MarLIN

information on the biology of species and the ecology of habitats found around the coasts and seas of the British Isles

Hediste diversicolor and Limecola balthica in littoral sandy mud

08-11-2016

Summary

UK and Ireland classification

Description

Mainly mid and lower shore sandy mud or mud in lower estuaries, sheltered bays and marine inlets, often subject to variable salinity. The main characterizing species are the ragworm Hediste diversicolor, the baltic tellin Limecola balthica, and the oligochaetes Tubificoides benedii and Tubificoides pseudogaster. Further polychaetes that are often common or abundant include Pygospio elegans, Streblospio shrubsolii, Tharyx killariensis, Aphelochaeta marioni, Capitella capitata and Manayunkia aestuarina. The oligochaete Heterochaeta costata and the mud shrimp Corophium volutator may be abundant. The spire shell Hydrobia ulvae is often common. Other species which occur in a significant proportion of samples include the polychaetes Eteone longa and Nephtys hombergii, and bivalves such as the cockle Cerastoderma edule and Abra tenuis. The sand gaper Mya arenaria is superabundant in about a quarter of the samples for this biotope. Mya arenaria is probably present in a higher proportion of areas of this biotope but may be missed in core samples due to its size. HedLim may occur on the mid/lower shore of lower estuarine shores, with HedLimEte or LimAre on the upper shore. HedLimScr, Hed.Str, NhomAph, and Hed.Cvol may be present on the same shore. Ulva spp. may form mats on the surface of the mud during the summer months, particularly in areas of nutrient enrichmnent. (Infromation from Connor et al., 2004; JNCC 2015).

Depth range

Lower shore, Mid shore

Additional information

-

Listed By

- none -

Further information sources

Search on:

Habitat review

Ecology

Ecological and functional relationships

  • The dominant and characterizing species in the biotope (Limecola balthica and Hediste diversicolor) are infaunal and display plasticity in their feeding methods (McLusky & Elliott, 1981; Nielsen et al., 1995). They are primarily deposit feeders but are able to switch to suspension feeding when conditions allow. For example, in the case of Hediste diversicolor, if phytoplankton concentrations are sufficiently high the species switches from predatory and surface deposit-feeding to suspension feeding. This behaviour is maintained as long as the phytoplankton concentration remains above a 'trigger level' of 1-3 µg chlorophyll a per litre (reviewed by Riisgard, 1994).
  • Obligate deposit feeders are also numerous in the biotope, e.g. Pygospio elegans, Aphelochaeta marioni, Arenicola marina and Hydrobia ulvae. Competition for resources is likely to occur between the deposit feeders. For example, densities of the amphipod, Corophium volutator, and the gastropod, Hydrobia ulvae, are strongly negatively correlated although the mechanism of the interaction is not well understood (Barnes & Hughes, 1992).
  • Furthermore, Corophium volutator and the infaunal annelid species in the biotope probably interfere strongly with each other. Adult worms probably reduce amphipod numbers by disturbing their burrows, while high densities of amphipods can prevent establishment of worms by consuming larvae and juveniles (Olafsson & Persson, 1986). Arenicola marina has been shown to have a strong negative effect on Corophium volutator due to reworking of sediment causing the amphipod to emigrate (Flach, 1992), and also on Pygospio elegans (Reise, 1985).
  • Suspension feeding bivalves , e.g. Cerastoderma edule, Mya arenaria and Abra tenuis, may occur where hydrodynamic conditions allow, i.e. in areas of stronger flow with coarser sediments.
  • Important epibenthic predators in the biotope include the shore crab, Carcinus maenas, and shrimps, e.g. Crangon crangon, which take infaunal populations of small bivalves, polychaetes and crustaceans (Elliot et al., 1998). Carcinus maenas has been shown to significantly reduce the numbers of Manayunkia aestuarina on mudflats (McClusky, 1989) and along with Crangon crangon may reduce the population of Corophium volutator in estuaries by more than 50% (Pihl, 1985).
  • Intertidal mudflats are important feeding areas for many fish species. The most significant predators are probably the flatfish, including sole, Solea solea, dab, Limanda limanda, flounder, Platichthys flesus, and plaice, Pleuronectes platessa, which feed on polychaetes and their tails (e.g. of Hediste diversicolor and Arenicola marina), bivalve young and siphons (e.g. of Limecola balthica) and crustaceans (see review by Elliot et al., 1998).
  • Mobile epibenthic predators are probably responsible for preventing the infauna reaching carrying capacity. For example, exclusion experiments by Reise (1985) on intertidal mudflats showed small fish and decapod crustaceans can significantly reduce numbers of Cerastoderma edule, Tubificoides benedii, Pygospio elegans, Aphelochaeta marioni, Eteone longa and Corophium volutator.
  • Carnivorous annelids, e.g. Eteone longa and Nephtys hombergii, operate at the trophic level below Carcinus maenas (Reise, 1985). They predate the smaller annelids and crustaceans in the biotope.
  • Macroalgae, such as Ulva lactuca and Ulva sp., probably only occur in the biotope where suitable hard substrata exist for attachment. However, Ulva lactuca may grow free floating in very sheltered conditions.

Seasonal and longer term change

Seasonal changes are likely to occur in the abundance of fauna in the biotope due to seasonal recruitment processes and variations in recruitment success. For example, in the case of Limecola balthica, Bonsdorff et al. (1995) reported juvenile density in the Baltic Sea following settlement in late summer to be 300,000/m² decreasing to a stable adult density of 1,000/m², and Ratcliffe et al. (1981) reported adult densities in the Humber Estuary, UK, between 5,000/m² and 40,000/m² depending on time since a successful spat fall. Variation in abundance is also very pronounced in the polychaete Aphelochaeta marioni. For example, in the Wadden Sea, peak abundance occurred in January (71,200 individuals per m²) and minimum abundance occurred in July (22,500 individuals per m²) following maximum spawning activity between May and July (Farke, 1979). However, the spawning period varies according to environmental conditions and so peak abundances will not necessarily occur at the same time each year. For example, Gibbs (1971) reported Aphelochaeta marioni spawning in late autumn in Stonehouse Pool, Plymouth Sound.
Some species make seasonal migrations in response to environmental conditions. For example, Beukema & De Vlas (1979) reported that 30% of the Limecola balthica population migrated into the subtidal during winter apparently in response to low temperatures. Migration is achieved by burrowing (Bonsdorff, 1984; Guenther, 1991) and/or floating (Sörlin, 1988) .
Macroalgal cover typically varies through the year due to changes in temperature and light availability. The green macroalgae in the biotope are likely to proliferate in the spring and summer and die back in the autumn and winter in conjunction with decreasing light levels and temperature and increased disturbance by storm events. Production by microphytobenthos and microalgae is also likely to be higher in spring and summer, increasing food availability for deposit feeders and suspension feeders in the biotope.
The biotope is likely to be susceptible to increased wave action during storms, particularly in winter. Storms may result in changes in sediment composition and washing out of infauna, leaving the biotope available for recolonization (see review by Hall, 1994). The infauna may be damaged by wave action, displaced from their preferred habitat and/or cast ashore, resulting in mortality. For example, Tamaki (1987) studied the passive transport by waves and tidal currents of the adults of 5 polychaete species. One species exhibited a landward shift in its centre of population during winter when the wave effects were most profound, and reoccupied its summer position by active migration of adults.

Habitat structure and complexity

  • The substratum in the biotope is uniform sediment with little structural diversity provided by either physiographic features or the biota. Some 3-dimensional structure is provided by the burrows of infauna such as Arenicola marina. Most species living within the sediment are limited to the area above the anoxic layer, the depth of which will vary depending on sediment particle size and organic content. However, the presence of burrows allows a larger surface area of sediment to become oxygenated, and thus enhances the survival of a considerable variety of small species (Pearson & Rosenberg, 1978).
  • Reworking of sediments by deposit feeders increases bioturbation and potentially causes a change in the substratum characteristics and the associated community (e.g. Rhoads & Young, 1970). For example, Widdows et al. (1998) reported that typical abundances of Limecola balthica increased sediment resuspension and/or erodability four fold and that there was a significant positive correlation between density of the species and sediment resuspension.
  • Where present, the macroalgae provide some structural complexity in the habitat, providing cover and sites for attachment of epifauna and epiphytes.

Productivity

  • Primary production in the biotope comes from benthic microalgae and water column phytoplankton (Elliott et al., 1998).
  • Production by benthic unicellular and filamentous algae in the littoral zone accounts for 0.2-1.3 g C/m² daily, depending on water clarity (Barnes & Hughes, 1992). Benthic microalgae are able to photosynthesize over a much wider range of light intensity than the planktonic species. There is little or no photoinhibition, adapting littoral species to the full sunlight experienced at low tide while still utilizing the very low light intensities at high tide (Barnes & Hughes, 1992).
  • Where present, macroalgae also contribute to primary production in the biotope. They exude considerable amounts of dissolved organic carbon which are taken up readily by bacteria and possibly by some larger invertebrates.
  • The majority of nutrients in the biotope are derived from allochthonous sources. Mudflats receive large inputs of nutrients, sediment and organic matter from the sea and land discharges of river water and sewage, resulting in high productivity despite low levels of primary production (Elliott et al., 1998).

Recruitment processes

The bivalves which characterize the biotope are capable of high recruitment and rapid recovery. For example, adult Limecola balthica spawn at least once a year and are highly fecund (Caddy, 1967). There is a planktotrophic larval phase which lasts up to 2 months (Fish & Fish, 1996) and so dispersal over long distances is potentially possible given a suitable hydrographic regime. Following settlement, development is rapid and sexual maturity is attained within 2 years (Gilbert, 1978; Harvey & Vincent, 1989). In addition to larval dispersal, dispersal of juveniles and adults occurs via burrowing (Bonsdorff, 1984; Guenther, 1991), floating (Sörlin, 1988) and probably via bedload transport (Emerson & Grant, 1991). It is expected therefore that recruitment can occur from both local and distant populations.
The infaunal polychaetes Hediste diversicolor, Arenicola marina and Aphelochaeta marioni have high fecundity and the eggs develop lecithotrophically within the sediment or at the sediment surface (Farke, 1979; Beukema & de Vlas, 1979). There is no pelagic larval phase and the juveniles disperse by burrowing. Recruitment must occur from local populations or by longer distance dispersal of postlarvae in water currents or during periods of bedload transport. For example, Davey & George (1986), found evidence that larvae of Hediste diversicolor were tidally dispersed within the Tamar Estuary over a distance of 3 km, as larvae were found on an intertidal mudflat which previously lacked a resident population of adults. Recruitment is therefore likely to be predictable if local populations exist but patchy and sporadic otherwise.
The deposit feeding gastropod Hydrobia ulvae appears to display plasticity in its developmental mechanism. Fish & Fish (1996) report planktotrophic development with a free-swimming larval phase lasting 3 weeks, while Pilkington (1971) stated that development occurred via a non-feeding benthic larvae which metamorphosed in just 3 days. It is possible that Hydrobia ulvae is able to change its developmental mechanism according to environmental conditions. If conditions are favourable, the eggs may hatch, develop directly and recruit locally. In more stressful conditions, it may benefit the individual to disperse its offspring more widely via a planktotrophic larva.
Recruitment of shallow burrowing infaunal species can depend on adult movement by bedload sediment transport and not just spat settlement and juvenile dispersal. Emerson & Grant (1991) investigated recruitment in Mya arenaria and found that bedload transport was positively correlated with clam transport. They concluded that clam transport at a high energy site accounted for large changes in clam density. Furthermore, clam transport was not restricted to storm events and the significance is not restricted to Mya arenaria recruitment. Many infauna, e.g. polychaetes, gastropods, nematodes and other bivalves, will be susceptible to movement of their substratum.

Time for community to reach maturity

No information was found concerning time taken for the community to reach maturity. However, the characterizing species are highly fecund and quick to grow and mature and so the community would be expected to reach maturity within 5 years.

Additional information

-

Preferences & Distribution

Habitat preferences

Depth Range Lower shore, Mid shore
Water clarity preferences
Limiting Nutrients No information found
Salinity Full (30-40 psu), Variable (18-40 psu)
Physiographic Enclosed coast / Embayment
Biological Zone Eulittoral
Substratum Mud, Sandy mud
Tidal
Wave Extremely sheltered, Sheltered, Very sheltered
Other preferences

Additional Information

Species composition

Species found especially in this biotope

    Rare or scarce species associated with this biotope

    -

    Additional information

    Sensitivity review

    Sensitivity characteristics of the habitat and relevant characteristic species

    Connor et al. (2004) summarise the biotope as being located mainly in mid and lower shore sandy mud or mud in lower estuaries, sheltered bays and marine inlets, often subject to variable salinity. The main characterizing species are the ragworm Hediste diversicolor, the baltic tellin Macoma balthica, and the oligochaetes Tubificoides benedii and Tubificoides pseudogaster. Further polychaetes that are often common or abundant include Pygospio elegansStreblospio shrubsoliiTharyx killariensisAphelochaeta marioni, Capitella capitata and Manayunkia aestuarina. The oligochaete Heterochaeta costata and the mud shrimp Corophium volutator may be abundant. The spire shell Hydrobia ulvae is often common. Other species which occur in a significant proportion of samples include the polychaetes Eteone longa and Nephtys hombergii, and bivalves such as the cockle Cerastoderma edule and Abra tenuis. The sand gaper Mya arenaria is superabundant in about a quarter of the samples for this biotope. Mya arenaria is probably present in a higher proportion of areas of this biotope, but may be missed in core samples due to its size.

    This review focuses on the ragworm Hediste diversicolor and the baltic tellin Macoma balthica as loss of these species would lead to loss of the biotope. Sandy mud or mud is also required to support these species and loss of these sediment types would also lead to loss of the biotope. Other species that are common or abundant are considered where they are likely to be affected by a pressure. 

    Resilience and recovery rates of habitat

    When impacted this biotope may recover through repair of damaged individuals, adult migration by mobile species and recolonization by pelagic larvae. The ability of postlarvae, larger juveniles, and adults of the key characterizing species Hediste diversicolor to swim, burrow and be carried by bedload transport can aid the rapid recolonization of disturbed sediments (Shull, 1997). Davey & George (1986), found evidence that larvae of Hediste diversicolor were tidally dispersed within the Tamar Estuary over a distance of 3 km.  Such passive dispersal alone suggested that recolonization of disturbed sediments, or populations removed from bait digging was likely to occur rapidly, depending upon larvae transport pathways.

    Generally Hediste diversicolor is reported to reach maturity between one and three years of age, like other Nereidae, Hediste diversicolor are monotelic, that is, they reproduce only once in their lifetime and then die (Olive & Garwood, 1981). Populations appear to show local characteristics in terms of spawning periods. Spawning may be limited to a short period in spring or extend over the summer. In the Thames Estuary, Dales (1950) reported specimens growing to maturity within one year, spawning in February, with some individuals surviving up to 18 months. Mettam et al. (1982), reported that Hediste diversicolor from the Severn Estuary matured rapidly in the spring and spawned at two years old. Olive & Garwood (1981), found that females in the Blyth Estuary, Northumberland, were in their second year before eggs began to appear, so most probably spawned in their third year.

    Mature males crawl around outside in search of a mature female and discharge sperm through the nephridia, directly outside her burrow. Direct contact between the sexes is not a necessity. Sperm is drawn into the burrow by females and fertilized eggs remain inside the burrow protected by the female. Both sexes die shortly after spawning. The trait to lay and protect eggs within a burrow is likely to increase the time populations recover from pressures such as sediment removal, as both adults and eggs will be affected. The pelagic larval dispersal phase is short (Scaps, 2002).

    In general, recovery of Hediste diversicolor populations from impacts appears to be relatively rapid. Recovery will be enhanced where adult migration (active or passive) can transport adults from adjacent, unimpacted habitats. Where a large area is severely impacted, however, recovery may require longer time-scales. The effects of a pipeline construction on benthic invertebrates were investigated using a Before/After impact protocol at Clonakilty Bay, West Cork, Ireland. Benthic invertebrates were sampled once before the excavation and at one, two, three and six months after the completion of the work.. An impact was obvious in the construction site in that no live invertebrates were found at one month after disturbance, but there followed a gradual recolonisation by Hediste diversicolor. At six months after the disturbance, there was no significant difference in the mean number of total individuals (of all species) per core sample amongst all study sites, but the apparent recovery in the impacted area was due to recovery of  Hediste diversicolor and Tubifex spp. (Lewis et al., 2002b).

    Bolam et al. (2004) experimentally simulated (in the field) the effect of dredged material emplacement (beach recharge) by manipulating defaunated sediments. Macrofaunal sampling was carried out after 1 week and after 1, 3, 6 and 12 months. Recolonisation patterns, although gradual, were found to be species specific with abundances of the polychaete Hediste diversicolor and the gastropod Hydrobia ulvae recovering to ambient levels within one week.

    The life history characteristics of give the species strong powers of recoverability. Adults spawn at least once a year and are highly fecund (Caddy, 1967). Females are capable of producing 10,000-30,000 eggs (MES, 2010). There is a planktotrophic larval phase which lasts up to 2 months (Fish & Fish, 1996) and so dispersal over long distances is potentially possible given a suitable hydrographic regime. Following settlement, development is rapid and sexual maturity is attained within 2 years (Gilbert, 1978; Harvey & Vincent, 1989). In addition to larval dispersal, dispersal of juveniles and adults occurs via burrowing (Bonsdorff, 1984; Guenther, 1991), floating (Sörlin, 1988) and probably via bedload transport (Emerson & Grant, 1991). It is expected therefore that recruitment can occur from both local and distant populations. Bonsdorff (1984) studied the recovery of a Limecola balthica (as Macoma balthica) population in a shallow, brackish bay in SW Finland following the removal of the substratum by dredging in the summer of 1976. Recolonization of the dredged area by Limecola balthica began immediately after the disturbance to the sediment and by November 1976, the Limecola balthica population had recovered to 51 individuals/m². One year later there was no detectable difference in the Limecola balthica population between the recently dredged area and a reference area elsewhere in the bay. In 1976, two generations could be detected in the newly established population indicating that active immigration of adults was occurring in parallel to larval settlement. In 1977, up to six generations were identified, giving further evidence of active immigration to the dredged area.

    A study by Verdelhos et al. (2011) suggested the existence of different life strategies within populations of Scrobicularia plana depending on temperature, latitudinal gradient and local habitat conditions. Higher-latitude populations usually exhibited low abundance values, shorter reproduction periods and lower growth rates. Conde et al. (2011) compared recruitment of the bivalve Scrobicularia plana to excavated and un-excavated control plots (expected to enhance the deposition of bivalve spat if the settlement of bivalves was the result of a passive process) at different shore levels in Portugal. Juveniles of both bivalve species were found to avoid excavated plots, showing significantly higher abundance in control plots. The data strongly suggested that recruited bivalves actively avoid unsuitable substrata.

    The polychaete Pygospio elegans have life history strategies that allow rapid colonization and population increase in  disturbed and defaunated patches where there is little competition from other species. Pygospio elegans exhibit a number of reproductive strategies (a trait known as poecilogony). Larvae may develop directly allowing rapid population increase in suitable patches or they may have a planktonic stage (allowing colonization of new habitats). Experimental defaunation studies have shown an increase in Pygospio elegans, higher than background abundances within 2 months, reaching maximum abundance within 100 days (Van Colen et al. 2008).  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 and sediments were disturbed by cockle movement (Desprez et al., 1992). Re-colonization of Pygospio elegans, was observed in 2 weeks by Dittmann et al. (1999) following a 1 month long defaunation of the sediment. However, McLusky et al. (1983) found that Pygospio elegans were significantly depleted for >100 days after harvesting (surpassing the study monitoring timeline). Ferns et al. (2000) found that tractor-towed cockle harvesting removed 83% of Pygospio elegans (initial density 1850 per m2).  In muddy sand habitats, Pygospio elegans had not recovered their original abundance after 174 days (Ferns et al., 2000). These results are supported by work by Moore (1991) who also found that cockle dredging can result in reduced densities of some polychaete species, including Pygospio elegans. Rostron (1995) undertook experimental dredging of sandflats with a mechanical cockle dredger, including a site comprised of stable, poorly sorted fine sands with small pools and Arenicola marina casts with some algal growths. At this site, post-dredging, there was a decreased number of Pygospio elegans with no recovery to pre-dredging numbers after six months.

    Tubificid populations tend to be large and to be constant throughout the year, although some studies have noticed seasonal variations (Giere & Pfannkuche, 1982). Many species, including Tubificoides benedii andBaltidrilus costata have a two-year reproductive cycle and only part of the population reproduces each season (Giere & Pfannkuche, 1982). Populations of Tubificoides benedii in the Fourth estuary have not demonstrated clear seasonality in recruitment (Bagheri & McLusky, 1982), although mature Tubificoides benedii (as Peloscolex benedeni) in the Thames Estuary were reported to occur in December with a maximum in late February (Hunter & Arthur, 1978), breeding worms increased from April and maximum cocoon deposition was observed in July (Hunter & Arthur, 1978).  Bolam and Whomersley (2003) observed faunal recolonization of fine sediments placed on saltmarsh as a beneficial use and disposal of fine grained dredged sediments. They found that tubificid oligochaetes began colonizing sediments from the first week following a beneficial use scheme involving the placement of fine-grained dredged material on a salt marsh in southeast England. The abundance of Tubificoides benedii recovered slowly in the recharge stations and required 18 months to match reference sites and those in the recharge stations prior to placement of sediments. The results indicate that some post-juvenile immigration is possible and that an in-situ recovery of abundance is likely to require more than 1 year.  Rapid recolonization has also been observed in the tubificid oligochaete  Baltidrilus costata (Tubifex costatus)  appeared  in upper sediment layers in experimentally defaunated patches (4m2) after 3 weeks (Gamenick et al., 1996).

    Resilience assessment. Resilience is assessed as ‘High’ (within 2 years), where resistance is ‘High’, ‘Medium’ or ‘Low’ based on Bolam et al. (2004), Bolam & Whomersley, (2003), Bonsdorff (1984), and Gamenick et al. (1996). Resilience is assessed as ‘Medium’ (2-10 years) where, resistance is ‘None’ and habitat recovery may also be required. 

    NB: The resilience and the ability to recover from human induced pressures is a combination of the environmental conditions of the site, the frequency (repeated disturbances versus a one off event) and the intensity of the disturbance.  Recovery of impacted populations will always be mediated by stochastic events and processes acting over different scales including, but not limited to, local habitat conditions, further impacts and processes such as larval-supply and recruitment between populations. Full recovery is defined as the return to the state of the habitat that existed prior to impact.  This does not necessarily mean that every component species has returned to its prior condition, abundance or extent but that the relevant functional components are present and the habitat is structurally and functionally recognizable as the initial habitat of interest. It should be noted that the recovery rates are only indicative of the recovery potential. 

     

    Hydrological Pressures

     ResistanceResilienceSensitivity
    High High Not sensitive
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High

    Hediste diversicolor and other important characterizing species are adapted to living within the intertidal zone where temperatures fluctuate. Some resistance to temperature fluctuations is achieved by burying within the sediment, which buffers against acute temperature changes over the tidal cycle.

    The geographic range of Hediste diversicolor (throughout north-west Europe on the Baltic Sea, North Sea and along Atlantic coasts to the Mediterranean) suggests that it is tolerant of a range of temperatures and a temperature increase at benchmark levels is unlikely to have an adverse effect on UK populations. Hediste diversicolor can tolerate temperatures from below zero under Baltic ice to high summer temperatures in Black Sea lagoons (>25°C) (Smith, 1977).

    Hediste diversicolor were not strongly affected by heat waves in an estuary in north western Portugal, where temperatures reached 40°C in intertidal pools (higher temperatures than experienced around UK and Irish coasts) (Dolbeth et al., 2011). Grilo et al., 2011) found that, at a Portuguese site, surface deposit feeders gradually decreased in periods of higher temperatures. However, sub-surface deposit feeders became dominant for up to 3 years after heat wave conditions had passed.

    Temperature change may adversely affect reproduction. Bartels-Hardege and Zeeck (1990) demonstrated that an increase from 12°C and maintenance of water temperature at 16°C induced reproduction in Hediste diversicolor specimens outside the normal period of spawning, and without a drop in temperature to simulate winter conditions the spawning period was prolonged and release of gametes was not synchronized. Poor synchronization of spawning could result in reduced recruitment, as gametes are wasted and mature specimens die shortly after gamete release.

    In Europe, Limecola balthica occurs as far south as the Iberian Peninsula and hence would be expected to tolerate higher temperatures than experienced in Britain and Ireland. Oertzen (1969) recorded that Limecola balthica (as Macoma balthica) could tolerate temperatures up to 49°C before thermal numbing of gill cilia occurred, presumably resulting in death. Ratcliffe et al., 1981) reported that Limecola balthica (as Macoma balthica) from the Humber Estuary, UK, tolerated 6 hours of exposure to temperatures up to 37.5°C with no mortality. Wilson (1981) show that the lethal temperatures for Limecola balthica change between seasons. Critical temperatures were studied for a Limecola balthica population in Dublin Bay, and a summer maximum of 37.5 and winter maximum of 27.5 were reported (Wilson, 1981).  Tolerances were also reported to change with height up the shore, which suggested tolerances adapt to variable conditions.

    Field observations and laboratory experiments showed that Limecola balthica spawns (criterion: 50% spent) in spring when the gradual increase of the mean (monthly averaged) water temperature surpasses 8.3oC. The success of spawning and recruitment is affected by the timing of the spring phytoplankton bloom and avoidance of the main settlement of the predator Crangon crangon on intertidal shores (Philippart et al., 2003). A mismatch in spawning cues due to an acute increase in temperature could result in low recruitment or recruitment failures

    Despite apparent adaptation to regional temperature ranges, Barda et al. (2014) studied populations of Limecola balthica (as Macoma balthica) in the Baltic Sea and found that increased temperature reduced growth rates. Beukema et al. (2014) also warn that increasing water temperatures as a result of global warming are likely to shorten the growing season (typically late winter to early spring) if warmer spring and summer water temperatures are experienced.  Jansen et al. (2007) suggest that temperature increases in the Spanish coast along the Bay of Biscay over the past 40 years caused loss of Limecola balthicapopulations, due to short-term but frequent exposure to >30° C in the Spanish estuaries, which induced elevated maintenance rates in Limecola balthica, and ultimately starvation. Repeated recruitment failure also occurred after mild winters in a comparable North Sea location, probably due to enhanced survival of predators(Beukema, 1992, Schueckel & Kroencke, 2013; Beukema et al., 2001).  As a result, Jansen et al. (2007) predict the southern limit of the species will progressively shift north if temperatures continue to rise.

    Pygospio elegans also shows a relationship between timing of reproduction and temperature. Gibson and Harvey (2000) in a study on asexual reproduction of Pygospio elegans in Nova Scotia, Canada found that temperature did not influence reproduction strategy (planktotrophy, lecithotrophy or asexual reproduction) but cite Anger (1984) that environmental conditions, including temperature influence timing of reproduction.

    Tubificoides benedii increased in abundance in mudflat habitats in Jade Bay, North Sea between 1930 and 2009 (Schueckel & Kroencke, 2013). Climate warming as well as decreasing nutrient loads and species introductions have occurred in the region since the 1970s, suggesting the species may adapt to temperature increases at benchmark pressures.Bamber & Spencer (1984) observed that Tubificoides were dominant species in an area affected by thermal discharge in the River Medway estuary.  Sediments were exposed to the passage of a temperature front of approximately 10oC between heated effluent and estuarine waters during the tidal cycles.Eteone longa and Pygospio elegans were summer visitors at the same sites (Bamber & Spencer, 1984) and are considered to be tolerant to this pressure at the benchmark.

    Higher temperatures have been implicated in the proliferation of trematode parasites which have caused mass mortalities in the snail Hydrobia ulvae (Jensen & Mouritsen, 1992), which is often abundant in this biotope. No other information was found on tolerance of component species to increased temperature. Nevertheless, an increase in temperature may indirectly affect some species as microbial activity within the sediments will be stimulated increasing oxygen consumption and promoting hypoxia.

    Sensitivity assessment. Typical surface water temperatures around the UK coast vary, seasonally from 4-19°C (Huthnance, 2010). It is likely that the important characteristic species are able to resist a long term increase in temperature of 2°C and may resist a short term increase of 5°C. Limecola balthica may retreat north as a result of long-term warming and climate change. However, the important characterizing species Hediste diversicolor are likely to survive a 5°C increase in temp for one month period, or 2°C for one year, although reproductive activities may be impacted. For instance, without colder winters spawning may not be synchronised and so recruitment would be reduced. A resistance of ‘High’, a resilience of ‘High’ (as longer lived later maturing species are present) and a sensitivity of ‘Not sensitive’ have been assigned.

    High High Not sensitive
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High

    A decrease in temperature has been shown to be beneficial to Hediste diversicolor as case studies report a reduction in numbers of the species’ predators. For instance, a severe winter in the Wadden Sea in 1995/1996 saw an increased abundance of Hediste diversicolor coincident with a reduction in the numbers of Carcinus maenus and Crangon crangon (Armonies et al., 2001). A similar increase in abundance was noted in the same area between 1978 and 1987 after a series of cold winters (mean Hediste diversicolor density increased from 24/m² to 151/m² respectively ) (Beukema, 1990).

    Decreased temperatures throughout the year may limit reproduction. Bartels-Hardege & Zeeck (1990) induced spawning in the laboratory, in specimens of Hediste diversicolor from tidal flats of the Jadebusen (North Sea), outside the normal spawning period of early spring. Temperatures were not lowered to simulate winter conditions but maintained at 16°C. Mature specimens appeared after four weeks and released gametes after a further four weeks according to a semilunar cycle. Reproduction was sustained for a period of four months. Such an extended spawning was witnessed on the Jadebusen following an unusually warm winter. Spawning occurred from February until May and was less synchronized. In contrast, the same population spawned within two months (February - March) following lower winter temperatures in another year. They concluded that not only a threshold temperature was important for synchronized spawning but the timing of the rise in temperature following winter was also a significant factor (Bartels-Hardege & Zeeck, 1990). A reduced rise in temperature is likely to limit this factor.

    Colder winter temperatures have been shown to benefit Limecola balthica population dynamics. Recruitment success increased following colder winters and repeated recruitment failure has occurred after mild winters in comparable North Sea location (Beukema, 1992, Schueckel & Kroencke, 2013; Beukema et al., 2001). In Friedrichskoog, Germany, König (1943) found a high accumulation of dead Cerastoderma edule biomass after a severe winter 1936/1937 but high numbers of Limecola balthica (80,000 individuals/m²) spat in following years (winter 1939). Winter water surface temperatures in the Wadden Sea (Netherlands) have increased 1.5°C since the 1980s (Oost et al., 2009). During milder winters greater body weight loss and production of fewer and smaller eggs has been observed in Limecola balthica (van der Meer et al., 2003). It is noted in the literature however, that reduced recruitment success during milder winters may also be due to increased predation as juvenile Crangon crangon have shown increased abundance in relation to milder winters (Beukema & Dekker, 2005).

    The geographical distribution of Limecola balthica suggests that it is very tolerant of low temperature. The species occurs in the Gulfs of Finland and Bothnia where the sea freezes for several months of the year (Green, 1968). It must, therefore, resist much lower temperatures than it experiences in Britain and Ireland. Furthermore, Limecola balthica was apparently unaffected by the severe winter of 1962/3, which severely affected  many other bivalve species (Crisp, 1964), and De Wilde (1975) noted that Limecola balthica kept at 0°C maintained a high level of feeding activity. It is likely, therefore, that in seas around the UK and Ireland Limecola balthica would be resist decreases in temperature at the pressure benchmark level.

    Eteone longa recorded as arctic-boreal, both from the Atlantic and the Pacific and occurs throughout the North Sea. Occurring in the arctic-boreal region suggests the species is resistant to a decrease in temperature in UK and Irish seas.

    Most littoral oligochaetes, including tubificids and enchytraeids, can survive freezing temperatures and can survive in frozen sediments (Giere & Pfannkuche, 1982). Tubificoides benedii (studied as Peloscolex benedeni) recovered after being frozen for several tides in a mudflat (Linke, 1939).

    Sensitivity assessment. The important characterizing species show limited impacts and, potentially, benefits to abundance and recruitment from decreases in temperature. Therefore, a 5°C decrease in temp for one month period, or 2°C for one year is likely to have limited negative impact on all characterizing species in the biotope, within British and Irish seas. Hence, resistance is assessed as ‘High’, resilience is assessed as ‘High’, and sensitivity as ‘Not Sensitive’.

    Low High Not sensitive
    Q: Low
    A: NR
    C: NR
    Q: High
    A: High
    C: Medium
    Q: Low
    A: Low
    C: Low

    The biotopes occurs in variable (18-35 ppt), reduced (18-30 ppt) and full salinity (30-35 ppt) (JNCC, 2015). Biotopes occurring in variable and reduced salinity are not considered sensitive to a change to full salinity as this falls within the natural habitat range. However, species may be sensitive to an increase in the salinity regime to hypersaline (>40 ppt).  Little evidence was found for tolerances of elevated salinity although populations are likely to be acclimated to short-term increased salinities in surface sediment layers as evaporation on hot days, or wind driven desiccation leads to increased salinities.

    Hediste diversicolor, the key characterizing species, occurs across all variant sub-biotopes and as such is resistant to the salinity range, from reduced to full, that the various sub-biotopes occur within. Hediste diversicolor is a euryhaline species, able to tolerate a range of salinities from fully marine  seawater down to 5 psu or less (Barnes, 1994).  Seitz (2011) found that the distribution of Limecola balthica (as Macoma balthica) across a salinity gradient between a minimum and maximum of 8.8psu to 19 psu in Cheaspeake Bay was not influenced by salinity. Instead, resource availability was the principal influence on Limecola balthica. McLusky & Allan (1976) reported that Limecola balthica (as Macoma balthica) failed to grow at 41 psu. It is likely that Limecola balthica would be tolerant of an increase in salinity category to fully marine but further increases to >40‰ are likely to affect growth and condition.

    Hylleberg (1975) also found that under controlled conditions of salinity ranging from 10 to 30ᵒ⁄ₒₒand temperatures ranging from 5 to 35° C, Hydrobia ulvae has maximal egestion at the combination of high salinity (30ᵒ⁄ₒₒ) and high temperature (30° C). The species would be likely to show high resistance to an increase in salinity from the reduced and variable conditions the biotope occurs within. 

    Pygospio elegans is common in both marine and brackish waters in the Schelde estuary (Netherlands) suggesting in European habitats the species tolerates a broad salinity range (Ysebaert et al., 1993). Studies of Pygospio elegans population structure in the Baltic Sea and North Sea also found larvae were not hampered by changes in salinity (Kesaniemi et al., 2012). Although case studies are lacking for British and Irish coasts, the existing evidence suggests Pygospio elegans would tolerate salinity changes from variable to full but may be sensitive to a change from full salinity to hypersaline (>40 ppt).

    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 dependent (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.

    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 adults and larvae. Biotope resistance is assessed as ‘Low’ as some adults may survive and acclimate. Biotope resilience (following a return to suitable habitat conditions) is assessed as ‘High’ and sensitivity is assessed as ‘Low’.

    Low High Low
    Q: High
    A: Medium
    C: Medium
    Q: High
    A: Low
    C: Medium
    Q: High
    A: Low
    C: Medium

    The biotope occurs in variable (18-35 ppt), redcued (18-30 ppt) and full salinity (30-35 ppt) (JNCC, 2015). Biotopes occuring in full salinity are considered tolerant of a change to variable or reduced salinity as this falls within the natural habitat range. The decrease in salinity assessed at the benchmark is to low salinity (<18 ppt). The available evidence (summarised below) suggests that the characterizing species are tolerant of a short-term decrease to low salinity but it is likely that for species such as Limecola balthica long-term reductions would lead to mortality.

    The key characterizing species Hediste diversicolor is known to tolerate low salinities below 18-24 psu and it has been shown to replace Arenicola marina in areas influenced by freshwater runoff or input (e.g. the head end of estuaries) (Barnes; 1994; Hayward, 1994). Lower salinities (<8 psu) can, however, have an adverse effect on Hediste diversicolor reproduction (Ozoh & Jones, 1990; Smith 1964). Fertilization in Hediste diversicolor is adapted to high salinity but not to low salinity below 7.63‰ (Ozoh & Jones, 1990). A decrease in salinity at the benchmark pressure (reduction to <18‰ ) may negatively impact recruitment and abundance if the dilution is close to that threshold.

    McLusky & Allan (1976) conducted salinity survival experiments with Limecola balthica (as Macoma balthica) over a period of 150 days. Survival times declined with decreased salinity. At 12 psu specimens survived 78 days, whilst specimens at 8.5 psu survived 40 days. Some specimens of Limecola balthica survived 2.5 days at 0.8 psu, which was apparently due to the animal’s ability to clamp its valves shut in adverse conditions. McLusky & Allan (1976) also reported that Limecola balthica failed to grow (increase shell length) at 15 psu. Limecola balthica is found in brackish and fully saline waters (Clay, 1967b) so may tolerate a state of flux. Its distribution in combination with the experimental evidence of McLusky & Allan (1976) suggests that Limecola balthica is likely to be resistant to  decreased salinity over a short period. A decline in salinity in the long term may have implications for the species viability in terms of growth, and the distribution of the species may alter as specimens at the extremes retreat to more favourable conditions. Metabolic function should, however, quickly return to normal when salinity returns to original levels. Decreased salinity may also affect the ability of Limecola balthica to tolerate contaminants such as heavy metals (see Bryant et al., 1985 & 1985a). Usually, contaminants become more toxic at low salinity (Langston, W.J. pers comm.).

    Muus (1967) revealed that Hydrobia ulvae did not crawl into water with a salinity lower than 9ᵒ⁄ₒₒ. Hylleberg (1975) also found that under controlled conditions of salinity ranging from 10 to 30ᵒ⁄ₒₒ and temperatures ranging from 5 to 35°C, shows that Hydrobia ulvae has maximal egestion at the combination of high salinity (30ᵒ⁄ₒₒ) and high temperature (30° C).

    Oligochaete dominated biotopes are recorded from a range of salinity regimes from full (LS.LSa.MoSa.Ol; LS.LSa.MoSa.Ol.FS), variable (SS.SMu.SMuVS.CapTubi) reduced (SS.SMu.SMuVS.CapTubi; LS.LMu.UEst.Tben ) and low (SS.SMu.SMuVS.LhofTtub) habitats (JNCC,2015). In very low salinities from <15 to 0 ‰ species such as Limnodrilus spp. and Tubifex tubifex are found (Giere & Pfannkuche, 1982).  A decrease in salinity at the pressure benchmark would probably result in replacement by oligochaete species more tolerant of lower salinities such as Limnodrilus hoffmeisteri and Tubifex tubifex that characterize the low salinity biotope SS.SMu.SMuVS.LhofTtub. Numerous studies suggest that Baltidrilus costata tolerates a wide range of salinities from 1‰ to 28‰ (Giere & Pfannkuche, 1982 and references therein), suggesting that this species is likely to still be present in the biotope.

    Sensitivity assessment. It is considered likely that a decrease in salinity at the pressure benchmark will lead to some species replacement by polychaetes and oligochaetes more tolerant of low salinity. Hediste diversicolor and oligochaetes are likely to remain but Limecola balthica is likely to reduce in low salinity conditions. A similar biotope could remain where salinities were close to 18 ppt but a severe reduction in salinity would probably lead to loss of the biotope. Resistance is therefore assessed as ‘Low’.  Resilience (following restoration of typical conditions)  is ‘High’ and sensitivity is assessed as ‘Low’. It should be noted that resistance would be lower, and sensitivity greater, where salinity was reduced to a level close to freshwater.

    Medium High Low
    Q: Low
    A: NR
    C: NR
    Q: High
    A: Low
    C: Medium
    Q: Low
    A: Low
    C: Low

    Hediste diversicolor characteristically inhabits littoral mudflats predominantly of clay (particles < 4 µm), silt (4-63 µm) and to a lesser extent, very fine sand (63-125 µm) (Jones et al., 2000). Highest abundances occur in very weak (negligible) to weak < 1 knot (<0.5 m/sec.) currents. These conditions are provided by this biotope, which occurs in extremely sheltered gravelly mud to gravelly sandy mud on the mid and lower shore.

    The type direction and speed of the currents control sediment deposition within an area. Finer sediment will fall to the substratum in weaker currents. An increase in water flow rate would entrain and maintain particles in suspension and erode the mud. As a result the scouring and consequent redistribution of components of the substratum would alter the extent of suitable habitat available to populations of Hediste diversicolor and other species in the biotope that prefer finer sediment. Recovery of Hediste diversicolor would be influenced by the length of time it would take for the potential habitat to return to a suitable state for recolonization by adult and juvenile specimens from adjacent habitats, and the establishment of a breeding population. Recolonization may take between one and three years, as populations differ in reaching maturity (Dales, 1950; Mettam et al., 1982; Olive & Garwood, 1981), from the time that the habitat again becomes suited to the species.

    Coarser sediments are likely to remain in areas of strongest flow velocity (where finer particles have been re-suspended). Species such as Pygospio elegans and other opportunist polychaetes that tolerate coarser particle size will possibly increase in abundance. Limecola balthica is likely to experience greater impact from increased water flow as the species thrives in low energy environments, such as the extremely sheltered areas that characterize the biotope (Tebble, 1976). Increased water flow rate is likely to influence the sediment characteristics in this biotope, primarily by re-suspending and preventing deposition of finer particles (Hiscock, 1983). This is likely to result in erosion of the preferred habitat, which may cause mortality of some portion of the population of Limecola balthica. Higher current velocity (18 cm/s 0.18 m/s) recorded in flume experiments conducted in the Isle of Sylt (North Sea) led to juvenile Limecola balthica being washed out of the sediment (Zuhlke & Reise, 1994). Green (1968) reported that, towards the mouth of an estuary where sediments became coarser and cleaner, Limecola balthica was replaced by another tellin species, Tellina tenuis.

    Sensitivity assessment. Limecola balthica abundance may be reduced if juveniles are washed from the substratum. Loss of mud content in some areas is possible and could lead to replacement by another species but this is unlikely at the pressure benchmark levels. An increase in flow velocity may alter the muddy sand sediments, resistance has been assessed as ‘Medium’, recoverability is assessed as ‘High’ and sensitivity is, therefore ‘Low’.

    High High Not sensitive
    Q: High
    A: Medium
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: Medium
    C: High

    The biotope is found on the lower, mid and upper shore (JNCC, 2015) and changes in emergence are unlikely to affect the biotope where it remains within an intertidal habitat. 

    Hediste diversicolor inhabits a burrow within the sediment which may be up to 0.3 m deep. The species retreats within the burrow during periods of exposure, protecting it from desiccation although increased emergence may cause a decline in the abundance of Hediste diversicolor at the upper limits of the intertidal zone, as they may become stressed by desiccation if the substrata begin to dry and are prone to more extremes of temperature. Hediste diversicolor is sufficiently mobile to gradually retreat back to damper substrata. Gogina et al. (2010) analysed patterns of benthic community distribution related to selected environmental parameters, including depth, in the western Baltic Sea with depths ranging from 0 m to 31 m. Hediste diversicolor displayed a preference for low-saline regions shallower than 18 m. Increased depth had the largest negative effect of all factors influencing distribution and abundance decreased with greater dept) Gogina et al. (2010).

    Limecola balthica occurs in the upper regions of the intertidal (Tebble, 1976) and is, therefore, likely to be tolerant of prolonged emergence. It is a bivalve and can close tightly by contraction of the adductor muscle, storing moisture inside the shell. The silty sediments in which the species lives have a high water content and are therefore resistant to desiccation. Furthermore, Limecola balthica is mobile and able to relocate in the intertidal by burrowing (Bonsdorff, 1984) or floating (Sörlin, 1988). It would be expected to react to an increase in emergence by migrating down the shore to its preferred position. There may be an energetic cost to this migration but it is not expected that mortality would result. Limecola balthica should quickly recover from the energetic cost of relocation. Limecola balthica occurs in the intertidal and sublittorally down to depths of 190 m (Olafsson, 1986), although is more abundant intertidally.  Hence, it would be expected to resist a decrease in emergence regime. However, a case study, predicting changes in biomass of Limecola balthica in the Humber estuary, UK (western North Sea) under expected sea level rise conditions displayed negative impacts. Coastal squeeze from sea level rise would produce steeper and more homogenous beach face profiles. Limecola balthica was predicted to be lower on steeper beach faces and biomass of Limecola balthica was predicted to decrease (Fujii & Raffaelli, 2008). 

    Tubificoides benedii is capable of penetrating the substratum to depths of 10 cm, shows a resistance to hypoxia and is often typified as an ‘opportunist’ that is adapted to the rapid environmental fluctuations and harsh conditions in estuaries (Gogina et al., 2010). Highest abundances were predicted by  Gogina et al. (2010) to be related to depth with an optimum of 10 m  to 20 m. The evidence  suggests that  abundance may be limited by a decrease in high water level or a change in time (increase) where substratum is not covered by the sea. An increase in the time the biotope is covered by the sea is likely to result in increased abundance of Tubificoides benedii.

    Sensitivity assessment. As intertidal species, the biotope and characterizing species are found at a range of shore heights and are considered relatively resistant to changes in emergence which do not alter the extent of the intertidal. An increase in emergence is likely to decrease the upper shore extent of Hediste diversicolor dominated biotopes at the land-ward extent of the intertidal as desiccation increases. A decrease in emergence under the benchmark pressure  is likely to extend the upper extent of the biotope as the species recolonize or migrate to favourable conditions. Biotope resistance is, therefore, assessed as ‘High’, recoverability is assessed as ‘High’ (by default) and the biotope is considered to be 'Not sensitive'.

    High High Not sensitive
    Q: Low
    A: NR
    C: NR
    Q: High
    A: High
    C: High
    Q: Low
    A: Low
    C: Low

    This biotope occurs in wave sheltered areas, where estimated wave categories range from sheltered, very sheltered and extremely sheltered (JNCC, 2015). The key characterizing species Hediste diversicolor is infaunal, inhabiting a burrow in which it seeks refuge from predators and may partially emerge to seek and capture food. An alteration of factors within the environment that increases wave exposure could cause erosion of the substrata and consequently, loss of habitat.

    Wave action stimulates Limecola balthica to start burrowing and individuals have been shown to continue burrowing for a longer period of time than in still water (Breum, 1970).  Limited zoobenthic biomass was recorded in areas exposed to strong currents and wave action (Beukema, 2002), limiting food availability, however,impacts from this pressure at the benchmark levels may be low for this biotope, as the biotope is limited to sheltered or extremely sheltered locations. Increases in wave action may therefore remain within the limits of the species tolerance but factors such as sediment redistribution may alter the physical biotope.

    Sensitivity assessment. Resistance to a change in nearshore significant wave height >3% but <5% of the two main characterizing species Hediste diversicolor and Limecola balthica is ‘High’, given that the biotope occurs in very sheltered locations and an increase in nearshore significant wave height of >3% but <5% would continue to result in sheltered conditions which are within the species tolerance limits. At the highest benchmark pressure (5% increase) the species exhibit ‘High’ resistance through their traits to live relatively deep in the sediment. Resilience (recoverability) is also ‘High’ and the biotope is considered to be ‘Not Sensitive’. Due to limited evidence, confidence in this assessment is Low.

    Chemical Pressures

     ResistanceResilienceSensitivity
    Not relevant (NR) Not relevant (NR) Not sensitive
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    The biotope is assessed as Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards. The following review discusses impacts at higher concentrations than the pressure benchmark.

    In Hediste diversicolor the acute toxicity is dependent on the rate of uptake of the metal, since this determines the speed with which the lethal dose is built up. The rate of intake is important because this determines whether the organism's detoxification mechanisms can regulate internal concentrations. The resistance of Hediste diversicolor is thought to be dependent on a complexing system which detoxifies the metal and stores it in the epidermis and nephridia (Bryan & Hummerstone, 1971; McLusky et al. 1986).

    Hediste diversicolor has been found successfully living in estuarine sediments contaminated with copper ranging from 20 µm Cu/g in low copper areas to >4000 µm Cu/g where mining pollution is encountered e.g. Restronguet Creek, Fal Estuary, Cornwall (Bryan & Hummerstone, 1971). Attempts to change the tolerance of different populations of Hediste diversicolor to different sediment concentrations of copper have shown that it is not readily achieved suggesting that increased tolerance to copper has a genetic basis (Bryan & Hummerstone, 1971; Bryan & Gibbs, 1983).

    Crompton (1997) reviewed the toxic effect concentrations of metals to marine invertebrates Annelid species, such as Hediste diversicolor were found to be at risk if metals exceeded the following concentrations during 4-14 days of exposure: >0.1 mg Hg l-1, > 0.01 mg Cu l-1, > 1 mg Cd l-1, >1 mg Zn l-1,>0.1 mg Pb l-1, >1 mg Cr L-1, >1 mg As l-1 and >10 mg Ni l-1. In general, for estuarine animals heavy metal toxicity increases as salinity decreases and temperature increases (McLusky et al., 1986). For example, Fernandez & Jones (1990) calculated 96 hour LC50 Zinc values for Hediste diversicolor at four salinities 5, 10, 17.5 and 30 psu at 12°C. The 96 hour LC50 at 17.5 psu and 12°C was 38 mg Zn l-1, while at 5 and 10 psu it was 7 and 19 mg Zn l-1 respectively. Toxicity decreased with increasing salinity. When salinity remained constant at 17.5 psu, but temperature varied, the following 96 hour LC 50 values for Zinc were recorded: 40 mg Zn l-1 at 6°C, 32 mg Zn l-1 at 12°C and 9.1mg Zn l-1 at 20°C. Toxicity increased with increasing temperature. Accumulation of zinc was also greater at the lowest salinities and when the temperature was highest at 20°C. In a parallel experiment, the presence of sediment was found to reduce toxicity and body accumulation of zinc in Hediste diversicolor. Recovery of this species would be influenced by the length of time it would take for the potential habitat to return to a suitable state (e.g. factors such as the decline of bioavailable metals within the marine environment), recolonization by adult and juvenile specimens from adjacent habitats, and the establishment of a breeding population. Since juveniles remain in the infauna throughout their development selection for metal tolerance can be expected to be operative from an early stage (Bryan & Gibbs, 1983).

    Scrobicularia plana can detect copper at a concentration of 0·01 ppm. The initial response is valve closure resulting in a rapid drop in heart rate. In concentrations of 0·05, 0·01 and, to a lesser extent, 0·1 ppm copper added to seawater, the clams begin to interact with the polluted water after 2–3 h. In 0·5 ppm, the valves remain closed and the heart rate is maintained at a low level over the 6-h exposure period. Mortality increases with time in 0·5 ppm copper concentration, reaching 50% in 5–7 days. In 0·05 and 0·01 ppm, no mortality was recorded over this period (Akberali & Black, 1980).Two experimental 30 day long static toxicity tests on small spat (2 to 3 mm in length) of Scrobicularia plana were undertaken on the effects of copper on the survival and burying activity in sand, of juveniles.  Results showed that:

    1)  exposure to up to 80 µg Cu 1.' did not result in increased mortalities with respect to the controls;

    2) Cu concentrations at 20 µg l-1and above also increased the burying time of juveniles by the end of the experiment; and 

    3) the no-observed-effect concentration (NOEC) for Cu was the lowest dose tested (i.e. 10 µg I-').  (Ruiz et al., 1994).

    Alterations of the burrowing behaviour of Scrobicularia plana, were studied in individuals exposed to soluble copper. Animals were exposed for 4 d to concentrations ranging from 25 to 150 μ g Cu l− 1. At the end of exposure, the burrowing kinetics in clean sediment were determined after 1 and 2 days. Even at the lowest tested concentrations, copper caused hypoactivity in organisms belonging to both species studied. Metabolical or physiological disturbances could be the cause of these impairments. Concentrations affecting burrowing behaviour were below those responsible for mortality in these species (Bonnard et al., 2009).

    Laboratory tests in clean water can be misleading as these do not reflect lowered toxicity in the marine environment due to the buffering effects of carbon and sulphide which render copper non-labile (not bioavailable) and the influence of water pH, hardness, temperature and salinity etc. Field surveys have found that Scrobicularia plana is present in the highly contaminated Fal Estuary where levels of copper and zinc are high (Bryan & Gibbs, 1983).

    Not relevant (NR) Not relevant (NR) Not sensitive
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    The biotope is assessed as Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards. The following review discusses impacts at higher concentrations than the pressure benchmark.

    The 1969 West Falmouth (America) spill of Grade 2 diesel fuel documents the effects of hydrocarbons in a sheltered habitat (Suchanek, 1993). The entire benthic fauna including Hediste diversicolor was eradicated immediately following the spill and remobilization of oil that continued for a period > 1 year after the spill, contributed to much greater impact upon the habitat than that caused by the initial spill. Effects are likely to be prolonged as hydrocarbons incorporated within the sediment by bioturbation will remain for a long time owing to slow degradation under anoxic conditions. Oil covering the surface and within the sediment will prevent oxygen transport to the infauna and promote anoxia as the infauna utilize oxygen during respiration. Although Hediste diversicolor is tolerant of hypoxia and periods of anoxia, a prolonged absence of oxygen will result in the death of it and other infauna. McLusky (1982) found that petrochemical effluents released from a point source to an estuarine intertidal mudflat, caused severe pollution in the immediate vicinity. Beyond 500 m distance the effluent contributed to an enrichment of the fauna in terms of abundance and biomass similar to that reported by Pearson & Rosenberg (1978) for organic pollution, and Hediste diversicolor was found amongst an impoverished fauna at 250 m from the discharge.

    Analysis of Scrobicularia plana collected in April 2008 from three estuaries along a pollution gradient (Goyen < Loire < Seine) found sub-lethal effects on clams including neurotoxicity and impairment of digestive enzyme activities (cellulase or amylase) in the Loire and Seine estuaries. The highest lactate dehydrogenase activity was registered in the Loire estuary, in parallel with enhanced levels of vanadium (a metal present in petroleum), likely as a consequence of a small oil spill that occurred one month before the sampling collection. The median size was significantly lower in clams exposed to direct (chemicals) or indirect (available food) effects in the most contaminated site. Burrowing behaviour was disturbed in clams from both of the Loire and Seine estuaries, a response probably due to physiological impairment rather than to avoidance of contaminated sediment. The activation of defence mechanisms towards metals (metallothionein) and other classes of contaminants (the biotransformation enzyme glutathione-S-transferase) do not ensure a total protection since a number of impairments were observed at the infra-organismal (AChE and digestive enzyme activities) and individual (burrowing behaviour) levels in relation to the degree of anthropogenic pressure. However, even in the most contaminated estuary (Seine), historical records do not show a consistent decrease of Scrobicularia plana populations (cited from Boldina-Cosqueric et al., 2010)

    Not relevant (NR) Not relevant (NR) Not sensitive
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    The biotope is assessed as Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards. The following review discusses impacts at higher concentrations than the pressure benchmark.

    Reports of the effects of synthetic chemicals on Hediste diversicolor illustrate that the intolerance of the species is highly dependent upon the molecular structure of the chemical, which determines the chemicals properties and use. For example:

    1. Collier & Pinn (1998) observed significant differences in both the abundance and biomass of a benthic community from the Ythan Estuary, Scotland, experiencing contamination by Ivermectin. Ivermectin is the 22,23-dihydro derivative of avermectin ß which has been shown to be highly efficient in the treatment of sea lice. Hediste diversicolor was the most intolerant species to Ivermectin in the benthic community studied. A rapid decline in both abundance and total biomass of Hediste diversicolor occurred within 7 days and with increasing concentration. An Ivermectin concentration of 8.0 mg m² caused 100% mortality within 14 days. Davies et al., (1998) modelled factors influencing the concentration of Ivermectin reaching the seabed which ranged from 2.2 to 6.6 mg m². The upper limit of this range was only slightly less than the concentrations found to be toxic by Collier & Pinn (1998) and Black et al. (1997). Davies et al.(1998) concluded that there was a significant risk to benthic organisms within a radius of 50 m of salmon farms utilizing Ivermectin and that Ivermectin could accumulate (half life of Ivermectin in marine sediments > 100 days) within the sediment beyond a single treatment and reach toxic levels.
    2. In contrast, Craig & Caunter (1990) examined the effects of the organosilicon compound, Polydimethylsiloxane (PDMS) in sediment on Hediste diversicolor. PDMS fluids are less dense than water and insoluble and form a discrete layer on the surface of the water. In an intertidal environment PDMS fluids are deposited upon the sediment surface at low tide and into contact with Hediste diversicolor. In laboratory tests, exposure to 10,000 mg PDMS per kg of sediment caused no deaths over 96 hours, and exposure to 1,000 mg PDMS per kg of sediment caused no deaths of Hediste diversicolor after 28 days.

    Recovery of this species would be influenced by the length of time it would take for the potential habitat to return to a suitable state (e.g. factors such as the rate of decay of the synthetic chemical within the marine environment), recolonization by adult and juvenile specimens from adjacent habitats, and the establishment of a breeding population. This may take between one and three years, as populations differ in reaching maturity (Dales, 1950; Mettam et al., 1982; Olive & Garwood, 1981), from the time that the habitat again becomes suited to the species.

    No evidence (NEv) Not relevant (NR) No evidence (NEv)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    Beasley & Fowler (1976) and Germain et al., (1984) examined the accumulation and transfers of radionuclides in Hediste diversicolor from sediments contaminated with americium and plutonium derived from nuclear weapons testing and the release of liquid effluent from a nuclear processing plant. Both concluded that the uptake of radionuclides by Hediste diversicolor was small. Beasley & Fowler (1976) found that Hediste diversicolor accumulated only 0.05% of the concentration of radionuclides found in the sediment. Both also considered that the predominant contamination pathway for Hediste diversicolor was from the interstitial water. However, there is insufficient information available on the biological effects of radionuclides to comment further upon the intolerance of this species to radionuclide contamination.

    Hutchins et al. (1998) described the effect of temperature on bioaccumulation by Limecola balthica of radioactive americium, caesium and cobalt, but made no comment on the intolerance of the species. Further, direct assessments of impacts at the benchmark pressure on benthic communities, and this biotope in particular were not found.

    Sensitivity assessment. No evidence. Insufficient evidence was available to complete and assessment.

    Not relevant (NR) Not relevant (NR) Not sensitive
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    Not sensitive at the pressure benchmark that assumes compliance with all relevant environmental protection standards

    High High Not sensitive
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High

    The habitats which Hediste diversicolor inhabits tend to have lower oxygen levels than other sediments. Hediste diversicolor is resistant to moderate hypoxia (Diaz & Rosenberg, 1995). Vismann (1990) demonstrated a mortality of only 15% during a 22 day exposure of Hediste diversicolor at 10% oxygen (ca. 2.8 mg O2 per litre). Hediste diversicolor is active at the sediment/water interface where hydrogen sulphide concentrations increase during periods of hypoxia. Vismann (1990) also demonstrated that the high tolerance of Hediste diversicolor to hypoxia in the presence of sulphide is enabled by elevated sulphide oxidation activity in the blood. Hediste diversicolor may also exhibit a behavioural response to hypoxia by leaving the sediment (Vismann, 1990) in the presence of sulphide. After 10 days of hypoxia (10% oxygen saturation) with sulphide (172-187 µmM) only 35% of Hediste diversicolor had left the sediment compared to 100% of Nereis virens. Laboratory experiments in the absence of sediments, found that Hediste diversicolor could survive hypoxia for more than 5 days and that it had a higher tolerance to hypoxia than Nereis virens, Nereis succinea and Nereis pelagica (Theede, 1973; Dries & Theede, 1974; Theede et al., 1973). Juvenile Hediste diversicolor survived hypoxic conditions for 4 days in laboratory conditions and combined hypoxia and increased sulphide (1 mmo1 1-') for 3 days (Gamenick et al., 1996). Post larvae Hediste diversicolor were the only life stage to show les tolerance to hypoxia, surviving for only 14 hr (Gamenick et al., 1996).

    Limecola balthica appears to be relatively tolerant of de-oxygenation. Brafield & Newell (1961) frequently observed that, in conditions of oxygen deficiency (e.g. less than 1 mg O2/l), Limecola balthica (as Macoma balthica) displayed survivability of low oxygen concentrations and shell growth continued (Jansson et al., 2015). Although, sub-lethal effects of hypoxia have been identified as individual Limecola balthica moved upwards to fully expose itself on the surface of the sand or buried at shallower depths, leaving them at greater risk of predation (Long et al., 2014). Specimens lay on their side with the foot and siphons retracted but with valves gaping slightly allowing the mantle edge to be brought into full contact with the more oxygenated surface water lying between sand ripples. In addition, Limecola balthica was observed under laboratory conditions to extend its siphons upwards out of the sand in to the overlying water when water was slowly deoxygenated with a stream of nitrogen. The lower the oxygen concentration became the further the siphons extended.

    This behaviour, an initial increase in activity stimulated by oxygen deficiency, is of interest because the activity of lamellibranchs is generally inhibited by oxygen deficient conditions (Brafield & Newell, 1961). Dries & Theede (1974) reported the following LT50 values for Limecola balthica (as Macoma balthica) maintained in anoxic conditions: 50 - 70 days at 5°C, 30 days at 10°C, 25 days at 15°C and 11 days at 20°C. Theede (1984) reported that Limecola balthica to resist extreme oxygen deficiency was mainly due to anaerobic metabolism. Macoma balthica is therefore very tolerant of hypoxia, although it may react by reducing metabolic activity and predation risk may increase. Metabolic function should quickly return to normal when oxygen levels are resumed and so recovery is expected.

    The characterizing oligochaetes and polychaetes within the biotope that display tolerance to hypoxia include Tubificoides benedii and Capitella capitata, while Pygospio elegans is highly sensitive to hypoxia (Gogina et al., 2010). Exposure to dissolved oxygen concentration of less than or equal to 2 mg/l for 1 week is likely to limit Pygospio elegans abundance, whilst having limited impact on Tubificoides benedii and Capitella capitata populations.

    Sensitivity assessment. Resistance to exposure to dissolved oxygen concentration of less than or equal to 2 mg/l for 1 week is assessed as ‘High’ for the characterizing species Hediste diversicolor and Limecola balthica. It is important to consider that other species that are common or abundant in the biotope may be impacted by decreased dissolved oxygen, such as Pygospio elegans and decreases in abundance of these species are likely. As this biotope is found in intertidal habitats oxygen levels will be recharged during the tidal cycle lowering exposure to this pressure for Pygospio elegans. Based on the reported tolerances for anoxia and intertidal habitat, biotope resistance is assessed as ‘High’ resilience is assessed as ‘High’ (by default) and the biotope is considered to be ‘Not sensitive’ at the benchmark level

    High High Not sensitive
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High

    This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations.  The benchmark is set at compliance with WFD criteria for good status, based on nitrogen concentration (UKTAG, 2014).  Primary production in the biotope will be limited to microalagae at the sediment surface, rather than macroalgae. Changes in primary production as a result of changes in nutrient enrichment are, therefore, not considered likely to directly alter the biotope.

    Aberson et al. (2016) found nutrient enrichment promotes surface deposit feeding in Hediste diversicolor, over suspension feeding and predation. At sewage-polluted sites in three estuaries in  SE  England Hediste diversicolor mainly consumed microphytobenthos, sediment organic matter and filamentous macroalgae Ulva  spp. At cleaner  sites  Hediste diversicolor relied more on suspension feeding and consumption of Spartina anglica (Aberson et al., 2016). Whilst suggesting adaptability to nutrient enrichment this behaviour will increase predation risk.

    Nutrient enrichment favours the growth of opportunistic green macro-algae blooms which can cause declines in some species and increases in others (Raffaelli, 2000). Evidence (Beukema, 1989; Reise et al., 1989; Jensen, 1992) suggested a doubling in the abundance of Hediste diversicolor in the Dutch Wadden Sea, accompanied by a more frequent occurrence of algal blooms that were attributed to marine eutrophication. Algae may be utilized by Hediste diversicolor in its omnivorous diet, so some effects of nutrient enrichment may be beneficial to this species. However, evidence for the effects of algal blooms stimulated by nutrient enrichment on Hediste diversicolor is not consistent. Raffaelli (1999) examined a 30 year data base to examine the effect of nutrient enrichment on an estuarine food web in Aberdeenshire, Scotland. This study displayed impacts to species characterizing the biotope from development of algal mats, the density and distribution of which was related to nutrient. In areas where algal biomass was greatest reduced invertebrate densities were recorded. The mud shrimp Corophium volutator showed the greatest decrease in density. Densities of Corophium volutator, Limecola balthica and Hediste diversicolor were lower in 1990 compared to 1964 at sites where macro-algal mats increased over the same period. Conversely, densities were on average higher in the upper reaches where macroalgal mats were generally absent before 1990 (Raffaelli, 1999). Capitella capitata and Pygospio elegans abundance were greater in areas that received greatest nutrient enrichment (Raffaelli, 1999). Long-term nutrient enrichment may, therefore, alter the biotope if high biomass of algal mats persists.

    Sensitivity assessment.  The benchmark is relatively protective and is not set at a level that would allow blooms of green algae on the sediment, based on this consideration and based on the lack of primary producers structuring the biotope, resistance is assessed as 'High' and resilience as 'High' (by default), so that the biotope is assessed as 'Not sensitive'.

     

    High High Not sensitive
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High

    All species apart from two that are abundant in the biotope are classed in a Marine Biotic Index as being indifferent to, tolerating or proliferating under organic enrichment conditions (Borja et al., 2000). Only Limecola balthica and Manayunkia aestuarina are recorded as being sensitive to organic enrichment.

    However, case studies display resilience of Limecola balthica populations to enrichment. Limecola balthica (as Macoma balthica) have been shown experimentally to be able to resist periods of up to 9 weeks under algal clover, their long siphon allowing them to reach oxygenated water although other bivalves decreased in abundance (Thiel et al., 1998). Organic enrichment from waste-water discharge in the Dutch Wadden Sea resulted in positive effects on Limecola balthicaabundance, biomass, shell growth and production. These effects were concluded to be due to increased food supply (Madsen & Jensen, 1987). Tubificoides benedii  and other oligochaetes are very tolerant of high levels of organic enrichment and often dominate sediments where sewage has been discharged or other forms of organic enrichment have occurred (Pearson & Rosenberg, 1978; Gray, 1971; McLusky et al., 1980). 

    Sensitivity assessment. At the benchmark levels, a resistance of ‘High’ as the main characterizing species Hediste diversicolor is tolerant of organic enrichment and an input at the pressure benchmark is considered unlikely to lead to gross pollution effects . A resilience of ‘High’ is assigned (by default) and the biotope is assessed as ‘Not sensitive’.

    Physical Pressures

     ResistanceResilienceSensitivity
    None Very Low High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High

    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.

    None Very Low High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: High

    This biotope and sub-biotopes is only found in sediment, in particular, gravelly sandy mud or gravelly mud (JNCC, 2015). The burrowing organisms characterizing this biotope, including Hediste diversicolor, and Limecola balthica would not be able to survive if the substratum type was changed to either a soft rock or hard artificial type.  Consequently, the biotope would be lost altogether if such a change occurred. 

    Sensitivity assessment. Biotope resistance is assessed as ‘None’, resilience is ‘Very low’ (as the change at the pressure benchmark is permanent) and biotope sensitivity is ‘High’.

    Low Very Low High
    Q: High
    A: Medium
    C: High
    Q: High
    A: High
    C: High
    Q: High
    A: Medium
    C: High

    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 biotope occurs in mud and sandy mud a change to finer sediments is not considered relevant as this falls within the natural habitat change (JNCC, 2015).

    An increase in gravel and a change to clean sands or coarse sediments is likely to have a more significant effect as sediment cohesion and ability to retain organic matter and water is reduced altering habitat suitability for burrowing polychaetes and amphipods and deposit feeders.

     

    Hediste diversicolor is infaunal and is reliant upon a muddy/sandy sediment in which to burrow.  Hediste diversicolor has been identified in other intertidal sediments including gravels, clays and even turf (Clay, 1967; Scaps, 2002), although abundance may be reduced in these habitats. Limecola balthica is likely to tolerate increased gravel content as sediment was not shown to affect burrowing (Tallqvist, 2001), however, growth, shell size and body mass were greatest in higher sand content sediment and lower in higher gravel content sediments (Azouzi et al. 2002), suggesting long term health and abundance may be affected by a permanent increase in  gravel content.

    Coarser sediments provide inhospitable conditions for colonizing infauna, although Pygospio elegans show greater tolerance of increased gravel content and are viewed as opportunistic species that are capable of exploiting these inhospitable conditions (Gray, 1981). Therefore, Pygospio elegans are likely to be less affected and even increase in abundance under a change in Folk class to gravelly mud (or a change from sandy mud to muddy sand, or gravelly muddy sand). Capitella capitata are likely to decrease in abundance as Degraer et al. (2006) found that Capitella capitata in the Belgium part of the North Sea were almost completely absent in sediments without mud. Similar species that prefer higher organic content may also show limited abundance in more gravelly sediments.

    Sensitivity assessment.  Case studies display decreasing abundance with increased gravel content (Hediste diversicolor) and reduced growth rates (Limecola balthica). Abundance of abundant polychaetes is likely to depend on each species tolerance of increasing gravel content, with species that can exploit the conditions increasing in abundance (Pygospio elegans) but other species decreasing in abundance. Resistance to a change in one Folk class is assessed as ‘Low’ as increased gravel content is likely to lead to reduced abundance of characterizing species and result in biotope reclassification to the mixed sediment biotope LS.LMx.GvMu . Resilience is assessed as ‘Very Low’ as a change at the benchmark is permanent. The sensitivity of the biotope overall is, therefore, considered to be ‘High’.

    None High Medium
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: Medium
    Q: High
    A: High
    C: Medium

    The substratum of this biotope consists of gravelly sandy mud or gravelly mud (Conner et al., 2004).  The characterizing infaunal species, including Hediste diversicolor, Eteone longa and Limecola balthica burrow into the sediment, to depths not exceeding 30 cm.  The process of extraction is considered to remove all biological components of the biotope group in the impact footprint and the sediment habitat.

    Sensitivity assessment. Resistance to extraction of substratum to 30 cm across the entire biotope is assessed as ‘None’ based on expert judgment but supported by the literature relating to the position of these species on or within the seabed and literature on impacts of dredging and bait digging activities (see penetration and disturbance pressure).  At the pressure benchmark the exposed sediments are considered to be suitable for recolonization almost immediately following extraction.  Recovery will be mediated by the scale of the disturbance and the suitability of the sedimentary habitat, biotope resilience is assessed as 'High' (based on recolonization by adults and pelagic larvae) and biotope sensitivity is assessed as 'Medium'.

    Medium High Low
    Q: High
    A: High
    C: Medium
    Q: High
    A: High
    C: Medium
    Q: High
    A: Medium
    C: Medium

    Muddy sand sediments, in general, tend to be cohesive although high levels of water content will reduce this and destabilise sediments. Sediment cohesion provides some sediment stabilisation to resist erosion following surface disturbance. The characterizing species associated with this biotope are infaunal and hence have some protection against surface disturbance, although siphons of bivalves and tubes of the sedentary polychaete Pygospio elegans, may project above the sediment surface. Damage to tubes and siphons would require repair. The snail Hydroia ulvae is present on the surface and abrasion may result in burial or damage to this speciesSurface 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).

    Sensitivity assessment. Resistance is assessed as ‘Medium', as abrasion is unlikely to affect high numbers of infaunal burrowing species such as the key characterizing species  Hediste diversicolor and the oligochaetes but bivalves, tube dwelling polychaetes and Hydrobia ulvae, may be reduced in abundance. Resilience is assessed as 'High' and biotope sensitivity is assessed as 'Low'.

    Low High Low
    Q: High
    A: High
    C: Medium
    Q: High
    A: High
    C: Medium
    Q: High
    A: High
    C: Medium

    As the characterizing species are burrowing species, the impact from damage to the sub-surface sea bed would be greater than damage to the sea bed surface only (see abrasion pressure). A number of studies have assessed the impacts of activities resulting in penetration and disturbance of sediments on the characterizing species in similar habitats. The characterizing species have some protective traits such as infaunal life habit, with deeper burrowing species less exposed. The shells of Limecola balthica provide some protection. Pygospio elegans inhabits fragile tubes at the sediment surface and Hydrobia ulvae crawl on the sediment, both species are likely to be vulnerable to penetration and disturbance of the sediment.

    The effects of a pipeline construction on benthic invertebrates were also investigated using a Before/After impact protocol at Clonakilty Bay, West Cork, Ireland. Benthic invertebrates were sampled once before the excavation and at one, two, three and six months after the completion of the work. Invertebrate samples were dominated by Hediste diversicolorScrobicularia plana and Tubifex spp. An impact was obvious in the construction site in that no live invertebrates were found at one month after disturbance, but there followed a gradual recolonisation by Hediste diversicolor. At six months after the disturbance there was no significant difference in the mean number of total individuals (of all species) per core sample amongst all study sites, but the apparent recovery in the impacted area was due to two taxa only, Hediste diversicolor and Tubifex spp. (Lewis et al., 2002).

    Ferns et al. (2000) studied effects of harvesting of cockles and reported a decline in muddy sands of 83% in Pygospio elegans (initial density 1850 m−2) when a mechanical tractor towed harvester was used in a cockle fishery. Pygospio elegans and Hydrobia ulvae were significantly depleted for >100 days after harvesting (surpassing the study monitoring timeline).

    Rees, 1978, (cited in Hiscock et al., 2002), assessed pipe laying activities. The pipe was laid in a trench dug by excavators and the spoil from the trenching was then used to bury the pipe. The trenching severely disturbed a narrow zone, but a zone some 50 m wide on each side of the pipeline was also disturbed by the passage of vehicles. The tracked vehicles damaged and exposed shallow-burrowing species such as the bivalves Cerastoderma edule and Limecola balthica, which were then preyed upon by birds. During the construction period, the disturbed zone was continually re-populated by mobile organisms, such as Hydrobia ulvae.

    Sensitivity assessment.  Resistance of the biotope is assessed as ‘Low’, although the significance of the impact for the bed will depend on the spatial scale of the pressure footprint.  Resilience is assessed as ‘High’, and sensitivity is assessed as ‘Low’.

    High High Not sensitive
    Q: Low
    A: NR
    C: NR
    Q: High
    A: High
    C: High
    Q: Low
    A: Low
    C: Low

    Changes in light penetration or attenuation associated with this pressure are not relevant to Hediste diversicolor and Limecola balthica biotopes. As the species live in the sediment they are also likely to be adapted to increased suspended sediment (and turbidity). However, alterations in the availability of food or the energetic costs in obtaining food or changes in scour could either increase or decrease habitat suitability for Hediste diversicolor, Limecola balthica as characterizing species and for other abundant species such as the oligochaetes Tubificoides benedii and Tubificoides pseudogaster.

    Hediste diversicolor characteristically inhabits estuaries where turbidity is typically higher than other coastal waters. Changes in the turbidity may influence the abundance of phytoplankton available as a food source that may be attained through filter feeding. Hediste diversicolor utilizes various other feeding mechanisms and, at the benchmark level, the likely effects of a change in one rank on the WFD scale are limited.

    Sensitivity assessment. The following sensitivity assessment relies on expert judgement, utilising evidence of species traits and distribution and therefore confidence has been assessed as low. Resistance is ‘High’ as no significant negative effects are identified and potential benefits from increased food resources may occur. Resilience is also ‘High’ as no recovery is required under the likely impacts. Sensitivity of the biotope is, therefore, assessed as ‘Not Sensitive’.

    Medium High Low
    Q: High
    A: High
    C: Medium
    Q: High
    A: High
    C: High
    Q: High
    A: High
    C: Medium

    The degree to which the characterizing species are able to resist this pressure depends primarily on species mobility, ability to survive within sediment without contact with the surface and ability to escape from the over-burden. Factors that affect the ability to regain the surface include grain size (Maurer et al., 1986), temperature and water content (Chandrasekara & Frid, 1998).

    Mobile polychaetes have been demonstrated to burrow through thick layers of deposits. Powilleit et al., (2009) studied the response of the polychaete Nephtys hombergii to smothering. This species successfully migrated to the surface of 32-41 cm deposited sediment layer of till or sand/till mixture and restored contact with the overlying water.   While crawling upward to the new sediment surfaces burrowing velocities of up to 20 cm/day were recorded for Nephtys hombergii. Similarly, Bijkerk (1988, results cited from Essink 1999) indicated that the maximal overburden through which species could migrate was 60 cm through mud for Nephtys and 90 cm through sand. No further information was available on the rates of survivorship or the time taken to reach the surface.

    Tubificoides spp. and other oligochaetes live relatively deeply buried and can tolerate periods of low oxygen that may occur following the deposition of a fine layer of sediment.  Tubificoides spp. showed some recovery through vertical migration following the placement of a sediment overburden 6cm thick on top of sediments (Bolam, 2011). Whomersley et al., (2010) experimentally buried plots on intertidal mudflats at two sites (Creeksea- Crouch Estuary, England and Blackness- lower Forth Estuary, Scotland), where Tubificoides benedii were dominant species. For each treatment anoxic mud was spread evenly to a depth of 4 cm on top of each treatment plot. The mud was taken from areas adjacent to the plots, and was obtained by scraping off the surface oxic layer and digging up the underlying mud from approximately 20 cm depth. Plots were subject to either low intensity treatments (burial every four weeks) or high (burial every two weeks). The experiment was carried out for 10 months at Creeksea and a year at Blackness. At Creeksea numbers of Tubificoides benedii increased in both burial treatments until the third month (high burial) and sixth month (low burial). At Blackness increased numbers ofTubificoides benedii  were found in both burial treatments after one month (Whomersley et al., 2010).

    Laboratory experiments have shown that the snail Hydrobia ulvae can rapidly resurface through 5cm thick fine deposits, although this ability is reduced where deposited sediments contain little water (Chandrasekara & Frid, 1998). 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).

    Limecola balthica is able to burrow both vertically and horizontally through the substratum. It is likely that Limecola balthica is not sensitive to smothering by a layer of sediment 5 cm thick as it is a mobile species able to burrow upwards and surface from a depth of 5 - 6 cm (Brafield & Newell, 1961; Brafield, 1963; Stekoll et al., 1980). Turk and Risk (1981) investigated the effect of experimentally induced sedimentation (through fences and boxes that induced sediment deposition on intertidal mudflats in the Bay of Fundy), of 1-3.5cm at a rate of 1.9-10.2 cm/month. The results showed that Limecola balthica was generally unaffected.

    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).

    The amphipod Corophium volutator may be sensitive to deposits at the pressure benchmark. Experimental fences placed on mudflats that caused sedimentation rates of 2-2.5 cm/month and reduced Corophium volutator densities from approximately 1700 m² to approximately 400 m². In areas without fences,Corophium volutator numbers increased from approximately 1700 per m² to 3500 per m² (Turk & Risk, 1981).

    In intertidal mudflats with similar characterizing species, experiments testing the effects of deposition of sediments typical of beach recharge, have found that recovery of biological assemblages is complete within two years (Bolam & Whomersley, 2003).

    Sensitivity assessment. As the exposure to the pressure is for a single discrete event, resistance is assessed as ‘Medium’ as some species associated with the biotope such as Corophium volutator and Pygospio elegans may decline but the biotope is likely to be recognizable within a week due to repositoing and migration of mobile species. Resilience is assessed as ‘High’ and sensitivity is assessed as ‘Low

    Low High Low
    Q: High
    A: High
    C: Medium
    Q: High
    A: Low
    C: Medium
    Q: High
    A: Low
    C: Medium

    Studies have found that beach ‘replenishment’ or ‘nourishment’ that involves the addition of sediments on beaches can have a number of impacts on the infauna (Peterson et al., 2000; Peterson et al., 2006). Impacts are more severe when the sediment added differs significantly in grain size or organic content from the natural habitat (Peterson et al., 2000).

    Hediste diversicolor inhabits depositional environments. It is capable of burrowing to depths of up to 0.3 m and reworking sub-surface modifications of its burrow through fine clays and sand. Smith (1955) found no appreciable difference in the population of a Hediste diversicolor colony which had been covered by several inches of sand through which the worms tunnelled. Mobile polychaetes have been demonstrated to burrow through thick layers of deposits. Powilleit et al., (2009) studied the response of the polychaete Nephtys hombergii to smothering. This species successfully migrated to the surface of 32-41 cm deposited sediment layer of till or sand/till mixture and restored contact with the overlying water.   While crawling upward to the new sediment surfaces burrowing velocities of up to 20 cm/day were recorded for Nephtys hombergii. Similarly, Bijkerk (1988, results cited from Essink 1999) indicated that the maximal overburden through which species could migrate was 60 cm through mud for Nephtys and 90 cm through sand. No further information was available on the rates of survivorship or the time taken to reach the surface.

    Witt et al. (2004) identified an increase in Limecola balthica (as Macoma balthica) abundance in areas of disposal of dredge waste spoil, possibly due to nutrient input at the disposal site. This suggests Limecola balthica responds opportunistically to this pressure.

    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).

    The amphipod Corophium volutator may be sensitive to deposits at the pressure benchmark. Experimental fences placed on mudflats that caused sedimentation rates of 2-2.5 cm/month and reduced Corophium volutator densities from approximately 1700 m² to approximately 400 m². In areas without fences,Corophium volutator numbers increased from approximately 1700 per m² to 3500 per m² (Turk & Risk, 1981).

    In intertidal mudflats with similar characterizing species, experiments testing the effects of deposition of sediments, typical of beach recharge have found that recovery of biological assemblages is complete within two years (Bolam & Whomersley, 2003).

    Sensitivity assessment. Deposition of up to 30 cm of fine material is likely to provide different impacts for the different species characterizing the biotope. Overall, although the characterizing species have some resistance to this to this pressure, but populations are likely to be reduced. Resistance to initial smothering is ‘Low’ Resilience is ‘High’ and biotope sensitivity is assessed as ‘Low’.

    Not Assessed (NA) Not assessed (NA) Not assessed (NA)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    Examples of the impact of specific marine litter, including cigarette butts and micro-plastics are also considered..

    Litter, in the form of cigarette butts has been shown to have an impact on ragworms. Hediste diversicolor showed increased burrowing times, 30% weight loss and a  >2 fold increase in DNA damage when exposed to water with toxicants (present in cigarette butts) in quantities 60 fold lower than reported from urban run-off (Wright et al., 2015). This UK study suggests health of infauna populations are negatively impacted by this pressure.

    Studies of other characterizing species in relation to micro plastics were not available. However, studies of sediment dwelling, sub surface deposit feeding worms, showed negative impacts from ingestion of micro plastics. For instance, Arenicola marina ingests micro-plastics that are present within the sediment it feeds within. Wright et al. (2013) carried out a lab study that displayed presence of micro-plastics (5% UPVC) significantly reduced feeding activity when compared to concentrations of 1% UPVC and controls. As a result, Arenicola marina showed significantly decreased energy reserves (by 50%), took longer to digest food, and as a result decreased bioturbation levels which would be likely to impact colonisation of sediment by other species, reducing diversity in the biotopes the species occurs within. Wright et al. (2013) also present a case study based on their results, that in the intertidal regions of the Wadden Sea, where Arenicola marina is an important ecosystem engineer, Arenicola marina could ingest 33 m² of micro-plastics a year.

    Sensitivity assessment. Marine litter in the form of cigarette butts or micro plastics health of populations of characterizing species may be impacted. Significant impacts have been shown in laboratory studies but impacts at biotope scales are still unknown. Evidence and confidence in the assessment is limited and this pressure is 'Not assessed'.

    No evidence (NEv) No evidence (NEv) No evidence (NEv)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    There is no evidence on effects of electric and magnetic fields on the characterizing species.

    Electric and magnetic fields generated by sources such as marine renewable energy device/array cables may alter behaviour of predators and affect infauna populations. Evidence is limited and occurs for electric and magnetic fields below the benchmark levels, confidence in evidence of these effects is very low.

    Field measurements of electric fields at North Hoyle wind farm, North Wales recorded 110µ V/m (Gill et al., 2009). Modelled results of magnetic fields from typical subsea electrical cables, such as those used in the renewable energy industry produced magnetic fields of between 7.85 and 20 µT (Gill et al. 2009; Normandeau et al. 2012). Electric and magnetic fields smaller than those recorded by in field measurements or modelled results were shown to create increased movement in potential predators of Hediste diversicolor, thornback ray Raja clavata and attraction to the source in catshark Scyliorhinus canicular (Gill et al. 2009).

    Flatfish species which are predators of many polychaete species including dab Limanda limanda and sole Solea solea have been shown to decrease in abundance in a wind farm array or remain at distance from wind farm towers (Vandendriessche et al., 2015; Winter et al. 2010). However, larger plaice increased in abundance (Vandendriessche et al., 2015). There have been no direct causal links identified to explain these results.

    Not relevant (NR) Not relevant (NR) Not relevant (NR)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    Species within the biotope can probably detect vibrations caused by noise and in response may retreat in to the sediment for protection. However, at the benchmark level the community is unlikely to be sensitive to noise and this therefore is ‘Not relevant’.

    Not relevant (NR) Not relevant (NR) Not relevant (NR)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    There is no direct evidence of effects of changes in incident light on the characterizing species of this biotope. All characterizing species live in the sediment and do not rely on light levels directly to feed or find prey so limited direct impact is expected.  As this biotope is not characterized by the presence of primary producers, shading is unlikely to alter the character of the habitat directly and this pressure is considered 'Not relevant'.

     

    More general changes to the productivity of the biotope may, however, occur. Beneath shading structures there may be changes in microphytobenthos abundance. Littoral mud and sand support microphytobenthos on the sediment surface and within the sediment. Mucilaginous secretions produced by these algae may stabilise fine substrata (Tait & Dipper, 1998).

     

    Shading will prevent photosynthesis leading to death or migration of sediment microalgae altering sediment cohesion and food supply to higher trophic levels. The impact of these indirect effects is difficult to quantify.

    Sensitivity assessment. Based on the direct impact, biotope resistance is assessed as ‘High’ and resilience is assessed as ‘High’ (by default) and the biotope is considered to be ‘Not sensitive’.

    High High Not sensitive
    Q: Low
    A: NR
    C: NR
    Q: High
    A: High
    C: High
    Q: Low
    A: Low
    C: Low

    Barriers that reduce the degree of tidal excursion may alter larval supply to suitable habitats from source populations. Barriers may also act as stepping stones for larval supply over greater distances (Adams et al., 2014).

    Conversely, the presence of barriers at brackish waters may enhance local population supply by preventing the loss of larvae from enclosed habitats to environments, which are unfavourable, reducing settlement outside of the population.

    If a barrier (such as a tidal barrier) incorporated renewable energy  devices such as tidal energy turbines, these devices may affect hydrodynamics and so migration pathways for larvae into and out of the biotope (Adams et al., 2014). Evidence on this pressure is limited.

    The trait of Hediste diversicolor to lay and protect eggs within a burrow is likely to limit the impact of barriers to movement on populations. The ability of postlarvae, larger juveniles, and adults of Hediste diversicolor to swim, burrow and be carried by bedload transport can aid the rapid recolonization of disturbed sediments (Shull, 1997).  Davey & George (1986), found evidence that larvae of Hediste diversicolor were tidally dispersed within the Tamar Estuary over a distance of 3 km. A barrier to movement is likely to limit colonization outside the enclosed area, but increase populations within the enclosed area  

    Capitella capitata and the associated species Pygospio elegans are capable of both benthic and pelagic dispersal. In the sheltered waters where this biotope occurs, with reduced water exchange, in-situ reproduction may maintain populations rather than long-range pelagic dispersal. As the tubificid oligochaetes that characterize this biotope have benthic dispersal strategies (via egg cocoons laid on the surface (Giere & Pfannkuche, 1982), water transport is not a key method of dispersal over wide distances. 

    Sensitivity assessment. Resistance to this pressure is assessed as 'High' and resilience as 'High' by default. This biotope is therefore considered to be 'Not sensitive'.

    Not relevant (NR) Not relevant (NR) Not relevant (NR)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    Not relevant’ to seabed habitats.  NB. Collision by interaction with bottom towed fishing gears and moorings are addressed under

    Not relevant (NR) Not relevant (NR) Not relevant (NR)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

    Characterizing species and have limited, visual perception, this pressure is therefore considered 'Not relevant'.

    Biological Pressures

     ResistanceResilienceSensitivity
    Not relevant (NR) Not relevant (NR) Not relevant (NR)
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR
    Q: NR
    A: NR
    C: NR

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

    Low Very Low High
    Q: High
    A: High
    C: High
    Q: Low
    A: NR
    C: NR
    Q: Low
    A: Low
    C: Low

    Intertidal sediments may be colonized by a number of invasive non-indigenous species. Invasive species that alter the character of the biotope or that predate on characterizing species are most likely to result in significant impacts. Intertidal flats may be colonized by the invasive non-indigenous species 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 and food chain changes (OSPAR, 2009b).

    In the Wadden Sea, the Pacific oyster Magallana gigas has colonized intertidal flats (Smaal et al., 2005). This species consumes pelagic larvae reducing recruitment (Smaal et al., 2005).   The most severe effects are likely to occur from impacts on sediment, where Magallana gigas create reefs on sedimentary flats that will prevent recruitment of juveniles and will restrict access of infauna to the sediment-water interface impacting respiration and feeding of bivalves such as Limecola balthica and Scrobicularia plana and polychaetes such as Pygospio elegans and disturbing the amphipod Corophium volutator. Burrowing infauna such as Hediste diversicolor and oligochaetes may persist within sediments but the overall character of the mixed sediment biotope would be altered.

    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. 2004, cited in Caldow et al. 2007).  Densities of Cerastoderma edule and Abra tenuis increased following the introduction of the Manila clam although the abundance of Scrobicularia plana and Limecola balthica declined (Caldow et al., 2005), although the decline of these species may have been caused by tri-butyl tin pollution (Langston et al., 2003) and may have facilitated the naturalization of the Manila clam.

    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.

    Sensitivity assessment. Intertidal muddy sands 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 biotope is considered to have ‘Low’ resistance and ‘Very low’ recovery (unless invasive species are removed). Biotope sensitivity is, therefore, assessed as ‘High’.

    High High Not sensitive
    Q: Low
    A: NR
    C: NR
    Q: High
    A: High
    C: High
    Q: Low
    A: Low
    C: Low

    Limited evidence was returned by searches on the effect on Hediste diversicolor of introduction of relevant microbial pathogens or metazoan disease vectors to an area where they are currently not present. Desrina et al. (2014) failed to induce infection of the ‘White Spot Shrimp’ virus that can cause large scale mortality in shrimp in Hediste diversicolor by both feeding and immersion. 

    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).  Parasites and disease are more likely to cause mortalities in populations that are subject to suboptimal 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).

    Sensitivity assessment. Biotope resistance is assessed as ‘High’ as mass mortalities of key characterizing species have not been reported, resilience is assessed as ‘High’ and sensitivity is, therefore assessed as ‘Not sensitive’.

    Low High Low
    Q: High
    A: Medium
    C: Medium
    Q: High
    A: Medium
    C: Medium
    Q: High
    A: Medium
    C: High

    Ragworms Hediste diversicolor are harvested as bait for angling while the abundant bivalves in the biotope Cerastoderma edule and Mya arenia are harvested commercially and recreationally. Removal of target species such as cockles Cerastoderma edule or bait digging for Hediste diversicolor is likely to impact the biotope. The extent of the impact will depend on the fishing / removal method and spatial extent.

    Hall & Harding (1997) examined the effects of hydraulic and tractor dredging of Cerastoderma edule on macrobenthic communities. They concluded that although significant mortality of Cerastoderma edule and other infauna occurred, recovery was rapid and the overall effects on populations was low. Hall & Harding (1997) found that Cerastoderma edule abundance had returned to control levels within about 56 days and Moore (1991) also suggested that recovery was rapid. Cotter et al. (1997) noted that tractor dredging reduced the Cerastoderma edule stock by 31-49% depending on initial density, while Pickett (1973) reported that hydraulic dredging removed about one third of the cockle fishery. Tractor dredging leaves visible tracks in the sediment, which can act as lines for erosion and accelerate erosion of the sediment (Moore, 1991; Gubbay & Knapman, 1999). In most cases subsequent settlement was good especially in areas of previously high population density, however, Franklin & Pickett (1978) noted that subsequent spat survival was markedly reduced. Cotter et al. (1997) reported appreciable loss of spat and juveniles, partly due increased predation of exposed juveniles. Pickett (1973) also noted reduced survivability of 1-2 year old cockles after hydraulic dredging. However, most studies concluded that the impact of mechanised dredging on cockle populations and macrofauna in the long term was low (Pickett, 1973; Franklin & Pickett, 1978; Cook, 1990; Moore, 1991; Cotter et al., 1997; Hall & Harding, 1997; Ferns et al., 2000). Time of year of exploitation will influence recovery and avoiding seasonal spawning or larval settlement periods is likely to reduce the time taken for recovery (Gubbay & Knapman, 1999). Cockle beds have been mechanically fished for decades but several beds are closed from time to time depending on settlement and recruitment to the population, which is sporadic. Recovery may take less than a year in years of good recruitment but longer in bad years.

    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., (1992) observed that Pygospio elegans established dense populations; when cockles returned these were lost within one year.

     

     

    Sensitivity assessment. Removal at a commercial or recreational scale is assessed as not affecting the entire extent of the biotope, but affecting patches within the biotope. Due to potential impacts on Hediste diversicolor populations, in particular females, and impacts on Cerastoderma edule populations the biotope is likely to be sensitive to this pressure. 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 the removal of cockles resulting in reduced competition for space and predation on larvae). Where sediments remain suitable cockles are likely to recolonise via adult migration, survival of small, discarded cockles or via larval recruitment.  In general fishing practices will be efficient at removing Hediste diversicolor and Cerastoderma edule. 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 ‘High’ (although Cerastoderma edule many not recover within this timescale due to episodic recruitment; see Cerastoderma edule dominated biotopes). Biotope sensitivity is assessed as ‘Low’. 

    Low High Low
    Q: Low
    A: NR
    C: NR
    Q: High
    A: Medium
    C: Medium
    Q: Low
    A: Low
    C: Low

    Direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures, while this pressure considers the ecological or biological effects of by-catch. Species in these biotopes, including the characterizing species, may be damaged or directly removed by static or mobile gears that are targeting other species (see abrasion and penetration pressures). Loss of these species would alter the character of the biotope resulting in re-classification, and would alter the physical structure of the habitat resulting in the loss of the ecosystem functions such as secondary production performed by these species.

    McLusky et al. (1983) found that Limecola balthica populations were unaffected in dug areas, following bait digging for lugworms suggesting the biotope would recover from this impact if it occurred over a limited spatial scale.  However, Hiddink (2003) shows density of Limecola balthica was reduced in areas in the Wadden sea (Netherlands) that had experienced suction dredging for cockles, which removes the surface sediment.  

    Incidental removal of the characterizing species would alter the character of the biotope and the delivery of ecosystem services such as secondary production and bioturbation. Populations of characterizing species, such asHediste diversicolor and Limecola balthica provide food for macroinvertebrates fish and birds.

    The removal of predators such as shrimp and crab may enhance recruitment of larvae of characterizing species (Beukema & Dekker, 2005). The physical effects of removal of other species such as polychaete worms targeted by bait diggers may, however, impact Eteone longaCerastoderma edule and other species associated with this biotope, through direct damage and removal and displacement. These direct effects of sediment disturbance are assessed in the physical damage sections.

    The removal of Hediste diversicolor  and Cerastoderma edule 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 Hediste diversicolor and Cerastoderma edule and other species as by-catch would alter the character of the biotope. As Cerastoderma edule and Limecola balthica and other abundant bivalve species are either sedentary or incapable of rapid evasive movements, resistance is assessed as ‘Low’. Resilience is assessed as ‘High’.  Cerastoderma edule may not be recovered in this timescale due to episodic recruitment but this is not considered likely to alter biotope classification(see Cerastoderma edule dominated biotopes) and sensitivity is therefore categorized as ‘Low’.  Physical damage to the sediment and other physical damage factors are considered in the abrasion and extraction pressures.

    Bibliography

    1. Mettam, C., Santhanam, V. & Havard, M.C.S., 1982. The oogenic cycle of Nereis diversicolor under natural conditions. Journal of the Marine Biological Association of the United Kingdom, 62, 637-645.

    2. Aberson, M., Bolam, S. & Hughes, R., 2016. The effect of sewage pollution on the feeding behaviour and diet of Hediste (Nereis diversicolor (OF Müller, 1776)) in three estuaries in south-east England, with implications for saltmarsh erosion. Marine Pollution Bulletin, 105 (1), 150-160.

    3. Adams, T.P., Miller, R.G., Aleynik, D. & Burrows, M.T., 2014. Offshore marine renewable energy devices as stepping stones across biogeographical boundaries. Journal of Applied Ecology, 51 (2), 330-338.

    4. Akberali, H.B. & Black, J.E., 1980. Behavioural responses of the bivalve Scrobicularia plana (da Costa) subjected to short-term copper (Cu II) concentrations. Marine Environmental Research, 4 (2), 97-107.

    5. Anger, K., 1975a. Quantitative studies on indicator species and communities. Merentutkimuslaitoksen Julkaisau. Helsinki, 239, 116-122.

    6. Anger, K., 1975b. On the influence of sewage pollution on inshore benthic communities in the south of Kiel Bay. Helgolander Wissenschaftliche Meeresuntersuchungen, 27, 408-438.

    7. Anger V., 1984. Reproduction in Pygospio-elegans Spionidae in relation to its geographical origin and to environmental conditions a preliminary report. Fischer, A. and H.-D. Pfannenstiel,  Fortschritte der Zoologie. pp. 45-52.

    8. Anonymous, 1999iii. UK Biodiversity Group: tranche 2 action plans: volume V- maritime species and habitats. , English Nature, Peterborough, UK.

    9. Armonies, W., Herre, E. & Sturm, M., 2001. Effects of the severe winter 1995/96 on the benthic macrofauna of the Wadden Sea and the coastal North Sea near the island of Sylt. Helgoland Marine Research, 55, (3), 170-175.

    10. Armonies, W., Herre, E. & Sturm, M., 2001. Effects of the severe winter 1995 / 1996 on the benthic macrofauna of the Wadden Sea and the coastal North Sea near the island of Sylt. Helgoland Marine Research, 55, 170-175.

    11. Attrill, M.J. ed., 1998. A rehabilitated estuarine ecosystem: The environment and ecology of the Thames estuary. Berlin: Springer Science & Business Media.

    12. Azouzi, L., Bourget, E. & Borcard, D., 2002. Spatial variation in the intertidal bivalve Macoma balthica: biotic variables in relation to density and abiotic factors. Marine Ecology Progress Series, 234, 159-170.

    13. Bagheri, E. & McLusky, D., 1982. Population dynamics of oligochaetes and small polychaetes in the polluted forth estury ecosystem. Netherlands Journal of Sea Research, 16, 55-66.

    14. Barda, I., Purina, I., Rimsa, E. & Balode, M., 2014. Seasonal dynamics of biomarkers in infaunal clam Macoma balthica from the Gulf of Riga (Baltic Sea). Journal of Marine Systems, 129, 150-156.

    15. Barnes, R.S.K. & Hughes, R.N., 1992. An introduction to marine ecology. Oxford: Blackwell Scientific Publications.

    16. Barnes, R.S.K., 1994. The brackish-water fauna of northwestern Europe. Cambridge: Cambridge University Press.

    17. Bartels-Hardege, H.D. & Zeeck, E., 1990. Reproductive behaviour of Nereis diversicolor (Annelida: Polychaeta). Marine Biology, 106, 409-412.

    18. Beasley, T.M. & Fowler, S.W., 1976. Plutonium and Americium: uptake from contaminated sediments by the polychaete Nereis diversicolor. Marine Biology, 38, 95-100.

    19. Beaumont, A.R., Newman, P.B., Mills, D.K., Waldock, M.J., Miller, D. & Waite, M.E., 1989. Sandy-substrate microcosm studies on tributyl tin (TBT) toxicity to marine organisms. Scientia Marina, 53, 737-743.

    20. Berthet, B., Mouneyrac, C., Amiard, C.J., Amiard-Triquet, C., Berthelot, Y., Le Hen, A., Mastain, O., Rainbow, S.P. & Smith, D.B., 2003. Accumulation and Soluble Binding of Cadmium, Copper, and Zinc in the Polychaete Hediste diversicolor from Coastal Sites with Different Trace Metal Bioavailabilities. Archives of Environmental Contamination and Toxicology, 45 (4), 468-478.

    21. Beukema, J.J., 1989. Long term changes in macrozoobenthic abundance on the tidal flats of the western part of the Dutch Wadden Sea. Helgolander Meeresuntersuchungen, 43 (3-4), 405-415.

    22. Beukema, J.J., 2002. Expected changes in the benthic fauna of Wadden Sea tidal flats as a result of sea-level rise or bottom subsidence. Journal of Sea Research, 47 (1), 25-39.

    23. Beukema, J.J. & De Vlas, J., 1979. Population parameters of the lugworm, Arenicola marina, living on tidal flats in the Dutch Wadden Sea. Netherlands Journal of Sea Research, 13, 331-353.

    24. Beukema, J.J. & Dekker, R., 2005. Decline of recruitment success in cockles and other bivalves in the Wadden Sea: possible role of climate change, predation on postlarvae and fisheries. Marine Ecology Progress Series, 287, 149-167.

    25. Beukema, J.J., 1990. Expected effects of changes in winter temperatures on benthic animals living in soft sediments in coastal North Sea areas. In Expected effects of climatic change on marine coastal ecosystems (ed. J.J. Beukema, W.J. Wolff & J.J.W.M. Brouns), pp. 83-92. Dordrecht: Kluwer Academic Publ.

    26. Beukema, J.J., 1992. Expected changes in the Wadden Sea benthos in a warmer world - lessons from periods with mild winters. Netherlands Journal of Sea Research, 30, 73-79.

    27. Beukema, J.J., 1995. Long-term effects of mechanical harvesting of lugworms Arenicola marina on the zoobenthic community of a tidal flat in the Wadden Sea. Netherlands Journal of Sea Research, 33, 219-227.

    28. Beukema, J.J., Cadee, G.C., Dekker, R. & Philippart, C.J.M., 2014. Annual and spatial variability in gains of body weight in Macoma balthica (L.): Relationships with food supply and water temperature. Journal of Experimental Marine Biology and Ecology, 457, 105-112.

    29. Beukema, J.J., Dekker, R., Essink, K. & Michaelis, H., 2001. Synchronized reproductive success of the main bivalve species in the Wadden Sea: causes and consequences. Marine Ecology Progress Series, 211, 143-155.

    30. Bijkerk, R., 1988. Ontsnappen of begraven blijven: de effecten op bodemdieren van een verhoogde sedimentatie als gevolg van baggerwerkzaamheden: literatuuronderzoek: RDD, Aquatic ecosystems.

    31. Black, K.D., Fleming, S. Nickell, T.D. & Pereira, P.M.F. 1997. The effects of ivermectin, used to control sea lice on caged farmed salmonids, on infaunal polychaetes. ICES Journal of Marine Science, 54, 276-279.

    32. Black, K.D., Fleming, S. Nickell, T.D. & Pereira, P.M.F. 1997. The effects of ivermectin, used to control sea lice on caged farmed salmonids, on infaunal polychaetes. ICES Journal of Marine Science, 54, 276-279.

    33. Boisson, F., Hartl, M.G.J., Fowler, S.W. & Amiard-triquet, C., 1998. Influence of chronic exposure to silver and mercury in the field on the bioaccumulation potential of the bivalve Macoma balthica. Marine Environmental Research, 45, 325-340.

    34. Bolam, S. & Whomersley, P., 2003. Invertebrate recolonization of fine-grained beneficial use schemes: An example from the southeast coast of England. Journal of Coastal Conservation, 9 (2), 159-169.

    35. Bolam, S., Whomersley, P. & Schratzberger, M., 2004. Macrofaunal recolonization on intertidal mudflats: effect of sediment organic and sand content. Journal of Experimental Marine Biology and Ecology, 306 (2), 157-180.

    36. Bolam, S.G., 2011. Burial survival of benthic macrofauna following deposition of simulated dredged material. Environmental Monitoring and Assessment, 181 (1-4), 13-27.

    37. Bolam, S.G. & Fernandes, T.F., 2002. Dense aggregations of tube-building polychaetes: response to small-scale disturbances. Journal of Experimental Marine Biology and Ecology, 269, 197-222.

    38. Boldina-Cosqueric, I., Amiard, J.-C., Amiard-Triquet, C., Dedourge-Geffard, O., Metais, I., Mouneyrac, C., Moutel, B. & Berthet, B., 2010. Biochemical, physiological and behavioural markers in the endobenthic bivalve Scrobicularia plana as tools for the assessment of estuarine sediment quality. Ecotoxicology and Environmental Safety, 73 (7), 1733-1741.

    39. Bonnard, M., Romeo, M. & Amiard-Triquet, C., 2009. Effects of copper on the burrowing behavior of estuarine and coastal invertebrates, the polychaete Nereis diversicolor and the bivalve Scrobicularia plana. Human and Ecological Risk Assessment, 15 (1), 11-26.

    40. Bonsdorff, E., 1984. Establishment, growth and dynamics of a Macoma balthica (L.) population. Limnologica (Berlin), 15, 403-405.

    41. Bonsdorff, E., Norrko, A. & Boström, C., 1995. Recruitment and population maintenance of the bivalve Macoma balthica (L.) - factors affecting settling success and early survival on shallow sandy bottoms. In Proceedings of the 28th European Marine Biology Symposium. Biology and ecology of shallow coastal waters (ed. A. Eleftheriou, A.D. Ansell and C.J. Smith). Olsen and Olsen.

    42. Borja, A., Franco, J. & Perez, V., 2000. A marine biotic index to establish the ecological quality of soft-bottom benthos within European estuarine and coastal environments. Marine Pollution Bulletin, 40 (12), 1100-1114.

    43. Bouchet, V.M. & Sauriau, P.-G., 2008. Influence of oyster culture practices and environmental conditions on the ecological status of intertidal mudflats in the Pertuis Charentais (SW France): A multi-index approach. Marine Pollution Bulletin, 56 (11), 1898-1912.

    44. Boyden, C.R., 1972. Behaviour, survival and respiration of the cockles Cerastoderma edule and C. glaucum in air. Journal of the Marine Biological Association of the United Kingdom, 52, 661-680.

    45. Braeckman, U., Foshtomi, M.Y., Gansbeke, D., Meysman, F., Soetaert, K., Vincx, M. & Vanaverbeke, J., 2014. Variable importance of macrofaunal functional biodiversity for biogeochemical cycling in temperate coastal sediments. Ecosystems, 17 (4), 720-737.

    46. Brafield, A.E. & Newell, G.E., 1961. The behaviour of Macoma balthica (L.). Journal of the Marine Biological Association of the United Kingdom, 41, 81-87.

    47. Brafield, A.E., 1963. The effects of oxygen deficiency on the behaviour of Macoma balthica. Animal Behaviour, 11, 245-346.

    48. Brenchley, G.A., 1981. Disturbance and community structure : an experimental study of bioturbation in marine soft-bottom environments. Journal of Marine Research, 39, 767-790.

    49. Breum, O., 1970. Stimulation of burrowing activity by wave action in some marine bivalves. Ophelia, 8 (1), 197-207.

    50. Brown, A.E., Burn, A.J., Hopkins, J.J. & Way, S.F., 1997. The habitats directive: selection of Special Areas of Conservation in the UK. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 270.

    51. Bryan G.W. & Hummerstone, L.G., 1971. Adaptation of the polychaete Nereis diversicolor to estuarine sediments containing high concentrations of heavy metals. I. General observations and adaption to copper. Journal of the Marine Biological Association of the United Kingdom, 51, 845-863

    52. Bryan, G.W. & Gibbs, P.E., 1983. Heavy metals from the Fal estuary, Cornwall: a study of long-term contamination by mining waste and its effects on estuarine organisms. Plymouth: Marine Biological Association of the United Kingdom. [Occasional Publication, no. 2.]

    53. Bryan, G.W. & Gibbs, P.E., 1991. Impact of low concentrations of tributyltin (TBT) on marine organisms: a review. In: Metal ecotoxicology: concepts and applications (ed. M.C. Newman & A.W. McIntosh), pp. 323-361. Boston: Lewis Publishers Inc.

    54. Bryan G.W. & Hummerstone, L.G., 1971. Adaptation of the polychaete Nereis diversicolor to estuarine sediments containing high concentrations of heavy metals. I. General observations and adaption to copper. Journal of the Marine Biological Association of the United Kingdom, 51, 845-863.

    55. Bryan, G.W., 1984. Pollution due to heavy metals and their compounds. In Marine Ecology: A Comprehensive, Integrated Treatise on Life in the Oceans and Coastal Waters, vol. 5. Ocean Management, part 3, (ed. O. Kinne), pp.1289-1431. New York: John Wiley & Sons.

    56. Bryant, V., Newbery, D.M., McLusky, D.S. & Campbell, R., 1985. Effect of temperature and salinity on the toxicity of arsenic to three estuarine invertebrates (Corophium volutator, Macoma balthica, Tubifex costatus). Marine Ecology Progress Series, 24, 129-137.

    57. Bryant, V., Newbery, D.M., McLusky, D.S. & Campbell, R., 1985a. Effect of temperature and salinity on the toxicity of nickel and zinc to two estuarine invertebrates (Corophium volutator, Macoma balthica). Marine Ecology Progress Series, 24, 139-153.

    58. Buffet, P.-E., Poirier, L., Zalouk-Vergnoux, A., Lopes, C., Amiard, J.-C., Gaudin, P., Risso-de Faverney, C., Guibbolini, M., Gilliland, D., Perrein-Ettajani, H., Valsami-Jones, E. & Mouneyrac, C., 2014. Biochemical and behavioural responses of the marine polychaete Hediste diversicolor to cadmium sulfide quantum dots (CdS QDs): Waterborne and dietary exposure. Chemosphere, 100, 63-70.

    59. Buffet, P.-E., Tankoua, O.F., Pan, J.-F., Berhanu, D., Herrenknecht, C., Poirier, L., Amiard-Triquet, C., Amiard, J.-C., Berard, J.-B., Risso, C., Guibbolini, M., Romeo, M., Reip, P., Valsami-Jones, E. & Mouneyrac, C., 2011. Behavioural and biochemical responses of two marine invertebrates Scrobicularia plana and Hediste diversicolor to copper oxide nanoparticles. Chemosphere, 84 (1), 166-174.

    60. Caddy, J.F., 1967. Maturation of gametes and spawning in Macoma balthica (L.). Canadian Journal of Zoology, 45, 955-965.

    61. Caldow, R., McGrorty, S.,West, A., Durell, S. E. A. le V. dit, Stillman, R.,Anderson, S. 2005. Macro-invertebrate fauna in the intertidal mudflats. In The ecology of Poole Harbour (eds J. Humphreys & V. May), pp. 91–108. Amsterdam, The Netherlands: Elsevier B.V.

    62. Caldow, R.W.G., Stillman, R.A., le V. dit Durell, S.E.A., West, A.D., McGrorty, S., Goss-Custard, J.D., Wood, P.J. & Humphreys, J., 2007. Benefits to shorebirds from invasion of a non-native shellfish. Proceedings of the Royal Society, B. 274, 1449 – 1455.

    63. Chandrasekara, W.U. & Frid, C.L.J., 1998. A laboratory assessment of the survival and vertical movement of two epibenthic gastropod species, Hydrobia ulvae, (Pennant) and Littorina littorea (Linnaeus), after burial in sediment. Journal of Experimental Marine Biology and Ecology, 221, 191-207.

    64. Chu, J.W. & Levin, L.A., 1989. Chu, J.W. & Levin, L.A., 1989. Photoperiod and temperature regulation of growth and reproduction in Streblospio benedicti (Polychaeta: Spionidae). Invertebrate Reproduction & Development, 15 (2), 131-142.

    65. Clay, E., 1967b. Literature survey of the common fauna of estuaries, 10. Macoma balthica and Tellina tenuis. Imperial Chemical Industries Limited, Brixham Laboratory, BL/A/705.

    66. Clay, E., 1967c. Literature survey of the common fauna of estuaries, 1. Cirratulus cirratusO.F. Müller. Imperial Chemical Industries Limited, Brixham Laboratory, PVM45/A/374.

    67. Coates, D.A., Deschutter, Y., Vincx, M. & Vanaverbeke, J., 2014. Enrichment and shifts in macrobenthic assemblages in an offshore wind farm area in the Belgian part of the North Sea. Marine Environmental Research, 95, 1-12.

    68. Coates, D.A., van Hoey, G., Colson, L., Vincx, M. & Vanaverbeke, J., 2015. Rapid macrobenthic recovery after dredging activities in an offshore wind farm in the Belgian part of the North Sea. Hydrobiologia, 756 (1), 3-18.

    69. Collie, J.S., Hall, S.J., Kaiser, M.J. & Poiner, I.R., 2000. A quantitative analysis of fishing impacts on shelf-sea benthos. Journal of Animal Ecology, 69 (5), 785–798.

    70. Collier, L.M. & Pinn, E.H., 1998. An assessment of the acute impact of the sea lice treatment Ivermectin on a benthic community. Journal of Experimental Marine Biology and Ecology, 230, 131-147.

    71. Conde, A., Novais, J.M. & Domínguez, J., 2011. A field experiment on the reproductive success of the invasive clam Mya arenaria (Bivalvia) in the Tagus estuary: coexistence with the native clam Scrobicularia plana. Scientia Marina, 75, 301-308.

    72. Conde, A., Novais, J.M. & Domínguez, J., 2013. Characterization of an estuarine environment by means of an index based on intertidal macrofauna. Marine Pollution Bulletin, 71 (1–2), 129-138.

    73. Connor, D.W., Allen, J.H., Golding, N., Howell, K.L., Lieberknecht, L.M., Northen, K.O. & Reker, J.B., 2004. The Marine Habitat Classification for Britain and Ireland. Version 04.05. Joint Nature Conservation Committee, Peterborough. www.jncc.gov.uk/MarineHabitatClassification.

    74. Connor, D.W., Brazier, D.P., Hill, T.O., & Northen, K.O., 1997b. Marine biotope classification for Britain and Ireland. Vol. 1. Littoral biotopes. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 229, Version 97.06., Joint Nature Conservation Committee, Peterborough, JNCC Report No. 230, Version 97.06.

    75. Cook, W., 1990. Studies on the effects of hydraulic dredging on cockle and other macroinvertebrate populations 1989-1990. North Western and North Wales Sea Fisheries Committee.

    76. Cotter, A.J.R., Walker, P., Coates, P., Cook, W. & Dare, P.J., 1997. Trial of a tractor dredger for cockles in Burry Inlet, South Wales. ICES Journal of Marine Science, 54, 72-83.

    77. Craig, N.C.D. & Caunter, J.E., 1990. The effects of polydimethylsiloxane (PDMS) in sediment on the polychaete worm Nereis diversicolor. Chemosphere, 21, 751-759.

    78. Crisp, D.J. (ed.), 1964. The effects of the severe winter of 1962-63 on marine life in Britain. Journal of Animal Ecology, 33, 165-210.

    79. Crompton T.R., 1997. Toxicants in the aqueous ecosystem. New York: John Wiley & Sons.

    80. Cunha, T., Hall, A. & Queiroga, H., 2005. Estimation of the Diopatra neapolitana annual harvest resulting from digging activity in Canal de Mira, Ria de Aveiro. Fisheries Research, 76 (1), 56-66.

    81. Da Fonseca-Genevois, V. & Cazaux, C., 1987. Streblospio benedicti Webster, 1879 (Annelida, Polychaeta) in the Loire Estuary - biology and ecology. Cahiers de Biologie Marine, 28 (2), 231-261.

    82. Dales, R. P., 1950. The reproduction and larval development of Nereis diversicolor O. F. Müller. Journal of the Marine Biological Association of the United Kingdom, 29, 321-360.

    83. Dame, R.F.D., 1996. Ecology of Marine Bivalves: an Ecosystem Approach. New York: CRC Press Inc. [Marine Science Series.]

    84. Davey, J.T. & George, C.L., 1986. Specific interactions in soft sediments: factors in the distribution of Nereis (Hediste) diversicolor in the Tamar Estuary. Ophelia, 26, 151-164.

    85. Davies, I.M, Gillibrand, P.A., McHenery, J.G. & Rae, G.H., 1998. Environmental risk of Ivermectin to sediment dwelling organisms. Aquaculture, 163, 29-46.

    86. de Wilde, P.A.W., 1975. Influence of temperature on behaviour, energy metabolism and growth of Macoma balthica (L.). In Ninth European Marine Biology Symposium (ed. H. Barnes), pp.239-256. Aberdeen University Press.

    87. De Wilde, P.A.W., 1975. Influence of temperature on behaviour, energy metabolism and growth of Macoma balthica (L.). In Barnes, e.H. In Ninth European Marine Biology Symposium Aberdeen University Press, pp. 239-256.

    88. Dedeh, A., Ciutat, A., Tran, D. & Bourdineaud, J.-P., 2014. DNA Alterations Triggered by Environmentally Relevant Polymetallic Concentrations in Marine Clams Ruditapes philippinarum and Polychaete Worms Hediste diversicolor. Archives of Environmental Contamination and Toxicology, 67, (4), 651-658.

    89. Degraer, S., Wittoeck, J., Appeltans, W., Cooreman, K., Deprez, T., Hillewaert, H., Hostens, K., Mees, J., Berge, V. & Vincx, M., 2006. The macrobenthos atlas of the Belgian part of the North Sea. Belgian Science Policy.

    90. Deslous-Paoli, J.-M., Lannou, A.-M., Geairon, P., Bougrier, S., Raillard, O. & Héral, M., 1992. Effects of the feeding behavior of Crassostrea gigas (Bivalve Molluscs) on biosedimentation of natural particulate matter. Hydrobiologia, 231 (2), 85-91.

    91. Desprez, M.H., Rybarczyk, H., Wilson, J.G., Ducrotoy, J.P., Sueur, F., Olivesi, R. & Elkaim, B., 1992. Biological impact of eutrophication in the Bay of Somme and the induction and impact of anoxia. Netherlands Journal of Sea Research, 30, 149-159.

    92. Desrina, Verreth, J.A.J., Vlak, J.M. & Verdegem, M.C.J., 2014. Hediste diversicolor (O.F. Mueller 1776) as a Possible Model to Study White Spot Syndrome Virus Infection in Polychaetes. Asian Fisheries Science, 27 (3), 199-211.

    93. Diaz, R.J. & Rosenberg, R., 1995. Marine benthic hypoxia: a review of its ecological effects and the behavioural responses of benthic macrofauna. Oceanography and Marine Biology: an Annual Review, 33, 245-303.

    94. Dittmann, S., Günther, C-P. & Schleier, U., 1999. Recolonization of tidal flats after disturbance. In The Wadden Sea ecosystem: stability, properties and mechanisms (ed. S. Dittmann), pp.175-192. Berlin: Springer-Verlag.

    95. Dolbeth, M., Cardoso, P.G., Grilo, T.F., Bordalo, M.D., Raffaelli, D. & Pardal, M.A., 2011. Long-term changes in the production by estuarine macrobenthos affected by multiple stressors. Estuarine Coastal and Shelf Science, 92 (1), 10-18.

    96. Dow, R.C., 1978. Size-selective mortalities of clams in an oil spill site. Marine Pollution Bulletin, 9, 45-48.

    97. Dow, R.L. & Wallace, D.E., 1961. The soft-shell clam industry of Maine. U.S. Fish and Wildlife Service, Department of the Interior, Circular no. 110., U.S.A: Washington D.C.

    98. Dries, R.R. & Theede, H., 1974. Sauerstoffmangelresistenz mariner Bodenvertebraten aus der West-lichen Ostsee. Marine Biology, 25, 327-233.

    99. Duinker, J.C., Hillebrand, M.T.J. & Boon, J.P., 1983. Organochlorines in benthic invertebrates and sediments from the Dutch Wadden Sea; identification of individual PCB components. Netherlands Journal of Sea Research, 17, 19-38.

    100. Elliot, M., Nedwell, S., Jones, N.V., Read, S.J., Cutts, N.D. & Hemingway, K.L., 1998. Intertidal sand and mudflats & subtidal mobile sandbanks (Vol. II). An overview of dynamic and sensitivity for conservation management of marine SACs. Prepared by the Scottish Association for Marine Science for the UK Marine SACs Project.

    101. Emerson, C.W. & Grant, J., 1991. The control of soft-shell clam (Mya arenaria) recruitment on intertidal sandflats by bedload sediment transport. Limnology and Oceanography, 36, 1288-1300.

    102. Essink, K., 1999. Ecological effects of dumping of dredged sediments; options for management. Journal of Coastal Conservation, 5, 69-80.

    103. Farke, H., 1979. Population dynamics, reproduction and early development of Tharyx marioni (Polychaeta, Cirratulidae) on tidal flats of the German Bight. Veroffentlichungen des Instituts fur Meeresforschung in Bremerhaven, 18, 69-99.

    104. Fernandez T.V. & Jones, N.V., 1990. The influence of salinity and temperature on the toxicity of zinc to Nereis diversicolor. Tropical Ecology, 31, 40-46

    105. Ferns, P.N., Rostron, D.M. & Siman, H.Y., 2000. Effects of mechanical cockle harvesting on intertidal communities. Journal of Applied Ecology, 37, 464-474.

    106. Fish, J.D. & Fish, S., 1996. A student's guide to the seashore. Cambridge: Cambridge University Press.

    107. Flach, E.C., 1992. Disturbance of benthic infauna by sediment-reworking activities of the lugworm Arenicola marina. Netherlands Journal of Sea Research, 30, 81-89.

    108. Franzen, N.C.M., 1995. Shear wave detection by Macoma balthica.

    109. Fujii, T. & Raffaelli, D., 2008. Sea-level rise, expected environmental changes, and responses of intertidal benthic macrofauna in the Humber estuary, UK. Marine Ecology Progress Series, 371, 23-35.

    110. Gamenick, I., Jahn, A., Vopel, K. & Giere, O., 1996. Hypoxia and sulphide as structuring factors in a macrozoobenthic community on the Baltic Sea shore: Colonization studies and tolerance experiments. Marine Ecology Progress Series, 144, 73-85.

    111. García - Arberas, L. & Rallo, A., 2002. Life Cycle, Demography and Secondary Production of the Polychaete Hediste ­diversicolor in a Non-Polluted Estuary in the Bay of Biscay. Marine Ecology, 23 (3), 237-251.

    112. Germain, P., Miramand, P. & Masson, M., 1984. Experimental study of long-lived radionuclide transfers (americium, plutonium, technetium) between labelled sediments and annelidae (Nereis diversicolor, Arenicola marina). In International symposium on the behaviour of long-lived radionuclides in the marine environment, (ed. A.Cigna & C. Myttenaere), pp. 327-341. Luxembourg: Office for Official Publications of the European Communities.

    113. Gibbs, P.E., 1971. Reproductive cycles in four polychaete species belonging to the family Cirratulidae. Journal of the Marine Biological Association of the United Kingdom, 51, 745-769.

    114. Gibson, G.D. & Harvey, J., 2000. Morphogenesis during asexual reproduction in Pygospio elegans Claparede (Annelida, Polychaeta). The Biological Bulletin, 199 (1), 41-49.

    115. Giere, O. & Pfannkuche, O., 1982. Biology and ecology of marine Oligochaeta, a review. Oceanography and Marine Biology, 20, 173-309.

    116. Giere, O. & Rhode, B., 1987. Anatomy and ultrastructure of the marine oligochaete Tubificoides benedii (Tubificidae), with emphasis on its epidermis-cuticle-complex. In Brinkhurst, R.O. and Diaz, R.J. (eds.). Aquatic Oligochaeta: Proceedings of the Third International Symposium on Aquatic Oligochaeta held in Hamburg, Germany September 29-October 4, 1985, Dordrecht: Springer Netherlands, pp. 159-159.

    117. Gilbert, M.A., 1978. Aspects of the reproductive cycle in Macoma balthica (Bivalvia). The Nautilus, 29, 21-24.

    118. Gill, A.B., Huang, Y., Gloyne-Philips, I., Metcalfe, J., Quayle, V., Spencer, J. & Wearmouth, V. (2009). COWRIE 2.0 Electromagnetic Fields (EMF) Phase 2: EMF-sensitive fish response to EM emissions from sub-sea electricity cables of the type used by the offshore renewable energy industry. Commissioned by COWRIE Ltd (project reference COWRIE-EMF-1-06)

    119. Gill, A.B., Huang, Y., Gloyne-Philips, I., Metcalfe, J., Quayle, V., Spencer, J. & Wearmouth, V., 2009. COWRIE 2.0 Electromagnetic Fields (EMF) Phase 2: EMF-sensitive fish response to EM emissions from sub-sea electricity cables of the type used by the offshore renewable energy industry. Commissioned by COWRIE Ltd (project reference COWRIE-EMF-1-06), 68.

    120. Gogina, M., Glockzin, M. & Zettler, M.L., 2010a. Distribution of benthic macrofaunal communities in the western Baltic Sea with regard to near-bottom environmental parameters. 1. Causal analysis. Journal of Marine Systems, 79 (1), 112-123.

    121. Gogina, M., Glockzin. M. & Zettler, M.L., 2010. Distribution of benthic macrofaunal communities in the western Baltic Sea with regard to near-bottom environmental parameters. 2. Modelling and prediction. Journal of Marine Systems, 80, 57-70. 

    122. Goss-Custard, J.D., Jones, R.E. & Newberry, P.E., 1989. The ecology of the Wash. 1. Distribution and diet of wading birds (Charadrii). Journal of Applied Ecology, 14, 681-700.

    123. Graf, G., 1989. Benthic-pelagic coupling in a deep-sea benthic community. Nature, 341 (6241), 437-439.

    124. Grant, J. & Thorpe, B., 1991. Effects of suspended sediment on growth, respiration, and excretion of the soft shelled clam (Mya arenaria). Canadian Journal of Fisheries and Aquatic Sciences, 48, 1285-1292.

    125. Gray, J.S., 1971. The effects of pollution on sand meiofauna communities. Thalassia Jugoslovica, 7, 76-86.

    126. Gray, J.S., 1981. The ecology of marine sediments. An introduction to the structure and function of benthic communities. Cambridge: Cambridge University Press.

    127. Gray, J.S., Clarke, K.R., Warwick, R.M. & Hobbs, G., 1990. Detection of initial effects of pollution on marine benthos - an example from the Ekofisk and Eldfisk oilfields, North Sea. Marine Ecology Progress Series, 66 (3), 285-299.

    128. Green, J., 1968. The biology of estuarine animals. Sidgwick and Jackson, London.

    129. Grilo, T.F., Cardoso, P.G., Dolbeth, M., Bordalo, M.D. & Pardal, M.A., 2011. Effects of extreme climate events on the macrobenthic communities' structure and functioning of a temperate estuary. Marine Pollution Bulletin, 62 (2), 303-311.

    130. Guenther, C.P., 1991. Settlement of Macoma balthica on an intertidal sandflat in the Wadden Sea. Marine Ecology Progress Series, 76, 73-79.

    131. Hailey, N., 1995. Likely impacts of oil and gas activities on the marine environment and integration of environmental considerations in licensing policy. English Nature Research Report, no 145., Peterborough: English Nature.

    132. Hall, S.J. & Harding, M.J.C., 1997. Physical disturbance and marine benthic communities: the effects of mechanical harvesting of cockles on non-target benthic infauna. Journal of Applied Ecology, 34, 497-517.

    133. Hall, S.J., 1994. Physical disturbance and marine benthic communities: life in unconsolidated sediments. Oceanography and Marine Biology: an Annual Review, 32, 179-239.

    134. Harvey, M. & Vincent, B., 1989. Spatial and temporal variations of the reproduction cycle and energy allocation of the bivalve Macoma balthica (L.) on a tidal flat. Journal of Experimental Marine Biology and Ecology, 129, 199-217.

    135. Hayward, P.J. 1994. Animals of sandy shores. Slough, England: The Richmond Publishing Co. Ltd. [Naturalists' Handbook 21.]

    136. Hiddink, J.G., 2003. Effects of suction-dredging for cockles on non-target fauna in the Wadden Sea. Journal of Sea Research, 50, 315-323.

    137. Hiscock, K., 1983. Water movement. In Sublittoral ecology. The ecology of shallow sublittoral benthos (ed. R. Earll & D.G. Erwin), pp. 58-96. Oxford: Clarendon Press.

    138. Hiscock, K., Tyler-Walters, H. & Jones, H., 2002. High level environmental screening study for offshore wind farm developments - marine habitats and species project. Marine Biological Association of the United Kingdom, Plymouth, AEA Technology, Environment Contract: W/35/00632/00/00, pp.

    139. Hunter, J., & Arthur, D.R., 1978. Some aspects of the ecology of Peloscolex benedeni Udekem (Oligochaeta: Tubificidae) in the Thames estuary. Estuarine and Coastal Marine Science, 6, 197-208.

    140. Hutchins, D.A., Stupakoff, I., Hook, S., Luoma, S.N. & Fisher, N.S., 1998. Effects of Arctic temperatures on distribution and retention of the nuclear waste radionuclides 241Am, 57Co and 137Cs in the bioindicator bivalve Macoma balthica. Marine Environmental Research, 45, 17-28.

    141. Hylleberg, J., 1975. The effect of salinity and temperature on egestion in mud snails (Gastropoda: Hydrobiidae). Oecologia, 21, 279-289.

    142. Jackson, M.J. & James, R., 1979. The influence of bait digging on cockle Cerastoderma edule, populations in north Norfolk. Journal of Applied Ecology, 16, 671-679.

    143. Jansen, J.M., Pronker, A.E., Bonga, S.W. & Hummel, H., 2007. Macoma balthica in Spain, a few decades back in climate history. Journal of Experimental Marine Biology and Ecology, 344 (2), 161-169.

    144. Jansson, A., Norkko, J., Dupont, S. & Norkko, A., 2015. Growth and survival in a changing environment: Combined effects of moderate hypoxia and low pH on juvenile bivalve Macoma balthica. Journal of Sea Research, 102, 41-47.

    145. Jensen, A.C., Humphreys, J., Caldow, R.W.G., Grisley, C. and Dyrynda, P.E.J., 2004. Naturalization of the Manila clam (Tapes philippinarum), an alien species, and establishment of a clam fishery within Poole Harbour, Dorset. Journal of the Marine Biological Association of the UK, 84,1069-1073.

    146. Jensen, K.T., 1992. Macrozoobenthos on an intertidal mudflat in the Danish Wadden Sea: Comparisons of surveys made in the 1930s, 1940s and 1980s. Helgolander Meeresuntersuchungen. Hamburg, 46 (4), 363-376.

    147. Jensen, K.T. & Mouritsen K.N., 1992. Mass mortality in two common soft bottom invertebrates, Hydrobia ulvae and Corophium volutator, the possible role of trematodes. Helgolander Meeresuntersuchungen, 46, 329-339.

    148. JNCC, 2015. The Marine Habitat Classification for Britain and Ireland Version 15.03. JNCC: JNCC. 2015(20/05/2015). jncc.defra.gov.uk/MarineHabitatClassification

    149. JNCC (Joint Nature Conservation Committee), 1999. Marine Environment Resource Mapping And Information Database (MERMAID): Marine Nature Conservation Review Survey Database. [on-line] http://www.jncc.gov.uk/mermaid

    150. Johnston, R., 1984. Oil Pollution and its management. In Marine Ecology: A Comprehensive, Integrated Treatise on Life in the Oceans and Coastal Waters vol. 5. Ocean Management, part 3 (ed. O. Kinne), pp.1433-1582. New York: John Wiley & Sons Ltd.

    151. Jones, L.A., Hiscock, K. & Connor, D.W., 2000. Marine habitat reviews. A summary of ecological requirements and sensitivity characteristics for the conservation and management of marine SACs. Joint Nature Conservation Committee, Peterborough. (UK Marine SACs Project report.). Available from: http://www.ukmarinesac.org.uk/pdfs/marine-habitats-review.pdf

    152. Kaiser, M., Broad, G. & Hall, S., 2001. Disturbance of intertidal soft-sediment benthic communities by cockle hand raking. Journal of Sea Research, 45 (2), 119-130.

    153. Kesaniemi, J.E., Geuverink, E. & Knott, K.E., 2012. Polymorphism in developmental mode and its effect on population genetic structure of a Spionid Polychaete, Pygospio elegans. Integrative and Comparative Biology, 52 (1), 181-196.

    154. Landner, I., Nilsson, K. & Rossenburg, R., 1977. Assessment of industrial pollution by means of benthic macrofauna surveys along the Swedish Baltic coast. Vatten, 33, 324-379.

    155. Langston, W.J., Chesman, B.S., Burt, G.R., Hawkins, S.J., Readman, J. & Worsfold, P., 2003. Characterisation of European Marine Sites. Poole Harbour Special Protection Area. Occasional Publication. Marine Biological Association of the United Kingdom, 12, 111.

    156. Lardicci, C., Ceccherelli, G. & Rossi, F., 1997. Streblospio shrubsolii (Polychaeta : Spionidae): temporal fluctuations in size and reproductive activity. Cahiers de Biologie Marine, 38 (3), 207-214.

    157. Latham, H., Sheehan, E., Foggo, A., Attrill, M., Hoskin, P. & Knowles, H., 2012. Fal and Helford Recreational Boating Study Chapter 1. Single block, sub‐tidal, permanent moorings: Ecological impact on infaunal communities due to direct, physical disturbance from mooring infrastructure. Falmouth Harbour Commissioners, UK.

    158. Leewis, L., Van Bodegom, P.M., Rozema, J. & Janssen, G.M., 2012. Does beach nourishment have long-term effects on intertidal macroinvertebrate species abundance? Estuarine, Coastal and Shelf Science, 113, 172-181.

    159. Levell, D., 1976. The effect of Kuwait Crude Oil and the Dispersant BP 1100X on the lugworm, Arenicola marina L. In Proceedings of an Institute of Petroleum / Field Studies Council meeting, Aviemore, Scotland, 21-23 April 1975. Marine Ecology and Oil Pollution (ed. J.M. Baker), pp. 131-185. Barking, England: Applied Science Publishers Ltd.

    160. Levin, L. & Creed, E., 1986. Effect of temperature and food availability on reproductive responses of Streblospio benedicti (Polychaeta: Spionidae) with planktotrophic or lecithotrophic development. Marine Biology, 92 (1), 103-113.

    161. Levin, L., Blair, N., DeMaster, D., Plaia, G., Fornes, W., Martin, C. & Thomas, C., 1997. Rapid subduction of organic matter by maldanid polychaetes on the North Carolina slope. Journal of Marine Research, 55 (3), 595-611.

    162. Lewis, L., Davenport, J. & Kelly, T., 2002b. A study of the impact of a pipeline construction on estuarine benthic invertebrate communities. Estuarine, Coastal and Shelf Science, 55 (2), 213-221.

    163. Lillicrap, A., Schaanning, M. & Macken, A., 2015. Assessment of the Direct Effects of Biogenic and Petrogenic Activated Carbon on Benthic Organisms. Environmental Science & Technology, 49 (6), 3705-3710.

    164. Lindsay, S.M., Jackson, J.L. & He, S.Q., 2007. Anterior regeneration in the spionid polychaetes Dipolydora quadrilobata and Pygospio elegans. Marine Biology, 150 (6), 1161-1172.

    165. Linke, O., 1939. Die Biota des Jadebusenwatts. Helgolander Wissenschaftliche Meeresuntersuchungen, 1, 201-348.

    166. Long, W.C., Seitz, R.D., Brylawski, B.J. & Lipcius, R.N., 2014. Individual, population, and ecosystem effects of hypoxia on a dominant benthic bivalve in Chesapeake Bay. Ecological Monographs, 84 (2), 303-327.

    167. Longshaw, M. & Malham, S.K., 2013. A review of the infectious agents, parasites, pathogens and commensals of European cockles (Cerastoderma edule and C. glaucum). Journal of the Marine Biological Association of the United Kingdom, 93 (01), 227-247.

    168. Luoma, S.N., Cain, D.J., Ho, K. & Hutchinson, A., 1983. Variable tolerance to copper in two species from San Francisco Bay. Marine Environmental Research, 10, 209-222.

    169. Maddock, A., 2008. UK Biodiversity Action Plan; Priority Habitat Descriptions. UK Biodiversity Action Plan, 94pp

    170. Madsen, P.B. & Jensen, K., 1987. Population dynamics of Macoma balthica in the Danish Wadden Sea in an organically enriched area. Ophelia, 27, 197-208.

    171. Maurer, D., Keck, R.T., Tinsman, J.C., Leatham, W.A., Wethe, C., Lord, C. & Church, T.M., 1986. Vertical migration and mortality of marine benthos in dredged material: a synthesis. Internationale Revue der Gesamten Hydrobiologie, 71, 49-63.

    172. McGreer, E.R., 1979. Sublethal effects of heavy metal contaminated sediments on the bivalve Macoma balthica(L.). Marine Pollution Bulletin, 10, 259-262.

    173. McLusky D.S., Bryant, V. & Campbell, R., 1986. The effects of temperature and salinity on the toxicity of heavy metals to marine and estuarine invertebrates. Oceanography and Marine Biology: an Annual Review, 24, 481-520

    174. McLusky, D., 1982. The impact of petrochemical effluent on the fauna of an intertidal estuarine mudflat. Estuarine, Coastal and Shelf Science, 14 (5), 489-499.

    175. McLusky, D.S.& Allan, D.G., 1976. Aspects of the biology of Macoma balthica (L.) from the estuarine Firth of Forth. Journal of Molluscan Studies, 42, 31-45.

    176. McLusky, D.S. & Elliott, M., 1981. The feeding and survival strategies of estuarine molluscs. In Feeding and survival strategies of estuarine organisms (ed. N.V. Jones & W.J. Wolff). Plenum Press.

    177. McLusky, D.S., 1982. The impact of petrochemical effluent on the fauna of an intertidal estuarine mudflat. Estuarine, Coastal and Shelf Science, 14, 489-499.

    178. McLusky, D.S., 1989. The Estuarine Ecosystem, 2nd ed. New York: Chapman & Hall.

    179. McLusky, D.S., Anderson, F.E. & Wolfe-Murphy, S., 1983. Distribution and population recovery of Arenicola marina and other benthic fauna after bait digging. Marine Ecology Progress Series, 11, 173-179.

    180. McLusky, D.S., Anderson, F.E. & Wolfemurphy, S., 1983. Distribution and population recovery of Arenicola marina and other benthic fauna after bait digging. Marine Ecology Progress Series, 11 (2), 173-179.

    181. McLusky, D.S., Bryant, V. & Campbell, R., 1986. The effects of temperature and salinity on the toxicity of heavy metals to marine and estuarine invertebrates. Oceanography and Marine Biology: an Annual Review, 24, 481-520.

    182. McLusky, D.S., Teare, M. & Phizachlea, P., 1980. Effects of domestic and industrial pollution on distribution and abundance of aquatic oligochaetes in the Forth estuary. Helgolander Wissenschaftliche Meeresuntersuchungen, 33, 384-392.

    183. Meire, P.M., 1993. The impact of bird predation on marine and estuarine bivalve populations: a selective review of patterns and underlying causes. In Bivalve filter feeders in estuarine and coastal ecosystem processes (ed. R.F. Dame). NATO ASI Series, Springer Verlag.

    184. MES, 2010. Marine Macrofauna Genus Trait Handbook. Marine Ecological Surveys Limited. http://www.genustraithandbook.org.uk/

    185. Mettam, C., Santhanam, V. & Havard, M.C.S., 1982. The oogenic cycle of Nereis diversicolor under natural conditions. Journal of the Marine Biological Association of the United Kingdom, 62, 637-645.

    186. Mitchell, I.M., 2006. In situ biodeposition rates of Pacific oysters (Crassostrea gigas) on a marine farm in Southern Tasmania (Australia). Aquaculture, 257 (1), 194-203.

    187. Mölsa, H., Hakkila, S. & Puhakka, M., 1986. Reproductive success of Macoma balthica in relation to environmental stability. Ophelia, Supplement 4, 167-178.

    188. Moore, J., 1991. Studies on the Impact of Hydraulic Cockle Dredging on Intertidal Sediment Flat Communities. A report to the Nature Conservancy Council from the Field Studies Council Research Centre, Pembroke, Wales, FSC/RC/4/91.

    189. Mouneyrac, C., Mastain, O., Amiard, J.C., Amiard-Triquet, C., Beaunier, P., Jeantet, A.-Y., Smith, B.D. & Rainbow, P.S., 2003. Trace-metal detoxification and tolerance of the estuarine worm Hediste diversicolor chronically exposed in their environment. Marine Biology, 143 (4), 731-744.

    190. Navarro, J.M. & Widdows, J., 1997. Feeding physiology of Cerastoderma edule in response to a wide range of seston concentrations. Marine Ecology Progress Series, 152, 175-186.

    191. Neves de Carvalho, A., Vaz, A.S.L., Sérgio, T.I.B. & Santos, P.J.T.d., 2013. Sustainability of bait fishing harvesting in estuarine ecosystems: Case study in the Local Natural Reserve of Douro Estuary, Portugal estuarinos: Caso de estudo na Reserva Natural Local do Estuário do Douro, Portugal. Revista de Gestão Costeira Integrada, 13 (2), 157-168.

    192. Newell, C.R. & Hidu, H., 1986. Species profiles: life histories and environmental requirements of coastal fishes and invertebrates (North Atlantic) . Softshell clam. http://www.nwrc.usgs.gov/wdb/pub/0168.pdf, 2000-10-02

    193. Newell, R., Seiderer, L. & Hitchcock, D., 1998. The impact of dredging works in coastal waters: a review of the sensitivity to disturbance and subsequent recovery of biological resources on the sea bed. Oceanography and Marine Biology: An Annual Review, 36, 127-178.

    194. Newell, R.C., Seiderer, L.J. & Hitchcock, D.R., 1998. The impact of dredging works in coastal waters: a review of the sensitivity to disturbance and subsequent biological recovery of biological resources on the sea bed. Oceanography and Marine Biology: an Annual Review, 36, 127-178.

    195. Nielsen, A.M., Eriksen, N.T., Iversen, J.J.L. & Riisgård, H.U., 1995. Feeding, growth and respiration in the polychaetes Nereis diversicolor (facultative filter-feeder) and Nereis virens (omnivorous) - a comparative study. Marine Ecology Progress Series, 125, 149-158.

    196. Norkko, A. & Bonsdorff, E., 1996. Population responses of coastal zoobenthos to stress induced by drifting algal mats. Marine Ecology Progress Series, 140, 141-151.

    197. Nugues, M., Kaiser, M., Spencer, B. & Edwards, D., 1996. Benthic community changes associated with intertidal oyster cultivation. Aquaculture Research, 27 (12), 913-924.

    198. Oertzen, J.A. Von., 1969. Erste Ergebrisse zur experimentellen ökologie von postglazialen Relikten (Bivalvia) der Ostsee. Limnologica (Berlin), 7, 129-137.

    199. Olafsson, E.B. & Persson, L.E., 1986. The interaction between Nereis diversicolor (Muller) and Corophium volutator (Pallas) as a structuring force in a shallow brackish sediment. Journal of Experimental Marine Biology and Ecology, 103, 103-117.

    200. Olafsson, E.B., 1986. Density dependence in suspension feeding populations of the bivalve Macoma balthica. A field experiment. Journal of Animal Ecology, 55, 517-526.

    201. Olive, P.J.W. & Garwood, P.R., 1981. Gametogenic cycle and population structures of Nereis (Hediste) diversicolor and Nereis (Nereis) pelagica from North-East England. Journal of the Marine Biological Association of the United Kingdom, 61, 193-213.

    202. Oost, A., Kabat, P., Wiersma, A. & Hofstede, J. (eds.), 2009. Climate. Thematic report no. 4.1. Wilhelmshaven, Germany: Common Wadden Sea Secretariat, Trilateral Monitoring and Assessment Group, pp. 16.

    203. OSPAR, 2009b. Background document for Intertidal mudflats. OSPAR Commission, Biodiversity Series, OSPAR Commission, London, 29 pp. http://www.ospar.org/documents?v=7186

    204. Ozoh, P.T.E. & Jones, N.N., 1990. Capacity adaptation of Hediste (Nereis) diversicolor embryogenesis to salinity, temperature and copper. Marine Environmental Research, 29, 227-243.

    205. Pearson, T.H. & Rosenberg, R., 1978. Macrobenthic succession in relation to organic enrichment and pollution of the marine environment. Oceanography and Marine Biology: an Annual Review, 16, 229-311.

    206. Pekkarinen, M., 1983. Seasonal changes in condition and biochemical constituents in the soft parts of Macoma balthica (Lamellibranchiata) in the Trarminne brackish water area (Baltic Sea). Annales Zoologici Fennici, 20, 311-322.

    207. Peterson, C.H., Bishop, M.J., Johnson, G.A., D'Anna, L.M. & Manning, L.M., 2006. Exploiting beach filling as an unaffordable experiment: benthic intertidal impacts propagating upwards to shorebirds. Journal of Experimental Marine Biology and Ecology, 338 (2), 205-221.

    208. Peterson, C.H., Hickerson, D.H. & Johnson, G.G., 2000. Short-term consequences of nourishment and bulldozing on the dominant large invertebrates of a sandy beach. Journal of Coastal Research, 368-378.

    209. Philippart, C.J., van Aken, H.M., Beukema, J.J., Bos, O.G., Cadée, G.C. & Dekker, R., 2003. Climate‐related changes in recruitment of the bivalve Macoma balthica. Limnology and Oceanography, 48 (6), 2171-2185.

    210. Pickett, G.D., 1973. The impact of mechanised harvesting on the Thames estuary cockle fishery. MAFF Laboratory Leaflet, no. 29, Fisheries Laboratory, Lowestoft, pp. 9.

    211. Picton, B.E. & Costello, M.J., 1998. BioMar biotope viewer: a guide to marine habitats, fauna and flora of Britain and Ireland. [CD-ROM] Environmental Sciences Unit, Trinity College, Dublin., http://www.itsligo.ie/biomar/

    212. Pihl, L., 1985. Food selection and consumption of mobile epibenthic fauna in shallow marine areas. Marine Ecology Progress Series, 22, 169-179.

    213. Pilkington, M.C., 1971. The veliger stage of Hydrobia ulvae (Pennant). Proceedings of the Malacological Society of London, 39, 281-287.

    214. Porcet-Henner, E. & Dugimont, T., 1992. Adaptability of the coccidian Coelotropha to parasitism. Developmental and Comparative Immunology, 16, 263-274.

    215. Powilleit, M., Graf, G., Kleine, J., Riethmuller, R., Stockmann, K., Wetzel, M.A. & Koop, J.H.E., 2009. Experiments on the survival of six brackish macro-invertebrates from the Baltic Sea after dredged spoil coverage and its implications for the field. Journal of Marine Systems, 75 (3-4), 441-451.

    216. Raffaelli, D., 1999. Nutrient enrichment and trophic organisation in an estuarine food web. Acta Oecologica, 20 (4), 449-461.

    217. Raffaelli, D., 2000. Trends in research on shallow water food webs. Journal of Experimental Marine Biology and Ecology, 250 (1-2), 223-232.

    218. Ragnarsson, S.Á. & Raffaelli, D., 1999. Effects of the mussel Mytilus edulis L. on the invertebrate fauna of sediments. Journal of Experimental Marine Biology and Ecology, 241 (1), 31-43.

    219. Ratcliffe, P.J., Jones, N.V. & Walters, N.J., 1981. The survival of Macoma balthica (L.) in mobile sediments. In Feeding and survival strategies of estuarine organisms (ed. N.V. Jones and W.J. Wolff), pp. 91-108. Plenum Press.

    220. Rees, E.I.S., 1978. Observations on the ecological effects of pipeline construction across the Lafan Sands. Report from University College of North Wales, Marine Science Laboratories, Menai Bridge. Nature Conservancy Council, Peterborough, CSD Report, No. 188 (Benthos Research Report, No. 78-1.)

    221. Reise, K., 1985. Tidal flat ecology. An experimental approach to species interactions. Springer-Verlag, Berlin.

    222. Reise, K., Herre, E., & Sturm, M. 1989. Historical changes in the benthos of the Wadden Sea around the island of Sylt in the North Sea. Helgoländer Meeresuntersuchungen, 43, 417-433.

    223. Rhoads, D. & Young, D., 1971. Animal-sediment relations in Cape Cod Bay, Massachusetts II. Reworking by Molpadia oolitica (Holothuroidea). Marine Biology, 11 (3), 255-261.

    224. Rhoads, D.C. & Young, D.K., 1970. The influence of deposit-feeding organisms on sediment stability and community trophic structure. Journal of Marine Research, 28, 150-178.

    225. Riisgård, H.U., 1994. Filter-feeding in the polychaete Nereis diversicolor: a review. Netherlands Journal of Aquatic Ecology, 28, 453-458.

    226. Rodrigues, A.M., Quintino, V., Sampaio, L., Freitas, R. & Neves, R., 2011. Benthic biodiversity patterns in Ria de Aveiro, Western Portugal: Environmental-biological relationships. Estuarine, Coastal and Shelf Science, 95 (2–3), 338-348.

    227. Rostron, D., 1995. The effects of mechanised cockle harvesting on the invertebrate fauna of Llanrhidian sands. In Burry Inlet and Loughor Estuary Symposium, pp. 111-117.

    228. Ruiz, J., Bryan, G. & Gibbs, P., 1994. Chronic toxicity of water tributyltin(TBT) and copper to spat of the bivalve Scrobicularia plana: Ecological implications. Marine Ecology Progress Series, 113 (1), 105-117.

    229. Rygg, B., 1970. Studies on Cerastoderma edule (L.) and Cerastoderma glaucum (Poiret). Sarsia, 43, 65-80.

    230. Sanders, H.L., 1978. Florida oil spill impact on the Buzzards Bay benthic fauna: West Falmouth. Journal of the Fisheries Board of Canada, 35, 717-730.

    231. Savari, A., Lockwood, A.P.M. & Sheader, M., 1991. Variation in the physiological state of the common cockle (Cerastoderma edule (L.)) in the laboratory and in Southampton Water. Journal of Molluscan Studies, 57, 33-34.

    232. Scaps, P., 2002. A review of the biology, ecology and potential use of the common ragworm Hediste diversicolor (O.F. Müller) (Annelida: Polychaeta). Hydrobiologia, 470, 203-218.

    233. Schueckel, U. & Kroencke, I., 2013. Temporal changes in intertidal macrofauna communities over eight decades: A result of eutrophication and climate change. Estuarine Coastal and Shelf Science, 117, 210-218.

    234. Schueckel, U. & Kroencke, I., 2013. Temporal changes in intertidal macrofauna communities over eight decades: A result of eutrophication and climate change. Estuarine Coastal and Shelf Science, 117, 210-218.

    235. SEEEC (Sea Empress Environmental Evaluation Committee), 1998. The environmental impact of the Sea Empress oil spill. Final Report of the Sea Empress Environmental Evaluation Committee, 135 pp., London: HMSO.

    236. Seitz, R.D., 2011. Gradient effects on structuring of soft-bottom benthic infauna: Macoma balthica and predation, recruitment, and food availability. Journal of Experimental Marine Biology and Ecology, 409 (1-2), 114-122.

    237. Sheehan, E., Coleman, R., Thompson, R. & Attrill, M., 2010. Crab-tiling reduces the diversity of estuarine infauna. Marine Ecology Progress Series, 411, 137-148.

    238. Sheehan, E.V., 2007. Ecological impact of the Carcinus maenas (L.) fishery'crab-tiling'on estuarine fauna. Ph.D. thesis,  University of Plymouth.

    239. Shull, D.H., 1997. Mechanisms of infaunal polychaete dispersal and colonisation in an intertidal sandflat. Journal of Marine Research, 55, 153-179.

    240. Silva, A.C.F., Tavares, P., Shapouri, M., Stigter, T.Y., Monteiro, J.P., Machado, M., Cancela da Fonseca, L. & Ribeiro, L., 2012. Estuarine biodiversity as an indicator of groundwater discharge. Estuarine Coastal and Shelf Science, 97, 38-43.

    241. Smaal, A., van Stralen, M. & Craeymeersch, J., 2005. Does the introduction of the Pacific oyster Crassostrea gigas lead to species shifts in the Wadden Sea? The comparative roles of suspension-feeders in ecosystems: Springer, 47, 277-289.

    242. Smaal, A.C., Vonck, A.P.M.A. & Bakker, M., 1997. Seasonal variation in physiological energetics of Mytilus edulis and Cerastoderma edule of different size classes. Journal of the Marine Biological Association of the United Kingdom, 77, 817-838.

    243. Smith, J.E., 1955. Salinity variation in interstitial water of sand at Kames Bay, Millport, with reference to the distribution of Nereis diversicolor Journal of the Marine Biological Association of the United Kingdom, 34, 33-46.

    244. Smith, J.E., 1964. On the early development of Nereis diversicolor in different salinities. Journal of Morphology, 114, 437-464.

    245. Smith, R.I., 1977. Physiological and reproductive adaptations of Nereis diversicolor to life in the Baltic Sea and adjacent waters. In Essays on polychaetous annelids (ed. D.J. Reish & R. Fauchald), pp. 373-390. Los Angeles: University of Southern California.

    246. Sommer, A., Klein, B. & Pörtner, H.O., 1997. Temperature induced anaerobiosis in two population of the polychaete worm Arenicola marina (L.). Journal of Comparative Physiology, series B, 167, 25-35.

    247. Sörlin, T., 1988. Floating behaviour in the tellinid bivalve Macoma balthica (L.). Oecologia, 77, 273-277.

    248. Sornin, J.-M., Feuillet, M., Heral, M. & Deslous-Paoli, J.-M., 1983. Effet des biodépôts de l'huître Crassostrea gigas (Thunberg) sur l'accumulation de matières organiques dans les parcs du bassin de Marennes-Oléron. Journal of Molluscan Studies, 49 (supp12A), 185-197.

    249. Stekoll, M.S., Clement, L.E. & Shaw, D.G., 1980. Sublethal effects of chronic oil exposure on the intertidal clam Macoma balthica. Marine Biology, 57, 51-60.

    250. Strömgren, T., 1979a. The effect of copper on the increase in length of Ascophyllum nodosum. Journal of Experimental Marine Biology and Ecology, 37, 153-159.

    251. Strömgren, T., 1980b. The effect of lead, cadmium and mercury on the increase in length of five intertidal Fucales. Journal of Experimental Marine Biology and Ecology, 43, 107-119.

    252. Suchanek, T.H., 1993. Oil impacts on marine invertebrate populations and communities. American Zoologist, 33, 510-523.

    253. Swennen, C. & Ching, H.L., 1974. Observations on the trematode Parvatrema affininis, causative agent of crawling tracks of Macoma balthica. Netherlands Journal of Sea Research, 8, 108-115.

    254. Tait, R.V. & Dipper, R.A., 1998. Elements of Marine Ecology. Reed Elsevier.

    255. Tallqvist, M., 2001. Burrowing behaviour of the Baltic clam Macoma balthica: effects of sediment type, hypoxia and predator presence. Marine Ecology Progress Series, 212, 183–191.

    256. Tamaki, A., 1987. Comparison of resistivity to transport by wave action in several polychaete species on an intertidal sand flat. Marine Ecology Progress Series, 37, 181-189.

    257. Taylor, A.C., 1976. Burrowing behaviour and anaerobisim in the bivalve Arctica islandica. Journal of the Marine Biological Association of the United Kingdom, 56, 95 - 109.

    258. Tebble, N., 1976. British Bivalve Seashells. A Handbook for Identification, 2nd ed. Edinburgh: British Museum (Natural History), Her Majesty's Stationary Office.

    259. Theede, H., 1973. Comparative studies on the influence of oxygen deficiency and hydrogen sulphide on marine bottom invertebrates. Netherlands Journal of Sea Research, 7, 244-252.

    260. Theede, H., 1984. Physiological approaches to environmental problems of the Baltic. Limnologica (Berlin), 15, 443-458.

    261. Theede, H., Ponat, A., Hiroki, K. & Schlieper, C., 1969. Studies on the resistance of marine bottom invertebrates to oxygen-deficiency and hydrogen sulphide. Marine Biology, 2, 325-337.

    262. Theede, H., Schaudinn, J. & Saffè, F., 1973. Ecophysiological studies on four Nereis species in the Kiel Bay. Oikos Supplementum, 15, 246-252,

    263. Thiel, M., Stearns, L. & Watling, L., 1998. Effects of green algal mats on bivalves in a New England mud flat. Helgoländer Meeresuntersuchungen, 52 (1), 15-28.

    264. Tillin, H.M., Hull, S.C. & Tyler-Walters, H., 2010. Development of a sensitivity matrix (pressures-MCZ/MPA features). Report to the Department of the Environment, Food and Rural Affairs from ABPmer, Southampton and the Marine Life Information Network (MarLIN) Plymouth: Marine Biological Association of the UK., Defra Contract no. MB0102 Task 3A, Report no. 22., London, 145 pp.

    265. Turk, T.R. & Risk, M.J., 1981. Invertebrate populations of Cobequid Bay, Bay of Fundy. Canadian Journal of Fisheries and Aquatic Sciences, 38, 642-648.

    266. UKTAG, 2014. UK Technical Advisory Group on the Water Framework Directive [online]. Available from: http://www.wfduk.org

    267. Van Colen, C., Montserrat, F., Vincx, M., Herman, P.M., Ysebaert, T. & Degraer, S., 2008. Macrobenthic recovery from hypoxia in an estuarine tidal mudflat. Marine Ecology-Progress Series, 372, 31-42.

    268. Van der Meer, J., Beukema, J.J. & Dekker, R., 2003. Large spatial variability in lifetime egg production in an intertidal Baltic tellin (Macoma balthica) population. Helgoland Marine Research, 56 (4), 274-278.

    269. Vandendriessche, S., Derweduwen, J. & Hostens, K., 2015. Equivocal effects of offshore wind farms in Belgium on soft substrate epibenthos and fish assemblages. Hydrobiologia, 756 (1), 19-35.

    270. Verdelhos, T., Cardoso, P., Dolbeth, M. & Pardal, M., 2011. Latitudinal gradients in Scrobicularia plana reproduction patterns, population dynamics, growth, and secondary production. Marine Ecology Progress Series, 442, 271-283.

    271. Villalba, A., Iglesias, D., Ramilo, A., Darriba, S., Parada, J.M., No, E., Abollo, E., Molares, J. & Carballal, M.J., 2014. Cockle Cerastoderma edule fishery collapse in the Ria de Arousa (Galicia, NW Spain) associated with the protistan parasite Marteilia cochillia. Diseases of Aquatic Organisms, 109 (1), 55-80.

    272. Vismann, B., 1990. Sulphide detoxification and tolerance in Nereis (Hediste) diversicolor and Nereis (Neanthes) virens (Annelida: Polychaeta). Marine Ecology Progress Series, 59, 229-238.

    273. Volkenborn, N. & Reise, K., 2006. Lugworm exclusion experiment: Responses by deposit feeding worms to biogenic habitat transformations. Journal of Experimental Marine Biology and Ecology, 330 (1), 169-179.

    274. Whomersley, P., Huxham, M., Bolam, S., Schratzberger, M., Augley, J. & Ridland, D., 2010. Response of intertidal macrofauna to multiple disturbance types and intensities – an experimental approach. Marine Environmental Research, 69 (5), 297-308.

    275. Widdows, J., Brinsley, M.D., Salkeld, P.N. & Elliott, M., 1998. Use of annular flumes to determine the influence of current velocity and bivalves on material flux at the sediment-water interface. Estuaries, 21, 552-559.

    276. Wilson, J.G., 1981. Temperature tolerance of circatidal bivalves in relation to their distribution. Journal of Thermal Biology, 6, 279-286.

    277. Winter, H., Aarts, G. & Van Keeken, O., 2010. Residence time and behaviour of sole and cod in the Offshore Wind farm Egmond aan Zee (OWEZ). IMARES Wageningen UR.

    278. Witt, J., Schroeder, A., Knust, R. & Arntz, W.E., 2004. The impact of harbour sludge disposal on benthic macrofauna communities in the Weser estuary. Helgoland Marine Research, 58 (2), 117-128.

    279. Wright, E.P., Kemp, K., Rogers, A.D. & Yesson, C., 2015. Genetic structure of the tall sea pen Funiculina quadrangularis in NW Scottish sea lochs. Marine Ecology, 36 (3), 659-667.

    280. Wright, S.L., Rowe, D., Thompson, R.C. & Galloway, T.S., 2013. Microplastic ingestion decreases energy reserves in marine worms. Current Biology, 23 (23), R1031-R1033.

    281. Ysebaert, T., Meire, P., Maes, D. & Buijs, J., 1993. The benthic macrofauna along the estuarine gradient of the Schelde estuary. Netherlands Journal of Aquatic Ecology, 27 (2-4), 327-341.

    282. Zebe, E. & Schiedek, D., 1996. The lugworm Arenicola marina: a model of physiological adaptation to life in intertidal sediments. Helgoländer Meeresuntersuchungen, 50, 37-68.

    283. Zuhlke, R. & Reise, K., 1994. Response of macrofauna to drifting tidal sediments. Helgolander Meeresuntersuchungen, 48 (2-3), 277-289.

    284. Zwarts, L. & Esselink, P., 1989. Versatility of male curlews Numenius arquata preying upon Nereis diversicolor: deploying contrasting capture modes dependent on prey availability. Marine Ecology Progress Series, 56, 255-269.

    Citation

    This review can be cited as:

    Tillin, H.M. & Rayment, W., 2016. [Hediste diversicolor] and [Limecola balthica] in littoral sandy mud. In Tyler-Walters H. and Hiscock K. (eds) Marine Life Information Network: Biology and Sensitivity Key Information Reviews, [on-line]. Plymouth: Marine Biological Association of the United Kingdom. Available from: http://www.marlin.ac.uk/habitat/detail/209

    Last Updated: 31/08/2016