|Researched by||Rebecca McQuillan & Dr Heidi Tillin||Refereed by||This information is not refereed.|
This biotope usually occurs on flats of medium fine sand and muddy sand, most often on the lower shore but sometimes also on waterlogged mid shores. The sand may contain a proportion of shell fragments or gravel. Lan can also occur on the lower part of predominantly rocky or boulder shores, where patches of sand or muddy sand occur between scattered boulders, cobbles and pebbles. Conditions may be tide-swept, and the sediment may be mobile, but the biotope usually occurs in areas sheltered from strong wave action. The sediment supports dense populations of the sand mason Lanice conchilega. Other polychaetes present are tolerant of sand scour or mobility of the sediment surface layers and include the polychaetes Anaitides mucosa, Eumida sanguinea, Nephtys hombergii, Scoloplos armiger, Aricidea minuta, Tharyx spp. and Pygospio elegans. The mud shrimp Corophium arenarium and the cockle Cerastoderma edule may be abundant. The baltic tellin Macoma balthica may be present. On boulder shores, and where pebbles and cobbles are mixed in with lower shore tide-swept sand with dense Lanice conchilega between the cobbles, the infaunal component is rarely sampled. The infaunal community under these circumstances, provided that the cobbles are not packed very close together, is likely to be similar to that in areas without the coarse material (JNCC, 2015).
No text entered.
The infauna established under the prevailing environmental conditions has the capacity to modify the sedimentary regime through activities that primarily effect the stability of and sedimentation within the habitat, for example by tube building, bioturbation, feeding behaviour and production of faeces and pseudofaeces. Such processes modify the habitat and increase the number of niches available for colonization (Elliott et al., 1998). Increased species diversity and abundance are known to occur around biogenic structures, such as the tubes of polychaetes, in otherwise relatively homogenous sedimentary habitats (Woodin, 1978). Studies of Lanice conchilega aggregations in the Wadden Sea (Zühlke et al., 1998; Dittmann, 1999; Zühlke, 2001) showed that tubes built by Lanice conchilega had significant effects on the distribution, density and diversity of other macrobenthic species and meiobenthic nematodes compared to sites with a lower density of Lanice conchilega or ambient sediment without biogenic structures. There is considerable interaction between the species in this biotope; some are listed below.
The occurrence of Lanice conchilega on tidal flats can be subject to high seasonal variations (Dittmann, 1999). Seasonal storms can cause displacement of the polychaete (Ropert & Dauvin, 2000) and in the intertidal the polychaete is known to be susceptible to severe winters (Strasser & Pielouth 2001). Populations of the cockle Cerastoderma edule are also periodically decimated by severe winter weather, and a high winter mortality is often followed by an exceptionally heavy spring spatfall (Hayward, 1994).
Biological production within intertidal sandbanks is highly variable being reliant on the quantity of nutrients being delivered or internally generated (Elliott et al., 1998). Some primary production comes from benthic microalgae and water column phytoplankton. The microphytobenthos in the interstices of the sand grains consist of unicellular eukaryotic algae and cyanobacteria that grow in the upper several millimetres of illuminated sediments, typically appearing only as a subtle brown or green shading (Elliott et al., 1998). The benthos is supported predominantly by pelagic production and by detrital materials emanating from the coastal fringe (Barnes & Hughes, 1992). According to Barnes & Hughes (1992) the amount of planktonic food reaching the benthos is related to:
In the relatively shallow waters around the British Isles secondary production in the benthos is generally high, but shows seasonal variation (Wood, 1987). Generally, secondary production is highest during summer months, when temperatures rise and primary productivity is at its peak. Spring phytoplankton blooms are known to trigger, after a short delay, a corresponding increase in productivity in benthic communities (Faubel et al., 1983). Some of this production is in the form of reproductive products.
Characterizing macrofauna of the biotope are iteroparous, meaning that they breed several times per lifetime. Whilst some of the infauna have a benthic and brooding mode of reproduction (e.g. amphipods and oligochaetes), most are broadcast spawners (Rasmussen, 1973). For instance, polychaete worms including Lanice conchilega, Nephtys spp. and spionid worms release their eggs and sperm into the water where, after fertilization and a relatively prolonged planktonic phase of development, metamorphose and commence a benthic habit. Recruitment of Nephtys species seems related to environmental conditions in central parts of the species range, marginal populations exhibit occasional reproductive failures, e.g. Nephtys cirrosa, which is a temperate species and reaches the northern limit of its range in the north of the British Isles. Populations of Nephtys cirrosa on the east and west coasts of northern Britain exhibit different reproductive patterns. In south west Scotland gravid adults breed every year in early autumn, whilst those on the east coast experience periods (e.g. over three years) of reproductive failure (Olive & Morgan, 1991). Bivalve populations typically show considerable pluriannual variations in recruitment, suggesting that recruitment is patchy and/or post settlement processes are highly variable (e.g. Dauvin, 1985). For instance, adults of Cerastoderma edule spawn in a short peak period over summer with remaining adults spawning over a protracted period, resulting in a short (ca. 3 month) period of peak settlement followed by generally declining numbers of recruits (Hancock, 1967; Seed & Brown, 1977). For further information refer to full MarLIN species reviews.
The time required for the community to reach maturity will be in part determined by the proximity of other source populations and the season during which a disturbance occurs. Recolonization by some groups is likely to be more rapid than others. For instance, diatoms may be transported by resuspension in the water column and by lateral sediment transport. The rapid colonization (within days) by diatoms establishes food resources for other species, usually nematodes, that subsequently colonize. Dittmann et al. (1999) observed that the number of nematode species returned to pre-impact levels within seven days following a month long disturbance. Polychaetes tend to rapid colonizers, and species recorded by Dittmann et al. (1999) within two weeks included the polychaetes Pygospio elegans, Polydora sp., Nephtys hombergii, Capitella capitata, Heteromastus filiformis, Eteone longa, Hediste diversicolor (as Nereis diversicolor) and Scoloplos armiger, and the molluscs Limecola balthica and Mytilus edulis. Next to polychaetes, amphipods e.g. Urothoe poseidonis, are also rapid colonizers owing to their mobility. However, species that did not recolonize within the period of subsequent monitoring (14 months) included Arenicola marina, Lanice conchilega and its commensal Malmgreniella lunulata. Although it is likely that these species would recolonize suitable substrata, settlement of Lanice conchilega, for instance, has been reported to be more successful in areas with existent adults than areas without (see full MarLIN review; Heuers & Jaklin, 1999). Strasser & Pielouth (2001) reported that establishment of a mature population took three years in the absence of an established population. Thus the time taken for the community to reach maturity is likely to be in the order of several years.
No text entered.
|Depth Range||Mid shore, Lower shore|
|Water clarity preferences||Field Unresearched|
|Limiting Nutrients||Field unresearched|
|Salinity preferences||Full (30-40 psu), Variable (18-40 psu)|
|Physiographic preferences||Open coast, Strait / sound|
|Biological zone preferences||Eulittoral, Lower eulittoral, Mid eulittoral|
|Substratum/habitat preferences||Fine clean sand, Medium clean sand, Mixed|
|Tidal strength preferences||Moderately Strong 1 to 3 knots (0.5-1.5 m/sec.), Strong 3 to 6 knots (1.5-3 m/sec.), Very Strong > 6 knots (>3 m/sec.), Very Weak (negligible), Weak < 1 knot (<0.5 m/sec.)|
|Wave exposure preferences||Extremely sheltered, Moderately exposed, Sheltered, Very sheltered|
|Other preferences||Tidal streams|
The biotope description and characterizing species are based on JNCC (2015). This biotope usually occurs on flats of medium fine sand and muddy sand and is characterized by dense populations of the tube building polychaete Lanice conchilega. The sensitivity assessments are based on Lanice conchilega as the loss or reduction of the Lanice conchilega population would result in a significant change to the character of the biotope and lead to biotope reclassification. The tubes of this species create physical structure in this biotope and modify habitat conditions, this species is therefore considered to be a key structuring/functional species. A range of other species typically found on sandy shores are typically present, including the polychaetes Nephtys hombergii, Scoloplos armiger, and Pygospio elegans, the common cockle Cerastoderma edule and the amphipod Corophium arenarium may be abundant. The sensitivities of these species are assessed generally. More detailed information is available for these species on this website from biotopes where these are key characterizing species.
Recovery characteristics of Lanice conchilega were assessed previously by Ager (2008) and Callaway et al. (2010). Lanice conchilega is sessile and larval recolonization is therefore the most important recovery mechanism. Water transport of adults, intact in tubes, has been observed after storms and this represents a potential colonisation mechanism. The tube itself can be rapidly repaired or rebuilt following damage under variable conditions (Nicolaidou, 2003). Yonow (1989) observed Lanice conchilega re-establishing tubes immediately after removal from the sediment into a suitable sediment in the laboratory.
The larvae of Lanice conchilega spend up to 60 days as plankton and therefore have a wide dispersal potential. Larvae preferentially settle on the tubes of Lanice conchilega but other suitable settlement surfaces might be selected, including artificial tubes (Heuers & Jaklin 1999) or bivalve shells (Herlyn et al., 2008). Given the preference for settlement on the tubes of adults, recovery of Lanice conchilega populations will be enhanced by the presence of adults that survive impacts.
Population recovery times appear to range between one and four years (Beukema, 1990; Heuers, 1998; Zühlke, 2001; Callaway et al., 2010). Strasser & Pielouth (2001) reported that larvae were observed to settle in areas where there were no adults and the population was subsequently re-established in three years. The length of recovery time was attributed to a decline in the wider meta-population that reduced larval supply (Strasser & Pieloth, 2001). In areas where there are dense reefs with suitable water currents recovery may therefore be more rapid due to the greater larval supply, perhaps taking 1-2 years.
Beukema (1990) reported that following removal of entire intertidal populations (following a cold winter), Lanice conchilega populations recovered rapidly (within 1 and 2 years) and total-biomass value increased faster as a consequence of generally high successful recruitment in the Wadden Sea. However, Heuers et al. (1998) observed Lanice conchilega re-established abundance within 3 to 4 years in the Spiekeroog area of the Wadden Sea following severe freezing. In two of three areas disturbed by cultivation of Manila clams, measurable re-colonization took place after one or two years (Callaway et al., 2010 and references therein). The results tally with other descriptions of Lanice conchilega’s re-colonization strategy,
Recovery of the associated species may be more rapid than the recovery of dense patches of Lanice conchilega. Patches of cockles are naturally more variable over space and time (Smaal et al., 2005) and beds are subject to either gradual declines as the population aged but inhibits recruitment or occasional mass mortalities that have been attributed to a number of causes (Burdon et al., 2014). Small disturbed patches may be rapidly infilled by movement of adult cockles by tidal currents and wave action or active migration of adults. Nephtys hombergii are also mobile and may migrate to disturbed patches. Associated species with opportunistic life strategies (small size, rapid maturation and short-lifespan of 1-2 years) include Chaetozone setosa, Pygospio elegans and Corophium arenarium. These are likely to recolonise disturbed areas first, however, the actual pattern will depend on recovery of habitat, season of occurrence and other factors.
Resilience assessment. Where the biotope is severely impacted and the entire assemblage is removed, recovery is assessed as ‘Medium’ as re-establishment of a reef may require up to 3 years, based on the observations of Strasser & Pielouth (2001). Where resistance is assessed as ‘Low’ or ‘Medium’, the presence of adults is considered likely to enhance settlement of juveniles and recovery is assessed as ‘High’ (within 2 years). These assessments are considered to equally apply to the associated macrofauna. Recovery of typical biomass and age-structured populations of longer-lived species such as Ceratoderma edule and Nepthys hombergii is likely to be ‘Medium’ where populations are removed or there is an element of habitat change.
NB: The resilience and the ability to recover from human induced pressures is a combination of the environmental conditions, the frequency (repeated 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.
Species found high on the shore are generally able to tolerate a greater range of temperatures than those confined to lower shore levels (Davenport & Davenport, 2005), however, the timing of site-specific factors such as low tides will influence local acclimation. For intertidal species increased temperatures may also result in desiccation when exposed (see changes in emergence pressure).
Limited information on the thermal tolerance of Lanice conchilega was found. Lanice conchilega occurs in temperate regions all around UK and Irish shores in the intertidal and shallow subtidal, populations also occur in the Mediterranean, in the Arabian Gulf and the Pacific and along the south east of the North Sea, but the species is absent from Arctic waters (Ropert & Dauvin, 2000; Connor et al., 2004; Degraer et al., 2008). The geographical range suggests that the species is adapted to variable temperature conditions but local populations are likely to be acclimated to prevailing temperatures.
Cerastoderma edule is found from Norway to Mauritania (Honkoop et al., 2008) and through the Baltic, Mediterranean and Black Sea (Longshaw & Malham, 2013). The eastern border of distribution is the Murmansk coast of the Barents sea (Genelt-Yanovskiy et al., 2010). The species is therefore likely to be exposed to warmer and colder water and air temperatures than experienced in the UK over its geographic range. Kristensen (1958) reported that Cerastoderma edule from the Dutch Wadden Sea have an upper temperature tolerance of 31°C for 24 hrs while Wilson (1981) reported an upper lethal temperature of 42.5°C. These temperatures are likely to exceed the pressure benchmark. Wilson (1993) concluded that Cerastoderma edule was probably tolerant of a long-term temperature rise of 2°C associated with climate change. Warmer temperatures during winter result in increased metabolic rate, hence depletion of energy reserves in a time of low food availability and may contribute to post winter mortality of adult cockles (Wilson & Elkaim, 1991). Therefore, the tolerance of Cerastoderma edule to temperature change will be dependent on season, so an acute, short-term temperature rise in summer or decrease in winter may be detrimental.
Nephtys hombergii has been reported from as far south as South Africa, suggesting the species can tolerate temperatures above a 5°C increase on UK and Irish coasts. Records are limited, but Emery & Stevensen (1957) found that Nephtys hombergii could withstand summer temperatures of 30-35°C. Pygospio elegans are summer visitors to an effluent exposed habitat in the UK where effluents are typically 8-12°C warmer than the receiving waters (Bamber & Spencer, 1984), and this species is considered tolerant of acute and chronic increases in temperature.
Sensitivity assessment. Typical surface water temperatures around the UK coast vary seasonally from 4-19 oC (Huthnance, 2010). The associated species are considered likely to be tolerant of acute and chronic increases in temperature at the pressure benchmark. Lanice conchilega, Nephtys hombergii and Cerastoderma edule have a wide geographic range. As species inhabiting the intertidal, they are likely to experiences rapid fluctuations in temperature over the tidal cycle. The lack of evidence for mass mortalities in very hot summers (compared with reports for low winter temperatures) suggest that the characterizing and associated species are likely to tolerate a chronic increase at the pressure benchmark (2°C for one year). An acute increase in temperature for one month may lead to changes in reproductive success and predation, particularly on spat and juveniles. Adults may, however, survive. Biotope resistance is therefore assessed as ‘High’ and residence is ‘High’ (by default), the biotope is therefore considered to be ‘Not sensitive’.
Many intertidal species are resistant of freezing conditions as they are exposed to extremes of low air temperatures during periods of emersion. They must also be able to cope with sharp temperature fluctuations over a short period of time during the tidal cycle. Species that occur in the intertidal are generally adapted to periods of freezing conditions during winter months and have evolved mechanisms of survival to withstand rapid cooling (Loomis, 1995; Davenport & Davenport, 2005).
Infaunal species such as Pygospio elegans and Nephtys hombergii and including Lanice conchilega may be protected to some extent for short periods by the ability to burrow deeper within the sediment. Prolonged freezing temperatures may, however, result in mortalities. In the German coastal area during 1978-79 winter, water temperature remained below 0 °C on 45 successive days and resulted in the eradication of Lanice conchilega. Up to July 1980 no recolonization of the area had taken place (Buhr, 1981). An intertidal population of Lanice conchilega, in the northern Wadden Sea, was wiped out during the severe winter of 1995/96 (Strasser & Pielouth, 2001), and Crisp (1964) described mortality of Lanice conchilega between the tidemarks but not at lower levels during the severe winter of 1962/63. These severe winters probably exceed the pressure benchmark. Beukema (1990) had noted that the recovery of most species including Lanice conchilega, was rapid (within 1 and 2 years) and total-biomass value increased faster as a consequence of generally high successful recruitment in the Wadden Sea. Whilst Heuers et al. (1998) observed Lanice conchilega re-established abundance within 3 to 4 years in the Spiekeroog area of the Wadden Sea following severe freezing.
High mortalities of Cerastoderma edule populations attributed to severe winters have been reported by many authors (Kristensen, 1958; Hancock & Urquhart, 1964; Beukema, 1990; Ducrotoy et al., 1991) and high shore populations are likely to be more vulnerable to extremes of temperature owing to their longer emergence time. Strasser et al. (2001) noted mortalities ranging on average between 80% and 100% in the high tidal zone. Mortalities that occurred during spring and summer may have constituted a time-lag effect of the preceding winter. Beukema (1990) suggested that Cerastoderma edule was more sensitive to low winter temperatures than other bivalves (e.g. Mya arenaria, and Mytilus edulis) which make up 50% of the of the macrobenthic biomass (Reise & Lackschewitz, 1998).
Nephtys hombergii are found as far north as the Barents Sea, and would be expected to be resistant to a 5°C decrease in temperature for one month period, or 2°C for one year. However, during the winter of 1978- 79, Nephtys hombergii in the River Tyne estuary (north-east England) did not spawn but instead reabsorbed the gametes, reducing the Nephtys hombergii population for subsequent years. In previous years, recruitment had been good and it was considered probable that this was a result of the extremely cold conditions experienced in 1978-1979 (Olive et al., 1981).
Sensitivity assessment. Resistance has been assessed to be ‘Low’ as evidence suggests that intertidal populations of Lanice conchilega and Cerastoderma edule and other polychaetes are likely to be vulnerable to acute decreases in temperature during winter, resilience is assessed as ‘High’ where some adults remain and sensitivity is assessed as ‘Low’.
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
This biotope occurs in full (30-35 ppt) and variable salinity biotopes (18-35 ppt). Biotopes occurring in variable salinity are not considered to be sensitivie to an increase to full salinity as this falls within the natural range. At the pressure benchmark, an increase from full to hypersaline (>40 ppt) is considered. Local populations may be acclimated to the prevailing salinity regime and may exhibit different tolerances, therefore caution should be used when inferring tolerances from populations in different regions.
No direct evidence was found to assess the sensitivity of this biotope. Salinities within tide swept habitats vary due to weather conditions such as rain fall at low tide and evaporation causing hypersaline conditions on hot days. Intertidal shores within estuarine environments can also experience considerable short-term changes in salinities. During summer, the Atlantic French coast rapidly increased in salinity and reached hypersaline conditions (48 ppt) in the upper sediment layer within 1.5 hrs (Juneau et al., 2015). However, increased salinity is not restricted to hot summer conditions. This can occur at moderate optimal temperatures (16-20°C), due to wind-driven desiccation in spring in the first 0.5 cm of the sediment where Lanice conchilega, Nephtys hombergii, Pygospio elegans and Scoloplos armiger and the cockle Cerastoderma edule occur. This may interfere with the sediment cohesion and composition and possibly the ability to retain pore-water at the surface to modulate the species activity (Paterson & Hagerthey, 2001). Lanice conchilega as a tube dwelling polychaete has the ability to retract deeper into the sediment to avoid unfavourable conditions.
Verdelhos et al. (2015) assessed the saline optimal ranges of Cerastoderma edule (a euryhaline species) from 0-35 ppt. High activity was initially observed in conditions above the optimal range (15, 30 and 35 ppt) for Cerastoderma edule, but activity decreased after 72 hrs. Results suggest that high saline conditions over short periods is tolerated. Responses involved reduced activity through osmo-regulations and behaviours such as valve closure to prevent exposure.
Sensitivity assessment. Species within the biotope may tolerate short periods of high salinity, however, prolonged exposure to hypersalinity may lead to changes in species richness, abundance and biomass and loss of characterizing and associated species. Sensitivity to this pressure is not assessed due to lack of evidence.
The subtidal biotope is found in full salinity (JNCC, 2015), and the intertidal biotope and associated species are found in full and variable salinity. It is therefore considered that a decrease in salinity at the pressure benchmark, from full to variable (18-35 ppt), would not result in loss of the biotope. Since the intertidal biotope is found in full and variable salinity (18-35 ppt), the biotope is not considered sensitive to a change from full to variable salinity, as this falls within the natural habitat distribution. However, the biotope may be sensitive to a change from variable to low salinity (<18 ppt).
There is limited information on the effects of decreased salinity on Lanice conchilega. Salinities within tide swept habitats vary due to weather conditions such as rain fall at low tide and evaporation causing hypersaline conditions on hot days. Intertidal shores within estuarine environments can also experience considerable short-term changes in salinities.
Substantial salinity fluctuations due to heavy rains leading to scouring are common in intertidal areas and shallow coastal areas inhabited by adult Cerastoderma edule. Kristensen (1958) reported death of young cockle spat (1-2 mm) in the Dutch Wadden Sea due to heavy rain, although adults were able to dig deeper to avoid the freshwater influx (reduced salinities).
Boyden & Russell (1972) stated that Cerastoderma edule prefers salinities between 15 and 35 psu, whilst Russell & Peterson (1973) reported lower median salinity limits of 12.5 psu and upper median salinity limits of 38.5 psu. Rygg (1970) noted that Cerastoderma edule did not survive 23 days exposure to <10 psu. Rygg (1970) also demonstrated that salinity tolerance was temperature dependant (after 3 days, 100% survival at 33 psu and 35-38°C, but 50% mortality occurred at 20 psu and 37°C and 100% mortality at 13 psu and 37°C). Wilson (1984) noted that Cerastoderma edule remained open during 1 hour exposure to salinities between 13.3 and 59.3 psu. It should be noted that the tolerances reported above depended on the duration of the experiment.
Gogina et al. (2010) reported the lower limit of Scoloplos armiger was resistant of salinities of 8 psu, yet mortalities have been observed in the Baltic region, suggesting a lower limit of 10.5 psu. However, it was noted that Pygospio elegans can survive in salinities as low as 2 psu.
Nephtys hombergii has been shown to be highly resistant of a wide range of salinities (as low as 16 psu) (Clarke & Haderile, 1960; Davey & George, 1986), extending from the Mediterranean to the western Baltic Sea (Hartmann-Schroéder, 1996). Nephtys hombergii is considered to be a brackish water species (Barnes, 1994), but where the species occurs in open coastal locations the species would have to tolerate salinities of 25 psu and above. Within a few months of the closure of a dam across the Krammer-Volkerak estuary in the Netherlands, Wolff (1971b) observed that species with pelagic larvae or a free-swimming phase, including Nephtys hombergii, expanded rapidly with a concomitant increase of salinity to 9-15 psu everywhere. Prior to the closure of the dam the estuary demonstrated characteristics of a typical 'salt-wedge' estuary with a salinity gradient from 0.3 to 15 psu. Hence, Nephtys hombergii is likely to survive increases in salinity within estuarine environments. Nephtys hombergii may still be found in fully marine locations but may be competitively inferior to other species of Nephtyidae (e.g. Nephtys ciliata and Nephtys hystricis) and occur in lower densities. An increase to fully marine (30-40 ‰) would therefore be likely to lead to a reduction in density of Nephtys hombergii.
Pygospio elegans has been found in brackish water in high abundances, with salinities as low as 2 psu (Bolam & Fernandes, 2003).
Sensitivity assessment. Little empirical evidence was found to assess sensitivity of Lanice conchilega to this pressure at the benchmark level and therefore the assessment was made based on the biotopes associated species. The available evidence suggests that adult Cerastoderma edule may survive a reduction in salinity to just below 18 psu and populations may become locally acclimated to reduced salinities. Other species such as Pygospio elegans and Nephtys hombergii are also likely to be tolerant of a reduction at the pressure benchmark. As a reduction may lead to some mortalities and a reduction in growth and reproductive success of the key characterizing species, biotope resistance is assessed as ‘Medium’, recovery is assessed as ‘High’ (following restoration of typical conditions) and resilience is ‘High’. Biotope sensitivity is therefore assessed as ‘Low’.
This biotope is found in areas exposed to tidal streams varying from very strong (>3 m/s) to negligible (JNCC, 2015). There is limited evidence to assess changes in water flow rates for >1 year at the benchmark level. It should be noted that Lanice conchilega occupies shores with variable exposed flow speeds ranging from sheltered (<0.5 m/sec) to very strong (>3 m/sec) (Connor et al., 2004). Laboratory flume studies have showed that Lanice conchilega feeding rates on diatoms alter according to flow (Denis et al., 2007). Feeding rates were optimal at 0.15 m/s and lower at flow speeds of 0.04 m/s and 0.27 m/s (Denis et al., 2007). Changes in current velocity at the pressure benchmark are relevant to feeding rates, however, impacts are likely to depend on food supply, population density and proportion of inorganic particles.
The motion of the water is the dominant factor in the transportation of sediment, facilitating erosion and redistribution of materials (Hjulström, 1939). This may change the sediment structure and have associated effects on the intertidal community. Lanice conchilega alter between passive and active feeding depending on the tidal regimes, i.e. in still water, they move their tentacles or in high velocities the fringed tentacles are supported. A switch from active to passive feeding behaviour was observed during the experiments between 4 and 8 cm/s (Denis et al., 2007). Moderate to high velocities of water flow have been reported to enhance settlement of Lanice conchilega larvae (Harvey & Bourget, 1997), increasing average particle size in favour of gravels and pebbles.
Increase particle movement is likely to increase sediment scour (see abrasion and sedimentation), re-suspension of material (see organic enrichment, and suspended sediment), and potentially a decrease in clearance in filter feeders. Populations of Lanice conchilega and Pygospio elegans may decline in part due to a lack of suitable substrata to build its tubes. The biotope would probably become dominated by water flow resistant species that prefer coarse substratum.
Reduced water flow is a factor that has been identified as affecting the density of Lanice conchilega, as recruitment to the benthos is reduced (Harvey & Bourget, 1995). The average grain size of the sediment would also be reduced and probably favour deposit feeders and detritivores, to the detriment of the suspension feeders.
Sensitivity assessment. A change in water flow rate at the pressure benchmark level of 0.1-0.2 m/s is considered to fall within the range of flow speeds experienced by the biotope. Resistance and resilience are assessed as ‘High’ and the biotope considered ‘Not sensitive’ to a change in water flow at the pressure benchmark level.
The biotope and characterizing species occur in the mid to low intertidal. All characterizing species would probably survive an increase in emergence. However, Lanice conchilega can only feed when immersed and therefore likely to experience reduced feeding opportunities. Over the course of a year, the resultant energetic cost is likely to cause some mortality. In addition, increased emergence is likely to increase the vulnerability to predation from shore birds. A decrease in emergence is likely to allow the biotope to extend its upper limit, where suitable substrata exist.
The dense Lanice biotope LS.LSa.MuSa.Lan on certain lower shores may be a littoral extension of the biotope SS.SCS.ICS.SLan. A decrease in emergence may lead to some shifts in species present but a Lanice conchilega dominated biotope is likely to persist.
Sensitivity assessment. Emergence is a key factor structuring intertidal biotopes (JNCC, 2015). The Lanice conchilega biotope is likely to shift distribution during periods of altered sea levels. Some changes in the structure of the biological assemblage would be predicted. Resistance is assessed as ‘Low’ and resilience is ‘High’, following habitat recovery and sensitivity is assessed as ‘Low’.
The biotope is recorded from locations that have variable wave exposures (moderately exposed to extremely sheltered) (Connor et al., 2004). The degree of wave exposure influences wave height, as in more exposed areas with a longer fetch waves would be predicted to be higher.
Sensitivity assessment. As this biotope is found across a range of wave exposed shores, resistance is assessed as ‘High’ and resilience as ‘High’ to a 3 to 5% change in wave height at the pressure benchmark. Therefore, the biotope is considered ‘Not sensitive’.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
'Not sensitive' at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Contamination at levels greater than the pressure benchmark may adversely affect the biotope. There is a lack of evidence to identify the tolerance of Lanice conchilega and other polychaetes (Scoloplos armiger, Pygospio elegans and Nephtys hombergii) and including the common cockle, Cerastoderma edule to transition metals. Bryan (1984) reported that short-term toxicity in polychaetes was highest to Hg, Cu and Ag, declined with Al, Cr, Zn and Pb with Cd, Ni, Co and Se being the least toxic. It was recorded that polychaetes have a range of tolerances to heavy metal levels of Cu, Zn, As and Sn being in the order of 1500-3500 µg/g. An analysis of organisms from Restronguet Creek revealed that Nephtys hombergii from the middle and lower reaches of the creek contained appreciably higher concentrations of Cu (2227 µg/g dry wt), Fe and Zn than comparable specimens of Hediste diversicolor. There was evidence to suggest that some metals were regulated and suggests that some species (Nephtys hombergii) had developed metal resistant populations as a functional genetic trait to Cu homeostasis (McQuillan et al., 2014). However, Alve (1991) and Ellison et al. (1986) concluded that there is a shift in the abundance of biodiversity away from the source of contamination, which is indicative of metal contamination (Stubbles, 1993).
Bryan (1984) stated that Hg is the most toxic metal to bivalves. Studies of Cerastoderma edule transplanted populations from polluted and uncontaminated sites resulted in 10-15% mortality within 63 days but 100% within 4 months at the Restronguet Creek (Bryan & Gibbs, 1983). Additionally, Cu and Zn are believed to inhibit the settlement of juvenile Cerastoderma edule, leading to patchy distributions (Langston et al., 2003).
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
‘Not sensitive’ at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Contamination at levels greater than the pressure benchmark may adversely influences the biotope. Suchanek (1993) reviewed the effects of oil spills and concluded that soft sediment polychaetes, bivalves and amphipods were particularly sensitive. Dauvin (2000) noted that 20 years after the Amoco Cadiz oil spill in 1978, Lanice conchilega was re-established between 1978 and 1984 but disappeared after 1985. A similar delayed response was observed by Sanders (1980) and Kingston et al. (1995) as a result of the Florida oil spill and the Braer oil spill (Gómez Gesteira & Dauvin, 2000).
The Amoco Cadiz oil spill resulted in reductions in abundance, biomass and production of the invertebrate communities, however, Nephtys hombergii and other polychaetes (cirratulids and capitellids) were largely unaffected by the Amoco Cadiz oil spill (Conan, 1982). The sediment rapidly recovered and in 1981, benthic recruitment occurred under normal conditions (Dauvin, 1998).
Savari et al. (1991a) observed the density and growth of Cerastoderma edule decreased with increasing hydrocarbon concentration. McLusky (1982) examined the intertidal mudflat fauna at Kinneil in the Forth Estuary that received petroleum, chemical and domestic effluents. Evidence suggests that soft sediment communities are highly susceptible of hydrocarbon contamination. In the littoral zone especially, oil spills resulting from tanker accidents are likely to be deposited directly on the sediment of the biotope, preventing oxygen transport to the substratum and oil pushed in to the substratum by tidal-pulsing will destabilize the sediment (Elliott et al., 1998). Samples obtained from mussel tissue, 11 months after the Aegean Sea spill (1992), showed evidence of contamination. Yet 2 years later (October 1995), there was no evidence of oil pollution in that area affected by the Aegean Sea spill (Gómez Gesteira & Dauvin, 2005), suggesting recovery.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
‘Not sensitive’ at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
No evidence concerning specific effects of synthetic chemical contaminants on Lanice conchilega or Nephtys hombergii was found. Boon et al. (1985) reported that Nephtys spp. in the North Sea accumulated organochlorines. However, based on total sediment analyses, organochlorine concentrations in Nephtys spp. did not correlate with the concentrations in the sediment which they inhabited. Specific deleterious effects of synthetic chemicals (Ivermecten, tri-butyl-tin (TBT)) have been reported for other polychaetes, e.g. Hediste diversicolor, Arenicola marina, Scoloplos armiger and Cirratulus cirratus (Collier & Pinn, 1998; Beaumont et al., 1989; Bryan & Gibbs, 1991). Beaumont et al. (1989) concluded that bivalves are particularly sensitive to TBT. For example, when exposed to 1-3 µg TBT/l, Cerastoderma edule suffered 100% mortality after two weeks. Bryan & Gibbs (1991) presented evidence that TBT caused recruitment failure in bivalves, due to either reproductive failure or larval mortality. Waldock et al. (1999) examined recovery of benthic infauna of the Crouch Estuary after a ban on the use of TBT on small boats. They observed marked increases in species diversity, especially of Ampeliscid amphipods and polychaetes (Tubificoides spp. and Aphelochaeta marioni) which mirrored the decline in sediment TBT concentration. Whilst a causal link could not be shown, the study by Waldock et al. (1999) suggested that crustacean and polychaete diversity may be inhibited by TBT contamination.
Polychaetes vary greatly in their tolerance of chemical contamination. The persistence of these chemical residues is highly dependent on the matrix and ambient environmental conditions. Generally, residues in water are less likely to be a long-term concern because of photodegendation and dilution to below biological significant concentrations. However, TBT has a high binding affinity to sediments and residues incorporated into the sediment tend to persist for longer periods (Austen & McEvoy, 1997; Huntington et al., 2006).
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
'Not sensitive' at the pressure benchmark that assumes compliance with all relevant environmental protection standards.
Niermann et al. (1990) reported that the abundance of Lanice conchilega in a subtidal fine sand community in the German Bight area, was significantly reduced during a period of hypoxia (1-3 mg/O2/dm³). However, the period of hypoxia lasted for about a month and, therefore, exceeds the pressure benchmark. The activities of Lanice conchilega oxygenate sediments (Forster & Graf, 1995) and this species may not, therefore, be adapted to anoxic/hypoxic conditions unlike burrowing infauna such as Nephtys spp. and oligochaetes.
Nilsson & Rosenberg (1994) summarized that Nephtys hombergii have the potential to be tolerant of severe hypoxia and hydrogen sulphide (Alheit, 1978; Arndt & Schiedek, 1997). Nephtys hombergii survived 5 days in pure artificial anoxic seawater, not shifting from the aerobic to anaerobic metabolic pathways before oxygen saturation decreases to 112% (Schottler, 1982). Other experiments demonstrated that 50% of Nephtys hombergii survived 23 days in almost anaerobic conditions (10.2 mg/l), whereas Nephtys ciliata survived 11 days in the same tests (Fallesen & Jorgensen, 1991).
Rosenberg et al. (1991) observed that Cerastoderma edule migrated to the sediment surface in response to reduced oxygen concentrations in the upper sediment layers and reported 100% mortality of Cerastoderma edule exposed to 0.5-1.0 ml/l oxygen for 43 days. Theede et al. (1969) reported 50% mortality after 4.25 days at 1.5 ml/l oxygen. Theede et al. (1969) added that Cerastoderma edule only survived 4 days’ exposure of <6.1 cm²/l of hydrogen sulphide, which is associated with anoxic conditions. Cerastoderma edule fatalities are likely to occur at the benchmark.
Sensitivity assessment. Although some species associated with the biotope, such as Nephtys hombergii, are tolerant of hypoxia and anoxia, while other species that occur in oxygenated sediments or at the surface may be more sensitive. As the biotope occurs in the intertidal, oxygen recharge during immersion may almeliorate the effects of hypoxia and anoxia. Biotope resistance is therefore assessed as ‘Medium’ and resilience as ‘High’ (following restoration of typical habitat conditions). Biotope resistance is therefore assessed as ‘Low’. Due to the lack of evidence for Lanice conchilega, confidence is assessed as ’Low’.
|Not relevant (NR)||Not relevant (NR)||Not sensitive|
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).
Moderate enrichment increases food supply, enhancing productivity and abundance. Mudflats can be sensitive to nutrient enrichment which can result in blooms of opportunistic ephemeral seaweeds such as Enteromorpha spp. Macoma balthica has been shown experimentally to be able to resist time periods of 9 weeks under algal clover as their long siphon allows them to reach oxygenated water, although other bivalves decreased in abundance (Thiel et al., 1998). Organic enrichment beneath oyster cultivation trestles and mussel cultivation sites and fish cages has led to community replacement/dominance by Cirratulid, Capitellid and Spionid polychaetes, particularly Manayunkia aesturina in mudflats, which characterize disturbed areas enriched in organic matter (Pearson & Rosenberg, 1978; Samuelson, 2001, see Bouchet & Saurier, 2008 for references for activities). Local hydrodynamics and topographic features will mediate or strengthen biodeposition effects. Gross effects will lead to anoxic, defaunated sediments which may be covered by sulphur reducing bacteria such as Beggiatoa spp. (Elliott et al., 1998). Diatom density may be reduced by organic enrichment potentially reducing the stability of mudflats (Elliott et al., 1998).
Sensitivity assessment. 'Not sensitive' at the pressure benchmark that assumes compliance with good status as defined by the WFD.
There is limited evidence to assess the effects of organic enrichment directly on Lanice conchilega, other polychaetes (Pygospio elegans, Scoloplos armiger and Nephtys hombergii) and Cerastoderma edule. The response of benthic communities and of the sediment to increased organic carbon can be determined by the physical composition of the substrata and the availability of dissolved oxygen.
Cromey et al. (1998) found that organic loading rates <36 g C/m²/year had little effect on the macrobenthic communities; the sediment became enriched up to inputs of 365 g C/m²/year and organic loading >500 g C/m²/year led to degraded conditions probably due to lower oxygen concentrations.
Lanice conchilega has been categorised by Borja et al. (2000) as AMBI Group II - 'Species indifferent to enrichment, always present in low densities with non-significant variations with time (from initial state, to slight unbalance)'. This assessment was reviewed by Gittenberger & Van Loon (2011) and changed to AMBI Group III - 'Species tolerant to excess organic matter enrichment. These species may occur under normal conditions, but their populations are stimulated by organic enrichment (slight unbalance situations)'. From the AMBI index group, Cerastoderma edule and Pygospio elegans are identified in the Ecological Group III reflecting feeding type and other traits: 'Tolerant of excessive organic matter enrichment and populations are stimulated by organic enrichment' (Borja et al., 2000).
Sensitivity assessment. There is little empirical evidence to quantify the effect of organic enrichment deposits of 100 g C/m²/yr on Lanice conchilega. Based on the ranges presented by Cromey et al. (1998), this benchmark pressure is unlikely to have an influence upon Lanice conchilega and associated species. The resistance of ‘High’ and a resilience of ‘High’ may also be assigned and by default the biotope is ‘Not sensitive’.
All marine habitats and benthic species are considered to have a resistance of ‘None’ to this pressure and to be unable to recover from a permanent loss of habitat (resilience is ‘Very Low’). Sensitivity within the direct spatial footprint of this pressure is therefore ‘High’. Although no specific evidence is described, confidence in this assessment is ‘High’ due to the incontrovertible nature of this pressure.
The species characterizing this biotope occur within the upper layers of the soft sediment (JNCC, 2015). This biotope does not occur on artificial substrata or rock habitats, although patches of sediment may support some of the species associated with this biotope. Any substratum other than the sediments on which this biotope is found would therefore lead to a loss of this biotope. Consequently, biotope resistance is assessed as 'None’, resilience is assessed as ‘Very Low’ (as the change, at the benchmark is permanent), and biotope sensitivity is assessed as ‘High’.
This biotope is found in medium to fine muddy sand and mixed sediments (Connor et al., 2004). The biotope would not be considered sensitive to a change to mixed sediments from muddy sands or vice versa, as the sediment would still be likely to support the biotope. A change to coarse sediments or clean sands could, however, lead to loss of habitat suitability for the associated species that are restricted to sediments with a high proportion of fine factions, such as the burrowing polychaetes.
The characterizing species Lanice conchilega occurs in a range of sediments, including clean sands and coarser, clean sediments and pockets of sediment within otherwise unsuitable habitats. It is therefore likely to still occur in habitats where there has been a change in Folk class at the pressure benchmark. However, abundances of Lanice conchilega and the species richness and abundances of other species present are likely to alter. Degraer et al. (2006) report that Lanice conchilega displays a preference for fine to medium-grained sediments (100 to 500 μm) with relatively high mud content (10 to 40%), and other sediment types are therefore likely to be sub-optimal.
Dense aggregations of Lanice conchilega facilitate the deposition of fine materials as the tubes reduce the velocity of flow. Some increase in fine sediment fraction to muds is likely to be tolerated. Sylvand (1995) highlighted the dominant presence of Lanice conchilega in areas of oyster cultivation where fine particles were deposited.
High densities of Nephtys hombergii were found in substrata of 0.3% particles >0.25 mm and 5.8% <0.125 mm in diameter, but tolerated up to 3.8% 0.25 mm and 2.2-15.9% <0.125 mm (Clark et al., 1962). Nephtys hombergii may be found in higher densities in muddy environments. An increase in gravel content, although tolerated by Nephtys hombergii, may lead to increased abundance of Nephtys cirrosa and decreased abundance of Nephtys hombergii. Degraer et al. (2006) summarized that the higher the medium grain size of the sediment the lower the relative occurrence of Nephtys hombergii and in grain sizes over 0.5 mm the species was absent in the Belgium part of the North Sea.
Habitat partitioning has been observed between Corophium volutator and Corophium arenarium on the basis of sediments, although modified by disturbance from other species (see removal of target species). Corophium arenarium is associated with areas of sandy sediment whilst Corophium volutator prefers muddier sediments (Watkins, 1941; Meadows, 1964; Flach, 1993), changes in sediment type may, therefore, lead to shifts in the species present.
Sensitivity assessment. Changes to increased finer sediments are likely to be tolerated, however, a change in the sediment Folk class classification at the pressure benchmark to gravel is likely to alter the abundance of Lanice conchilega and the composition of the species assemblage, resulting in loss of this biotope. Resistance is therefore assessed as 'None' and resilience as 'Very Low' (as the change at the pressure benchmark is permanent). The biotope is considered to have 'High' sensitivity to this pressure.
Sedimentary communities are likely to be highly intolerant of substratum removal, which will lead to partial or complete defaunation, exposure of underlying sediment which may be anoxic and/or of a different character or bedrock and lead to changes in the topography of the area (Dernie et al., 2003). Any remaining species, given their new position at the sediment/water interface, may be exposed to conditions to which they are not suited, i.e. unfavourable conditions. Newell et al. (1998) state that removal of 0.5 m depth of sediment is likely to eliminate benthos from the affected area.
Removal of 30 cm of sediment will remove species that occur at the surface and within the upper layers of sediment. The extraction of sediment would remove the characterizing species Lanice conchilega, and the associated species present, including Cerastoderma edule, which is found to a depth of 5 cm and Nephtyid species and other polychaetes, such as Scoloplos armiger and Pygospio elegans that burrow between 5 and 20 cm into the sediment (Schӧttler, 1982; Pedersen, 1991; Kruse et al., 2004).
Hydrodynamics and sedimentology (mobility and supply) influence the recovery of soft sediment habitats (Van Hoey et al., 2008). Recovery of the sedimentary habitat would occur via infilling, some recovery of the biological assemblage may take place before the original topography is restored, if the exposed, underlying sediments are similar to those that were removed.
Sensitivity assessment. Extraction of 30 cm of sediment will remove the characterizing biological component of the biotope. Resistance is assessed as ‘None’ and biotope resilience is assessed as ‘Medium’ as some sediment recovery may be required before recovery processes begin. Biotope sensitivity is therefore assessed as ‘Medium’.
The characterizing species of this biotope (Lanice conchilega) has robust, flexible tubes and may retract below the surface. They are also able to rapidly rebuild or repair tubes (Nickolaidou, 2003). These characteristics reduce exposure to this pressure and enhance recovery. Rabaut et al. (2008) studied fisheries impacts at the species level in temperate sandy bottom areas. A controlled field manipulation experiment was designed focusing on areas with high densities of Lanice conchilega (i.e. Lanice conchilega reefs). A treatment zone was exposed to a one-off experimental trawling and the impact on, and recovery of, the associated fauna was investigated for a period of 9 days post-impact. Community analysis showed a clear impact on associated species such as Eumida sanguinea, followed by a relatively quick recovery. The passage of a single beam trawl did not significantly alter the density of Lanice conchilega (Rabaut et al., 2008).
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. Surface compaction can collapse burrows and reduce the pore space between particles, decreasing penetrability and reducing stability and oxygen content (Sheehan, 2007). The tops of burrows may be damaged and repaired subsequently at energetic cost to their inhabitants. Experiments have shown that areas subject to compaction from trampling tend to have reduced species abundance and diversity (Sheehan, 2007; Rossi et al., 2007). Sheehan (2007) proposed that following compaction, organisms avoid or emigrate from affected areas.
Rossi et al. (2007) investigated the effects of continuous trampling (a source of abrasion) on intertidal soft sediments, with on average of 5 people for 3-5 hrs, twice a month between March and September (2005). The abundance of Cerastoderma edule was reduced presumably by sediment compression, leading to asphyxia in adults. Rossi et al. (2007) conducted the investigation during spawning periods, when juveniles where present in the water column. There was therefore little observable effect on juvenile Cerastoderma edule abundances due to continuous recruitment. Rossi et al. (2007) noted that during the growing season, recovery can be fast but in the long-term trampling increased the recruitment of Limecola balthica at the expense of Cerastoderma edule.
Abrasion is likely to reduce the density of Corophium spp. by emigration and increased mortality, as this species is sensitive to sediment disturbance from bioturbating species (Flach, 1993).
Exposure to abrasion may vary according to life stage. For example, the egg cocoons of Scoloplos armiger and Nephtys hombergii are laid on the sediment surface before hatching and entering into their planktonic stages of development. The juveniles found a few millimetres below the surface (Reise, 1979; Kruse et al., 2004) are sensitive to the effects of abrasion.
Sensitivity assessment. The experiments by Rabaut et al. (2008) suggest that Lanice conchilega has ‘High’ resistance to abrasion, however, other associated species may be more impacted. Biotope resistance to a single abrasion event is assessed as ‘High’ based on the key characterizing species Lanice conchilega. There may be some damage to tubes but this is expected to be sub-lethal and tubes are likely to be rapidly repaired. Recovery from impacts of associated species is predicted to be ‘High’ and the biotope is assessed as ‘Not sensitive’.
The characterizing species of this biotope (Lanice conchilega) has robust, flexible tubes and may retract below the surface. They are also able to rapidly rebuild or repair tubes (Nickolaidou, 2003). These characteristics reduce exposure to this pressure and enhance recovery.
Rabaut et al. (2008) studied fisheries impacts at the species level in temperate sandy bottom areas. A controlled field manipulation experiment was designed focusing on areas with high densities of Lanice conchilega (i.e. Lanice conchilega reefs). A treatment zone was exposed to a one-off experimental trawling, and the impact on and recovery of the associated fauna was investigated for a period of 9 days post-impact. Community analysis showed a clear impact followed by a relatively quick recovery. The passage of a single beam trawl did not significantly alter the density of Lanice conchilega. Rabaut et al. (2009) also studied the direct mortality of Lanice conchilega as a consequence of sustained physical disturbance at varying frequencies to reflect the effect of beam trawl fisheries. Research was based on a laboratory experiment in which four different disturbance regimes were applied (disturbance every other 12, 24 and 48 h and no fishing disturbance as a control). Survival dropped significantly after 10 and 18 days (with a disturbance frequency of every 12 and 24 h, respectively). The results indicate that Lanice conchilega is relatively resistant to physical disturbance but that reef systems can potentially collapse under continuous high frequency disturbance.
Ferns et al. (2000) investigated the effect of tractor dredging for cockles on an intertidal muddy sand at Burry Inlet, South Wales mechanical cockle harvesting. The tubes of Lanice conchilega were damaged but this damage was rapidly repaired. Associated species were more vulnerable and a decline of 31% in populations of Scoloplos armiger (initial density 120/m2) was recorded in muddy sands and an 83% decline in Pygospio elegans (initial density 1850/m2). Pygospio elegans were significantly depleted for >100 days after harvesting (surpassing the study monitoring timeline). Whilst Nephtys hombergii and Scoloplos armiger demonstrated recovery >50 days after harvesting in muddy sands. Cerastoderma edule recovered more quickly than those in muddy sand with a more structured community, which included Pygospio elegans and Lanice conchilega in clean sands (Ferns et al., 2000).
Hand raking for cockles on intertidal silty sandflats, using rakes that penetrated the surface by 5-10 cm, resulted in a three-fold increase in the damage rate of cockles compared to control plots and, in the short-term, led to a relative decrease in the overall abundance of fauna (Kaiser et al., 2001). After 56 days, the small (9 m2) plots had recovered but the larger (36 m2) plots remained in an altered state. Results collected over a year after the disturbance suggested that while effects of hand-raking may be significant within a year, they are unlikely to persist beyond this time-scale unless there are larger long-lived species present within the community (Kaiser et al., 2001). Effects may be longer lasting where sediment damage occurs without recovery. A study in the Dutch Wadden Sea showed that suction-dredging for cockles (Cerastoderma edule) led to a significant long-term reduction in settlement and stocks of the target bivalve species (Piersma et al., 2001). Analysis of sediment characteristics before and after dredging showed an increase in median grain size and a reduction of silt content, and that these changes were most pronounced in the area dredged for cockles. Sediment characteristics only returned to pre-impact conditions 8-11 years after the suction dredging.
Sensitivity assessment. The experiments by Rabaut et al. (2008, 2009) suggest that Lanice conchilega has ‘High’ resistance to abrasion and penetration, however, other associated species may be more impacted. Biotope resistance to a single abrasion event is assessed as ‘High’ based on the key characterizing species Lanice conchilega. There may be some damage to tubes but this is expected to be sub-lethal and tubes are likely to be rapidly repaired. Recovery from impacts of associated species is predicted to be ‘High’ and the biotope is assessed as ‘Not sensitive’.
Production within the biotope is predominantly secondary, derived from detritus and to some extent phytoplanktonic production. Characteristic infauna do not require light and therefore the effects of increased turbidity on light attenuation are not directly relevant. However, an increase in turbidity may affect primary production in the water column and therefore reduce the availability of phytoplankton as food but phytoplankton would also be transported in to the biotope from distant areas, so the effect of increased turbidity may be mitigated. As soon as light levels return to normal, phytoplanktonic primary production would increase, the species would resume optimal feeding, so recoverability has been assessed to be very high. A decrease in suspended organic particles may reduce food supply to Lanice conchilega, however, this species can switch from suspension to deposit feeding and vice versa (Buhr & Winter, 1977). Supply of materials needed to build the tubes may be affected with a decrease in suspended solids but wave action would be likely to continue to re-suspend and transport sediments.
Sensitivity assessment. A decrease in turbidity and hence increased light penetration may result in increased phytoplankton production and hence increased food availability for suspension feeders, including Lanice conchilega and Cerastoderma edule. Therefore, reduced turbidity may be beneficial. In areas of high suspended sediment, a decrease may result in improved condition and recruitment due to a reduction in the clogging of filtration apparatus of suspension feeders and an increase in the relative proportion of organic particulates. However, a decrease in suspended organic particles in some areas may reduce food availability for deposit feeders resulting in lower growth or reduced energy for reproduction.
Where increased turbidity results from organic particles then subsequent deposition may enhance food supply for suspension and deposit feeders within the biotope. Alternatively, if turbidity results from an increase in suspended inorganic particles then energetic costs may be imposed on these species as sorting and feeding becomes less efficient, reducing growth rates and reproductive success. Lethal effects are considered unlikely given the occurrence of Lanice conchilega and Cerastoderma edule and other associated species in estuaries where turbidity is frequently high from suspended organic and inorganic matter. Resistance and resilience are therefore assessed as 'High' and the biotope is considered to be 'Not sensitive'.
The tube of Lanice conchilega protrudes above the sediment by 1–4 cm, an adaptation to trap suspended particles (Zühlke, 2001). Furthermore, Ziegelmeier (1952) showed that the tubes of Lanice conchilega increased the height with increasing sedimentation so that it could continue feeding and metabolising aerobically. This enables Lanice conchilega to switch between deposit and suspension feeding (Buhr, 1976). Therefore, Lanice conchilega is suitably adapted to discrete events of sedimentation, characteristic of sandy, muddy sand intertidal flats. Lanice conchilega has dominated areas of Manila clam cultivation where protective netting had led to an increase in sedimentation (Spencer et al.,1998) and in areas where oysters were cultivated and sedimentation increased (Sylvand, 1995).
Cerastoderma edule has short siphons and needs to keep in contact with the surface of the sediment. It will quickly burrow to the surface if covered by as little as 2 cm of sediment (Richardson et al., 1993b; Cotter et al., 1997). When buried to 5 cm depth in a series of siltation experiments most cockles returned to the surface but few were able to reposition to the surface if buried at 10 cm depth (Jackson & James, 1979). None had died after 72 hours. Additional burial experiments under 10 cm of sediment assessed movement and survival after 3, 6 and 9 days in two sediment types (a mix of surface mud and sand in two ratios 9:1 and 1:9). The movement towards the surface was slower in the predominantly muddy sediment and all cockles died between 3 and 6 days. Substantial mortality resulted in the predominantly sandy mixture, although some cockles were able to move towards the surface and survive for 9 days (Jackson & James, 1979).
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 pseudofaeces (Sornin et al., 1983; Deslous-Paoli et al., 1992; Mitchell, 2006; Bouchet & Sauriau, 2008). Pygospio elegans is known to decline in areas following re-deposition of very fine particulate matter (Rhoads & Young, 1970; 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). Nephtys species are highly mobile within the sediment. Vader (1964) observed that Nephtys hombergii relocated throughout the tidal cycle and is unlikely to be affected by smothering with sediment consistent with that of the habitat.
Sensitivity assessment. Based upon the ability of Lanice conchilega to increase tube elevation above the sediment surface when deposition has increased, it is likely that Lanice conchilega is resistant to the effects of a single discrete event of deposition of 5 cm. Although Cerastoderma edule and Pygospio elegans may be impacted, the Lanice conchilega biotope is unlikely to be altered from a single event. Sensitivity to continuous events will depend on tidal hydrodynamics, and spawning and recruitment, prior to the event (see recover/resilience rates). As the benchmark refers to a single discrete event, resistance is ‘High’ and resilience is ‘High’, therefore this biotope is ‘Not sensitive’.
No evidence was found to assess sensitivity of Lanice conchilega to heavy sedimentation. It is likely that intertidal flats may sometimes be subject to sedimentation following storm events but no species-specific examples were found. Ziegelmeier (1952) showed that Lanice conchilega increased the height of its tube top with increasing sedimentation so that it could continue feeding and respire. However, heavy sedimentation of 30 cm during a single event may have a severe effect on Lanice conchilega and the associated species.
Cerastoderma edule has short siphons and needs to keep in contact with the surface of the sediment. Bait digging (for large polychaete worms) disturbs the sediment down to a depth of 30 cm (Jackson & James, 1979) and leads to repositioning within sediment or burial from piled sediments. Intensification of bait digging on intertidal flats on the north Norfolk coast for lugworm (Arenicola marina) and ragworm (Hediste diversicolor) was associated with declines in the Cerastoderma edule populations (Jackson & James, 1979). Experimental simulation of bait digging (sediment dug over to a depth of 30 cm with a garden fork) led to high mortalities of cockles in dug areas rather than undug areas (48% mortality in 9 days to a maximum of 85% after 11 days), probably due to smothering (Jackson & James, 1979). The observation was tested experimentally in the laboratory by burying 3 size-classes of cockles at 0, 5 or 10 cm depth in a mix of oxidated and deeper anoxic sands (mixed as a ratio of 3:1). Movements were recorded after 24, 48 and 72 hours. When buried to 5 cm depth, most cockles returned to the surface but few were able to reposition to the surface if buried at 10 cm depth. None had died after 72 hours. Additional burial experiments under 10 cm of sediment assessed movement and survival after 3, 6 and 9 days in two sediment types (a mix of surface mud and sand in two ratios 9:1 and 1:9). The movement towards the surface was slower in the predominantly muddy sediment and all cockles died between 3 and 6 days. Substantial mortality resulted in the predominantly sandy mixture, although some cockles were able to move towards the surface and survive for 9 days.
Sensitivity assessment. The addition of 30 cm of sediment would prevent Cerastoderma edule from extending siphons to the surface. It is unlikely that this species could emerge from this depth of sediment, although some individuals may survive and sediment may be rapidly removed by tide and wave action. It is likely, however, that there would be considerable mortality of Cerastoderma edule and other species, including Lanice conchilega. Resistance is assessed as ‘Low’ and resilience as ‘High’. Sensitivity is therefore assessed as ‘Low’. Although some polychaetes may be able to reposition following sedimentation at the pressure benchmark, this will depend on the characteristics of the overburden, and sedentary species such as Pygospio elegans are likely to suffer high levels of mortality.
|Not Assessed (NA)||Not assessed (NA)||Not assessed (NA)|
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
No information was found concerning the intolerance of the biotope or the characterizing species to noise. The siphons of bivalves and palps of polychaetes are likely to detect vibrations and would probably be withdrawn as a predator avoidance mechanism. However, it is unlikely that the biotope and characterizing species will be affected by noise or vibrations caused by noise at the level of the benchmark and this pressure is assessed as ‘Not relevant’.
|No evidence (NEv)||No evidence (NEv)||No evidence (NEv)|
No evidence. As this feature is not characterized by the presence of primary producers, it is not considered that shading would alter the character of the habitat directly. However, there may be changes in microphytobenthos abundance. Littoral muddy sands support microphytobenthos on the sediment surface and within the sediment. The microphytobenthos consists of unicellular eukaryotic algae and cyanobacteria that grow within the upper several millimetres of illuminated sediments, typically appearing only as a subtle brownish or greenish shading. Mucilaginous secretions produced by these algae may stabilise fine substrata (Tait & Dipper, 1998). The biomass of the benthic microalgae often exceeds that of the phytoplankton in the overlying waters (MacIntyre et al., 1996) such that benthic microalgae play a significant role in system productivity and trophic dynamics, as well as habitat characteristics such as sediment stability. Shading will prevent photosynthesis leading to death or migration of sediment microalgae, altering sediment cohesion and food supply to higher trophic levels.
Barriers that reduce the degree of tidal excursion may alter larval supply to suitable habitats from source populations. 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. Resistance to this pressure is assessed as 'High' and resilience as 'High' by default. This biotope is therefore considered to be 'Not sensitive'.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
‘Not relevant’ to seabed habitats. NB. Collision by grounding vessels is addressed under ‘surface abrasion’.
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
|Not relevant (NR)||Not relevant (NR)||Not relevant (NR)|
Characterizing and associated species are not cultivated or transplanted and this pressures is considered 'Not relevant'.
Intertidal flats may be colonized by the invasive non-indigenous species Crepidula fornicata and the pacific oyster Crassostrea gigas. The two species have not only attained considerable biomasses from Scandinavian to Mediterranean countries but have also generated ecological consequences such as alterations of benthic habitats and communities or food chain changes (OSPAR, 2009).
In the Wadden Sea, the Pacific oyster Magallana gigas has colonized intertidal flats (Smaal et al., 2005). This species may consume larvae including the pelagic larvae of Cerastoderma edule and Limecola balthica and other species, reducing recruitment (Smaal et al., 2005). Troost et al. (2009) investigated whether Magallana gigas had a competitive feeding advantage over Cerastoderma edule and concluded that differences in feeding current characteristics were small and not considered to lead to significant differences in feeding efficiency. 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.
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). Densities of Cerastoderma edule and Abra tenuis have increased since the introduction of the Manila clam, but the abundance of Scrobicularia plana and Limecola balthica has declined (Caldow et al., 2005). However, 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) and 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’.
No evidence was found for microbial pathogens, parasites or diseases of Lanice conchilega or other characterizing polychaetes.
A review of parasites, pathogens and commensals identified a range of agents impacting European cockles, including 50 conditions affecting Cerastoderma edule (Longshaw & Malham, 2013). Cockles are hosts to viruses, bacteria, fungi, Apicomplexa , Amoeba, Ciliophora, Perkinsozoa, Haplosporidia, Cercozoa, Turbellaria, Digenea, Nematoda, Crustacea and Nemertea. Mortalities are associated particularly with digeneans and some protistan infections; parasites may limit growth, reduce fecundity and alter burrowing behaviour (Longshaw & Malham, 2013). A number of examples of conditions associated with mass mortalities of Cerastoderma edule are presented below. Parasites and disease are more likely to cause mortalities in populations that are subject to suboptimal conditions or other stressors such as hot summers or cold winters (Longshaw & Malham, 2013).
The Cerastoderma edule catchment in Galicia (NW Spain) was previously a productive cockle fishery, however, cockle mortality rate increased sharply in this bed in April 2012, reaching 100% in May 2012. Marteiliosis, which was first detected in February 2012 and reached 100% prevalence in April 2012, was identified as the most probable cause. Marteiliosis had never been detected in Galician cockles. Extensive surveillance of the Galician coast in May to July 2012 detected Marteiliosis in most cockle beds of the Ría de Arousa, whereas it was not found in other rías; 2 months later, the cockle catch in the Ría de Arousa became negligible as a result of a Martelia cochillia protozoan infection (Villalba et al., 2014).
Sensitivity assessment. Based on the evidence for Cerastoderma edule, it is likely that parasitic infection may indirectly alter the species composition of the biotope, however, there is no evidence of pathogens influencing Lanice conchilega and the other characterizing polychaetes. Resistance is assessed as ‘High’ as the biotope classification would be unlikely to change and removal of Cerastoderma edule may increase the extent of Lanice conchilega patches and the abundance of species that are sensitive to its sediment disturbing activities, such as Pygospio elegans and Corophium arenarium. Resilience is therefore assessed as ‘High’ and the biotope is considered to be ‘Not sensitive’.
The direct, physical impacts of removal are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of target species on this biotope. The cockle Cerastoderma edule may be targeted by recreational and commercial fishers and the polychaete Nephtys hombergii may be targeted by bait collectors. The removal of target species (Cerastoderma edule) may result in changes to the sediment structure (see physical disturbances pressures).
Dense populations of Cerastoderma edule on intertidal flats support commercial fisheries in several areas of the UK and the species is also harvested throughout Europe. Harvesting may use mechanical methods (e.g. tractor dredges or hydraulic suction dredging) or hand collection using rakes or other methods. The method of harvesting cockles will influence the proportions that are removed and damaged. Pickett (1973) found that intense dredging for a short period on a bed of cockle spat had little effect on survival and growth, although Cook (1991) found that impacts on small cockles from dredging were variable, with little reduction one year but a reduction in density observed the following year. Cotter et al. (1997) assessed the catch rates and damage and mortality of Cerastoderma edule resulting from experimental tractor dredging at the Burry Inlet (Wales). Stocks of adult cockles were reduced by 31 and 49% in low and high density areas respectively. Similarly, mechanical cockle harvesting in muddy sand reduced the abundance of Cerastoderma edule by ca 34%. Populations had not recovered to their original abundance after 174 days (Ferns et al., 2000).
Removal of adult cockles by harvesting or other factors allows other species to establish. Following experimental removal of large adult Cerastoderma edule by Cesar & Frid (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. This biotope is characterized by Lanice conchilega, the removal of Cerastoderma edule and Nephthys hobergii may affect the biotope through physical damage (see abrasion and penetration pressures) but their loss would not lead to biotope reclassification. Resistance is therefore assessed as ‘Medium’ as removal would lead to some alteration of the biological assemblage through changes in abundance and biomass and through changes in competition for space and sediment disturbance (Cerastoderma edule) and predation (Nephtys hombergii). The densities of prey species and species sensitive to sediment disturbance such as Corophium arenarium may increase. Removal of competitors and predators may enhance recruitment of Lanice conchilega. Where sediments remain suitable cockles are likely to recolonise via adult migration, survival of small, discarded cockles or via larval recruitment. Resilience is assessed as ‘High’ so that biotope sensitivity is assessed as ‘Low’.
Species within this biotope may be removed or damaged by static or mobile gears that are targeting other species. These direct, physical impacts are assessed through the abrasion and penetration of the seabed pressures. The sensitivity assessment for this pressure considers any biological/ecological effects resulting from the removal of non-target species on this biotope.
Lanice conchilega (sand mason worm) is the dominant polychaete in this biotope. It qualifies as an ‘ecosystem engineer’ in that it changes and/or creates a habitat, which affects the abundance of other species (Jones et al., 1994, 1997). The tube which Lanice conchilega builds provides structure and stabilises the sediment (Jones & Jago, 1993). The tubes also obstruct the activities of predatory burrowers enabling other sedentary animals to establish themselves (Wood, 1987). The burrows and tubes allow oxygenated water to penetrate into the sediment, the oxic upper layer of sediment to support shallow burrowing species such as amphipods (Ampelisca spp., Bathyporeia spp. & Gammarus spp.) and small Crustacea. Lanice conchilega may therefore increase species richness and abundance through habitat modification (Rabaut et al., 2007; Zuhlke, 2001). Dense aggregates of Lanice conchilega have the potential to enhance food supply to birds and fish due to increased biodiversity (Rabaut et al., 2010; Godet et al., 2008; Van Hoey et al., 2008; Callaway et al., 2010; De Smet et al., 2013).
Sensitivity assessment. The removal of non-target species may result in changes to the biological community and hence the classification of the assemblage type as assessed in the biotope. Incidental removal of Lanice conchilega would alter the biotope classification and most likely result in changes to the habitat and associated species. As Lanice conchilega is sedentary and present at the surface, resistance is ‘Low’ and resilience is ‘High’. Biotope sensitivity is therefore assessed as ‘Low’. This assessment considers the ecological effect of removal, the abrasion pressure assessment provides evidence for the robustness of tubes and the species may be resistant to removal by bottom trawls deployed at low intensity (e.g. single passes).
Ager, O.E.D. 2008. Lanice conchilega Sand mason. 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/species/detail/1642
Alheit, J., 1978. Distribution of the polychaete genus Nephtys: a stratified random sampling survey. Kieler Meeresforschungen, 4, 61-67.
Alve, E., 1991. Benthic foraminifera in sediment cores reflecting heavy metal pollution in Sorfjord, Western Norway. The Journal of Foraminiferal Research, 21 (1), 1-19.
Amos, C., Li, M. & Sutherland, T., 1998. The contribution of ballistic momentum flux to the erosion of cohesive beds by flowing water. Journal of Coastal Research, 14 (2), 564-569.
Ansell, A.D., Barnett, P.R.O., Bodoy, A. & Masse, H., 1981. Upper temperature tolerances of some European Mollusca III. Cardium glaucum, C. tuberculata and C. edule. Marine Biology, 65, 177-183.
Arndt, C. & Schiedek, D., 1997. Nephtys hombergii, a free living predator in marine sediments: energy production under environmental stress. Marine Biology, 129, 643-540.
Austen, M.C. & McEvoy, A.J., 1997. Experimental effects of tributyltin (TBT) contaminated sediment on a range of meiobenthic communities. Environmental Pollution, 96 (3), 435-444.
Bamber, R.N. & Spencer, J.F. 1984. The benthos of a coastal power station thermal discharge canal. Journal of the Marine Biological Association of the United Kingdom, 64, 603-623.
Bandeira, S.O., 1995. Marine botanical communities in southern Mozambique: Sea grass and seaweed diversity and conservation. Ambio, 24, 506-509.
Barnes, R.S.K. & Hughes, R.N., 1992. An introduction to marine ecology. Oxford: Blackwell Scientific Publications.
Barnes, R.S.K., 1994. The brackish-water fauna of northwestern Europe. Cambridge: Cambridge University Press.
Beadle, L.C., 1937. Adaptation to changes of salinity in the Polychaetes. Journal of Experimental Biology, 14 (1), 56-70.
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.
Bergström, P., 2014. Blue Oceans with Blue Mussels-Management and planning of mussel farming in coastal ecosystems. Ph.D. thesis, Biological and Environmental Sciences, University of Gothenburg.
Beukema, J.J., 1987. Influence of the predatory polychaete Nephtys hombergii on the abundance of other polychaetes. Marine Ecology Progress Series, 40, 95-107.
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.
Beukema, J.J., De Bruin, W. & Jansen, J.J.M., 1978. Biomass and species richness of the macrobenthic animals living on the tidal flats of the Dutch Wadden Sea: Long-term changes during a period of mild winters. Netherlands Journal of Sea Research, 12, 58-77.
Bolam, S.F. & Fernandes, T.F., 2003. Dense aggregations of tube-building polychaetes: response to small-scale disturbances. Journal of Experimental Marine Biology and Ecology, 269 (2), 197-222.
Bolam, S.G., Fernandes, T.F., Read, P. & Raffaelli, D., 2000. Effects of macroalgal mats on intertidal sandflats: an experimental study. Journal of Experimental Marine Biology and Ecology, 249 (1), 123-137.
Boon, J.P., Zantvoort, M.B., Govaert, M.J.M.A. & Duinker, J.C ., 1985. Organochlorines in benthic polychaetes (Nephtys spp.) and sediments from the southern North Sea. Identification of individual PCB components. Netherlands Journal of Sea Research, 19, 93-109.
Borja, Á., Rodríguez, J.G., Black, K., Bodoy, A., Emblow, C., Fernandes, T.F., Forte, J., Karakassis, I., Muxika, I. & Nickell, T.D., 2009. Assessing the suitability of a range of benthic indices in the evaluation of environmental impact of fin and shellfish aquaculture located in sites across Europe. Aquaculture, 293 (3), 231-240.
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.
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.
Boyden, C.R. & Russel, P.J.C., 1972. The distribution and habitat range of the brackish water cockle (Cardium (Cerastoderma) edule) in the British Isles. Journal of Animal Ecology, 41, 719-734.
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.
Brandt, G., Fleßner, J., Glaser, D. et al., 1995. Dokumentation zur hydrographischen Frühjahrs-Meßkampagne 1994 der ökosystemforschung Niedersächsisches Wattenmeer im Einzugsgebeit der otzmer Balje. Hyrographie Nr. 8, Nieders. Landesamt für ökologie - Forschungstelle Küste, Norderney.
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.
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.
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.]
Bryan, G.W. & Gibbs, P.E., 1991. Impact of low concentrations of tributyltin (TBT) on marine organisms: a review. In: Metal ecotoxicology: concepts and applications (ed. M.C. Newman & A.W. McIntosh), pp. 323-361. Boston: Lewis Publishers Inc.
Bryan, G.W., 1984. Pollution due to heavy metals and their compounds. In Marine Ecology: A Comprehensive, Integrated Treatise on Life in the Oceans and Coastal Waters, vol. 5. Ocean Management, part 3, (ed. O. Kinne), pp.1289-1431. New York: John Wiley & Sons.
Buhr, K-J., 1976. Suspension-feeding and assimilation efficiency in Lanice conchilega. Marine Biology, 38, 373-383.
Buhr, K.J., 1981. Effects of the cold winter 1978/79 on the macrobenthos of the Lanice-association in the Weser Estuary. Veroffentlichungen des Instituts fur Meeresforschung in Bremerhaven, 19, 115–131.
Buhr, K.J. & Winter, J.E., 1977. Distribution and maintenance of a Lanice conchilega association in the Weser estuary (FRG), with special reference to the suspension-feeding behaviour of Lanice conchilega. In Proceedings of the Eleventh European Symposium of Marine Biology, University College, Galway, 5-11 October 1976. Biology of Benthic Organisms (ed. B.F. Keegan, P.O. Ceidigh & P.J.S. Boaden), pp. 101-113. Oxford: Pergamon Press.
Burdon, D., Callaway, R., Elliott, M., Smith, T. & Wither, A., 2014. Mass mortalities in bivalve populations: A review of the edible cockle Cerastoderma edule (L.). Estuarine, Coastal and Shelf Science, 150, 271-280.
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.
Callaway, R., 2003a. Long-term effects of imitation polychaete tubes on benthic fauna: they anchor Mytilus edulis (L.) banks. Journal of Experimental Marine Biology and Ecology, 283 (1), 115-132.
Callaway, R., 2006. Tube worms promote community change. Marine Ecology Progress Series, 308, 49-60.
Callaway, R., Desroy, N., Dubois, S.F., Fournier, J., Frost, M., Godet, L., Hendrick, V.J. & Rabaut, M., 2010. Ephemeral Bio-engineers or Reef-building Polychaetes: How Stable are Aggregations of the Tube Worm Lanice conchilega (Pallas, 1766)? Integrative and Comparative Biology, 50 (2), 237-250.
Carey, D.A., 1983. Particle resuspension in the benthic boundary layer induced by flow around polychaete tubes. Canadian Journal of Fisheries and Aquatic Sciences, 40 (Suppl. 1), 301-308.
Carey, D.A., 1987. Sedimentological effects and palaeoecological implications of the tube-building polychaete Lanice conchilega Pallas. Sedimentology, 34, 49-66.
Cesar, C.P. & Frid, C.L., 2012. Benthic disturbance affects intertidal food web dynamics: implications for investigations of ecosystem functioning. Marine Ecology Progress Series, 466, 35.
Clark, R.B. & Haderlie, E.C., 1960. The distribution of Nephtys cirrosa and Nephtys hombergii of the south western coasts of England and Wales. Journal of Animal Ecology, 29, 117-147.
Clark, R.B., Alder, R.R. & McIntyre, A.D., 1962. The distribution of Nephtys sp. on the Scottish coast. Journal of Animal Ecology, 31, 359-372.
Collier, L.M. & Pinn, E.H., 1998. An assessment of the acute impact of the sea lice treatment Ivermectin on a benthic community. Journal of Experimental Marine Biology and Ecology, 230, 131-147.
Conan, G., 1982. The long-term effects of the Amoco Cadiz oil spill. Philosophical Transactions of the Royal Society of London B, 297, 323-333.
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.
Connor, D.W., Allen, J.H., Golding, N., Howell, K.L., Lieberknecht, L.M., Northen, K.O. & Reker, J.B., 2004. The Marine Habitat Classification for Britain and Ireland. Version 04.05. Joint Nature Conservation Committee, Peterborough. www.jncc.gov.uk/MarineHabitatClassification.
Connor, D.W., 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.
Connor, D.W., Dalkin, M.J., Hill, T.O., Holt, R.H.F. & Sanderson, W.G., 1997a. Marine biotope classification for Britain and Ireland. Vol. 2. Sublittoral biotopes. Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06., Joint Nature Conservation Committee, Peterborough, JNCC Report no. 230, Version 97.06.
Cook, W., 1991. Studies on the effects of hydraulic dredging on cockle and other macroinvertebrate populations 1989-1990. LANCASTER UNIV., LANCASTER(UK). 1991.
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.
Cotter, A.J.R., Walker, P., Coates, P., Cook, W. & Dare, P.J., 1997. Trial of a tractor dredger for cockles in Burry Inlet, South Wales. ICES Journal of Marine Science, 54, 72-83.
Crisp, D.J. (ed.), 1964. The effects of the severe winter of 1962-63 on marine life in Britain. Journal of Animal Ecology, 33, 165-210.
Cromey, C., Black, K., Edwards, A. & Jack, I., 1998. Modelling the deposition and biological effects of organic carbon from marine sewage discharges. Estuarine, Coastal and Shelf Science, 47 (3), 295-308.
Cryer, M., Whittle, B.N. & Williams, K., 1987. The impact of bait collection by anglers on marine intertidal invertebrates. Biological Conservation, 42, 83-93.
Dame, R.F.D., 1996. Ecology of Marine Bivalves: an Ecosystem Approach. New York: CRC Press Inc. [Marine Science Series.]
Dauvin, J.-C., Thiébaut, E., Gesteira, J.L.G., Ghertsos, K., Gentil, F., Ropert, M. & Sylvand, B., 2004. Spatial structure of a subtidal macrobenthic community in the Bay of Veys (western Bay of Seine, English Channel). Journal of Experimental Marine Biology and Ecology, 307 (2), 217-235.
Dauvin J C. & Ruellet, T., 2007. Polychaete/amphipod ratio revisited. Marine Pollution Bulletin, 55, 215–224.
Dauvin, J.C., 1985. Dynamics and production of a population of Venus ovata (Pennant) (Mollusca-Bivalvia) of Morlaix Bay (western English Channel). Journal of Experimental Marine Biology and Ecology, 91, 109-123.
Dauvin, J.C., 1998. The fine sand Abra alba community of the Bay of Morlaix twenty years after the Amoco Cadiz oil spill. Marine Pollution Bulletin, 36, 669-676.
Dauvin, J.C., 2000. The muddy fine sand Abra alba - Melinna palmata community of the Bay of Morlaix twenty years after the Amoco Cadiz oil spill. Marine Pollution Bulletin, 40, 528-536.
Davenport, J. & Davenport, J.L., 2005. Effects of shore height, wave exposure and geographical distance on thermal niche width of intertidal fauna. Marine Ecology Progress Series, 292, 41-50.
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.
Davies, C.E. & Moss, D., 1998. European Union Nature Information System (EUNIS) Habitat Classification. Report to European Topic Centre on Nature Conservation from the Institute of Terrestrial Ecology, Monks Wood, Cambridgeshire. [Final draft with further revisions to marine habitats.], Brussels: European Environment Agency.
De Smet, B., Godet, L., Fournier, J., Desroy, N., Jaffré, M., Vincx, M. & Rabaut, M., 2013. Feeding grounds for waders in the Bay of the Mont Saint-Michel (France): the Lanice conchilega reef serves as an oasis in the tidal flats. Marine Biology, 160 (4), 751-761.
Degraer, S., Verfaillie, E., Willems, W., Adriaens, E., Vincx, M. & Van Lancker, V., 2008. Habitat suitability modelling as a mapping tool for macrobenthic communities: An example from the Belgian part of the North Sea. Continental Shelf Research, 28 (3), 369-379.
Degraer, S., Wittoeck, J., Appeltans, W., Cooreman, K., Deprez, T., Hillewaert, H., Hostens, K., Mees, J., Vanden Berghe, E. & Vincx, M., 2006. The macrobenthos atlas of the Belgian part of the North Sea. Belgian Science Policy, Brussel.
Denis, L., Desroy, N. & Ropert, M., 2007. Ambient flow velocity and resulting clearance rates of the terebellid polychaete Lanice conchilega (Pallas, 1766). Journal of Sea Research, 58 (3), 209-219.
Dernie, K.M., Kaiser, M.J., Richardson, E.A. & Warwick, R.M., 2003. Recovery of soft sediment communities and habitats following physical disturbance. Journal of Experimental Marine Biology and Ecology, 285-286, 415-434.
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.
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.
Dittmann, S., 1999. Biotic interactions in a Lanice conchilega dominated tidal flat. In The Wadden Sea ecosystem, (ed. S. Dittmann), pp.153-162. Germany: Springer-Verlag.
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.
Ducrotoy, C.R., Rybarczyk, H., Souprayen, J., Bachelet, G., Beukema, J.J., Desprez, M., Dõrjes, J., Essink, K., Guillou, J., Michaelis, H., Sylvand, B., Wilson, J.G., Elkaïm, B. & Ibanez, F., 1991. A comparison of the population dynamics of the cockle (Cerastoderma edule) in North-Western Europe. In Proceedings of the Estuarine and Coastal Sciences Association Symposium, ECSA 19, 4-8 September 1989, University of Caen, France. Estuaries and Coasts: Spatial and Temporal Intercomparisons, pp. 173-184. Denmark: Olsen & Olsen.
Dyrynda, P., 1995. Impacts of bait dragging on the seabed within Poole Harbour. Report to Southern Sea District Fisheries Committee from the Marine Environmental Research Group, University of Wales, Swansea.
Dyrynda, P. & Lewis K., 1995. Ecological studies within the Crymlyn Burrows SSSI (Swansea Bay, Wales): impact of mechanised cockle harvesting. Marine Environment Research Group Report, University of Swansea, Swansea, UK.
Eckman, J.E., 1985. Flow perturbation by a protruding animal tube affects sediment bacterial recolonization. Journal of Marine Research, 43, 419-435.
Eckman, J.E., Nowell, A.R.M. & Jumars, P.A., 1981. Sediment destabilization of animal tubes. Journal of Marine Research, 39, 361-374.
Eleftheriou, A. & McIntyre, A.D., 1976. The intertidal fauna of sandy beaches-a survey of the Scottish coast. Scottish Fisheries Research Report, 6, 1-61.
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.
Elliott, M. & Quintino, V., 2007. The estuarine quality paradox, environmental homeostasis and the difficulty of detecting anthropogenic stress in naturally stressed areas. Marine Pollution Bulletin, 54 (6), 640-645.
Elliott, M., 1994. The analysis of macrobenthic community data. Marine Pollution Bulletin, 28, 62-64.
Elliott, M., 1998. Summary of effects of commercial fisheries on estuarine ecosystems: a European perspective. , Unpublished report to SCOR working group 105, Halifax NS, March 1998. P57.
Ellison, R.L., Broome, R. & Ogilvie, R., 1986. Foraminiferal response to trace metal contamination in the Patapsco River and Baltimore Harbour, Maryland. Marine Pollution Bulletin, 17 (9), 419-423.
Emery, K.O. & Stevenson, R.E., 1957. Estuaries and lagoons. In Treatise on marine ecology and paleoecology.1. Ecology, (ed. J.W. Hedgpeth), USA: Geological Society of America.
Everett, R.A., 1991. Intertidal distribution of infauna in a central California lagoon: the role of seasonal blooms of macroalgae. Journal of Experimental Marine Biology and Ecology, 150, 223-247.
Fallesen, G. & Jørgensen, H.M., 1991. Distribution of Nephtys hombergii and Nephtys ciliata (Polychaeta: Nephtyidae) in Århus Bay, Denmark, with emphasis on the severe oxygen deficiency. Ophelia, Supplement 5, 443-450.
Faubel, A., Hartig, E. & Thiel, H., 1983. On the ecology of the benthos of sublittoral sediments, Fladen Ground, North Sea. 1. Meiofauna standing stock and estimation of production. Meteor Forschungsergebnisse, 36, 35-48.
Fauchald, K., (2007). World Register of Polychaeta. available online at http://www.marinespecies.org/polychaeta
Fernandez-Gonzalez, V., Martinez-Garcia, E. & Sanchez-Jerez, P., 2016. Role of fish farm fouling in recolonisation of nearby soft-bottom habitats affected by coastal aquaculture. Journal of Experimental Marine Biology and Ecology, 474, 210-215.
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.
Flach, E.C., 1993. The distribution of the amphipod Corophium arenarium in the Dutch Wadden Sea- relationships with sediment composition and the presence of cockles and lugworms. Netherlands Journal of Sea Research, 31 (3), 281-290.
Flach, E.C. & De Bruin, W., 1994. Does the activity of cockles, Cerastoderma edule (L.) and lugworms, Arenicola marina (L.), make Corophium volutator Pallas more vulnerable to epibenthic predators: a case of interaction modification? Journal of Experimental Marine Biology and Ecology, 182, 265-285.
Flach, E.C., 1996. The influence of the cockle, Cerastoderma edule, on the macrozoobenthic community of tidal flats of the Wadden Sea. Marine Biology, 17, 87-98.
Forster, S. & Graf, G., 1995. Impact of irrigation on oxygen flux into the sediment: intermittent pumping by Callianassa subterranea and "piston pumping" by Lanice conchilega. Marine Biology, 123, 335-346.
Franklin, A. & Pickett, G.D., 1978. Studies of the indirect effects of fishing on stocks of cockles, Cardium edule, in the Thames estuary and Wash. Fisheries Research, Technical Report, no. 42, MAFF, Lowestoft, pp. 9.
Galbraith, H., Jones, R., Park, R., Clough, J., Herrod-Julius, S., Harrington, B. & Page, G., 2002. Global climate change and sea level rise: potential losses of intertidal habitat for shorebirds. Waterbirds, 25 (2), 173-183.
Genelt-Yanovskiy, E., Poloskin, A., Granovitch, A., Nazarova, S. & Strelkov, P., 2010. Population structure and growth rates at biogeographic extremes: a case study of the common cockle, Cerastoderma edule (L.) in the Barents Sea. Marine Pollution Bulletin, 61 (4), 247-253.
Gesteira, J.G. & Dauvin, J.-C., 2005. Impact of the Aegean Sea oil spill on the subtidal fine sand macrobenthic community of the Ares-Betanzos Ria (Northwest Spain). Marine Environmental Research, 60 (3), 289-316.
Gesteira, J.G. & Dauvin, J.-C., 2000. Amphipods are good bioindicators of the impact of oil spills on soft-bottom macrobenthic communities. Marine Pollution Bulletin, 40 (11), 1017-1027.
Gittenberger, A. & Van Loon, W.M.G.M., 2011. Common Marine Macrozoobenthos Species in the Netherlands, their Characterisitics and Sensitivities to Environmental Pressures. GiMaRIS report no 2011.08.
Godet, L., Toupoint, N., Olivier, F., Fournier, J. & Retière, C., 2008. Considering the Functional Value of Common Marine Species as a Conservation Stake: The Case of Sandmason Worm Lanice conchilega (Pallas 1766) (Annelida, Polychaeta) Beds. AMBIO: A Journal of the Human Environment, 37 (5), 347-355.
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.
Gomez Gesteira, J.L. & Dauvin, J.C., 2000. Amphipods are good bioindicators of the impact of oil spills on soft-bottom macrobenthic communities. Marine Pollution Bulletin, 40 (11), 1017-1027.
Graf, G., 1989. Benthic-pelagic coupling in a deep-sea benthic community. Nature, 341 (6241), 437-439.
Gray, J.S., 1981. The ecology of marine sediments. An introduction to the structure and function of benthic communities. Cambridge: Cambridge University Press.
Gray, J.S., Wu R.S.-S. & Or Y.Y., 2002. Effects of hypoxia and organic enrichment on the coastal marine environment. Marine Ecology Progress Series, 238, 249-279.
Hancock, D.A. & Urquhart, A.E., 1964. Mortalities of edible cockles (Cardium edule L.) during the severe winter of 1962-64. Journal of the Marine Biological Association of the United Kingdom, 33, 176-178.
Hancock, D.A., 1967. Growth and mesh selection in the edible cockle (Cardium edule L.). Journal of Applied Ecology, 4, 137-157.
Hartmann - Schroeder, G., 1996. Annelida, Borstenwuermer, Polychaeta, Jena: Fischer.
Harvey, M. & Bourget, E.,1997. Recruitment of marine invertebrates onto arborescent epibenthic structures: active and passive processes acting at different spatial scales. Marine Ecology Progress Series, 153, 203‐215.
Harvey, M. & Bourget, E., 1995. Experimental evidence of passive accumulation of marine bivalve larvae on filamentous epibenthic structures. Limnology and Oceanography, 40, 94-104.
Hayward, P.J. 1994. Animals of sandy shores. Slough, England: The Richmond Publishing Co. Ltd. [Naturalists' Handbook 21.]
Helmuth, B. & Denny, M.W., 2003. Predicting wave exposure in the rocky intertidal zone: Do bigger waves always lead to larger forces? Limnology and Oceanography, 48 (3), 1338-1345.
Helmuth, B., Harley, C.D., Halpin, P.M., O'Donnell, M., Hofmann, G.E. & Blanchette, C.A., 2002. Climate change and latitudinal patterns of intertidal thermal stress. Science, 298 (5595), 1015-1017.
Herlyn, M., Millat, G. & Petersen, B., 2008. Documentation of sites of intertidal blue mussel (Mytilus edulis L.) beds of the Lower Saxonian Wadden Sea, southern North Sea (as of 2003) and the role of their structure for spatfall settlement. Helgoland Marine Research, 62 (2), 177-188.
Heuers, J. & Jaklin, S., 1999. Initial settlement of Lanice conchilega. Senckenbergiana Maritima, 29 (suppl.), 67-69.
Heuers, J., 1998. Ansiedlung, Dispersion, Rekrutierung und Störungen als strukturierende Faktoren benthischer Gemeinschaften im Eulitoral. Dissertation, Universität Bonn.
Heuers, J., Jaklin, S., Zülkhe, R., Dittmann, S., Günther, C-P., Hildenbrandt, H. & Grimm, V., 1998. A model on the distribution and abundance of the tube-building polychaete Lanice conchilega (Pallas, 1766) in the intertidal of the Wadden Sea. Verhandlungen Ges Ökologie, 28, 207-215.
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.
Hjulström, F., 1939. Transportation of detritus by moving water: Part 1. Transportation. Recent Marine Sediments, a Symposium (ed. P.D. Trask), pp. 5-31. Dover Publications, Inc.
Honkoop, P.J., Berghuis, E.M., Holthuijsen, S., Lavaleye, M.S. & Piersma, T., 2008. Molluscan assemblages of seagrass-covered and bare intertidal flats on the Banc d'Arguin, Mauritania, in relation to characteristics of sediment and organic matter. Journal of Sea Research, 60 (4), 255-263.
Huntington, T., Roberts, H., Cousins, N., Pitta, V., Marchesi, N., Sanmamed, A. & Brockie, N., 2006. Some Aspects of the Environmental Impact of Aquaculture in Sensitive Areas. Report to the DG Fish and Maritime Affairs of the European Commission. Poseidon Aquatic Resource Management Ltd.
Huthnance, J., 2010. Ocean Processes Feeder Report. London, DEFRA on behalf of the United Kingdom Marine Monitoring and Assessment Strategy (UKMMAS) Community.
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.
Jennings, S. & Kaiser, M.J., 1998. The effects of fishing on marine ecosystems. Advances in Marine Biology, 34, 201-352.
JNCC, 2015. The Marine Habitat Classification for Britain and Ireland Version 15.03. JNCC: JNCC. 2015(20/05/2015). jncc.defra.gov.uk/MarineHabitatClassification
Jones, C.G., Lawton, J.H. & Shackak, M., 1994. Organisms as ecosystem engineers. Oikos, 69, 373-386.
Jones, C.G., Lawton, J.H. & Shackak, M., 1997. Positive and negative effects of organisms as ecosystem engineers. Ecology, 78, 1946-1957.
Jones, J., Duerdoth, C., Collins, A., Naden, P. & Sear, D., 2014. Interactions between diatoms and fine sediment. Hydrological Processes, 28 (3), 1226-1237.
Jones, S.E. & Jago, C.F., 1993. In situ assessment of modification of sediment properties by burrowing invertebrates. Marine Biology, 115, 133-142.
Juneau, P., Barnett, A., Méléder, V., Dupuy, C. & Lavaud, J., 2015. Combined effect of high light and high salinity on the regulation of photosynthesis in three diatom species belonging to the main growth forms of intertidal flat inhabiting microphytobenthos. Journal of Experimental Marine Biology and Ecology, 463, 95-104.
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.
Kaiser, M., Clarke, K., Hinz, H., Austen, M., Somerfield, P. & Karakassis, I., 2006. Global analysis of response and recovery of benthic biota to fishing. Marine Ecology Progress Series, 311, 1-14.
Kesäniemi, J.E., Mustonen, M., Boström, C., Hansen, B.W. & Knott, K.E., 2014. Temporal genetic structure in a poecilogonous polychaete: the interplay of developmental mode and environmental stochasticity. BMC Evolutionary Biology, 14 (1), 1-16.
Kessler, M., 1963. Die Entwicklung von Lanice conchilega (Pallas) mit besonderer Berücksichtigung der Lebensweise. Helgolander Wissenschaftliche Meeresuntersuchungen, 8, 425-476.
Kingston, P., 1974. Some observations on the effects of temperature and salinity upon the growth of Cardium edule and Cardium glaucum larvae in the laboratory. Journal of the Marine Biological Association of the United Kingdom, 54, 309-317.
Kingston, P.F., Dixon, I.M.T., Hamilton, S. & Moore, D.C., 1995. The impact of the Braer oil spill on the macrobenthic infauna of the sediments off the Shetland Islands. Marine Pollution Bulletin, 30 (7), 445-459.
Kirwan, M.L., Guntenspergen, G.R., D'Alpaos, A., Morris, J.T., Mudd, S.M. & Temmerman, S., 2010. Limits on the adaptability of coastal marshes to rising sea level. Geophysical Research Letters, 37 (23).
Kochmann, J., Buschbaum, C., Volkenborn, N. & Reise, K., 2008. Shift from native mussels to alien oysters: differential effects of ecosystem engineers. Journal of Experimental Marine Biology and Ecology, 364 (1), 1-10.
Kristensen, I., 1958. Differences in density and growth in a cockle population in the Dutch Wadden Sea. Archives Néerlandaises de Zoologie, 12, 351-453.
Kruse, I., Strasser, M. & Thiermann, F., 2004. The role of ecological divergence in speciation between intertidal and subtidal Scoloplos armiger (Polychaeta, Orbiniidae). Journal of Sea Research, 51, 53-62.
Kuhl, H., 1972. Hydrography and biology of the Elbe Estuary. Oceanography and Marine Biology: an Annual Review, 10, 225-309.
Lackschewitz, D. & Reise, K., 1998. Macrofauna on flood delta shoals in the Wadden Sea with an underground association between the lugworm Arenicola marina and the amphipod Urothoe poseidonis. Helgolander Meeresuntersuchungen, 52, 147-158.
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.
Laverock, B., Smith, C.J., Tait, K., Osborn, A.M., Widdicombe, S. & Gilbert, J.A., 2010. Bioturbating shrimp alter the structure and diversity of bacterial communities in coastal marine sediments. The ISME journal, 4 (12), 1531-1544.
Levin, L.A. & Gage, J.D., 1998. Relationships between oxygen, organic matter and the diversity of bathyal macrofauna. Deep Sea Research Part II: Topical Studies in Oceanography, 45 (1), 129-163.
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.
Loomis, S.A., 1995. Freezing tolerance of marine invertebrates. Oceanography and Marine Biology: an Annual Review, 33, 337-350.
MacIntyre, H.L. & Cullen, J.J., 1996. Primary production by suspended and benthic microalgae in a turbid estuary: time-scales of variability in San Antonio Bay, Texas. Marine Ecology Progress Series, 145 (1), 245-268.
MacIntyre, H.L., Geider, R.J. & Miller, D.C., 1996. Microphytobenthos: The ecological role of the "secret garden" of unvegetated, shallow-water marine habitats. 1. Distribution, abundance and primary production. Estuaries, 19, 186-201.
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.
McLaughlin, E., Portig, A. & Johnson, M.P., 2007. Can traditional harvesting methods for cockles be accommodated in a Special Area of Conservation? ICES Journal of Marine Science: Journal du Conseil, 64 (2), 309-317.
McLusky, D., Anderson, F. & Wolfe-Murphy, S., 1983. Distribution and population recovery of Arenicola marina and other benthic fauna after bait digging. Marine Ecology Progress Series, 11 (2), 173-179.
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.
McLusky, D.S., 1989. The Estuarine Ecosystem, 2nd ed. New York: Chapman & Hall.
McLusky, D.S., Bryant, D.M. & Elliott, M., 1992. The impact of land‐claim on macrobenthos, fish and shorebirds on the forth estuary, eastern Scotland. Aquatic Conservation: Marine and Freshwater Ecosystems, 2 (3), 211-222.
McQuillan, J.S., Kille, P., Powell, K. & Galloway, T.S., 2014. The regulation of copper stress response genes in the Polychaete Nereis diversicolor during prolonged extreme copper contamination. Environmental science & technology, 48 (22), 13085-13092.
Meadows, P., 1964. Substrate selection by Corophium species: the particle size of substrates. The Journal of Animal Ecology, 33, 387-394.
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.
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.
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.
Nicolaidou, A., 2003. Observations on the re-establishment and tube construction by adults of the polychaete Lanice conchilega. Journal of the Marine Biological Association of the United Kingdom, 83 (06), 1223-1224.
Niermann, U., 1996. Fluctuation and mass occurrence of Phoronis muelleri (Phoronidea) in the south-eastern North Sea during 1983-1988. Senckenbergiana Maritima, 28, 65-79.
Niermann, U., Bauerfeind, E., Hickel, W. & Westernhagen, H.V., 1990. The recovery of benthos following the impact of low oxygen content in the German Bight. Netherlands Journal of Sea Research, 25, 215-226.
Nilsson, H.C. & Rosenberg, R., 1994. Hypoxic response of two marine benthic communities. Marine Ecology Progress Series, 115, 209-217.
Nugues, M., Kaiser, M., Spencer, B. & Edwards, D., 1996. Benthic community changes associated with intertidal oyster cultivation. Aquaculture Research, 27 (12), 913-924.
Olive, P.J.W. & Morgan, P.J., 1991. The reproductive cycles of four British intertidal Nephtys species in relation to their geographical distribution (Polychaeta: Nephtyidae). Ophelia, Supplement 5, 351-361.
Olive, P.J.W., Garwood, P.R., Bentley, M.G. & Wright, N., 1981. Reproductive success, relative abundance and population structure of two species of Nephtys in an estuarine beach. Marine Biology, 63 (2), 189-196.
OSPAR Commission. 2009. Background document for Modiolus modiolus beds. OSPAR Commission Biodiversity Series. OSPAR Commission: London. Available from: http://www.ospar.org/documents?v=7193
Papaspyrou, S., Gregersen, T., Cox, R.P., Thessalou-Legaki, M. & Kristensen, E., 2005. Sediment properties and bacterial community in burrows of the ghost shrimp Pestarella tyrrhena (Decapoda: Thalassinidea). Aquatic Microbial Ecology, 38 (2), 181-190.
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.
Pedersen, T.F., 1991. Metabolic adaptations to hypoxia of two species of Polychaeta, Nephtys ciliata and Nephtys hombergii. Journal of Comparative Physiology B, 161 (2), 213-215.
Petersen, B. & Exo, K.M., 1999. Predation of waders and gulls on Lanice conchilega tidal flats in the Wadden Sea. Marine Ecology Progress Series, 178, 229-240.
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.
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/
Piersma, T., Koolhaas, A., Dekinga, A., Beukema, J.J., Dekker, R. & Essink, K., 2001. Long‐term indirect effects of mechanical cockle‐dredging on intertidal bivalve stocks in the Wadden Sea. Journal of Applied Ecology, 38 (5), 976-990.
Poggiale, J.C. & Dauvin, J.C., 2001. Long term dynamics of three benthic Ampelisca (Crustacea - Amphipoda) populations from the Bay of Morlaix (western English Channel) related to their disappearance after the Amoco Cadiz oil spill. Marine Ecology Progress Series, 214, 201-209.
Rabaut, M., Braeckman, U., Hendrickx, F., Vincx, M. & Degraer, S., 2008. Experimental beam-trawling in Lanice conchilega reefs: Impact on the associated fauna. Fisheries Research, 90 (1), 209-216.
Rabaut, M., Guilini, K., Van Hoey, G., Vincx, M. & Degraer, S., 2007. A bio-engineered soft-bottom environment: The impact of Lanice conchilega on the benthic species-specific densities and community structure. Estuarine, Coastal and Shelf Science, 75 (4), 525-536.
Rabaut, M., Van de Moortel, L., Vincx, M. & Degraer, S., 2010. Biogenic reefs as structuring factor in Pleuronectes platessa (Plaice) nursery. Journal of Sea Research, 64 (1), 102-106.
Rabaut, M., Vincx, M. & Degraer, S., 2009. Do Lanice conchilega (sandmason) aggregations classify as reefs? Quantifying habitat modifying effects. Helgoland Marine Research, 63 (1), 37-46.
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.
Rasmussen, E., 1973. Systematics and ecology of the Isefjord marine fauna (Denmark). Ophelia, 11, 1-507.
Rees, E.I.S., Nicholaidou, A. & Laskaridou, P., 1977. The effects of storms on the dynamics of shallow water benthic associations. In Proceedings of the 11th European Symposium on Marine Biology, Galway, Ireland, October 5-11, 1976. Biology of Benthic Organisms, (ed. B.F. Keegan, P. O'Ceidigh & P.J.S. Boaden), pp. 465-474.
Reise, K. & Lackschewitz, D., 1998. Benthos des Wattenmeeres zwischen Sylt und Rømø. Ökosystem Wattenmeer: Austausch-, Transport-und Stoffumwandlungsprozesse (Ch Gätje, K Reise, Hrsg), pp.55-62, Springer, Berlin.
Reise, K., 1979. Spatial configurations generated by motile benthic polychaetes. Helgoländer Wissenschaftliche Meeresuntersuchungen, 32, 55-72.
Reise, K., Bouma, T.J., Olenin, S. & Ysebaert, T., 2009. Coastal habitat engineers and the biodiversity in marine sediments. Helgoland Marine Research, 63 (1),1-2.
Remeikaitė-Nikienė, N., Lujanienė, G., Malejevas, V., Barisevičiūtė, R., Žilius, M., Garnaga-Budrė, G. & Stankevičius, A., 2016. Distribution and sources of organic matter in sediments of the south-eastern Baltic sea. Journal of Marine Systems, 157, 75-81.
Rhoads, D.C., 1974. Organism–sediment relations on the muddy seafloor. Oceanography and Marine Biology: An Annual Review, 12, 263-300.
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.
Richardson, C.A., Ibarrola, I. & Ingham, R.J., 1993b. Emergence pattern and spatial distribution of the common cockle Cerastoderma edule. Marine Ecology Progress Series, 99, 71-81.
Ropert, M. & Dauvin, J-C., 2000. Renewal and accumulation of a Lanice conchilega (Pallas) population in the baie des Veys, western Bay of seine. Oceanologica Acta, 23, 529-546.
Rosenberg, R., Hellman, B. & Johansson, B., 1991. Hypoxic tolerance of marine benthic fauna. Marine Ecology Progress Series, 79, 127-131.
Rossi, F., Forster, R., Montserrat, F., Ponti, M., Terlizzi, A., Ysebaert, T. & Middelburg, J., 2007. Human trampling as short-term disturbance on intertidal mudflats: effects on macrofauna biodiversity and population dynamics of bivalves. Marine Biology, 151 (6), 2077-2090.
Russell, P.J.C. & Petersen, G.H., 1973. The use of ecological data in the elucidation of some shallow water European Cardium species. Malacologia, 14, 223-232.
Rygg, B., 1970. Studies on Cerastoderma edule (L.) and Cerastoderma glaucum (Poiret). Sarsia, 43, 65-80.
Samuelson, G.M., 2001. Polychaetes as indicators of environmental disturbance on subarctic tidal flats, Iqaluit, Baffin Island, Nunavut Territory. Marine Pollution Bulletin, 42 (9), 733-741.
Sanders, H.L., Grassle, J.F., Hampson, G.R., Morse, L.S., Garner-Price, S. & Jones, C.C., 1980. Anatomy of an oil spill: long-term effects from the grounding of the barge Florida off West Falmouth, Massachusetts. Journal of Marine Research, 38 (2), 265-380.
Savari, A., Lockwood, A.P.M. & Sheader, M., 1991a. 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.
Savari, A., Lockwood, A.P.M. & Sheader, M., 1991b. Effects of season and size(age) on heavy metal concentration of the common cockle (Cerastoderma edule (L.)) from Southampton Water. Journal of Molluscan Studies, 57, 33-44.
Schöttler, U., 1982. An investigation on the anaerobic metabolism of Nephtys hombergii (Annelida: Polychaeta). Marine Biology, 71 (3), 265-269.
Schories, D. & Reise, K., 1993. Germination and anchorage of Enteromorpha spp. In sediments of the Wadden Sea. Helgolander Meeresuntersuchungen, 47, 275-285.
Schubert, A. & Reise, K., 1987. Predatory effects of Nephtys hombergii on other polychaetes in tidal flat sediments. Marine Ecology Progress Series, 34, 117-124.
Sebens, K.P., 1991. Habitat structure and community dynamics in marine benthic systems. In Habitat structure, (ed. S.S. Bell), pp. 211-234. Chapman & Hall.
Seed, R. & Brown, R.A., 1977. A comparison of the reproductive cycles of Modiolus modiolus (L.), Cerastoderma (=Cardium) edule (L.), and Mytilus edulis L. in Strangford Lough, Northern Ireland. Oecologia, 30, 173-188.
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.
Seitzinger, S.P., 1988. Denitrification in freshwater and coastal marine ecosystems: ecological and geochemical significance. Limnology and Oceanography, 33 (4part2), 702-724.
Sheehan, E.V., 2007. Ecological impact of the Carcinus maenas (L.) fishery'crab-tiling'on estuarine fauna. Ph.D. thesis, University of Plymouth.
Sinderman, C.J., 1990. Principle diseases of marine fish and shellfish, 2nd edition, Volume 2. Diseases of marine shellfish. Academic Press, 521 pp.
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.
Snelgrove, P.V., Butman, C.A. & Grassle, J.F., 1995. Potential flow artifacts associated with benthic experimental gear: deep-sea mudbox examples. Journal of Marine Research, 53 (5), 821-845.
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.
Spencer, B.E., Kaiser, M.J. & Edwards, D.B., 1998. Intertidal clam harvesting: benthic community change and recovery. Aquaculture Research, 29 (6), 429-437.
Stief, P., 2013. Stimulation of microbial nitrogen cycling in aquatic ecosystems by benthic macrofauna: mechanisms and environmental implications. Biogeosciences, 10 (12), 7829-7846.
Strasser, M. & Pielouth, U., 2001. Recolonization pattern of the polychaete Lanice conchilega on an intertidal sandflat following the severe winter of 1995/96. Helgoland Marine Research, 55, 176-181.
Strasser, M., Reinwald, T. & Reise, K., 2001. Differential effects of the severe winter of 1995/96 on the intertidal bivalves Mytilus edulis, Cerastoderma edule and Mya arenaria in the Northern Wadden Sea. Helgoland Marine Research, 55 (3), 190-197.
Stubbles, S., 1993. Recent benthic Foraminiferida as indicators of pollution in Restronguet Creek, Cornwall. Proceedings-USSHER Society, 8, 200-200.
Suchanek, T.H., 1993. Oil impacts on marine invertebrate populations and communities. American Zoologist, 33, 510-523.
Sylvand B., 1995. La baie des Veys(Littoral occidental de la baie de Seine, Manche), 1972-1993: structure et évolution à long terme d'un écosystème benthique intertidal de substrat meuble sous influence estuarienne. Thèse de Doct. d'Etat.: Université de Caen, 409 pp.
Tait, R.V. & Dipper, R.A., 1998. Elements of Marine Ecology. Reed Elsevier.
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.
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.
Tillin, H. & Tyler-Walters, H., 2014. Assessing the sensitivity of subtidal sedimentary habitats to pressures associated with marine activities. Phase 2 Report – Literature review and sensitivity assessments for ecological groups for circalittoral and offshore Level 5 biotopes. JNCC Report No. 512B, 260 pp. Available from: www.marlin.ac.uk/publications
Tillin, H.M. & Hull, S.C., 2013c. Tools for Appropriate Assessment of Fishing and Aquaculture Activities in Marine and Coastal Natura 2000 sites. Report III: Intertidal and Subtidal Muddy Sands and Sandy Muds. Report No. R.2071. Report by ABPmer for the Marine Institute (Galway).
Toupoint, N., Godet, L., Fournier, J., Retière, C. & Olivier, F., 2008. Does Manila clam cultivation affect habitats of the engineer species Lanice conchilega (Pallas, 1766)?. Marine Pollution Bulletin, 56 ( 8), 1429-1438.
Troost, K., 2010. Causes and effects of a highly successful marine invasion: case-study of the introduced Pacific oyster Crassostrea gigas in continental NW European estuaries. Journal of Sea Research, 64 (3), 145-165.
Troost, K., Stamhuis, E.J., van Duren, L.A. & Wolff, W.J., 2009. Feeding current characteristics of three morphologically different bivalve suspension feeders, Crassostrea gigas, Mytilus edulis and Cerastoderma edule, in relation to food competition. Marine Biology, 156 (3), 355-372.
UKTAG, 2014. UK Technical Advisory Group on the Water Framework Directive [online]. Available from: http://www.wfduk.org
Vader, W.J.M., 1964. A preliminary investigation in to the reactions of the infauna of the tidal flats to tidal fluctuations in water level. Netherlands Journal of Sea Research, 2, 189-222.
Van Hoey, G., Guilini, K., Rabaut, M., Vincx, M. & Degraer, S., 2008. Ecological implications of the presence of the tube-building polychaete Lanice conchilega on soft-bottom benthic ecosystems. Marine Biology, 154 (6), 1009-1019.
Verdelhos, T., Marques, J.C. & Anastácio, P., 2015. The impact of estuarine salinity changes on the bivalves Scrobicularia plana and Cerastoderma edule, illustrated by behavioral and mortality responses on a laboratory assay. Ecological Indicators, 52, 96-104.
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.
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.
Waldock, R., Rees, H.L., Matthiessen, P. & Pendle, M.A., 1999. Surveys of the benthic infauna of the Crouch Estuary (UK) in relation to TBT contamination. Journal of the Marine Biological Association of the United Kingdom, 79, 225 - 232.
Wang Y., Sheng, H.F., He, Y., Wu, J.Y., Jiang, Y.X., Tam, N.F. & Zhou, H.W., 2012. Comparison of the levels of bacterial diversity in freshwater, intertidal wetland, and marine sediments by using millions of illumina tags. Applied Environmental Microbiology, 78 (23), 8264-71.
Watkin, E.E., 1941. The yearly life cycle of the amphipod, Corophium volutator. The Journal of Animal Ecology, 10, 77-93.
Willems, W., Goethals, P., Van den Eynde, D., Van Hoey, G., Van Lancker, V., Verfaillie, E., Vincx, M. & Degraer, S., 2008. Where is the worm? Predictive modelling of the habitat preferences of the tube-building polychaete Lanice conchilega. Ecological Modelling, 212 (1–2), 74-79.
Wilson, J., 1984. Assessment of the effect of short term salinity changes on the acute oxygen consumption of Cerastoderma balthica and Tellina tenuis from Dublin Bay. Journal of Life Sciences Royal Dublin Society, 5 (1), 57-63.
Wilson, J.G. & Elkaim, B., 1997. Seasonal and geographical differences in oxygen consumption with temperature of Cerastoderma glaucum (Poiret) and a comparison with C. edule (L.). Estuarine, Coastal and Shelf Science, 45 (5), 571-577.
Wilson, J.G. & Elkain, B., 1991. Tolerances to high temperature of individual bivalves and the effect of geographic distribution, position on the shore and season. Journal of the Marine Biological Association of the United Kingdom, 71, 169-177.
Wilson, J.G., 1981. Temperature tolerance of circatidal bivalves in relation to their distribution. Journal of Thermal Biology, 6, 279-286.
Wilson, J.G., 1983. The uptake and accumulation of Ni by Cerastoderma edule and its effect on mortality, body condition and respiration rate. Marine Environmental Research, 8, 129-148.
Wilson, J.G., 1993. Climate change and the future for the cockle Cerastoderma edule in Dublin Bay - an exercise in prediction modelling. In Biogeography of Ireland: past, present, and future, (ed. M.J. Costello & K.S. Kelly), pp. 140-149. Occasional publication of the Irish Biogeographical Society, no. 2. Dublin: Irish Biogeographical Society Ltd.
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.
Wolff, W.J., 1971b. Changes in intertidal benthos communities after an increase in salinity. Thalassia Jugoslavica, 7, 429-434.
Wood, E.M., 1987. Subtidal Ecology. London: Edward Arnold.
Woodin, S.A., 1978. Refuges, disturbance and community structure: a marine soft bottom example. Ecology, 59, 274-284.
Yonow, N., 1989. Feeding observations on Acteon tornatilis (Linnaeus) (Opisthobranchia: Acteonidae). Journal of Molluscan Studies, 55, 97-102.
Zühlke, R., 2001. Polychaete tubes create ephemeral community patterns: Lanice conchilega (Pallas, 1766) associations studied over six years. Journal of Sea Research, 46, 261-272.
Zühlke, R., Blome, D., van Bernem, K.H. & Dittmann, S., 1998. Effects of the tube-building polychaete Lanice conchilega (Pallas) on benthic macrofauna and nematodes in an intertidal sandflat. Senckenbergiana Maritima, 29, 131-138.
Ziegelmeier, E., 1952. Beobachtungen über den Röhrenbauvon Lanice conchilega (Pallas) im Experiment und am natürlichen Standort. Helgolander Wissenschaftliche Meeresuntersuchungen, IV. 107-129.
This review can be cited as:
Last Updated: 01/06/2016