud sediments, creating new habitats for an increasing abundance of infaunal and epibenthic mobile species (Kochmann et al., 2008; Lejart & Hily, 2011; Zwerschke et al., 2018). Results have shown 38% of species present in the oyster reefs on mud were characteristic of rocky substratum habitats (Lejart & Hily, 2011).

In the Bay of Brest, Pacific oyster reefs had a higher diversity and species richness than surrounding mud habitats, including the mud underneath the reefs, where the population was dominated by carnivores rather than suspension the feeders found on the mudflats (Lejart & Hily, 2011; Herbert et al., 2012). In addition, in muddy habitats around the UK, Ireland and Northern France, macrofaunal diversity increased as Pacific oyster density increased but epifaunal diversity decreased as oyster densities increased (Zwerschke et al., 2018). It was suggested that the decrease in epifaunal diversity was due to the decrease in settlement space and an increase in habitat fragmentation because of dense oyster assemblages (Zwerschke et al., 2018).

Green & Crowe (2014) examined the effects of Magallana gigas density in experimental plots (0.25 m²) in Lough Swilly and Lough Foyle, Ireland. The number of species and species diversity increased with oyster cover on mudflats, depending on site and duration. The assemblage also changed due to the increased abundance of barnacles and bryozoans on the oyster shells and polychaetes within the sediment (Green & Crowe, 2014). Zwerschke et al. (2020) suggested that Pacific oyster beds could replace the ecosystem services provided by native oysters, in areas where native oysters had been lost. Morgan et al. (2021) suggested that the smothering of sediment habitats could prevent fish and bird species from feeding on infauna like worms, molluscs, and crustaceans. Also, the development of tidepools within mixed Pacific oyster and blue mussel reefs in soft sediment intertidal sites has been observed in the Wadden Sea, which can create new microhabitats within the reefs (Weniger et al., 2022).

Pacific oysters have been found to reduce the proportion of fine particles and increase the proportion of large particles in the mud under the reef (Lejart & Hily, 2011). The evidence suggests that Pacific oyster reefs change sediment characteristics, by affecting nutrient cycling and increasing the organic content of sediment, sand-to-silt ratio and levels of porewater ammonium (Kochmann et al., 2008; Padilla, 2010; Wagner et al., 2012 cited in Tillin et al., 2020; Green & Crowe, 2014; Herbert et al., 2012, 2016; Zwerschke et al., 2020; Hansen et al., 2023).  Zwerschke et al. (2020) found no significant differences in nutrient cycling rates of native oyster beds or Magallana gigas beds or their associated benthic communities, in experimental plots in Ireland. Persistent changes in the rates of nutrient cycling were driven by the density and presence of oysters (Zwerschke et al., 2020).

The deposition of faeces and pseudo-faeces by Magallana gigas can increase the toxic levels of sulphide in sediments and associated hypoxic sediment conditions, which can reduce photosynthesis and growth in eelgrass (Kelly & Volpe, 2007). Faecal deposition and hypoxia have also been suggested to explain a reduction in species diversity in the sediment underlying high-density oyster reefs (Green & Crowe, 2013, 2014; Herbert et al., 2016). However, Lejart & Hily (2011) observed no organic or silt enrichment by Pacific oysters in mud beneath oyster reefs in the Bay of Brest, and no significant difference in the amount of organic matter found in the mud underneath oyster reefs and on bare mud not colonized by the oyster. The biodeposits excreted by the oyster may be washed away by powerful tides and currents seen in the Bay of Brest and the effects of organic enrichment at oyster reefs might be minimal due to wave action (Lejart & Hily, 2011).

The American slipper limpet Crepidula fornicata was introduced to the UK and Europe in the 1870s from the Atlantic coasts of North America with imports of the eastern oyster Crassostrea virginica. It was recorded in Liverpool in 1870 and the Essex coast in 1887-1890. It has spread through expansion and introductions along the full extent of the English Channel and into the European mainland (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Helmer et al., 2019; Hinz et al., 2011; McNeill et al., 2010; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015).

Crepidula fornicata is recorded from shallow, sheltered bays, lagoons and estuaries or the sheltered sides of islands, in variable salinity (18 to 40) although it prefers ca 30 (Tillin et al., 2020). Larvae require hard substrata for settlement. It prefers muddy gravelly, shell-rich, substrata that include gravel, or shells of other Crepidula, or other species e.g., oysters, and mussels. It is highly gregarious and seeks out adult shells for settlement, forming characteristic ‘stacks’ of adults. But it also recorded in a wide variety of habitats including clean sands, artificial substrata, Sabellaria alveolata reefs and areas subject to moderately strong tidal streams (Blanchard, 1997, 2009; Bohn et al., 2012, 2013a, 2013b, 2015; De Montaudouin et al., 2018; Hinz et al., 2011; Powell-Jennings & Calloway, 2018; Preston et al., 2020; Stiger-Pouvreau & Thouzeau, 2015; Tillin et al., 2020).

High densities of Crepidula fornicata cause ecological impacts on sedimentary habitats. The species can form dense carpets that can smother the seabed in shallow bays, changing and modifying the habitat structure. At high densities, the species physically smothers the sediment, and the resultant build-up of silt, pseudofaeces, and faeces is deposited and trapped within the bed (Tillin et al., 2020, Fitzgerald, 2007, Blanchard, 2009, Stiger-Pouvreau & Thouzeau, 2015). The biodeposition rates of Crepidula are extremely high and once deposited, form an anoxic mud, making the environment suitable for other species, including most infauna (Stiger-Pouvreau & Thouzeau, 2015, Blanchard, 2009). For example, in fine sands, the community is replaced by a reef of slipper limpets, that provide hard substrata for sessile suspension-feeders (e.g., sea squirts, tube worms and fixed shellfish), while mobile carnivorous microfauna occupy species between or within shells, resulting in a homogeneous Crepidula dominated habitat (Blanchard, 2009). Blanchard (2009) suggested the transition occurred and became irreversible at 50% cover of the limpet. De Montaudouin et al. (2018) suggested that homogenization occurred above a threshold of 20-50 Crepidula /m².

Impacts on the structure of benthic communities will depend on the type of habitat that Crepidula colonizes. De Montaudouin & Sauriau (1999) reported that in muddy sediment dominated by deposit-feeders, species richness, abundance and biomass increased in the presence of high densities of Crepidula (ca 562 to 4772 ind./m²), in the Bay of Marennes-Oléron, presumably because the Crepidula bed provided hard substrata in an otherwise sedimentary habitat. In medium sands, Crepidula density was moderate (330-1300 ind./m²) but there was no significant difference between communities in the presence of Crepidula. Intertidal coarse sediment was less suitable for Crepidula with only moderate or low abundances (11 ind./m²) and its presence did not affect the abundance or diversity of macrofauna. However, there was a higher abundance of suspension–feeders and mobile Crustacea in the absence of Crepidula (De Montaudouin & Sauriau, 1999). The presence of Crepidula as an ecosystem engineer has created a range of new niche habitats, reducing biodiversity as it modifies habitats (Fitzgerald, 2007). De Montaudouin et al. (1999) concluded that Crepidula did not influence macroinvertebrate diversity or density significantly under experimental conditions, on fine sands in Arcachon Bay, France. De Montaudouin et al. (2018) noted that the limpet reef increased the species diversity in the bed, but homogenised diversity compared to areas where the limpets were absent. In the Milford Haven Waterway (MHW), the highest densities of Crepidula were found in areas of sediment with hard substrata, e.g., mixed fine sediment with shell or gravel or both (grain sizes 16-256 mm) but, while Crepidula density increased as gravel cover increased in the subtidal, the reverse was found in the intertidal (Bohn et al., 2015). Bohn et al. (2015) suggested that high densities of Crepidula in high-energy environments were possible in the subtidal but not the intertidal, suggesting the availability of this substratum type is beneficial for its establishment. Hinz et al. (2011) reported a substantial increase in the occurrence of Crepidula off the Isle of Wight, between 1958 and 2006, at a depth of ca 60 m, on hard substrata (gravel, cobbles, and boulders), swept by strong tidal streams. Presumably, Crepidula is more tolerant of tidal flow than the oscillatory flow caused by wave action which may be less suitable (Tillin et al., 2020).

The availability of hard substrata (e.g., gravel) may only restrict initial colonization as higher densities of Crepidula function as substrata for subsequent colonization (Thieltges et al., 2004; Blanchard, 2009). However, Bohn et al. (2015) noted that Crepidula occurred at low density or was absent in areas of homogenous fine sediment and areas dominated by boulders. Bohn et al. (2015) suggested that wave action (exposure) probably prevented the establishment of large numbers of Crepidula in high-energy areas. Blanchard (2009) noted that sandy areas in the Bay of Saint-Mont Michel were not colonized by Crepidula because of surface sand mobility. Thieltges et al. (2003) also noted that storm events removed some clumps of mussels and presumably Crepidula onto tidal flats where they disappeared, which caused their abundance to fluctuate. Similarly, Crepidula was absent from sandy substrata in Swansea Bay but was most abundant in the shelter of the breakwater at the Swansea east site (Powell-Jennings & Calloway, 2018). Powell-Jennings & Calloway (2018) noted that Crepidula is killed by sudden burial and possibly burial due to deposition, which could mitigate Crepidula density. In addition, in the MHW, the highest densities of Crepidula were found in areas of sediment with hard substrata, e.g., mixed fine sediment with shell, or gravel or both but, while Crepidula density increased as gravel cover increased in the subtidal, the reverse was found in the intertidal (Bohn et al., 2015). However, gravel formed the base of most stacks of Crepidula in the intertidal, which suggested that initial colonization occurred on available hard substrata (i.e., gravel) in the absence of adult shells of Crepidula (Bohn et al., 2015).

Crepidula is recorded from the lower intertidal to ca 160 m in depth but it is most common in the shallow subtidal and low water springs (Blanchard, 1997; Thieltges et al., 2003; Bohn et al., 2012, 2015; Hinz et al., 2011; OBIS, 2023; Tillin et al., 2020). Bohn et al. (2012, 2013a, 2013b, 2015) suggested that extreme conditions in intertidal limited its upward distribution due to early post-settlement mortality. It reached its highest densities on the lower shore (below ca 0.7 m) and was absent from the high tidal level (ca 1.8 m) in the MHW (Bohn et al., 2015). Bohn et al. (2013b) noted that Crepidula spat in their experimental intertidal panels suffered high mortality of 78-100% during emersion by low water spring tides. Thieltges et al. (2003) noted that Crepidula abundance at the intertidal to the subtidal transition zone (ca 21/ m²) was significantly higher than in the upper, mid, and lower intertidal ca <3/ m²). Similarly, Diederich & Pechenik (2013) noted that Crepidula densities were not significantly different in the low intertidal (+0.2 m) and shallow subtidal (-1 m) but became lower at +0.4 and were absent above +0.6 m in Bissel Cove, Rhode Island where the mean high water was +1.38 m. They reported that intertidal adults experienced temperatures of ca 42°C, which were 15°C higher than subtidal adults. However, there was no significant difference in the tolerance of subtidal and intertidal adults with a lethal range of 33-37°C after three hours in the laboratory. Diederich & Pechenik (2013) suggested that adult Crepidula were living close to their upper thermal limit in Rhode Island and would be driven into the subtidal due to climate change. Diederich et al. (2015) reported that most juvenile Crepidula died after aerial exposure under laboratory conditions (20°C, 75% relative humidity), while adults from the intertidal and subtidal survived (26°C, 75% relative humidity). Franklin et al. (2023) noted that the body mass index of adult Crepidula did not decrease significantly in winter months in New Hampshire, USA, but did decrease in spring and summer, probably due to its investment in reproduction. 

The density of Crepidula populations in northern Europe (Germany, Denmark, and Norway) was significantly lower (ca <100/ m2) than in southern waters. Thieltges et al. (2004) reported that the population of Crepidula was affected strongly by cold winters in the Wadden Sea. The winters of 2001 and 2003 resulted in ca 56-64% mortality of intertidal Crepidula and up to 97% on one mussel bed, compared to only 11-14% in southern areas without frost. Crepidula almost vanished from the Wadden Sea after the 1978/79 winter and took ten years to recover due to moderate winters which regularly affected the population. Similarly, 25% mortality was observed in Crepidula populations on the south coast of the UK after the extreme 1962/63 winter (Crisp, 1964, Bohn et al., 2012). Thieltges et al. (2003) suggested that global warming may allow Crepidula populations to become more abundant in northern Europe. 

The Manila clam (Tapes philippinarium), which was introduced to Poole Harbour for aquaculture in 1998, has become a naturalised population on the intertidal mudflats (occurring at densities of 60 clams/m² in some locations within the harbour (Jensen et al. 2007, cited in Caldow et al. 2007). Densities of Cerastoderma edule and Abra tenuis had increased since the introduction of the Manila clam although the abundance of Scrobicularia plana and Macoma balthica declined (Caldow et al. 2005). The burrowing lifestyle of the characterizing cirratulids and other infaunal polychaetes may confer some protection from changes to the sediment surface and may provide some new habitat (as this species has been found among oyster banks).  The predatory veined whelk (Rapana venosa) and Hemigrapsus takinei are not established in the UK (although Hemigrapsus takinei has been recorded at two locations) could become significant predators of Cerastoderma edule and other species associated with the biotope in the future. 

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009). Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024). In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Michin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

Although a widespread invader, Didemnum vexillum has a limited ability for natural dispersal since the pelagic larvae remain in the water column for a short time (up to 36 hours). Therefore, it has a short dispersal phase that can allow the species to build localized populations (Herborg et al., 2009; Vercaemer et al., 2015; Holt, 2024). However, Bullard et al. (2007) suggested that Didemnum vexillum can form new colonies asexually by fragmentation. Colonies can produce long tendrils from an encrusting colony, which can fragment, disperse and settle, attaching to suitable hard substrata elsewhere (Bullard et al., 2007; Lambert, 2009; Stefaniak & Whitlatch, 2014). A fragmented colony can spread naturally for up to three weeks transported by ocean currents, attached to floating seaweed, seagrass or other floating biota, or as free-floating spherical colonies (Bullard et al., 2007; Lengyel et al., 2009; Stefaniak & Whitlatch, 2014; Holt, 2024). Fragments can reattach to suitable substrata within six hours of contact. Fragments have the potential to disperse around 20 km before reattachment (Lengyel et al., 2009). Valentine et al. (2007a) reported that colonies of Didemnum vexillum enlarged by 6 to 11 times by asexual budding after 15 days and enlarged 11 to 19 times after 30 days. Valentine et al. (2007a) concluded fragments could successfully grow, survive, and help to spread Didemnum vexillum.

While natural fragmentation of tendrils is thought to allow Didemnum vexillum to invade longer distances and increase its dispersal potential, Stefaniak & Whitlatch (2014) found that only one tendril out of 80 reattached to the flat, bare substrata used in their study, because tendrils required an extensive (at least eight hour) period of contact to reattach. Stefaniak & Whitlatch (2014) suggested that once fragmented from a colony, the success of tendril reattachment was limited, and reattachment was not a major contributor to the invasive success of Didemnum vexillum. However, Stefaniak & Whitlatch (2014) also found that larvae-packed tendril fragments may increase natural dispersal distance, reproduction, and invasive success of Didemnum vexillum, and increase the distance larvae can travel. Not all colonies produce tendrils at all locations.

Human-meditated transport via aquaculture facilities, boat hulls, commercial fishing vessels, and ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijkstra et al., 2007; Griffith et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009). For example, in oyster farms in British Columbia, large fragments of Didemnum sp. come off oyster strings when they are pulled out of water and other fragments can be pulled off oysters and mussels and thrown back into the water, which is likely to aid dispersal of the invasive species (Bullard et al., 2007). Dijkstra et al. (2007) hypothesised that Didemnum sp. was introduced to the Gulf of Maine with oyster aquaculture in the Damariscotta River and transported via Pacific oysters.

Didemnum vexillum was likely introduced into the UK from northern Europe or Ireland via poorly maintained or not antifouled vessels, movement of contaminated shellfish stock and aquaculture equipment, or via marine industries such as oil, gas, renewables, and dredging (Holt, 2024). Recent evidence from genetic material suggests that human-mediated dispersal, between marinas and shellfish culture sites, is the most likely pathway for connectivity of Didemnum vexillum populations throughout Ireland and Britain (Prentice et al., 2021; Holt, 2024). Didemnum vexillum can disperse away from artificial substrata, invading and colonizing natural substrata in surrounding areas (Tillin et al., 2020). Holt (2024) noted that Didemnum vexillum had not spread as far as feared in the UK since it was first recorded. The current evidence of Didemnum vexillum’s ability to spread on natural habitats in this area is sparse and often conflicting, complicated by genetics and its apparent variable habitat preferences and tolerances and its variable ability to adapt to ‘new’ conditions (Holt 2024).

The seasonal growth cycle of Didemnum vexillum is influenced by temperature (Valentine et al., 2007a). In warmer months (June and July) colonies may be large and well-developed encrusting mats. Populations experience more rapid growth from July to September sometimes continuing into December. Colonies begin to decline in health and ‘die-off’ when temperatures drop below 5°C during winter months from around October to April (Gittenberger, 2007; Valentine et al., 2007a; Herborg et al., 2009). Cold water months cause colonies to regress and reduce in size, yet they often regenerate as temperatures warm (Griffith et al., 2009; Kleeman, 2009, Mercer et al., 2009), although some populations may not survive winter at all (Dijkstra et al., 2007). The early growth phase, from May to July, is initiated by smaller colonies developing from remnants of colonies that survived the cold water (Valentine et al., 2007a). The seasonal growth cycle is also likely influenced by location. For example, the Didemnum sp. growth cycle for colonies in Sandwich tide pool (temperature range from -1 °C to 24 °C, with daily fluctuations), probably does not occur in deep offshore subtidal habitats in Georges Bank (annual temperature range from 4 °C to 15°C, and daily fluctuations are minimal) (Valentine et al., 2007a).  Larval release and recruitment typically occur between 14 to 20°C and slow or cease below 9 to 11°C as summer ends (Griffith et al., 2009; Mckenzie et al., 2017). In New Zealand, recruitment occurs from November to July, where the highest average temperatures were recorded in February (18 to 22°C) and the lowest average temperatures were recorded in July (9 to 10°C) (Fletcher et al., 2013a). In this New Zealand study, higher water temperatures were associated with a higher level of recruitment (Fletcher et al., 2013a).

Didemnum vexillum requires suitable hard substrata for successful settlement and the establishment of colonies. It can grow quickly and establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems because of its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock gravel, pebble, cobble, or boulders or artificial substrata such as a variety of maritime structures such as pontoons, docks, wood and metal pilings, chains, ropes and moorings, plastic and ship hulls and at aquaculture facilities (Valentine et al., 2007 a&b; Bullard et al., 2007; Griffith et al., 2009; Lambert, 2009; Tagliapietra et al., 2012; Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent, and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020).

Didemnum vexillum has the ability to rapidly overgrow and displace on other sessile organisms such as other colonial ascidians (Ciona intestinalis, Styela clava, Ascidiella aspera, Botrylloides violaceus, Botryllus schlosseri, Diplosoma listerianium and Aplidium spp.), bryozoan, hydroids, sponges (Clione celata and Halichrondria sp.), anemone (Diadumene cincta), calcareous tube worms, eelgrass (Zostera marina), kelp (Laminaria spp. and Agarum sp.), green algae (Codium fragile subsp. fragile), red algae (Plocamium, Chondrus crispus and bush weed Agardhiella subulata), brown algae (Ascophyllum nodosum, Sargassum, Halidrys, Fucus evanescens and Fucus serratus), calcareous algae (Corallina officinalis), mussels (Mytilus galloprovincialis, Perna canaliculus  and Mytilus edulis), barnacles, oysters (Magallana gigas, Ostrea edulis and Crassostrea virginica), sea scallops (Placopecten magellanicus), or dead shells (Dijkstra et al., 2007; Gittenberger, 2007; Valentine et al., 2007a; Valentine et al., 2007b; Griffith et al., 2009; Carman & Grunden, 2010; Dijkstra & Nolan, 2011; Groner et al., 2011; Hitchin, 2012; Tagliapietra et al., 2012; Minchin & Nunn, 2013; Gittenberger et al., 2015; Long & Groholz, 2015; Vercaemer et al., 2015).

Once established, Didemnum vexillum can expand rapidly, taking over most available hard substrata, which studies have hypothesized may alter species diversity and community composition and may decrease species abundance and biodiversity. Gittenberger (2007) stated that at this site, Didemnum sp. could cover around 95% of hard substrata locally leaving little space for recruitment and growth of other species. On Georges Bank, USA, Didemnum vexillum has altered the benthic community (Lengyel et al., 2009; Tillin et al., 2020). The pebble gravel substrata on Georges Bank is important to the success and survival of haddock (Melanogrammus aeglefinus) and Atlantic cod (Gadus morhua), and the settlement of sea scallop larvae (Placopecten magellanicus). Therefore, the invasion of Didemnum vexillum and its ability to change the habitat complexity of the seafloor, may in turn negatively impact the benthic community (Lengyel et al., 2009). In Georges Bank, Lengyel et al. (2009) analysed photographs of the seabed and suggested that Didemnum vexillum outcompeted other epifaunal and macrofaunal species. Benthic changes were seen in hydroids, the second most abundant epifaunal species at the location, which were overgrown by the invasive tunicate and negatively correlated with the percentage cover of Didemnum vexillum (Lengyel et al., 2009). The number of non-colonial macrofauna was also negatively related to the percentage cover of Didemnum vexillum (Lengyel et al., 2009). Dredge samples revealed clear differences in benthic species composition and revealed a significant difference in the species abundance before and after the colonization of Didemnum vexillum (Lengyel et al., 2009). Invasion of Didemnum vexillum also provided a new habitat for species not normally present, such as two polychaete species Nereis zonata and Harmothoe extenuate, changing the species composition. The increase in abundance of polychaetes Nereis zonata and Harmothoe extenuate were also seen in dredge samples collected from Georges Bank (Valentine et al., 2007b).

In contrast, some studies have suggested that potentially the overgrowth of Didemnum vexillum has little impact on benthic communities. In Long Island Sound, USA, Mercer et al. (2009) found the total abundance and richness of native epifaunal and infaunal species were either not different or significantly higher in samples taken inside Didemnum vexillum mats compared with samples collected outside the mats. While the mats did lead to subtle changes in community structure and shifts in species dominance, the authors suggested that benthic species may use Didemnum vexillum mats as a novel habitat and species living beneath the mats may use it for shelter and protection from epibenthic predators (Mercer et al., 2009). The predator protection could explain the high abundance of infaunal invertebrates found under the mats as well as the reduced abundance of crabs and demersal fish predators in areas dominated by Didemnum vexillum compared to uncolonized areas (Mercer et al., 2009). In addition, dredge samples taken from Georges Bank found 15 polychaete species and seven bivalve species living beneath the Didemnum vexillum mat (Valentine et al., 2007b). The comparisons of 85 benthic megafauna collected from dredge samples before and after Didemnum sp. became abundant, in Georges Bank fishing ground, showed slight changes in abundance but changes to the invertebrate species composition were statistically marginally insignificant (Valentine et al., 2007b).

In contrast to Didemnum vexillum’s preference for sheltered conditions, established colonies observed in Georges Bank and Long Island Sound were exposed to moderately strong tidal currents (1 to 2 knots; ca 0.5 to 1 m/s recorded at both sites) that may mobilise sediment (Valentine et al., 2007b; Mercer et al., 2009; Tillin et al., 2020). However, Valentine et al. (2007b) describe the substratum as immobile, presumably consolidated gravel, cobbles and pebbles. Kleeman (2009), stated that the presence of a consistent mild wave action or ‘swash zone’ appears to favour Didemnum sp. establishment in the intertidal. Although some evidence suggests that waves and currents can facilitate the fragmentation and spread of Didemnum vexillum (Mckenzie et al., 2017), the tidal current velocities at some sites where Didemnum vexillum has been reported (for example, New England, where current velocities reach up to around 3 m/s) is lower than the current velocity required for the dislodgement of Didemnum vexillum fragments (around 7.6 m/s) (Reinhardt et al., 2012). This suggests that not all tidal currents are likely to dislodge Didemnum vexillum fragments. When on boat hulls the species can experience higher current velocities which is enough to cause dislodgement (Reinhardt et al., 2012). 

The Sandwich tide pools (USA) were subject to air exposure at low tide, and daily changes in water depth and temperatures (Valentine et al., 2007a). Didemnum vexillum colonies survived exposure to air at low tides for a short time (not exceeding two hours) during rapid colony growth in the summer months of July to September (Valentine et al., 2007a). However, parts of the large established colonies, which were artificially exposed to air for two to three hours in October, were observed desiccated or predated on by grazing periwinkles 30 days later, in the winter month of November (Valentine et al., 2007a). They suggested that the invasive tunicates’ ability to tolerate exposure to air varies with the seasonal growth cycle. Didemnum vexillum also tolerated emersion in Kent, as colonies on the mid-shore at Reculver flourish and survive in air exposure for up to three hours per cycle during springs (Hitchin, 2012). Hitchin (2012) suggested the porous nature of the sandstone boulders the species colonized retained water. The Kent shore was sheltered but held water due to its shallow slope and flats, which may allow Didemnum sp. to survive in the low to mid-shore. There is evidence that Didemnum vexillum died when exposed to air for more than 6 hours (Laing et al., 2010).

Sensitivity assessment. Intertidal gravelly sandy muds may be exposed to invasive species which can alter the character of the habitat (primarily Crepidula fornicata at the sublittoral fringe and Magallana gigas), leading to re-classification of this biotope. The above evidence suggests that Crepidula fornicata could colonize mixed sediment environments, typical of this biotope, due to the presence of gravel, pebbles, cobbles or any other hard substrata that can be used for larvae settlement (Tillin et al., 2020). Therefore, Crepidula has the potential to colonize, and modify the habitat and its associated community due to the introduction of Crepidula shell biomass, silt, pseudofaeces and faeces (Blanchard, 2009; Tillin et al., 2020). Bohn et al. (2015) demonstrated that Crepidula had a preference for gravelly habitats, while De Montaudouin & Sauriau (1999) and Bohn et al. (2015) noted that Crepidula densities were low in intertidal coarse sediments. This habitat is very sheltered to extremely sheltered from wave action, which is also suitable for Crepidula colonization. There may be higher densities of Crepidula in the lower shore examples of the biotope, but the densities may be lower in the mid-shore, and it may be absent from the upper shore due to the unsuitable extreme conditions in the intertidal zone preventing Crepidula post-settlement recruitment and mitigating colonization (Bohn et al. 2015). Therefore, resistance to colonization by Crepidula fornicata is assessed as ‘Low’ and resilience is assessed as ‘Very low’, so the biotope sensitivity is assessed as ‘High’. The confidence in the assessment is 'Low' because the sensitivity of this biotope to Crepidula is potentially site-specific, there is a risk of its introduction by artificial means, and the is a lack of direct evidence of Crepidula being reported to occur in the biotope. 

The above evidence suggests that this biotope could be suitable for the colonization of Magallana gigas due to the presence of gravel and other hard substrata required for successful settlement and establishment (Kochmann, 2012; Kochmann et al., 2013; Mckinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). The majority of records (ca 87%) of this biotope occur on the lower shore and Magallana gigas populations have a preference for the lower shore and shallow subtidal (ca 10 m), suggesting the oyster could colonize the lower shore extent or examples of the biotope in high densities. The distribution of Magallana gigas can overlap with Cerastoderma edule and other native bivalve species (Troost, 2010; Herbert et al., 2012, 2016), However, Cerastoderma edule is generally found at higher shore elevations compared to the lower shore habitats preferred by the Pacific oyster, with greater densities of Cerastoderma edule found above the level of Magallana gigas reef development (Troost, 2010; Herbert et al., 2012, 2016). Therefore, resistance to colonization by Magallana gigas is assessed as 'Low'. Resilience is assessed as 'Very low' as the Magallana gigas population would need to be removed for recovery to occur. Hence, sensitivity is assessed as 'High'. The confidence in the assessment is 'Low' because the sensitivity of this biotope to Magallana gigas is potentially site-specific.

There is no evidence of Didmenum vexillum colonizing this biotope in the UK. Didemnum vexillum requires hard substrata for successful colonization, therefore, it could colonize the mixed sediments typical of this biotope. Didemnum vexillum has a preference for wave-sheltered conditions and it has been recorded in the lower intertidal. There is no evidence of Didemnum vexillum interacting with cockles but it is known to overgrow and displace other bivalve species, such as mussels (Mytilus galloprovincialis, Perna canaliculus and Mytilus edulis), barnacles, oysters (Magallana gigas, Ostrea edulis and Crassostrea virginica), and sea scallops (Placopecten magellanicus). If Didemnum vexillum was able to colonize and form extensive mats in this biotope it may cause minor changes to the community composition as seen in George Bank. Valentine et al. (2007b) found slight changes in megafauna abundance before and after Didemnum sp. became abundant in Georges Bank, but changes to the invertebrate species composition were statistically marginally insignificant. Therefore, a resistance of 'Medium' (some mortality, <25%) is suggested as a precaution. Resilience is likely to be 'Very low' as Didemnum vexillum would need to be physically removed to allow recovery. Hence, sensitivity to invasion by Didemnum is assessed as 'Medium'. 

Introduction of microbial pathogens

Benchmark. The introduction of relevant microbial pathogens or metazoan disease vectors to an area where they are currently not present (e.g. Martelia refringens and Bonamia, Avian influenza virus, viral Haemorrhagic Septicaemia virus). Further detail

Evidence

A recent review of parasites, pathogens and commensals identified a range of agents impacting European cockles, including 50 conditions affecting Cerastoderma edule (Longshaw & Malham, 2013). Cockles are hosts to viruses, bacteria, fungi, Apicomplexa , Amoeba, Ciliophora, Perkinsozoa, Haplosporidia, Cercozoa, Turbellaria, Digenea, Nematoda, Crustacea and Nemertea.  Mortalities are associated particularly with digeneans and some protistan infections; parasites may limit growth, reduce fecundity and alter burrowing behaviour (Longshaw & Malham, 2013). A number of examples of conditions associated with mass mortalities of Cerastoderma edule are presented below.  Parasites and disease are more likely to cause mortalities in populations that are subject to subptimal conditions or other stressors such as hot summers or cold winters (Longshaw & Malham, 2013).  Experimental infection of Cerastoderma edule with a trematode parasite showed that effects differed depending on habitat conditions (Wegeberg & Jensen, 2003). Infected Cerastoderma edule reared in sub-optimal conditions lost more body weight than infected cockles in more optimal habitats and did not regain condition when placed in higher shore habitats where immersion and food supply was limited.  Infected cockles placed on lower shore sites with longer emersion times regained condition despite the infection and were equivalent to controls. The impact of trematodes is therefore mediated by habitat conditions and in some instances may have no effect (Wegeberg & Jensen, 2003).

Infestation by a trematode parasite Cercaria cerastodermae impairs the burrowing ability of Cerastoderma edule and was identified as the likely cause of a mass mortality of cockles in Scandinavian waters in 1991 (Jonsson & André, 1992). Another trematode parasite Gymnophallis choledochus may castrate Cerastoderma edule, reducing reproduction and recruitment and indirectly leading to population declines (Thieltges, 2006). Boyden (1972) reported castration of 13% of the cockle population in the River Couch estuary due to infestation with larval digenetic trematodes.

An unidentified amoeba, measuring 18–20 mm in diameter, was described from the sub-epithelial gill tissues of Cerastoderma edule from Portugal (Azevedo, 1997). The amoeba was associated with haemocytic infiltration and necrosis of host cells. Affected cockles were found gaping at the surface and the infection was considered to be responsible for mass mortalities.

Cockles also suffer from disseminated neoplasia-a leukaemia like disease associated with mass mortalities. Cerastoderma edule from Ireland have been reported to be especially susceptible (Barber et al., 2004). High mortalities of cockles observed in northwest Spain in 1997 were associated with a higher prevalence (up to 84%) of disseminated neoplasia compared to control areas (4% prevalence) not experiencing mortality (Villalba et al. 2001).

Other species characterizing the biotope may be infected by microbial pathogens and parasites. Nearly all Aphelochaeta marioni (as Tharyx marioni individuals from Stonehouse Pool in Plymouth were infected with a sporozoan parasite of the Gonospora genus but no evidence was found that the animal was adversely affected by its presence (Gibbs, 1971).

Sensitivity assessment. The available evidence suggests that the key characterizing species Cerastoderma edule is susceptible to a range of pathogens and parasites. The effects of these may be exacerbated by stressors such as  thermal stress (amongst others). As evidence exists for mass mortalities, resistance is assessed as ‘Low’ and resilience is assessed as ‘Medium’. Sensitivity is therefore categorised as ‘Medium’.

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

The sedimentary biotope and characterizing and associated species may be disturbed and damaged by static or mobile gears that are targeting Cerastoderma edule or other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of target species on this biotope. Dense populations of Cerastoderma edule on intertidal flats support commercial fisheries in several areas of the UK and the species is also harvested throughout Europe. Harvesting may use mechanical methods (e.g. tractor dredges or hydraulic suction dredging) or hand collection using rakes or other methods. The fishery is managed through local byelaws and local target size classes are set (usually cockles below 20 mm are not collected). The commercial importance of the fishery has stimulated research on impacts on cockles and the sedimentary habitat and associated species. In some habitats removal of Arenicola marina and Hediste diversicolor for fishing bait may occur but these species are not considered to characterize this biotope and the removal of these as target species is not considered in this review; the relevant biotopes (LS.LSa.MuSa.LimAre; SS.SMu.IFiMu.Are; LS.LMu.MEst.HedMac), characterized by these species contain more information.

The physical effects of harvesting on this species are addressed in the physical disturbance sections. Removal of Cerastoderma edule (cockles) by targeted harvesting may result in an altered community and may alter the character and reduce the spatial extent of the Cerastoderma edule and polychaetes in littoral muddy sand biotope.

The method of harvesting cockles will influence the proportions that are removed and damaged. Pickett (1973) found that intense dredging for a short period on a bed of cockle spat had little effect on survival and growth although Cook (1991) found that impacts on small cockles from dredging were variable, with little reduction one year but a reduction in density observed the following year. Cotter et al. (1997) assessed the catch rates and damage and mortality of Cerastoderma edule resulting from experimental tractor dredging at the Burry Inlet (Wales). Stocks of adult cockles were reduced by 31 and 49% in low and high density areas respectively. Similarly mechanical cockle harvesting in muddy sand reduced the abundance of Cerastoderma edule by ca 34%. Populations had not recovered their original abundance after 174 days (Ferns et al. 2000).

Following size sorting (either mechanically or by hand), undersized cockles are deposited on the sediment surface. Damage rates and survival rates of harvested and discarded cockles and rates of reburrowing and displacement have been examined in a number of studies.  Cook (1991) reported overall damage rates of 11-14% of rejects from rotary riddles on three hydraulic section dredgers operating. Undersized and rejected cockles may be stunned where these suffer prolonged vibrations from passage through mechanical gear and sorters, this can delay reburrowing, leading to increased predation and/or distribution by tidal waves and currents (Coffen-Smout & Rees, 1998).

The sediment on which discards are deposited affects burrowing. Experimental displacement to stimulate harvesting impacts, found that cockles deposited in pools are more able to rebury while none of those deposited on drained (and hence hard) sands were able to reburrow (Coffen-Smout & Rees, 1999). Greater proportions of smaller cockles than medium or large reburrow, so that larger cockles are more likely to be displaced by tides (Coffen-Smnout & Rees, 1999). Cockles that were transported up to 200m on the flood tide could reburrow if habitats in the new position were suitable (Coffen-Smout & Rees, 1999).

Hand raking for cockles on intertidal silty sandflats, using rakes that penetrated the surface by 5-10cm, resulted in a three-fold increase in the damage rate of cockles compared to control plots and, in the short-term, led to a relative decrease in the overall abundance of fauna (Kaiser et al., 2001). After 56 days the small (9 m²) plots had recovered but the larger (36 m²) plots remained in an altered state. Results collected over a year after the disturbance suggested that while effects of hand-raking may be significant within a year, they are unlikely to persist beyond this time-scale unless there are larger long-lived species present within the community (Kaiser et al., 2001).

The presence of dense cockle beds inhibit the establishment of other benthic species through space and resource competition, disturbance (Flach, 1996) and consumption of larvae (Andre et al., 2003) and changes in sediment characteristics (). Removal of adult cockles by harvesting or other factors allows other species to establish. Following experimental removal of large adult Cerastoderma edule by Frid & Casear (2012) sediments showed increased biodiversity and assemblages dominated by traits common to opportunist taxa at a species-poor shore at Warton Sands, Morecambe Bay, and a more diverse shore at Thurstaston, Dee estuary. The movements of cockles disturb and exclude the amphipod Corophium volutator and other species (Flach, 1996; Flach & de Bruin, 1994) the removal of cockles may therefore allow this species to colonize intertidal flats. During periods of low cockle density, Desprez et al., 91992) observed that Pygospio elegans established dense populations; when cockles returned these were lost within one year.

It should be noted that removal of Cerastoderma edule by targeted harvesting may lead to wider ecological effects through starvation of shore birds in winter. This has been observed in the Dutch Wadden Sea (Smit et al., 1998). This effect is not directly of significance to this biotope and is not considered within the assessment.

Sensitivity assessment. Removal of the key characterizing species Cerastoderma edule by targeted harvesting would alter the character of the biotope and result in reclassification. The abundance of other soft-sediment infauna (particularly opportunist species such as Pygospio elegans and Capitella capitata) may increase in disturbed patches in the short-term as a result of removal of cockles resulting in reduced competition for space and predation (on larvae). Where sediments remain suitable cockles are likely to recolonize via adult migration, survival of small, discarded cockles or via larval recruitment.  In general fishing practices will be efficient at removing this species, resistance is therefore assessed as ‘Low’ (removal is not considered to be total as smaller individuals are not retained and harvesting is unlikely to be 100% efficient at removing larger cockles). Resilience is assessed as ‘Medium’, so that sensitivity is assessed as ‘Medium’. Recovery will be influenced by a range of factors as outlined in the resilience section. Small patches are likely to be in-filled by adult cockle movement, large patches will recover through larval recruitment, which again is subject to many factors, and may be improved by the removal of adult cockles. However, as Cerastoderma edule recruitment is episodic, a recovery of ‘Medium’ to represent recovery of age-classes from broad-scale removal was considered appropriate. Biotope sensitivity is therefore assessed as ‘Medium’. Although some experiments have shown rapid recovery, the plots used in experiments are small and subject to low levels of harvesting compared to intertidal flats that are harvested at larger scales and where patches may be re-worked over a season.

Removal of non-target species

Benchmark. Removal of features or incidental non-targeted catch (by-catch) through targeted fishery, shellfishery or harvesting at a commercial or recreational scale. Further detail

Evidence

The sedimentary biotope and characterizing and associated species may be disturbed and damaged by static or mobile gears that are targeting other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. A number of studies have assessed the impact of Cerastoderma edule removal on non-target infauna (Hall & Harding, 1998; Hiddink, 2003). However, these studies are relevant to physical damage and are discussed in those sections. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of non-target species on this biotope. Dense beds of Cerastoderma edule occur in situations where the faunal assemblage is species-diverse and productive, but also where very few taxa are present (Cesar, 2012). These observations suggest that Cerastoderma edule populations are not dependent on other invertebrate species and are therefore unlikely to be impacted by ecological/biological effects from removal of other species.  The removal of predators such as shrimp and crab may enhance recruitment and survival of larvae (Beukema & Dekker, 2005; Sanchez-Salazar et al., 1987). The physical effects of removal of other species such as polychaete worms targeted by bait diggers may, however, impact Cerastoderma edule and other species associated with this biotope, through direct damage, smothering (Jackson & James, 1979) and removal and displacement. These direct effects of sediment disturbance are assessed in the physical damage sections.

The removal of Cerastoderma edule (as by-catch from another fishery) and other associated species would alter the biotope from the description and change community structure (diversity, biomass and abundance), potentially altering ecosystem function and the delivery of ecosystem goods and services (including the supply of food to fish and birds).

Sensitivity assessment. The assessment considers whether the removal of characterizing and associated species as by-catch would impact the biotope. Lethal damage to and removal of Cerastoderma edule and other species as by-catch would alter the character of the biotope through changes in the structure of the biological assemblage (changes to species richness, abundance and biomass). As Cerastoderma edule  and other species are either sedentary or incapable of rapid evasive movements, biotope resistance is assessed as ‘Low’. Resilience is assessed as ‘Medium’ based on Cerastoderma edule and sensitivity is therefore categorized as ‘Medium’.  Physical damage to the sediment and other physical damage factors are considered in the abrasion and extraction pressures