The Marine Life Information Network

Temperature increase (local)

Benchmark. A 5°C increase in temperature for one month, or 2°C for one year (Temperature change pressure definition).

Evidence

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

The geographic range of Hediste diversicolor (throughout north-west Europe on the Baltic Sea, North Sea and along Atlantic coasts to the Mediterranean) suggests that it is tolerant of a range of temperatures, and a temperature increase at benchmark levels is unlikely to have an adverse effect on UK populations. Hediste diversicolor can tolerate temperatures from below zero under Baltic ice to high summer temperatures in Black Sea lagoons (>25°C) (Smith, 1977) and in the Moroccan Oualidia lagoon (experiencing a temperature range of 14 to 27°C) (El Asri et al., 2018; Chaouti et al., 2019). In Gorgan Bay (Caspian Sea), Hediste diversicolor was amongst the most abundant species and maintained stable abundance across seasonal temperature variations from around 9.59 to 29.17°C (Zaferani et al., 2017). In the Ria de Averio coastal lagoon (Portugal), Hediste diversicolor inhabits the intertidal mudflats, which experience summer mean temperatures close to 23 to 31°C and a maximum habitat temperature of 35 to 37°C in 2019 to 2020 (Fernandes et al., 2023).

Temperature is a key factor influencing the growth of Hediste diversicolor, with growth increasing as temperature increases (Galasso et al., 2018; Aguado-Giménez et al., 2023; Malzahn et al., 2023; Villena-Rodriguez et al., 2025). Aguado-Giménez et al. (2023) reported that the optimal temperature for Hediste growth was 24.9°C, with growth declining at temperatures above this optimum and mortality tending to increase in exposure to temperatures above 22°C. Similarly, Villena-Rodriguez et al. (2025) found that the highest specific growth rates were recorded at temperatures more than 14.3°C. In a controlled laboratory experiment, Aguado-Giménez et al. (2025) found that Hediste diversicolor exhibited the highest feeding rates and growth rates at 22°C, while feeding efficiency decreased with increasing temperature. No mortality occurred during the study, in which individuals were maintained at temperatures of 14, 18 and 22°C for a total of 14 days (two feeding and fasting cycles) (Aguado-Giménez et al., 2025). Malzahn et al. (2023) reported an increased growth rate with increasing temperature, but survival decreased from 100% survival at 5.8°C and around 70% survival at 17.1°C.

Oxygen consumption in Hediste diversicolor also increased with temperature and body size, indicating an increase in metabolic rate (Galasso et al., 2018; Farrell et al., 2024). Seasonal changes in temperature can influence the bioturbation activity of Hediste diversicolor (Morelle et al., 2024), and Farrell et al. (2024) found that increasing temperatures from 15 to 20°C increases bioturbation rates. Temperature can also influence Hediste diversicolor burrowing depth, which is usually less than 40cm but can extend to 50 to 60 cm during cold winters (Gilbert et al., 2021).

Temperature change may adversely affect reproduction. Bartels-Hardege & Zeeck (1990) demonstrated that an increase from 12°C and maintenance of water temperature at 16°C induced reproduction in Hediste diversicolor specimens outside the normal period of spawning, and without a drop in temperature to simulate winter conditions the spawning period was prolonged and release of gametes was not synchronized. Poor synchronization of spawning could result in reduced recruitment, as gametes are wasted and mature specimens die shortly after gamete release. Wang et al. (2020) found that Hediste diversicolor embryonic development increased with a temperature increase from 6.1 to 21.2°C, but high temperatures (21.2 and 24.5°C) resulted in abnormal embryo growth and 100% embryo mortality occurred (at 21.2°C).

Madiera et al. (2021a) suggested that Hediste diversicolor was highly tolerant of long-lasting heatwaves and was able to increase its thermal tolerance in order to survive. Hediste diversicolor were not strongly affected by heat waves in an estuary in northwestern Portugal, where temperatures reached 40°C in intertidal pools (higher temperatures than experienced around UK and Irish coasts) (Dolbeth et al., 2011). Grilo et al. (2011) found that, at a Portuguese site, surface deposit feeders gradually decreased in periods of higher temperatures. However, sub-surface deposit feeders became dominant for up to three years after heat wave conditions had passed. This evidence was supported in laboratory experiments that exposed Hediste diversicolor to a combination of three temperatures (24°C control, 27°C warming conditions, 30°C heatwave conditions) and two salinities (20 and 30) for 28 days (Madiera et al., 2021b; Fernandes et al., 2021). Under heatwave conditions, Hediste diversicolor showed signs of physiological stress and shifts in fatty acid content, but extreme stress levels were not reached, growth was maintained, and survival was not significantly affected by temperature (Madiera et al., 2021b; Fernandes et al., 2021). Fernandes et al. (2023) found that after one month of acclimation to higher heatwave temperatures (a +6 °C increase) Hediste diversicolor individuals could shift its Critical Thermal Maximum (CTMax) or upper thermal limits by 0.81°C.

Madiera et al. (2021b) and Fernandes et al. (2021; 2023) reported Hediste diversicolor individuals exposed to a salinity of 20 for 28 days were more sensitive to increases in temperature and heatwave conditions than individuals exposed to a salinity of 30. Hediste diversicolor can acclimate to increased temperatures when in optimal salinity conditions. It was concluded that temperature effects are dependent on salinity and duration of exposure (Madiera et al., 2021b; Fernandes et al., 2021).

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

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

On the other hand, some evidence has suggested that Macoma balthica might be sensitive and have a low tolerance to temperature increase (Beukema et al., 2009; Garcia et al., 2016). Despite apparent adaptation to regional temperature ranges, increased temperature reduced growth rates and weight gain in populations of Macoma balthica in the Baltic Sea (Beukema et al., 2009; Barda et al., 2014; Beukema et al., 2017a; Beukema & Dekker, 2020). Beukema et al. (2014) also warn that increasing water temperatures as a result of global warming are likely to shorten the growing season (typically late winter to early spring) if warmer spring and summer water temperatures are experienced.

In the Baltic Sea, Macoma balthica annual mortality rates were significantly higher at high temperatures in summer and winter (Beukema et al., 2009; Beukema et al., 2017b;c; Beukema & Dekker, 2020). Jansen et al. (2007) suggest that temperature increases in the Spanish coast along the Bay of Biscay over the past 40 years caused loss of Macoma balthica populations, due to short-term but frequent exposure to >30° C in the Spanish estuaries, which induced elevated maintenance rates in Macoma balthica, and ultimately starvation.

Repeated recruitment failure also occurred after mild winters in a comparable North Sea location, probably due to enhanced survival of predators, particularly shrimp and shore crabs feeding on post-larvae Macoma balthica (Beukema, 1992; Beukema et al., 2001; Schueckel & Kroencke, 2013; Beukema et al., 2017b). In the Baltic Sea, recruitment was more successful after cold winters than warm (Beukema et al., 2009; Beukema et al., 2017c; Beukema & Dekker, 2020). As a result, Jansen et al. (2007) predicted that the southern limit of the species would progressively shift north if temperatures continued to rise. In long-term monitoring (44 years) of Macoma balthica populations in the Wadden Sea, Beukema et al. (2017b;c) showed that warming alone could not explain population decline because, although recruitment and adult survival declined in warm winters, a sharp population decline in the mid-1990s and rapid spread of mortality between sites was consistent with a disease-like outbreak.

Examining two models, Ehrnsten et al. (2019a) found that increasing bottom-water temperatures reduced Macoma balthica (reported as Limecola balthica) biomass by increasing metabolic costs and increasing degradation of the food available. Therefore, warming weakened the otherwise positive biomass response to increased organic matter sedimentation.

Staniek et al. (2025) found 15-day exposure to short-term heatwave conditions, +2.8° C above ambient temperatures (22.2° C) in a mild heatwave treatment and +4.4° C above ambient temperatures in a strong heatwave treatment, did not affect the survival (97 to 99%) or size of Macoma balthica, and an increase in total biomass was observed. However, the strong heatwave treatment did cause a significant decrease in the chlorophyll and respiration rates in Macoma balthica, suggesting a potential long term negative effect of heatwaves.

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

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

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

Consistent with the above evidence, Garcia et al. (2016) reviewed the temperature sensitivity of the main intertidal benthic taxa in the Severn Estuary (UK), and concluded that Hediste diversicolor is unlikely to be sensitive to temperature, whereas Macoma balthica was the only taxa likely to be affected.

Sensitivity assessment. Typical surface water temperatures around the UK coast vary, seasonally from 4 to 19°C (Huthnance, 2010). It is likely that the important characteristic species are able to resist a long-term increase in temperature of 2°C and may resist a short-term increase of 5°C. Macoma balthica may retreat north as a result of long-term warming and climate change. However, the important characterizing species Hediste diversicolor are likely to survive a 5°C increase in temp for one month period, or 2°C for one year, although reproductive activities may be impacted. For instance, without colder winters spawning may not be synchronised and so recruitment would be reduced. Similarly, warmer winters might result in recruitment failure in Macoma balthica, although longer-term studies could not identify the effects of climate warming.  Therefore, a resistance of ‘Medium’ is suggested to represent local changes in population due to recruitment failure, especially in Macoma.  Therefore, resilience is assessed as  ‘High’ and sensitivity is assessed as  ‘Low’ . However, confidence in the assessment is ‘Low’ due to variation in effects provided in the evidence.

Temperature decrease (local)

Benchmark. A 5°C decrease in temperature for one month, or 2°C for one year (Temperature change pressure definition).

Evidence

A decrease in temperature has been shown to be beneficial to Hediste diversicolor as case studies report a reduction in numbers of the species’ predators. For instance, a severe winter in the Wadden Sea in 1995/1996 saw an increased abundance of Hediste diversicolor coincident with a reduction in the numbers of Carcinus maenus and Crangon crangon (Armonies et al., 2001). A similar increase in abundance was noted in the same area between 1978 and 1987 after a series of cold winters (mean Hediste diversicolor density increased from 24/m² to 151/m² respectively) (Beukema, 1990). Similarly, El Asri et al. (2018) found that Hediste diversicolor remained the dominant and most abundant polychaete species in the Oualidia Lagoon during the winter (16.9 and 19.9°C).

Seasonal changes in temperature can influence the bioturbation activity of Hediste diversicolor (Morelle et al., 2024), and Farrell et al. (2024) found that increasing temperatures from 15 to 20°C increases bioturbation rates. Temperature can also influence Hediste diversicolor burrowing depth, which are usually less than 40 cm but can extend to 50 to 60 cm during cold winters (Gilbert et al., 2021).

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

Colder winter temperatures have been shown to benefit Macoma balthica population dynamics. Recruitment success increased following colder winters and repeated recruitment failure has occurred after mild winters in comparable North Sea location (Beukema, 1992, Schueckel & Kroencke, 2013; Beukema et al., 2001). In Friedrichskoog, Germany, König (1943) found a high accumulation of dead Cerastoderma edule biomass after a severe winter 1936/1937 but high numbers of Macoma balthica (80,000 individuals/m²) spat in following years (winter 1939). Winter water surface temperatures in the Wadden Sea (Netherlands) have increased 1.5°C since the 1980s (Oost et al., 2009). During milder winters greater body weight loss and production of fewer and smaller eggs has been observed in Macoma balthica (van der Meer et al., 2003). However,  reduced recruitment success during milder winters may also be due to increased predation as juvenile Crangon crangon have shown increased abundance in relation to milder winters (Beukema, 1992; Beukema et al., 2001; Beukema & Dekker, 2005; Schueckel & Kroencke, 2013; Beukema et al., 2017b). In the Baltic Sea, recruitment was more successful after cold winters than warm (Beukema et al., 2009; Beukema et al., 2017c; Beukema & Dekker, 2020).

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

Eteone longa is recorded as arctic-boreal, both from the Atlantic and the Pacific and occurs throughout the North Sea. Occurring in the arctic-boreal region suggests the species is resistant to a decrease in temperature in UK and Irish seas. Most littoral oligochaetes, including tubificids and enchytraeids, can survive freezing temperatures and can survive in frozen sediments (Giere & Pfannkuche, 1982). Tubificoides benedii (studied as Peloscolex benedeni) recovered after being frozen for several tides in a mudflat (Linke, 1939).

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

Salinity increase (local)

Benchmark. An increase in one MNCR salinity category above the usual range of the biotope or habitat (Salinity regime change pressure definition).

Evidence

This biotope occurs in variable (18 to 35 ppt), reduced (18 to 30 ppt) and full salinity (30 to 35 ppt) (JNCC, 2015). For example, this biotope largely dominates intertidal assemblages in Poole Harbour, UK, in reduced salinity, with a mean median salinity from 27.3 to 27.4 ppt (Clarke et al., 2018). Biotopes occurring in variable and reduced salinity are not considered sensitive to a change to full salinity as this falls within the natural habitat range. However, species may be sensitive to an increase in the salinity regime to hypersaline (>40 ppt). Ruso et al. (2007) examined the effects of brine discharge on soft-bottom communities and found that at sites close to the discharge where salinity exceeded 39 psu, assemblages dominated by Polychaeta, Crustacea and Mollusca were replaced by communities characterized by nematodes. Evidence suggests that increased salinity can alter assemblage structure and reduce polychaete abundance, richness and diversity up to 400 m from the discharge site, although sensitivity to brine discharge differs among polychaete families (Ruso et al., 2007; 2008; Roberts et al., 2010). Little evidence was found for tolerances of elevated salinity, although populations are likely to be acclimated to short-term increased salinities in surface sediment layers, as evaporation on hot days, or wind-driven desiccation leads to increased salinities.

Hediste diversicolor, the key characterizing species, occurs across all variant sub-biotopes and from reduced to full salinity. Hediste diversicolor is a euryhaline species (El Asri et al., 2018; Krupa & Grishaeva, 2019), typical of brackish waters (Chaouti et al., 2019; Scaps, 2002 cited in Bettoso et al., 2024), but able to tolerate a range of salinities from fully marine seawater (40.0 psu) down to 5 psu or less (Barnes, 1994; Murray et al., 2017; Krupa & Grishaeva, 2019).

In the Lagoon of Marano and Grado (Adriatic Sea), rising salinities between 1993 to 1995 and 2008 to 2021 was associated with a decrease frequency of Hediste diversicolor in euhaline (salinity of 30 to 40) and polyhaline (salinity of 20 to 30) water bodies, with individuals occurring mainly in mesohaline (salinity of 5 to 20) areas in the Lagoon (Bettoso et al., 2024). A muddy assemblage dominated by Macoma balthica and Hediste diversicolor occurs in the polyhaline Lateral Bank and the polyhaline Great Mudflat of the Seine estuary (Lécuyer et al., 2024).

Evidence has shown that Hediste diversicolor growth was significantly affected by salinity and temperature, Villena-Rodriguez et al. (2025) found the highest specific growth rates were recorded at salinities more than 35 psu.

Seitz (2011) found that the distribution of Macoma balthica across a salinity gradient between a minimum and maximum of 8.8 psu to 19 psu in Chesapeake Bay was not influenced by salinity. Instead, resource availability was the principal influence on Macoma balthica. McLusky & Allan (1976) reported that Macoma balthica failed to grow at 41 psu. It is likely that Macoma balthica would be tolerant of an increase in salinity category to fully marine but further increases to >40‰ are likely to affect growth and condition.

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

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

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

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

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

Salinity decrease (local)

Benchmark. A decrease in one MNCR salinity category above the usual range of the biotope or habitat (Salinity regime change pressure definition detail).

Evidence

This biotope occurs in variable (18 to 35 ppt), reduced (18 to 30 ppt) and full salinity (30 to 35 ppt) (JNCC, 2015). For example, this biotope largely dominates intertidal assemblages in Poole Harbour, UK, in reduced salinity, with a mean median salinity from 27.3 to 27.4 ppt (Clarke et al., 2018). Therefore, the decrease in salinity assessed at the benchmark is from reduced to  low salinity (<18 ppt). The available evidence (summarised below) suggests that the characterizing species are tolerant of a short-term decrease to low salinity but it is likely that for species such as Macoma balthica long-term reductions would lead to mortality.

Hediste diversicolor, the key characterizing species, occurs across all variant sub-biotopes and as such is resistant to the salinity range, from reduced to full. Hediste diversicolor is a euryhaline species (El Asri et al., 2018; Krupa & Grishaeva, 2019), typical of brackish waters (Chaouti et al., 2019; Scaps, 2002 cited in Bettoso et al., 2024), but able to tolerate a range of salinities from fully marine seawater (40.0 psu) down to 5 psu or less (Barnes, 1994; Murray et al., 2017; Krupa & Grishaeva, 2019).

Field observations provide evidence of Hediste diversicolor tolerance to a range of salinities. For example, in Jade Bay, Wadden Sea Hediste diversicolor is a characteristic species in a community near the tidal gates, which can experience salinities as low as 18 (Schückel et al., 2015). Krupa & Grishaeva (2019) reported Hediste diversicolor as one of the dominant macrozoobenthos, distributed across a wide salinity gradient in the Small Aral Sea, with high biomass of the polychaete recorded at low (1.0 to 1.5 psu) to increased salinity (15.0 to 20.0 psu). The salinity limits and optimal salinity range for Hediste diversicolor in the Aral Sea were reported as 1.0 to 35.3 psu and 1.0 to 27.0 psu, respectively (Krupa & Grishaeva, 2019).

In the Lagoon of Marano and Grado (Adriatic Sea), rising salinities between 1993 to 1995 and 2008 to 2021 were associated with a decrease frequency of Hediste diversicolor in euhaline (salinity of 30 to 40) and polyhaline (salinity of 20 to 30) water bodies, with individuals occurring mainly in mesohaline (salinity of 5 to 20) areas in the Lagoon (Bettoso et al., 2024).

In experiments mimicking heatwave conditions, Madiera et al. (2021b) and Fernandes et al. (2021; 2023) reported Hediste diversicolor individuals exposed to a salinity of 20 for 28 days were more sensitive to increases in temperature and heatwave conditions than individuals exposed to a salinity of 30. Hediste diversicolor can acclimate to increased temperatures when in optimal salinity conditions. It was concluded that temperature effects are dependent on salinity and duration of exposure (Madiera et al., 2021b; Fernandes et al., 2021).

Evidence has shown that Hediste diversicolor growth was significantly affected by salinity and temperature, Villena-Rodriguez et al. (2025) found the highest specific growth rates were recorded at salinities more than 35 psu.

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

Macoma balthica is found in brackish and fully saline waters (Clay, 1967b) so may tolerate salinity fluctuations, including decreases in salinity. For example, it can acclimatise in extremely brackish conditions in the Baltic Sea and is able to grow in oligohaline water between 0.5 to 3.0 ppt (Jansen et al., 2009 summary only; Pourmozaffar et al., 2020). Gebruk et al. (2023) found Macoma balthica dominated the macrozoobenthic biomass at most studied stations along a salinity gradient in Pechora Bay, where salinity increased from estuarine to near euhaline conditions (average around 2.6 to 26.3 psu surface salinity and average around 10.7 to 29.4 psu near bottom salinity). In tidal channels in Jade Bay, Wadden Sea Macoma balthica is recorded as a characteristic species in subtidal communities exposed to a constant salinity of 29 to 30, but also in communities near the tidal gates which can experience low salinity of 18 (Schückel et al., 2015).

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

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

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

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

Water flow (tidal current) changes (local)

Benchmark. A change in peak mean spring bed flow velocity of between 0.1 m/s and 0.2 m/s for more than one year (Water flow pressure definition). 

Evidence

Hediste diversicolor characteristically inhabits littoral mudflats predominantly of clay (particles < 4 µm), silt (4 to 63 µm) and to a lesser extent, very fine sand (63 to 125 µm) (Jones et al., 2000). Highest abundances occur in very weak (negligible) to weak < 1 knot (<0.5 m/sec.) currents. These conditions are provided by this biotope, which mud or sandy mud from the upper to lower shore. For example, in tidal channels in Jade Bay (Wadden Sea), Hediste diversicolor was recorded as a characteristic species of the tidal gates community, which is associated with a lower tidal current velocity (mean 0.09 m/s and max 0.96 m/s) than the other tidal channels (Schückel et al., 2015). In Pool Harbour, UK, Hediste diversicolor is found exposed to mean maximum velocities from 0.13 to 0.25 m/s (Clarke et al., 2018). In the Mondego estuary (Portugal), Hediste diversicolor dominated communities in the South arm which is characterized by weaker hydrodynamic conditions (Van Der Linden et al., 2016).Brun et al. (2021) reported Hediste diversicolor dominated intertidal muddy habitats in Cadiz Bay (Spain), characterized by high water flow velocities (up to 75 cm/s or 0.75 m/s), where seagrass canopies altered local hydrodynamics and food availability.

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

Coarser sediments are likely to remain in areas of strongest flow velocity (where finer particles have been re-suspended). Species such as Pygospio elegans and other opportunist polychaetes that tolerate coarser particle size will possibly increase in abundance.

Macoma balthica is likely to experience greater impact from increased water flow as the species thrives in low energy environments, such as the extremely sheltered areas that characterize the biotope (Tebble, 1976). Increased water flow rate is likely to influence the sediment characteristics in this biotope, primarily by re-suspending and preventing deposition of finer particles (Hiscock, 1983). This is likely to result in erosion of the preferred habitat, which may cause mortality of some portion of the population of Macoma balthica. Higher current velocity (18 cm/s or 0.18 m/s) recorded in flume experiments conducted in the Isle of Sylt (North Sea) caused juvenile Macoma balthica to be washed out of the sediment (Zuhlke & Reise, 1994). Green (1968) reported that, towards the mouth of an estuary where sediments became coarser and cleaner, Macoma balthica was replaced by another tellin species, Tellina tenuis. However, field studies have shown that Macoma balthica can be more abundant in macrofauna communities adapted to unstable sediments (Schückel et al., 2015). In Jade Bay (Wadden Sea), Macoma balthica had a higher mean abundance in a community associated with higher current velocity (mean 0.39 m/s and max 1.34 m/s) than in a community associated with lower tidal current velocity (mean 0.09 m/s and max 0.96 m/s) (Schückel et al., 2015). It is important to note that the mean abundances per site were low in both communities (4 and 2 individuals per site, respectively) (Schückel et al., 2015).

Sensitivity assessment. This biotope is characterized by mud and sandy mud sediments, typical of low energy (limited tidal flow, sheltered from wave action). Therefore, an increase in water flow by 0.1 to 0.2 m/s (the benchmark) may erode fine sediments, depending on local conditions. Hence, the sediment may transition to muddy sand. While the characteristic species occur in a wider range water flow rates, the change in sediment type may result in subtle changes in the community and hence reclassification of the biotope, for example, to HedMx. Also, Macoma balthica abundance may be reduced if juveniles are washed from the substratum due to increased water flow. A decrease in water flow is unlikely to be significant. Therefore, an increase in flow velocity may alter the muddy sand sediments at the benchmark level, and resistance has been assessed as ‘Low’. Resilience is assessed as ‘High’ and sensitivity is, therefore ‘Low’.

Emergence regime changes

Benchmark.  1) A change in the time covered or not covered by the sea for a period of ≥1 year, or 2) an increase in relative sea level or decrease in high water level for ≥1 year. (Emergence regime change pressure definition).

Evidence

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

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

Macoma balthica occurs in the upper regions of the intertidal (Tebble, 1976) and is, therefore, likely to be tolerant of prolonged emergence. Philippe et al. (2016) found that Macoma balthica biomass and abundance increased with increasing emersion time, with the highest biomass and abundance found in sites experiencing the longest emersion time.

It is a bivalve and can close tightly by contraction of the adductor muscle, storing moisture inside the shell. The silty sediments in which the species lives have a high-water content and are therefore resistant to desiccation. Furthermore, Macoma balthica is mobile and able to relocate in the intertidal by burrowing (Bonsdorff, 1984) or floating (Sörlin, 1988). It would be expected to react to an increase in emergence by migrating down the shore to its preferred position. There may be an energetic cost to this migration but it is not expected that mortality would result. Macoma balthica should quickly recover from the energetic cost of relocation. 

Macoma balthica also occurs in the intertidal and sublittoral zones down to depths of 190 m (Olafsson, 1986), although it is more abundant intertidally.  Hence, it would be expected to resist a decrease in emergence regime. However, a case study, predicting changes in biomass of Macoma balthica in the Humber estuary, UK (western North Sea) under expected sea level rise conditions displayed negative impacts. Coastal squeeze from sea level rise would produce steeper and more homogenous beach face profiles. Macoma balthica was predicted to be lower on steeper beach faces and the biomass of Macoma balthica was predicted to decrease (Fujii & Raffaelli, 2008). 

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

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

Wave exposure changes (local)

Benchmark. A change in near shore significant wave height of >3% but <5% for more than one year (Wave action pressure definition). 

Evidence

This biotope occurs in wave sheltered areas, where estimated wave categories range from sheltered, very sheltered and extremely sheltered (JNCC, 2015, 2022). The key characterizing species Hediste diversicolor is infaunal, inhabiting a burrow in which it seeks refuge from predators and may partially emerge to seek and capture food. An alteration of factors within the environment that increases wave exposure could cause erosion of the substrata and consequently, loss of habitat.

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

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

Transition elements & organo-metal contamination

Benchmark. Exposure of marine species or habitat to one or more relevant Transitional metal or organometal (e.g. TBT) contaminants via uncontrolled releases or incidental spills (Transitional metals and organometals pressure definition). 

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Transitional metal or organometal' contaminants on selected polychaete, mollusc, and amphipod species are summarized below. The full 'evidence reviews' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Transitional metal or organometal' contaminant examined, together with an overall pressure assessment. 

Transitional metals. Hediste spp. and Nereis spp. (often as Hediste (Nereis) diversicolor) were the most studied genera in the 'evidence review' and contributed 73.5% of the results on the effects of ‘transitional metals and organometals’ in the evidence reviewed. The effect of ‘metal' exposure varied between studies depending on the metal and its chemical form, and the environmental conditions used in the study. In general, heavy metal toxicity increases for estuarine animals as salinity decreases and temperature increases (McLusky et al., 1986). Nevertheless, exposure to ‘metal’ contaminants was reported to result in ‘Severe’ mortality in 8% of the results from Hediste (Nereis) diversicolor and Nereis virens. ‘Significant’ mortality was reported in 48.8% of the results, but sublethal effects were reported in 26.7% of the results. Overall, the most toxic metals (in terms of the reported mortality) were copper, zinc, cadmium, lead, mercury, and silver. Exposure to vanadium was also reported to result in ‘significant’ mortality, but in a single study.

In Hediste diversicolor, the acute toxicity is dependent on the rate of uptake of the metal since this determines the speed with which the lethal dose is built up. The rate of intake is important because this determines whether the organism's detoxification mechanisms can regulate internal concentrations. The resistance of Hediste diversicolor is thought to be dependent on a complexing system which detoxifies the metal and stores it in the epidermis and nephridia (Bryan & Hummerstone, 1971; McLusky et al., 1986). Hediste diversicolor has been found successfully living in estuarine sediments contaminated with copper ranging from 20 µM Cu/g in low copper areas to >4000 µM Cu/g where mining pollution is encountered, e.g. Restronguet Creek, Fal Estuary, Cornwall (Bryan & Hummerstone, 1971). Attempts to change the tolerance of different populations of Hediste diversicolor to different sediment concentrations of copper have shown that it is not readily achieved, which suggests that increased tolerance to copper has a genetic basis (Bryan & Hummerstone, 1971, 1973; Bryan & Gibbs, 1983). Since juveniles remain in the infauna throughout their development, selection for metal tolerance can be expected to be operative from an early stage (Bryan & Gibbs, 1983).

Overall, the evidence suggests the resistance of Hediste spp. and Nereis spp. ranges from ‘None’ to ‘Low’ for most of the ‘transitional metals’ examined, except silver. Therefore, as resilience is probably ‘Medium’, the ‘worst-case’ sensitivity is assessed as ‘Medium’. However, the toxicity of metals varies with the environmental conditions, and local populations can adapt to long-term contamination.

‘Severe’ or ‘significant’ mortality was reported in 52% of the results from studies in the evidence review of the effects of ‘Transitional metals and organometal’ exposure on Macoma spp. depending on the exposure concentration or duration. Copper and cadmium were reported to result in ‘Severe’ mortality. Arsenic, chromium, mercury, silver, zinc, and nickel were reported to result in ‘significant’ mortality, depending on concentration, duration, and environmental conditions. The remaining metals were reported to result in no mortality or sublethal effects. Barite (in the form of drilling mud barite) was shown to cause 100% mortality of Macoma balthica within 12 days at a depth of 2- and 3-mm dosage, but may have been due to physical damage to their gill filaments rather than chemical toxicity (Barlow & Kingston, 2001). Overall, the evidence suggests that the worst-case resistance of Macoma spp. to ‘transitional metals’ exposure is ‘None’. Therefore, resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’.

‘Severe’ or ‘significant’ mortality was reported in 45% of the results from studies in the evidence review of the effects of ‘Transitional metals and organometal’ exposure on Scrobicularia plana depending on the exposure concentration or duration. Copper was reported to result in ‘Severe’ mortality, while cadmium, lead, silver, and zinc were reported to result in ‘significant’ mortality, depending on concentration, duration, and environmental conditions. Laboratory tests in clean water can be misleading as these do not reflect lowered toxicity in the marine environment due to the buffering effects of carbon and sulphide, which render copper non-labile (not bioavailable) and the influence of water pH, hardness, temperature and salinity, etc. Field surveys have found that Scrobicularia plana is present in the highly contaminated Fal Estuary, where levels of copper and zinc are high (Bryan & Gibbs, 1983).  Nevertheless, the evidence suggests that the worst-case resistance of Scrobicularia spp. to ‘transitional metals’ exposure is ‘None’. Therefore, resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’.

Organometals. The evidence on the effects of tributyltin was limited to two papers, one of which was not accessible (Walsh et al., 1984). Beaumont et al. (1989) introduced TBT into three microcosms at high (1-3 µg/l) concentrations and three microcosms at low (0.06-0.17 µg/l) concentrations. Non-introduced Nereis diversicolor, Arenicola marina, and Eteone sp. occurred in the low-level TBT and the control treatments, but not in the high-level TBT treatments. High mortalities of Nereis diversicolor were recorded in all microcosms, including the control, so the results were inconclusive. Overall, only sublethal effects were reported in the evidence reviewed. Therefore, the sensitivity of Nereis spp. to TBT is assessed as ‘Not sensitive’ but with ‘Low’ confidence due to the lack of evidence.

Beaumont et al. (1989) also examined the effects of tributyltin (TBT) exposure in Macoma balthica. However, no mortality was reported, and some juvenile Macoma recruited to the low (0.06-0.17 µg/l) TBT treatment mesocosm. Therefore, the worst-case sensitivity of Macoma spp. to TBT exposure is assessed as ‘Not sensitive’, but confidence in the assessment is ‘Low’ due to the limited number of studies reviewed.

The effects of tributyltin (TBT) on Scrobicularia plana were investigated by six articles (Beaumont et al., 1989; Ruiz et al., 1994a&b, 1995 a,b&c) and reviewed by Langston (2020). Beaumont et al. (1989) reported that Scrobicularia plana mortalities increased with time in the high TBT treatments (1-3 µg/l), with 100% mortality after 10 weeks of exposure. Ruiz et al. (1994) reported an LC50 of Scrobicularia plana spat of <1.3 µg/l TBT and that TBT reduced their growth rate significantly at 50, 125, 250 and 500 ng/l TBT. Ruiz et al. (1995) reported that exposure to TBT levels of >125 ng/l resulted in significant mortalities (>50%) of Scrobicualria plana pediveligers and negligible shell growth of individuals. Larval shell growth was reduced significantly and was abnormal at the lowest concentration tested (50 ng/l TBT). Ruiz et al. (1995b) reported that exposure to TBT caused abnormal larval development at 9%, 51% and 85% in the 125, 250, and 500 ng/l treatments, respectively. The mean length of the D-larvae produced in the 250 and 500 ng/l treatments was reduced compared to D-larvae in the controls and at the lowest TBT treatment. 

Langston (2020) noted that Scrobicularia plana numbers declined in heavily TBT contaminated areas during the 1980s when TBT concentrations peaked, e.g., the Solent area of the English Channel. Recruitment patterns and abundances recovered slowly over 25 years after the introduction of TBT regulatory measures. In the Solent area, the sex ratios of Scrobicularia plana were skewed towards males (2:1) where TBT sediment concentrations were highest. Langston (2020) noted that such masculinization of the clams by TBT probably reduced larval production and adversely affected reproduction in Scrobicularia plana. Therefore, the worst-case resistance of Scrobicularia plana to TBT exposure is assessed as ‘None’. Hence, resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’.

Nanoparticulate metals. Sublethal effects and ‘no’ mortality were reported in all but one of the studies reviewed that examined the effects of nanoparticulate metals on polychaetes. The nanoparticulates did result in reduced burrowing speed or activity, where an effect was reported. While a reduction in burrowing ability may reduce feeding or increase the susceptibility to predation, no direct mortality was reported. However, Cozzari et al. (2015) reported that silver nanoparticles caused 8% and 16% mortalities in the 2.5 and 10 µg/g treatments after 4 days, respectively, and the 5 µg/g treatments caused 12% mortality by day seven. Therefore, the worst-case resistance of Nereis spp. to nanoparticulate metals is assessed as ‘Medium’, resilience as ‘High’ and sensitivity as ‘Low’ but with ‘Low’ confidence due to the lack of evidence.

Dai et al. (2013) investigated the effects of silver, silver oxide nanoparticulates, copper, and copper oxide nanoparticulates on Macoma balthica. Clams were exposed to sediment spiked with 200 µg/g of silver or copper for 35 days. No significant effects on mortality, condition index, or burrowing behaviour were observed for any of the metal forms. Therefore, the sensitivity of Macoma spp. to nanoparticulate copper or silver is assessed as ‘Not sensitive’, but confidence in the assessment is ‘Low’ due to the limited number of studies reviewed. 

Nine articles in the evidence review examined the effects of nanoparticulate metals on Scrobicularia plana. Eight of the articles reported sublethal effects, e.g., reduced burrowing activity, because of nanoparticulate metal exposure (Buffet et al., 2011, 2012, 2013a&b, 2014; Pan et al., 2015; Bertrand et al., 2016). Eight of the articles reported sublethal effects, e.g., reduced burrowing activity, because of nanoparticulate metal exposure (Buffet et al., 2011, 2012, 2013a&b, 2014; Pan et al., 2015; Bertrand et al., 2016). However, Scola et al. (2021) reported that exposure to 500 µg/g copper nanoparticulates resulted in 100% mortality within 30 days of exposure. Therefore, the worst-case resistance of Scrobicularia plana to nanoparticulate metals is assessed as ‘None’ but with ‘Low’ confidence due to limited evidence of mortality. Hence, resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’.

The evidence for the effects of 'transitional metals and organometals' on the Nephtys spp., Eteone spp. and Tubificoides spp. was limited to single observations from four studies. In all cases, only sublethal results were reported. Therefore, the sensitivity of Nephtys spp., Eteone spp. and Tubificoides spp. to ‘Transitional metals and organometals’ is assessed as ‘Not sensitive’ but with ‘Low’ confidence due to the lack of evidence.

Overall sensitivity assessment of 'Transitional metals and organometals'. Hediste diversicolor and Macoma balthica are the important characteristic species that define this biotope. The evidence collated suggests that both Hediste spp. and Macoma spp. have been reported to experience 'severe' or 'significant' mortality due to exposure to a range of transitional metals, depending on the concentration of the metals, duration of exposure, and local conditions such as temperature and salinity. Hediste diversicolor and molluscs are reported to be able to detoxify metal contamination within their tissue, and Hediste diversicolor can also genetically adapt to local metal contamination. Therefore, the evidence suggests that the worst-case resistance of this biotope to ‘transitional metals’ exposure is ‘None’, so resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’. However, confidence in the assessment is 'Medium' due to the variation in response between the metals studied, changes in toxicity due to local conditions, and the potential for localised adaptation. Where present, Scrobicularia plana populations may be adversely affected by TBT or other endocrine disruptors.  The reported effects of nanoparticulate metals on the resident polychaete and bivalve populations suggest that exposure could result in population decline, either due to direct mortality (as in Scrobicularia) or a reduction in feeding and burrowing activity and the potential increase in predation. However, the evidence is limited. 

Hydrocarbon & PAH contamination

Benchmark. Exposure of marine species or habitat to one or more relevant hydrocarbon or polyaromatic hydrocarbon (PAH) contaminants via uncontrolled releases or incidental spills (Hydrocarbon & PAH pressure definition).

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Hydrocarbon and PAH' contaminants on selected polychaete and mollusc species are summarized below. The full 'evidence reviews' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Hydrocarbon and PAH' contaminant examined, together with an overall pressure assessment. 

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

The Amoco Cadiz oil spill resulted in reduced abundance, biomass, and production of the affected 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).

Hydrocarbons and PAHs. Lewis et al. (2008) reported that both the WAF of crude oil and the PAH fluoranthene adversely affected fertilization success in Neries virens, while the WAF caused 'significant' larval mortality. No mortality due to PAH exposure was reported in the evidence reviewed. McLusky & Martins (1998) investigated the long-term effects of petrochemical discharge on the faunal composition of an estuarine mudflat over a 20-year period. The abundance of Neries diversicolor varied over the 20 years, but no significant trends were recorded.  Therefore, the resistance of Nereis spp. (and Hediste spp.) to petrochemical hydrocarbons and PAHs is assessed as ‘Low’ based on the worst-case results reviewed. Hence, resilience is ‘High’ and sensitivity is assessed as ‘Low’ on the assumption that fertilization failure could result in long-term population decline, but with ‘Low’ confidence due to the limited evidence reviewed. 

Stekoll et al. (1980) exposed the Macoma balthica, to Prudhoe Bay crude oil in flowing seawater for six months at three concentrations; low 0.03 mg/l, medium 0.3 mg/l and high 3.0 mg/l and concluded that chronic exposure of Macoma balthica to oil-in-seawater concentrations even as low as 0.03 mg/l would in time lead to population decreases. The individuals in this study were not subjected to any of the stresses that normally occur in their natural environment on mudflats, such as changes in salinity, temperature, oxygen availability and wave action, therefore, exposure of Macoma balthica to oil under field conditions may result in higher mortality. Shaw et al. (1976) also reported mortality of Macoma balthica caused by exposure to crude oil following an experimental application of oil at a concentration of 1.2 µl oil/cm² and 5.0 µl oil/cm² to sediments, which equated to oil spills of one ton /20 km² and one ton/100 km². 'Significant' mortalities were observed after only two days following the application of the oil at a concentration of 5.0 µl oil/cm². Some specimens of Macoma balthica survived the application of oil in these experiments but were weakened (Shaw et al., 1976). Therefore, the worst-case resistance of Macoma spp. to petroleum hydrocarbons is assessed as ‘Low’, resilience as ‘High’ and sensitivity as ‘Low’ but with ‘Low’ confidence due to the limited number of studies reviewed.

Only one article (Farke & Gunther, 1984) examined the effects of dispersants on Macoma balthica. No mortality was reported, and it was unclear what effects the dispersant had on the population based on the data alone. Therefore, the evidence was not adequate to support an assessment of its sensitivity to dispersants.  No results on the impacts of ‘Hydrocarbons and PAHs’ on Scrobicularia plana were found in the evidence reviewed. 

Overall sensitivity assessment of this pressure. The evidence reviewed suggests that the important characteristic species Hediste diversicolor (especially larvae) and Macoma balthica could experience 'significant' mortality due to exposure to petroleum hydrocarbons. Oil spills have the potential to significantly affect benthic fauna in sheltered soft sediments, but the effects vary depending on location and the extent, duration and type of oil spilt.  Therefore, the resistance of this biotope to petrochemical hydrocarbons and PAHs is assessed as ‘Low’ based on the worst-case results reviewed in Hediste spp. (and Nereis  spp.) and Macoma spp. Hence, resilience is ‘High’ and sensitivity is assessed as ‘Low’ but with ‘Low’ confidence due to the limited evidence reviewed. 

Synthetic compound contamination

Benchmark. Exposure of marine species or habitat to one or more synthetic compound contaminants via uncontrolled releases or incidental spills (Synthetic compound contamination pressure definition).

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Synthetic compounds' contaminants on selected polychaete and mollusc species are summarized below. The full 'evidence reviews' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Synthetic compounds'' contaminant examined, together with an overall pressure assessment. 

Pesticides/biocides. Seven articles examined the effects of pesticides/biocides on Hediste (Nereis) diversicolor. For example, Collier & Pinn (1998) reported that Nereis diversicolor was the most sensitive of the species in their experiment, as 8.0 mg/m² Ivermectin caused 100% mortality within 14 days. Mayor et al. (2008) reported that the sea-lice insecticide emamectin benzoate had significant effects on the survival of Hediste diversicolor with an LC50 value of 1,367.71 µg/kg (wet sediment). Scaps et al. (1997) reported that by day 21, 20% of Nereis diversicolor exposed to 10^-6 and 10^-8 M Malathion died, and 33.5 and 62.5% mortality occurred in the 10^-8 and 10^-6 M parathion-ethyl treatments. Only sublethal effects were reported in the other pesticides reviewed. Underwood & Paterson (1993) reported that Nereis diversicolor was absent from areas of sediment treated with formaldehyde. Although not studied directly, the observation suggests that Nereis diversicolor was excluded from the treatment site due to the toxicity of the formaldehyde or the lack of microalgal food. Therefore, the worst-case resistance of Nereis spp. and Hediste spp. to ‘Pesticides/biocides’ is assessed as ‘None’, resilience as ‘Medium’ and sensitivity is assessed as ‘Medium’. The confidence is assessed as ‘low’ due to the limited evidence and the variation in toxicity between species and chemicals tested.

Three articles in the evidence review examined the effects of ‘Pesticides/biocides’ on Macoma spp. Armstrong & Millemann (1974) exposed Macoma nasuta to Sevin at 15, 20, 25, and 30 mg/l for 96 hours. After 96 hours of exposure, around 50% of the clams had withdrawn one or both siphons, with estimated 48-hour and 96-hour EC50s of 27.5 and 17 mg/l Sevin, respectively. But no mortality was reported. Boese et al. (1990) reported sublethal effects on the physiology of Macoma spp. exposed to 5.2 – 7.8 µg/l hexachlorobenzene for 3-7 days. Dumbauld et al. (2001) examined the effect of pesticide treatment of oyster sites on benthic infauna. No significant effect on Macoma spp. abundance was observed 24 hours, 2 weeks, 1 month and 1 year after the application of Carbaryl at 5.6 kg/ha. The average density of Macoma spp. was significantly different from the control plots at 51 days post-exposure to 8.4 kg/ha Carbaryl, but there were no significant differences after 2 days or 1 year (Dumbauld et al., 2001). Therefore, the worst-case sensitivity of Macoma spp. to the ‘Pesticides/biocides’ tested is assessed as ‘Not sensitive’ since only sublethal or transient effects were reported. However, confidence in the assessment is ‘Low’ due to the limited number of studies and pesticides tested.

Three articles in the evidence review examined the effects of ‘Pesticides/biocides’ on Scrobicularia plana. Akberali & Black (1980) reported significant mortality in Scrobicularia plana exposed to the highest concentration of Sevin (a carbamate) with LT50s of 9 days at 10,000 µg/l and 15 days at 5,000 µg/l. No significant mortality occurred at 1,000 µg/l or in the control. Goncalves et al. (2016) reported that Scrobicularia plana of both size classes were more sensitive to the herbicide Primextra® Gold TZ than Cerastoderma edule, with 96-hour LC50 values of 13.26 mg/l (L), 5.54 mg/l (S) and 100% mortality at 60 mg/l. Gutiérrez et al. (2019) reported that S-metolachlor was the most toxic to Scrobicularia plana with 96-hour LC50s of 40.702 mg/l for large individuals and 41.517 mg/l for small, and  96-hour LC50s for Terbuthylazine of 118.59 mg/l for large individuals and 108.42 mg/l for small.  Therefore, the worst-case resistance of Scrobicularia plana to the pesticides tested is assessed as ‘None’, resilience as ‘Medium’ and sensitivity as ‘Medium’.

Pharmaceuticals. Eight articles in the evidence review examined the effects of several ‘Pharmaceuticals’ on Hediste (Nereis) diversicolor. For example, Maranho et al. (2014) reported that the survival of the Hediste diversicolor was negatively correlated with concentrations of Carbamazepine and 17α-ethynylestradiol. 'Significant' mortality (>25%) was observed in the Carbamazepine (0.05 and 0.5 µg/l), Fluoxetine (0.001 µg/l) and 17α-ethynylestradiol (0.01 and 0.1 µg/l) treatment groups. Pires et al. (2016) reported that Hediste diversicolor experienced 8.3% mortality at concentrations of 0.3 and 3.0 µg/l after 28 days of exposure to carbamazepine, 25% mortality at 6.0 µg/l and 16.7% mortality at 18.0 µg/l after 28 days of exposure to carbamazepine. 'Some' mortality occurred after exposure to the beta-blocker Propranolol (Maranho et al., 2014). However, the other studies reported only sublethal effects, including a reduction in burrowing activity.  Therefore, the worst-case resistance of Nereis spp. and Hediste spp. to ‘Pharmaceuticals is assessed as ‘Low’, resilience as ‘High’ and sensitivity is assessed as ‘Low’. The confidence is assessed as ‘Low’ due to the limited evidence and the variation in toxicity between the chemicals studied. 

Six articles in the evidence review examined the effects of ‘Pharmaceuticals’ on Scrobicularia plana. Almeida et al. (2017b) examined the effect of the anticonvulsant Carbamazepine on Scrobicularia plana but did not detect significant effects. Freitas et al. (2015) reported 11% mortality in Scrobicularia plana from a mercury contaminated site when exposed to 3 µg/l of Carbamazepine at pH 7.8. Freitas et al. (2015b) reported 10% mortality exposed to Carbamazepine 0.3 µg/l for 28 days. Freitas et al. (2016) reported 33% mortality of Scrobicularia plana after 28 days in 3 µg/l Carbamazepine treatment and in the 3 µg/l Carbamazepine and low pH (pH 7.1) treatment. There was no mortality in the control treatment, but 22% mortality in the low pH exposure. Therefore, the worst-case resistance of Scrobicularia plana to the anticonvulsant Carbamazepine is assessed as ‘Low’, resilience as ‘Medium’ and sensitivity as ‘Medium’ but with ‘Low’ confidence due to the limited evidence.

Langston et al. (2007) examined the effects of endocrine disrupting chemicals (EDCs) on Scrobicularia plana. Langston et al. (2007) examined the effects of the human hormone 17α-ethinyloestradiol (EE2) or 17β-oestradiol (E2) alone, and in mixtures with other EDCs, that is 17β-oestradiol (E2), 17α-ethinyloestradiol (EE2), octylphenol (OP) and nonylphenol (NP) in sediment for six weeks. In low-level exposures, 44% of males were observed to have ovotestis condition. In the high-level exposures, recovery rates of the clams were too low for statistical analysis of intersex. Therefore, the sensitivity of Scrobicularia plana to the endocrine disrupting chemicals tested is assessed as ‘Not sensitive’ because only sublethal (intersex) effects were reported. Langston (2020) reported that this intersex (feminisation of male clams) was widespread in the southwest of England, with 60% of males from the Bristol Channel and Severn Estuary exhibiting the condition, although the male to female sex ratio was ca 1:1.  Some of the highest levels of intersex were observed in the Mersey Estuary, which is highly industrialised and reported to exhibit elevated EDCs in flounders (Langston, 2020). However, the implications for fecundity and the population were unknown (Langston, 2020). 

Overall sensitivity assessment for this pressure. Hediste spp., Nereis spp., and Scrobicularia plana were reported to experience severe or significant mortality due to exposure to many but not all of the pesticides studied, depending on the concentration and duration of exposure, while only sublethal effects were reported in Macoma spp. albeit based on fewer articles. Hediste diversicolor was also reported to experience 'significant' or 'some' mortality after exposure to the pharmaceuticals tested. Scrobicularia plana was also reported to develop intersex due to EDCs, but the effects at the population level were still unknown. Therefore, the worst-case resistance of this biotope, based on the effect of pesticides/biocides on the important characteristic species Hediste spp.,  is assessed as ‘None’, resilience as ‘Medium’ and sensitivity is assessed as ‘Medium’. Where present, Scrobicularia plana may also be adversely affected by pesticide/biocide contamination. The confidence is assessed as ‘Low’ due to the limited evidence and the variation in toxicity between species and chemicals tested.

Radionuclide contamination

Benchmark. An increase in 10µGy/h above background levels (Radionuclides contamination pressure definition).

Evidence

Beasley & Fowler (1976) and Germain et al. (1984) examined the accumulation and transfers of radionuclides in Hediste diversicolor from sediments contaminated with americium and plutonium derived from nuclear weapons testing and the release of liquid effluent from a nuclear processing plant. Both concluded that the uptake of radionuclides by Hediste diversicolor was small. Beasley & Fowler (1976) found that Hediste diversicolor accumulated only 0.05% of the concentration of radionuclides found in the sediment. Both also considered that the predominant contamination pathway for Hediste diversicolor was from the interstitial water. However, there is insufficient information on the biological effects of radionuclides to comment further upon the intolerance of this species to radionuclide contamination. Hutchins et al. (1998) described the effects of temperature on bioaccumulation by Macoma balthica of radioactive americium, caesium and cobalt, but did not comment on the intolerance of the species. Further, direct assessments of impacts at the benchmark pressure on benthic communities, and this biotope, in particular, were not found. Insufficient evidence was available to complete a sensitivity assessment.

Introduction of other substances

Benchmark. Exposure of marine species or habitat to one or more relevant "other" substances (solid, liquid or gas) contaminants via uncontrolled releases or incidental spills (Introduction of other substances pressure definition). 

Evidence

The results of the Rapid Evidence Assessment on the effects of 'Other substances' contaminants on selected polychaete and mollusc species are summarized below. The full 'evidence reviews' should be consulted for details of the studies examined and their results. A sensitivity assessment is provided for each type or source of 'Other substances' contaminant examined, together with an overall pressure assessment. 

De Marchi et al. (2017) investigated the effects of different multi-walled carbon nanotubes (MWCNTs) (at 0.01, 0.10 and 1.00 mg/l) on Diopatra neapolitana and Hediste diversicolor. The effects on physiological and biochemical performance were assessed after 28 days of exposure. The respiration rate of Hediste diversicolor was measured after 28 days of exposure. Exposure to 0.01 mg/l MWCNTs increased the respiration rate compared to the controls. However, respiration rates significantly decreased at 0.1 mg/L MWCNTs compared to the control, but respiration rates increased at the highest tested concentration (1.00 mg/L) of MWCNTs. Mortality of Hediste diversicolor individuals exposed to 0.01, 0.10 and 1.00 mg/l was 11% at each of the concentrations tested. In the control treatment, 100% survival was recorded after bioassay.

Pires et al. (2022) investigated the effects of graphene oxide (GO) nanosheets on the behavioural, physiological, and biochemical responses of Hediste diversicolor. Polychaetes were exposed to a range of concentrations of graphene oxide nanosheets (10, 100, 1000, and 10,000 µg/l) for 28 days. The study assessed the effects on the behaviour, feeding activity, mucus production, regenerative capacity, antioxidant status, biochemical damage, and metabolism. Body regeneration was significantly influenced by GO exposure, with all individuals exposed to GO exhibiting reductions in the number of regenerated segments when compared to the controls. Feeding activity was influenced by GO exposure, with increased feeding times for exposed individuals. The segregation of mucus was significantly higher in individuals exposed to GO compared to the controls. Burrowing rates of polychaetes exposed to GO were significantly slower than those in the control, with 20-35% of individuals in the highest tested concentrations unable to burrow by the end of the 30-minute assay. Mortality was around 40% in the GO-exposed treatments, with 30% mortality in the 10 and 100 µg/l treatments, but only 5% in the highest tested concentration of 10,000 µg/l. No mortality occurred in the control.

Pires et al. (2016) investigated the biochemical effects of single and combined exposure to Carbamazepine and caffeine on Hediste diversicolor. Polychaetes were exposed to a range of concentrations of carbamazepine (0.3, 3.0, 6.0 and 9.0 µg/l) and caffeine (0.5, 3.0, and 18.0 µg/l) for 28 days. 8.3% of mortality at concentrations 0.3 and 3.0 µg/l after 28 days of exposure to carbamazepine, 25% at 6.0 µg/l and 16.7% at 18.0 µg/l. Specimens experienced 8.3% of mortality after 28 days of exposure to caffeine at 0.5 and 18.0 µg/l, and polychaetes exposed to 0.3 µg/l carbamazepine + 0.5 µg/l caffeine had 8.3% mortality. No mortality was recorded in the 3 µg/l caffeine or 3 µg/l caffeine + 6 µg/l carbamazepine treatment.

Muller-Karanassos et al. (2021) investigated the effects of environmental concentrations of antifouling paint particles on sediment-dwelling invertebrates. Adult ragworms and cockles were exposed to three types of antifouling paint particles (APP), two biocidal (‘historic’ and ‘modern’) and one biocide-free (‘silicone’). Two laboratory-based 18-day and 5-day exposure experiments were carried out. The APPs ranged in particle size and included varying concentrations of Cu, Sn, Pb, Hg, and Zn. Trial experiments carried out using the maximum environmental APP concentration (18.8 g/l) caused 100% mortality of all ragworms and cockles in the modern treatment within six days. In the 18-day exposure, antifouling paint particle concentrations were 4.2 g/l for the historic biocidal treatment, 3.0 g/l for the modern biocidal treatment, and 2.1 g/l for the non-biocidal silicone treatment. The burrowing rate of the ragworms was reduced by 29% in the modern biocidal treatment. However, there were no significant differences between treatments. Ragworms decreased in weight and feeding rates significantly, but significant differences were only seen between the modern biocidal treatment and the control. Modern biocidal antifouling paint particles were used at concentrations ranging from 0 to 30 g/l (ragworms) and 0 to 6 g/l (cockles) to estimate the 5-day LC50 exposure. The 5-day LC50 values were 19.9 g/l for the ragworms and 2.3 g/l for cockles. The 5-day EC50 values were 14.6 g/l for the ragworms and 1.4 g/l for cockles.

Therefore, Hediste diversicolor probably has a resistance of ‘Low’ to graphene oxide nanosheets and ‘Medium’ resistance to multi-walled carbon nanotubes. Hence, resilience is probably ‘High’ and sensitivity is assessed as ‘Low’ but with ‘Low’ confidence. The above evidence suggests that Hediste diversicolor has a ‘Low’ sensitivity to caffeine exposure, but confidence in the assessment is ‘Low’ due to the lack of evidence. Muller-Karanassos et al. (2021) suggest that antifouling paint particles remain toxic in the environment. Therefore, the resistance of Hediste diversicolor to APPs is assessed as ‘None’. Hence, resilience is assessed as ‘Medium’ and sensitivity as ‘Medium’, but confidence in the assessment is ‘Low’ due to the lack of evidence.

Caldwell et al. (1975) exposed Macoma balthica to 100, 330, 1,000, 3,300, and 10,000 µg/l hydrogen sulphide for 96 hours. The longer the clams were exposed to hydrogen sulphide, the lower the concentration was required to cause 50% mortality. The LC50 at 24, 48, and 96 hours were 10,000, 8,000 and 6,000 µg/l, respectively. Crecelius (1979) examined the effect of bromate on Macoma inquinata and reported 100% mortality after 72 hours at 880 mg/l bromate. Therefore, the worst-case resistance of Macoma spp. to the inorganic chemicals tested is assessed as ‘None’, resilience as ‘Medium’ and sensitivity as ‘Medium’ but with ‘Low confidence due to the limited evidence.

Wastewater discharge was shown to reduce the abundance of Macoma cumana in the affected area (Bergayou et al., 2019). Exposure to pulverised fuel ash was reported to result in 20% mortality of Macoma balthica in the 100% PFA treatment (Jenner & Bowmer, 1990). Therefore, the worst-case resistance of Macoma balthica to pulverised fuel ash exposure is assessed as ‘Medium’, resilience as ‘High’, and sensitivity as ‘Low’, but with ‘Low’ confidence. However, its resistance to wastewater discharge may be ‘Medium’ and its sensitivity ‘Low’, depending on the nature of the contaminants involved, but confidence is ‘Low’ due to the lack of evidence.

Rida et al. (2004) investigated the lethal effects of sediment from the Guadalquivir estuary following the Aznalcollar mining spill. Toxicity tests using dilutions (0.3, 1.8, 7.9, 20, and 32%) of toxic mud and sediment from environmental stations were tested using the amphipod Ampelisca brevicornis and the clam Scrobicularia plana. Two bioassays were performed to assess the toxicity of the mud dilutions (96-hour and 10-day). The results showed the amphipods to be more sensitive to the spill than the clams. At dilutions of toxic mud above 1.8%, there was 100% mortality of amphipods. For the clams, no mortalities were observed at toxic mud dilutions of 0.3% and 1.8% within the first 72 hours, and by the end of the bioassay (96 hours), there were no significant differences in mortality among the environmental stations. However, at dilutions above 1.8%, significant mortality occurred. At the dilutions of 20% and 32% toxic mud, 100% mortality of clams occurred by the end of the 96-hour exposure. At the dilution of 7.9%, mortality of around 75% occurred by the end of the 96-hour bioassay. The 96-hour LC50 of the toxic mud to the clams was calculated at 3.25%.

Silva et al. (2012) investigated the effects of fish farm effluents on benthic community structure and biomarker responses of the clam Scrobicularia plana. The benthic fauna and clam samples were collected from five sites following a gradient of contamination from the aquaculture effluent to the control site. The numbers of species, abundance, richness, and Shannon diversity were the biodiversity indicators calculated for each sampling site. The morphological and reproductive status of clams was assessed using the condition factor and gonadosomatic index, respectively. Benthic biodiversity indicators were significantly negatively correlated with organic matter. The condition factor significantly increased at sites nearest to the fish farm effluent compared to the control. The gonadosomatic index was significantly positively correlated with the distance to fish farm effluent and negatively correlated with organic matter.

Overall, wastewater discharge was shown to reduce the abundance of Scrobicularia plana in the affected area (Bergayou et al., 2019), while Silva et al. (2012) reported changes in its condition. However, Rida et al. (2004) reported that the exposure to toxic mud from a mine spill resulted in ‘Significant’ mortality of Scrobicularia plana at dilutions above 1.8%; the 96-hour LC50 of the toxic mud to the clams was calculated at 3.25%. However, the exact nature of the contaminants in each mixture is unclear. Therefore, the worst-case resistance of Scrobicularia plana to mine spill effluent exposure is assessed as ‘Low’, resilience as ‘Medium’ and sensitivity as ‘Medium’ but with ‘Low’ confidence. However, its resistance to wastewater discharge may be ‘Medium’ and its sensitivity ‘Low’, depending on the nature of the contaminants involved, but confidence is ‘Low’ due to the lack of evidence.

Overall sensitivity assessment for this pressure.  A variety of chemicals fall within this pressure definition, so it is difficult to compare the individual toxicity of different chemicals between the species within the biotope. The important characteristic species, Hediste diversicolor and Macoma balthica, were reported to experience mortality after exposure to multi-walled carbon nanotubes, graphene oxide nanosheets, caffeine, antifouling paint particles, and bromate. Scrobicularia plana was reported to experience mortality after exposure to mining spill effluent and sublethal effects from fish farm effluent.  Overall,  the worst-case resistance of this biotope (based on Hediste diversicolor and Macoma spp.) to the inorganic chemicals tested is assessed as ‘None’, resilience as ‘Medium’ and sensitivity as ‘Medium’ but with ‘Low confidence due to the limited evidence and variation in the effects of the between the chemicals tested.

De-oxygenation

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

Evidence

The habitats which Hediste diversicolor inhabits tend to have lower oxygen levels than other sediments. Hediste diversicolor is resistant to moderate hypoxia (Diaz & Rosenberg, 1995). Vismann (1990) demonstrated a mortality of only 15% during a 22-day exposure of Hediste diversicolor at 10% oxygen (ca. 2.8 mg O₂ per litre). Hediste diversicolor is active at the sediment/water interface where hydrogen sulphide concentrations increase during periods of hypoxia. Vismann (1990) also demonstrated that the high tolerance of Hediste diversicolor to hypoxia in the presence of sulphide is enabled by elevated sulphide oxidation activity in the blood. Hediste diversicolor may also exhibit a behavioural response to hypoxia by leaving the sediment (Vismann, 1990) in the presence of sulphide. After 10 days of hypoxia (10% oxygen saturation) with sulphide (172 to 187 µmM), only 35% of Hediste diversicolor had left the sediment compared to 100% of Nereis virens.

However, in Gamo Lagoon (Japan), Kanaya et al. (2016) Hediste spp. was amongst the most dominant taxa, but densities were lower at sites characterized by significantly high hydrogen sulphide, and the dominant polychaetes (including Hediste spp.) decreased or sometimes disappeared during summer months when sediment hydrogen sulphide increased. In the warmer summer month, dissolved oxygen often decreases to near 0 mg/l at night (Kanaya et al., 2016). In an in situ enclosed experiment, adult Hediste spp. colonized mud with sulphide removed in significantly higher densities than sulfidic mud (Kanaya et al., 2016). Juveniles had the same result but the differences were not significantly different. This experiment showed that recolonization by the opportunistic polychaetes can occur rapidly after sediment recovery (e.g., within 37 days in the experiment). Kanaya et al. (2016) concluded that where sediments contained hydrogen sulphide, the dominant polychaetes (including Hediste spp.) mainly occupied the thin oxidized surface layer, indicating as hydrogen sulphide increased during summer months, individuals deeper in the sediment died or moved upwards towards the surface. The enclosure experiment also indicated the positive effects of organic-rich mud for the settlement of dominant polychaetes, such as Hediste spp., when hydrogen sulphide is absent or removed.

Laboratory experiments in the absence of sediments found that Hediste diversicolor could survive hypoxia for more than five days and that it had a higher tolerance to hypoxia than Nereis virens, Nereis succinea and Nereis pelagica (Theede, 1973; Dries & Theede, 1974; Theede et al., 1973). Juvenile Hediste diversicolor survived hypoxic conditions for four days in laboratory conditions and combined hypoxia and increased sulphide (1 mmol/l) for three days (Gamenick et al., 1996). Post larvae Hediste diversicolor were the only life stage to show less tolerance to hypoxia, surviving for only 14 hours (Gamenick et al., 1996).

Koop-Jakobsen et al. (2017) observed oxygen in a Hediste diversicolor burrow can fluctuate between 0 and 20.4 umol/l O₂, and reached anoxic conditions approximately three times per hour during a five-hour monitoring period. This suggests that Hediste diversicolor may regularly experience short-term anoxic conditions in the burrows made in the sediment, and could be able to tolerate fluctuating oxygen levels.

Macoma balthica appears to be relatively tolerant of deoxygenation. Brafield & Newell (1961) frequently observed that, in conditions of oxygen deficiency (e.g. less than 1 mg O₂/l), Macoma balthica survived low oxygen concentrations and shell growth continued. Similarly, Jansson et al. (2015) found juvenile Macoma balthica exhibited the highest growth rate and significantly higher survival in exposure to low oxygen concentrations (3.0 mg/l) for 29 days in experimental conditions. An LC50 of dissolved oxygen of Macoma balthica was reported as 1.7 mg/l for a 28-day experiment (Long et al., 2008 cited in Song et al., 2024).

Examining two models, Ehrnsten et al. (2019a) described Macoma balthica (as Limecola) as being less sensitive to hypoxia compared to other deposit feeders and predators in the model; however, its biomass increased once oxygen levels increased. Field evidence in the Baltic Sea supported this, as Macoma balthica was present at sites described as intermittently hypoxic, but occurred in higher abundances at sites which remained normoxic all year round (Gammal et al., 2017). However, sites that experienced prolonged seasonal hypoxia (0.0 mg/l and 0.2 mg/l) Macoma balthica was absent from the benthic macrofauna communities at sites that experienced prolonged seasonal hypoxia (0.0 mg/l and 0.2 mg/l), suggesting that prolonged hypoxia can reduce local communities (Gammal et al., 2017).

Sub-lethal behaviour and physiological effects of hypoxia have been identified and observed (Villnas et al., 2019). Brafield & Newell (1961) frequently observed that in conditions of oxygen deficiency (e.g. less than 1 mg O2/l) Macoma balthica moved upwards to fully expose itself on the surface of the sand. Specimens lay on their side with the foot and siphons retracted but with valves gaping slightly allowing the mantle edge to be brought into full contact with the more oxygenated surface water lying between sand ripples. In addition, Macoma balthica was observed under laboratory conditions to extend its siphons upwards out of the sand into the overlying water when water was slowly deoxygenated with a stream of nitrogen. The lower the oxygen concentration became the further the siphons extended (Brafield & Newell, 1961; Long et al., 2014; Villnas et al., 2019; Song et al., 2024). Migration to the surface due to hypoxia may leave it at greater risk of predation (Long et al., 2008; Long et al., 2014; Villnas et al., 2019; Song et al., 2024). In one study, these behaviours were observed after short-term (3-day) recurring hypoxic disturbance, but recovered quickly by reburying once oxic conditions were re-established (Villnas et al., 2019). Under recurring hypoxic disturbance, impacts were observed to affect juveniles before adults. However, long-term (30 days) of uninterrupted hypoxia, increasingly reduced the bivalve population and the majority of Macoma balthica individuals exposed to did not rebury at all (Villnas et al., 2019). Long et al. (2014) also provided evidence that hypoxia decreases individual Macoma balthica growth, causes local extinction of populations and reduces egg production by 40%, while increasing protein investment per egg (cited in Song et al., 2024).

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

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

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

Nutrient enrichment

Benchmark. Increased levels of the elements nitrogen, phosphorus, silicon, and iron in the marine environment compared to background concentrations (Nutrient enrichment pressure definition).

Evidence

This pressure relates to increased levels of nitrogen, phosphorus and silicon in the marine environment compared to background concentrations. Primary production in the biotope will be limited to microalgae at the sediment surface, rather than macroalgae. Changes in primary production as a result of changes in nutrient enrichment are, therefore, not considered likely to directly alter the biotope.

The species characterizing this biotope, (e.g. Hediste diversicolor and Macoma balthica) are recognised in current literature as important bioturbators and key bio-irrigator species that contribute significantly to nutrient cycling, sediment oxygenation and bioremediation of nutrients, making them a major contributor to benthic ecosystem functioning (Martinez-Garcia et al., 2015; Clare et al., 2016; Fang et al., 2019; 2021; Dolbeth et al., 2019; Ehrnsten et al., 2019b; Gilbert et al., 2021; Wyness et al., 2021; Farrell et al., 2024; Lehuen et al., 2024; Morelle et al., 2024; Bhuiyan et al., 2025; Gammal et al., 2025).

Evidence has shown that Hediste diversicolor can successfully feed on, survive in, and exhibit sustained positive growth in biodeposits and nutrient-rich aquaculture sludge composed of organic waste from farmed species (Wang et al., 2019; Bergström et al., 2017; 2019; Pombo et al., 2020; Anglade et al., 2023a; Malzahn et al., 2023; Santos et al., 2025). As a result, Hediste diversicolor can decrease or remove particulate organic matter from aquaculture effluents and bioremediate waste into valuable biomass (Jerónimo et al., 2020; 2021; Santos et al., 2025). This species can consume waste and convert it into high-value compounds, such as lipids and fatty acids, assimilating nutrients including carbon, nitrogen and phosphorus that would otherwise be lost to the external environment (Wang et al., 2019; Pombo et al., 2020; Anglade et al., 2023a; 2023b; 2024).

Experiments on Hediste diversicolor have shown increased population densities and high reproductive success associated with elevated organic matter from aquaculture effluent (Jerónimo et al., 2020). It has been suggested that successful bioturbation processes are important to prevent build up in organic matter (Jerónimo et al., 2020).

In laboratory experiments, Murray et al. (2017) showed that Hediste diversicolor can alter sediment particle reworking in response to algal (Ulva intestinalis) enrichment. Particle reworking increased in moderate enrichment but reduced in higher levels of enrichment. During this experiment, changes in the overlying water nutrients (nitrogen, ammonium, phosphorus and silicate) were influenced by the algal enrichment and salinity, and differed between treatments with and without the presence of Hediste diversicolor (Murray et al. 2017). This indicates that Hediste diversicolor can continue to bioturbate the sediment and influence sediment-water nutrient cycling under organically enriched conditions.

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

Touhami et al. (2025) reported Hediste diversicolor characterized moderately enriched communities in the Loukkos estuary, at sites associated with urban discharge and agricultural runoff with elevated nutrient loads (especially nitrate). Bergayou et al. (2019) examined the changes in intertidal macrobenthic communities in the Oued Souss estuary (Morocco) following the cessation of untreated wastewater discharge. The study compared surveys from 2001 to 2002, during discharge and 2003, after cessation. Hediste diversicolor was a dominant species in the community with very high abundance and densities during the discharge period, and contributed strongly to total community biomass (Bergayou et al., 2019). After discharge stopped, Hediste diversicolor abundance and density increased, but its biomass contribution decreased. This was consistent with reduced organic enrichment after wastewater inputs stopped (Bergayou et al., 2019).

Singer et al. (2023) found that long-term de-eutrophication in the East-Frisian Wadden Sea reduced nutrient inputs, primary production, and the supply of particulate organic matter to the seabed, leading to a substantial decline in benthic food resources. This decrease in organic enrichment was linked to a decrease in the abundance and biomass of Macoma balthica (reported as Limecola) and Hediste diversicolor. This shows the species characterising this biotope are adapted to organically enriched environments and rely on the organic matter supply.

Despite Hediste diversicolor’s ability to bioturbate nutrients and organically enriched sediment, Martinez-Garcia et al. (2015) reported that under high organic enrichment (created using fish faecal pellets), Hediste diversicolor began to die once Beggiatoa mats appeared (from day 14) in enriched muddy sediments. Although individuals were replaced, the replacements died again within 1 to 2 days. This indicates that the enriched sediment had become increasingly toxic and sulfidic, which was unsuitable for Hediste diversicolor (Martinez-Garcia et al., 2015).

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

Sensitivity assessment.  The above evidence suggests that both Hediste diversicolor and Macoma balthica increase in abundance in nutrient and organic-enriched conditions, and may decrease in abundance when nutrient enrichment is removed. Bioturbation by Hediste diversicolor may also mitigate the effects of nutrient and organic enrichment. However, high enrichment that results in anoxic conditions and/or bacterial mats was reported to result in mortality in Hediste. The effects of algal mats, themselves caused by eutrophication, was inconsistent.  Therefore, resistance is assessed as ‘High’ based that Hediste and Macoma can benefit from nutrient or organic enrichment, but with ‘Low’ confidence due to inconsistences in some of the evidence. Hence, resilience is assessed as 'High' (by default), so that the biotope is assessed as 'Not sensitive'.

Organic enrichment

Benchmark. A deposit of 100 gC/m²/yr (Organic enrichment pressure definition).

Evidence

All species apart from two that are abundant in the biotope are classed in a Marine Biotic Index as being indifferent to, tolerating or proliferating under organic enrichment conditions (Borja et al., 2000). Evidence shows that Hediste diversicolor positively correlates with high sediment organic matter, and is often a dominant species in macrofaunal communities in organically enriched environments (El Asri et al., 2018). In Jade Bay (Wadden Sea), Hediste diversicolor was recorded in macrofauna communities associated with high total organic (1.77%) (Schückel et al., 2015). 

Evidence has shown that Hediste diversicolor can successfully feed on, survive in, and exhibit sustained positive growth in biodeposits and nutrient-rich aquaculture sludge composed of organic waste from farmed species (Wang et al., 2019; Bergström et al., 2017; 2019; Pombo et al., 2020; Anglade et al., 2023a; Malzahn et al., 2023; Santos et al., 2025). As a result, Hediste diversicolor can decrease or remove particulate organic matter from aquaculture effluents and bioremediate waste into valuable biomass (Jerónimo et al., 2020; 2021; Santos et al., 2025). This species can consume waste and convert it into high-value compounds, such as lipids and fatty acids, assimilating nutrients including carbon, nitrogen and phosphorus that would otherwise be lost to the external environment (Wang et al., 2019; Pombo et al., 2020; Anglade et al., 2023a; 2023b; 2024).

Experiments on Hediste diversicolor have shown increased population densities and high reproductive success associated with elevated organic matter from aquaculture effluent (Jerónimo et al., 2020). It has been suggested that successful bioturbation processes are important to prevent build up in organic matter (Jerónimo et al., 2020).

In laboratory experiments, Murray et al. (2017) showed that Hediste diversicolor can alter sediment particle reworking in response to algal (Ulva intestinalis) enrichment. Particle reworking increased in moderate enrichment but reduced in higher levels of enrichment. During this experiment, changes in the overlying water nutrients (nitrogen, ammonium, phosphorus and silicate) were influenced by the algal enrichment and salinity, and differed between treatments with and without the presence of Hediste diversicolor (Murray et al. 2017). This indicates that Hediste diversicolor can continue to bioturbate the sediment and influence sediment-water nutrient cycling under organically enriched conditions.

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

Touhami et al. (2025) reported Hediste diversicolor characterized moderately enriched communities in the Loukkos estuary, at sites associated with urban discharge and agricultural runoff with elevated nutrient loads (especially nitrate). Bergayou et al. (2019) examined the changes in intertidal macrobenthic communities in the Oued Souss estuary (Morocco) following the cessation of untreated wastewater discharge. The study compared surveys from 2001 to 2002, during discharge and 2003, after cessation. Hediste diversicolor was a dominant species in the community with very high abundance and densities during the discharge period, and contributed strongly to total community biomass (Bergayou et al., 2019). After discharge stopped, Hediste diversicolor abundance and density increased, but its biomass contribution decreased. This was consistent with reduced organic enrichment after wastewater inputs stopped (Bergayou et al., 2019).

Singer et al. (2023) found that long-term de-eutrophication in the East-Frisian Wadden Sea reduced nutrient inputs, primary production, and the supply of particulate organic matter to the seabed, leading to a substantial decline in benthic food resources. This decrease in organic enrichment was linked to a decrease in the abundance and biomass of Macoma balthica (reported as Limecola) and Hediste diversicolor. This shows the species characterising this biotope are adapted to organically enriched environments and rely on the organic matter supply.

Despite Hediste diversicolor’s ability to bioturbate nutrients and organically enriched sediment, Martinez-Garcia et al. (2015) reported that under high organic enrichment (created using fish faecal pellets), Hediste diversicolor began to die once Beggiatoa mats appeared (from day 14) in enriched muddy sediments. Although individuals were replaced, replacements died again within 1 to 2 days. This indicates that the enriched sediment had become increasingly toxic and sulfidic, which was unsuitable for Hediste diversicolor (Martinez-Garcia et al., 2015).

In the Marine Biotic index, Macoma balthica was assigned as ‘species very sensitive to organic enrichment and present under unpolluted conditions (initial state)’ (Borja et al., 2000). However, case studies display resilience of Macoma balthica populations to enrichment. Macoma balthica have been shown experimentally to be able to resist periods of up to nine weeks under algal clover, their long siphon allowing them to reach oxygenated water although other bivalves decreased in abundance (Thiel et al., 1998). Organic enrichment from waste-water discharge in the Dutch Wadden Sea resulted in positive effects on Macoma balthica abundance, biomass, shell growth and production. These effects were concluded to be due to increased food supply (Madsen & Jensen, 1987). Tubificoides benedii and other oligochaetes are very tolerant of high levels of organic enrichment and often dominate sediments where sewage has been discharged or other forms of organic enrichment have occurred (Pearson & Rosenberg, 1978; Gray, 1971; McLusky et al., 1980). In Jade Bay (Wadden Sea), Macoma balthica is recorded in macrofauna communities associated with total organic carbon content between 0.82% to 1.77% (Schückel et al., 2015).

Gittenberger & van Loon (2011) in the development of an AMBI index to assess disturbance (including organic enrichment), assigned Hediste diversicolor and Macoma balthica to their Ecological Group III ‘species tolerant to excess organic matter enrichment. These species may occur under normal conditions, but their populations are stimulated by organic enrichment (slight unbalance situations). They are surface deposit-feeding species, as tubicolous spionids’.

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

Physical loss (to land or freshwater habitat)

Benchmark. A permanent loss of existing saline habitat within the site (Physical loss pressure definition). 

Evidence

All marine habitats and benthic species are considered to have a resistance of ‘None’ to this pressure and to be unable to recover from a permanent loss of habitat (resilience is ‘Very Low’).  Sensitivity within the direct spatial footprint of this pressure is therefore ‘High’. Although no specific evidence is described confidence in this assessment is ‘High’, due to the incontrovertible nature of this pressure.

Physical change (to another seabed type)

Benchmark. Permanent change from sedimentary or soft rock substrata to hard rock or artificial substrata, or vice versa (Physical change in subtratum type pressure definition).

Evidence

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

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

Physical change (to another sediment type)

Benchmark. Permanent change in one Folk class (based on UK SeaMap simplified classification) (Physical change in sediment type pressure definition). 

Evidence

The benchmark for this pressure refers to a change in one Folk class.  The pressure benchmark originally developed by Tillin et al. (2010) used the modified Folk triangle developed by Long (2006) which simplified sediment types into four categories: mud and sandy mud, sand and muddy sand, mixed sediments and coarse sediments.  The change referred to is, therefore, a change in sediment classification rather than a change in the finer-scale original Folk categories (Folk, 1954).  The biotope occurs in mud and sandy mud a change to finer sediments is not considered relevant as this falls within the natural habitat change (JNCC, 2015).

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

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

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

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

Habitat structure changes - removal of substratum (extraction)

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

Evidence

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

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

Abrasion / disturbance of the surface of the substratum or seabed

Benchmark. Damage to surface features (e.g. species and physical structures within the habitat) (Surface abrasion/disturbance pressure definition).

Evidence

Muddy sand sediments, in general, tend to be cohesive, although high levels of water content will reduce this and destabilise sediments. Sediment cohesion provides some sediment stabilisation to resist erosion following surface disturbance. The characterizing species associated with this biotope are infaunal and hence have some protection against surface disturbance, although siphons of bivalves and tubes of the sedentary polychaete Pygospio elegans, may project above the sediment surface. Damage to tubes and siphons would require repair. The snail Hydroia ulvae is present on the surface and abrasion may result in burial or damage to this species. Surface compaction can collapse burrows and reduce the pore space between particles, decreasing penetrability and reducing stability and oxygen content (Sheehan, 2007). Trampling (3 times a week for 1 month) associated with bait digging reduced the abundance and diversity of infauna (Sheehan, 2007; intertidal muds and sands).

Clarke et al. (2018) examined pump-scoop dredging in Poole Harbour using a Before-After-Control-Impact (BACI) design with three study sites; a long-term heavily dredged area, a newly opened dredged area and a control with no dredging. A high abundance of Hediste diversicolor was reported in heavily dredged sites exposed to long term high fishing intensity, which had an estimated 67.94 % of disturbed area (Clarke et al., 2018). At the newly opened dredged site, exposed to an estimated 68.03% of disturbed sediment, Hediste diversicolor abundance significantly increased between June (pre-dredging) to November (after opened) (Clarke et al., 2018). This suggested that as a scavenger, Hediste diversicolor could benefit from short-term physical disturbance from dredging, as it reduces competition for food and space (Clarke et al., 2018). Streblospio shrubsolii density also dramatically increased at the newly dredged site compared to other sites (Clarke et al., 2018). Despite differences found between sites, Clarke et al. (2018) described all sites as “moderately disturbed”, according to the AZTI Marine Biotic Index (AMBI).

Long-term monitoring of the Wadden Sea found that cockle dredging had no significant effect on the recruitment of Macoma balthica (reported as Limecola balthica), and survival of recruits did not differ between fishing and non-fishing years (Beukema & Dekker, 2018; 2020; Van Der Meer & Folmer, 2023).

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

Penetration or disturbance of the substratum subsurface

Benchmark. Damage to sub-surface features (e.g. species and physical structures within the habitat) (Sub-surface penetration pressure definition).

Evidence

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

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

Clarke et al. (2018) examined pump-scoop dredging in Poole Harbour using a Before-After-Control-Impact (BACI) design with three study sites; a long-term heavily dredged area, a newly opened dredged area and a control with no dredging. A high abundance of Hediste diversicolor was reported in heavily dredged sites exposed to long term high fishing intensity, which had an estimated 67.94 % of disturbed area (Clarke et al., 2018). At the newly opened dredged site, exposed to an estimated 68.03% of disturbed sediment, Hediste diversicolor abundance significantly increased between June (pre-dredging) to November (after opened) (Clarke et al., 2018). This suggested that as a scavenger, Hediste diversicolor could benefit from short-term physical disturbance from dredging, as it reduces competition for food and space (Clarke et al., 2018). Streblospio shrubsolii density also dramatically increased at the newly dredged site compared to other sites (Clarke et al., 2018). Despite differences found between sites, Clarke et al. (2018) described all sites as “moderately disturbed”, according to the AZTI Marine Biotic Index (AMBI).

Long-term monitoring of the Wadden Sea found that cockle dredging had no significant effect on the recruitment of Macoma balthica (reported as Limecola balthica) and survival of recruits did not differ between fishing and non-fishing years (Beukema & Dekker, 2018; 2020; Van Der Meer & Folmer, 2023).

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

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

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

Changes in suspended solids (water clarity)

Benchmark. A change in one rank on the WFD (Water Framework Directive) scale, e.g. from clear to intermediate for one year (Suspended sediment pressure definition).

Evidence

Changes in light penetration or attenuation associated with this pressure are not relevant to Hediste diversicolor and Macoma balthica biotopes. As the species live in the sediment, they are also likely to be adapted to increased suspended sediment (and turbidity). Evidence has shown that suspended particulate organic matter (SPOM) can be an important food source, supporting growth in estuarine bivalves (Jung et al., 2019). For example, in the western Wadden Sea, Macoma balthica relies on SPOM in March, but shows seasonal flexibility and can shift feeding modes in summer months when freshwater run off, providing the SPOM, is low (Jung et al., 2019).

This indicates that alterations in the availability of food or the energetic costs in obtaining food or changes in scour could either increase or decrease habitat suitability for characterizing species Macoma balthica and Hediste diversicolor, and for other abundant species such as the oligochaetes Tubificoides benedii and Tubificoides pseudogaster.

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

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

Smothering and siltation rate changes (light)

Benchmark. ‘Light’ deposition of up to 5 cm of fine material added to the seabed in a single discrete event (Smothering pressure definition).

Evidence

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

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

Hediste diversicolor inhabits a burrow within the sediment, which can be from 10 cm to 30 or 40 cm in depth (Gogina et al., 2010b; Clare et al., 2016; Gilbert et al., 2021; Bhuiyan et al., 2025). Ahlmann et al. (2025) examined the effects of sand capping on Hediste diversicolor under short-term laboratory conditions. In sand-capped conditions and burial by 7 to 8 cm of coarse sand, Hediste diversicolor showed low recovery (11% of individuals recovered), compared to the recovery of 83% in uncapped conditions. The reduced recovery of individuals is likely due to impaired burrow integrity and ventilation (Ahlmann et al., 2025).

Macoma balthica is able to burrow both vertically and horizontally through the substratum. It is likely that Macoma balthica is not sensitive to smothering by a layer of sediment 5 cm thick as it is a mobile species able to burrow upwards and surface from a depth of 5 to 6 cm up to 10 cm (Brafield & Newell, 1961; Brafield, 1963; Stekoll et al., 1980; Bhuiyan et al., 2025). Turk & Risk (1981) investigated the effect of experimentally induced sedimentation (through fences and boxes that induced sediment deposition on intertidal mudflats in the Bay of Fundy), of 1 to 3.5 cm at a rate of 1.9 to 10.2 cm/month. The results showed that Macoma balthica was generally unaffected.

Examining two models, Ehrnsten et al., (2019a) found that increased organic matter sedimentation and increased food availability generally increased Macoma balthica (reported as Limecola balthica) biomass. However, this positive effect was counteracted by higher-bottom temperatures, because increased temperatures increased degradation of the food available and metabolic cost, reducing biomass in response to sedimentation.

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

Laboratory experiments have shown that the snail Hydrobia ulvae can rapidly resurface through 5 cm thick fine deposits, although this ability is reduced where deposited sediments contain little water (Chandrasekara & Frid, 1998). Field experiments where 10 cm of sediment were placed on intertidal sediments to investigate the effects of the beneficial use of dredged materials found that the abundance of Hydrobia ulvae had returned to ambient levels within 1 week (Bolam et al., 2004).

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

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

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

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

Smothering and siltation rate changes (heavy)

Benchmark. ‘Heavy’ deposition of up to 30 cm of fine material added to the seabed in a single discrete event (Smothering pressure definition).

Evidence

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

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

Hediste diversicolor inhabits depositional environments and is capable of burrowing to depths up to 30 to 40 cm deep (Gogina et al., 2010b; Gilbert et al., 2021), reworking sub-surface modifications of its burrow through fine clays and sand. Ahlmann et al. (2025) examined the effects of sand capping on Hediste diversicolor under short-term laboratory conditions. In sand-capped conditions and burial of 7 to 8 cm of coarse sand, Hediste diversicolor showed low recovery (11% of individuals recovered), compared to the recovery of 83% in uncapped conditions. The reduced recovery of individuals is likely due to impaired burrow integrity and ventilation (Ahlmann et al., 2025). Smith (1955) found no appreciable difference in the population of a Hediste diversicolor colony that had been covered by several inches of sand through which the worms tunnelled.

Macoma balthica is able to burrow both vertically and horizontally through the substratum. It is likely that Macoma balthica is not sensitive to smothering by a layer of sediment 5 cm thick as it is a mobile species able to burrow upwards and surface from a depth of 5 to 6 cm up to 10 cm (Brafield & Newell, 1961; Brafield, 1963; Stekoll et al., 1980; Bhuiyan et al., 2025). Turk & Risk (1981) investigated the effect of experimentally induced sedimentation (through fences and boxes that induced sediment deposition on intertidal mudflats in the Bay of Fundy), of 1 to 3.5 cm at a rate of 1.9 to 10.2 cm/month. The results showed that Macoma balthica was generally unaffected.

Examining two models, Ehrnsten et al., (2019a) found that increased organic matter sedimentation and increased food availability generally increased Macoma balthica (reported as Limecola balthica) biomass. However, this positive effect was counteracted by higher-bottom temperatures, because increased temperatures increased degradation of the food available and metabolic cost, reducing biomass in response to sedimentation.

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

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

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

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

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

Litter

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

Evidence

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

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

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

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

Electromagnetic changes

Benchmark. A local electric field of 1 V/m or a local magnetic field of 10 µT (Electromagnetic pressure definition).

Evidence

Evidence on the effect of electromagnetic fields (EMFs) on benthic organisms is still severely lacking. No studies examining the effect of EMFs on macroalgae were found. Some studies have investigated the effect of anthropogenically induced EMFs on benthic invertebrates at intensities ranging between 2 nT and 40 mT, which is often much higher than in-situ measurements from subsea cables. While some report changes to behaviour, physiology, reproduction, development, immunology, cytotoxicity and orientation, others demonstrate no effect from exposure to the EMF (Albert et al., 2020; Hutchison et al., 2020), depending on the study species and duration and intensity of exposure. There have been no studies investigating the effect of EMFs at the population or community level for benthic organisms and there is a lack of evidence on the effects of electric and magnetic fields many of the characteristic species within this biotope.

Electric and magnetic fields generated by sources such as marine renewable energy device/array cables may alter the behaviour of predators and affect infauna populations. Evidence of effects of behaviour was found in Jakubowska et al. (2019) study which found Hediste diversicolor did not exhibit avoidance or attraction behaviour in response to exposure to low frequency EMF (1 mT) for eight days, but burrowing activity did increase, with individuals spending most of the exposure period buried in the sediment. The authors suggested that individuals exposed to EMF frequently resided deeper in the sediment layers while also increasing the rate of vertical movements within the sediment, indicating an effect on bioturbation potential. There was also a limited effect on Hediste diversicolor physiology, as food consumption and respiration were not affected by EMF exposure, but ammonia excretion rate significantly reduced (Jakubowska et al., 2019).

Further laboratory evidence found 12-day exposure of 1 mT (50 Hz) EMF to Hediste diversicolor and Macoma balthica (reported as Limecola), caused a significant increase in the frequencies of genotoxicity and cytotoxicity endpoints, which can potentially affect the integrity of genetic material and cause a variety of diseases (Stankevičiūte et al., 2019). The strongest responses were seen in Macoma balthica, as six out of eight endpoints had increased, and Stankevičiūte et al. (2019) concluded that this bivalve species could be the most suitable bioindicator for EMF induced effects. However, the survival of Hediste diversicolor and Macoma balthica in EMF-exposed treatments was 100% and 92%, respectively, with individuals spending most of the exposure time buried in the sediment (Stankevičiūte et al., 2019).

Field measurements of electric fields at North Hoyle wind farm, North Wales recorded 110µ V/m (Gill et al., 2009). Modelled results of magnetic fields from typical subsea electrical cables, such as those used in the renewable energy industry, produced magnetic fields of between 7.85 and 20 µT (Gill et al. 2009; Normandeau et al. 2012). Electric and magnetic fields smaller than those recorded by in field measurements or modelled results were shown to create increased movement in potential predators of Hediste diversicolor, thornback ray Raja clavata and attraction to the source in catshark Scyliorhinus canicular (Gill et al. 2009). Flatfish species, which are predators of many polychaete species including dab Limanda limanda and sole Solea solea have been shown to decrease in abundance in a wind farm array or remain at distance from wind farm towers (Vandendriessche et al., 2015; Winter et al. 2010). However, larger plaice increased in abundance (Vandendriessche et al., 2015). There have been no direct causal links identified to explain these results.

Sensitivity assessment. Recent evidence suggested that the species characterizing this biotope were able to survive short-term exposure to EMF (Stankevičiūte et al., 2019). However, low frequency EMF could cause sublethal cytotoxic and genotoxic effects, although at levels a hundred-fold higher than the benchmark. At present, there is ‘Insufficient evidence’ on which to base an assessment. 

Underwater noise changes

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

Evidence

Species within the biotope can probably detect vibrations caused by noise and in response may retreat in to the sediment for protection. Wang et al. (2022) laboratory experiments demonstrated that low-frequency noise (10 to 500 Hz of added noise) had no significant effect on the burial depth or bioturbation rates of Macoma balthica (reported as Limecola balthica). However, anti-burrowing behaviour was observed that indicated a potential stress response, which could reduce its bioturbation potential (Wang et al., 2022).

Sensitivity assessment. However, at the benchmark level the community is unlikely to be sensitive to noise and given the lack of data and evidence. Sensitivity is therefore recorded as ‘Insufficient evidence’.

Introduction of light or shading

Benchmark. A change in incident light via anthropogenic means (Introduced light or shade pressure definition).

Evidence

All characterizing species live in the sediment and do not rely on light levels directly to feed or find prey so limited direct impact is expected. As this biotope is not characterized by the presence of primary producers, shading is unlikely to alter the character of the habitat directly.

More general changes to the productivity of the biotope may, however, occur. Beneath shading structures there may be changes in microphytobenthos abundance. Littoral mud and sand support microphytobenthos on the sediment surface and within the sediment. Mucilaginous secretions produced by these algae may stabilise fine substrata (Tait & Dipper, 1998). Shading will prevent photosynthesis leading to death or migration of sediment microalgae altering sediment cohesion and food supply to higher trophic levels. The impact of these indirect effects is difficult to quantify.

Since 2016, research on artificial light at night (ALAN) has expanded considerably in the marine and coastal environment. Light was previously assumed to be of low ecological significance in subtidal and intertidal habitats, but there is now evidence that ALAN is widespread in the marine environment, with biologically relevant levels of light penetrating to depths of up to 50 m (Davies et al., 2020; Smyth et al., 2021). ALAN can alter biological processes across taxa and at multiple levels of organisation. Documented responses include disruption of diel and circalunar rhythms, changes in activity and foraging, altered predator–prey interactions, shifts in community composition, and impacts on algal growth and phenology (Davies et al., 2014, 2015; Gaston et al., 2017; Tidau et al., 2021; Lynn et al., 2022; Marangoni et al., 2022; Miller & Rice, 2023; Ferretti et al., 2025). Evidence for benthic habitats and assemblages specifically is beginning to emerge (e.g. Trethewy et al., 2023; Schaefer et al., 2025), but remains limited and fragmented, often focusing on single taxa or short-term experiments. Mortality thresholds, long-term consequences, and responses at the biotope scale are rarely addressed, and there are major gaps around indirect effects such as trophic cascades or habitat modification.

Sensitivity assessment.  Shading may decrease the availability of microphytobenthos, however, it is not the only food source utilized by the characteristic species. Artificial light is unlikely to affect any but the shallowest biotopes. Given the rapid expansion of the evidence base but the continuing lack of data at the level of individual biotopes, resistance and resilience cannot be robustly assessed. Therefore, sensitivity is recorded as ‘Insufficient evidence’.

Barrier to species movement

Benchmark. A permanent or temporary barrier to species movement over ≥50% of water body width or a 10% change in tidal excursion (Barrier to species movement pressure definition).

Evidence

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

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

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

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

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

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

Death or injury by collision

Benchmark. Injury or mortality from collisions of biota with both static or moving structures due to 0.1% of tidal volume on an average tide, passing through an artificial structure (Death for collision pressure definition).

Evidence

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

Visual disturbance

Benchmark. The daily duration of transient visual cues exceeds 10% of the period of site occupancy by the feature (Visual disturbance pressure definition). 

Evidence

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

Genetic modification & translocation of indigenous species

Benchmark. Translocation of indigenous species or the introduction of genetically modified or genetically different populations of indigenous species may result in changes in the genetic structure of local populations, hybridization, or a change in community structure (Translocation pressure definition).

Evidence

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

Introduction of microbial pathogens

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

Evidence

Limited evidence was returned by searches on the effect on Hediste diversicolor of introduction of relevant microbial pathogens or metazoan disease vectors to an area where they are currently not present. Desrina et al. (2014) failed to induce infection of the ‘White Spot Shrimp’ virus that can cause large scale mortality in shrimp in Hediste diversicolor by both feeding and immersion. Evidence has suggested that Hediste diversicolor is an intermediate host of parasite Dichelyne minutus (Pronkina et al., 2017).

Dairain et al. (2020) reported that disseminated neoplasia, a leukaemia-like disease common amongst marine bivalves, was highly prevalent (21 to 89%) in Macoma balthica (recorded as Limecola balthica) in the Dutch Wadden Sea, and the local prevalence of the disease increased with higher Macoma balthica density. This high disease prevalence may affect population dynamics by increasing mortality and could explain the decline in adult survival in the European Wadden Sea over the past two decades. Further evidence is required to conclude that disseminated neoplasia causes mortality in Macoma balthica, but current literature shows that the disease usually increases mortality amongst affected species.

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).  Parasites and disease are more likely to cause mortalities in populations that are subject to suboptimal conditions or other stressors such as hot summers or cold winters (Longshaw & Malham, 2013).  Experimental infection of Cerastoderma edule with a trematode parasite showed that effects differed depending on habitat conditions (Wegeberg & Jensen, 2003). Infected Cerastoderma edule reared in sub-optimal conditions lost more body weight than infected cockles in more optimal habitats and did not regain condition when placed in higher shore habitats where immersion and food supply was limited.  Infected cockles placed on lower shore sites with longer emersion times regained condition despite the infection and were equivalent to controls. The impact of trematodes is therefore mediated by habitat conditions and in some instances may have no effect (Wegeberg & Jensen, 2003).

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

Removal of target species

Benchmark. Removal of species targeted by fishery, shellfishery or harvesting at a commercial or recreational scale (targeted removal pressure definition).

Evidence

Ragworms Hediste diversicolor are harvested as bait for angling while the abundant bivalves in the biotope Cerastoderma edule and Mya arenia are harvested commercially and recreationally. Removal of target species such as cockles Cerastoderma edule or bait digging for Hediste diversicolor is likely to impact the biotope. Evidence has shown that hand digging of bait worms such as Hediste diversicolor can disturb the characteristics of the intertidal benthic community, can release heavy metals and nutrients from the sediment into the water, and harvesting may affect prey availability of non-target and/or target species (Pombo et al., 2022). The extent of the impact will depend on the fishing / removal method and spatial extent.

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

Following experimental removal of large adult Cerastoderma edule by Frid & Casear (2012) sediments showed increased biodiversity and assemblages dominated by traits common to opportunist taxa at a species-poor shore at Warton Sands, Morecambe Bay, and a more diverse shore at Thurstaston, Dee estuary. The movements of cockles disturb and exclude the amphipod Corophium volutator and other species (Flach, 1996; Flach & de Bruin, 1994) the removal of cockles may, therefore, allow this species to colonize intertidal flats. During periods of low cockle density, Desprez et al. (1992) observed that Pygospio elegans established dense populations; when cockles returned these were lost within one year.

Sensitivity assessment. Removal at a commercial or recreational scale is assessed as not affecting the entire extent of the biotope, but affecting patches within the biotope. Due to potential impacts on Hediste diversicolor populations, in particular females, and impacts on Cerastoderma edule populations the biotope is likely to be sensitive to this pressure. The abundance of other soft-sediment infauna (particularly opportunist species such as Pygospio elegans and Capitella capitata may increase in disturbed patches in the short-term as a result of the removal of cockles resulting in reduced competition for space and predation on larvae). Where sediments remain suitable cockles are likely to recolonize via adult migration, survival of small, discarded cockles or via larval recruitment.  In general fishing practices will be efficient at removing Hediste diversicolor and Cerastoderma edule. Therefore, resistance is assessed as ‘Low’ (removal is not considered to be total as smaller individuals are not retained and harvesting is unlikely to be 100% efficient at removing larger cockles). Resilience is assessed as ‘High’ (although Cerastoderma edule many not recover within this timescale due to episodic recruitment; see Cerastoderma edule dominated biotopes). Biotope sensitivity is assessed as ‘Low’. 

Removal of non-target species

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

Evidence

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

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

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

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

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

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

The American slipper limpet, Crepidula fornicata

Evidence

Intertidal sediments may be colonized by a number of invasive non-indigenous species. Invasive species that alter the character of the biotope or that predate on characterizing species are most likely to result in significant impacts. Intertidal flats may be colonized by the invasive non-indigenous species Crepidula fornicata and the Pacific oyster Magallana gigas. The two species have not only attained considerable biomasses from Scandinavian to Mediterranean countries but have also generated ecological consequences such as alterations of benthic habitats and communities and food chain changes (OSPAR, 2009b).

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

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

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

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

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

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

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

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 above evidence suggests that this biotope is unsuitable for the colonization of Crepidula fornicata due to a lack of gravel, shells, or any other hard substrata used for larvae settlement (Tillin et al., 2020). Despite the sheltered to extremely sheltered conditions of the habitat that would otherwise be suitable for Crepidula, the mobility of the sediment is unsuitable and makes it unlikely for Crepidula to become established. There may be higher densities of Crepidula in the lower shore examples of the biotope, but the densities may be lower in the mid-shore, and it may be absent from the upper shore due to the unsuitable extreme conditions in the intertidal zone preventing Crepidula post-settlement recruitment and mitigating colonization (Bohn et al. 2015). In addition, the reduced salinity levels in this biotope might prevent colonization at high densities, because the majority of evidence records of Crepidula occur in salinities from 30 to 35 psu (OBIS, 2023). The biotope is recorded in the Wash which already harbours Crepidula but is noted for its sediment mobility (Quinn, 2018). In addition, Powell-Jennings & Calloway (2018) noted that Crepidula is killed by sudden burial and possibly burial due to deposition, which could mitigate Crepidula density. Therefore, resistance to colonization by Crepidula fornicata is assessed as 'High' and resilience as 'High' so the biotope is assessed as 'Not sensitive'. The confidence in the assessment is 'Low' because the sensitivity of this biotope to Crepidula is potentially site-specific, there is a risk of its introduction by artificial means, and there is a lack of direct evidence of Crepidula being reported to occur in the biotope.

The carpet sea squirt, Didemnum vexillum

Evidence

The carpet sea squirt Didemnum vexillum (syn. Didemnum vestitum; Didemnum vestum) is a colonial ascidian with rapidly expanding populations that have invaded most temperate coastal regions around the world (Kleeman, 2009; Stefaniak et al., 2012; Tillin et al., 2020). It is an ‘ecosystem engineer’ that can change or modify invaded habitats and alter biodiversity (Griffith et al., 2009; Mercer et al., 2009).

Didemnum vexillum has colonized and established populations in the northeast Pacific, Canadian and USA coast; New Zealand; France, Spain, and the Wadden Sea, Netherlands; the Mediterranean Sea and Adriatic Sea (Bullard et al., 2007; Coutts & Forrest, 2007; Dijkstra et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Lambert, 2009; Hitchin, 2012; Tagliapietra et al., 2012; Gittenberger et al., 2015; Vercaemer et al., 2015; Mckenzie et al., 2017; Cinar & Ozgul, 2023; Holt, 2024).

In the UK, Didemnum vexillum has colonized Holyhead marina and Milford Haven, Wales; the west coast of Scotland (marinas around Largs, Clyde, Loch Creran and Loch Fyne), South Devon (Plymouth, Yealm, and Dartmouth estuaries), the Solent, northern Kent, Essex, and Suffolk coasts (Griffith et al., 2009; Lambert, 2009; Hitchin, 2012; Michin & Nunn, 2013; Bishop et al., 2015; Mckenzie et al., 2017; Tillin et al., 2020, Holt, 2024; NBN, 2024).

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

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

Human-meditated transport via aquaculture facilities, boat hulls, commercial fishing vessels, ballast water is probably the most important vector that has aided the long-distance dispersal of Didemnum vexillum and explains its prevalence in harbours and marinas (Bullard et al., 2007; Dijstra et al., 2007; Griffiths et al., 2009; Herborg et al., 2009). Fragmentation of colonies during transport or human disturbance (such as trawling or dredging) could indirectly disperse the species and enable it to find suitable conditions for establishment (Herborg et al., 2009).

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

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

Didemnum vexillum requires suitable hard substrata for successful settlement and the establishment of colonies. It can grow quickly and can establish large colonies of dense encrusting mats on a variety of hard substrata (Valentine et al., 2007a; Griffith et al., 2009; Lambert, 2009; Groner et al., 2011; Cinar & Ozgul, 2023). Mats can be up to several meters in area, covering large portions of the seafloor (Mercer et al., 2009). Gittenberger (2007) stated that invasive Didemnum sp. was a threat to native ecosystems by its ability to overgrow virtually all hard substrata present. Suitable hard substrata can include rocky substrata such as bedrock gravel, pebble, cobble, or boulders (Tillin et al., 2020). Didemnum vexillum has been reported colonizing these types of hard substrata in the USA, Canada, northern Kent and the Solent (Bullard et al., 2007; Valentine et al., 2007a; Valentine et al., 2007b; Hitchin, 2012; Vercaemer et al., 2015; Tillin et al., 2020). 

There are few observations of Didemnum vexillum on soft bottom habitats as evidence suggests it is unable to establish or grow easily on mud, mobile sand or other unstable substrata, and it is vulnerable to smothering by fine sediment (Bullard et al., 2007; Valentine et al., 2007a; Griffith et al., 2009). The species is usually found established in areas where the colony is protected from sedimentation and wave action (Valentine et al., 2007b; Mckenzie et al., 2017; Tillin et al., 2020). For example, at Georges Bank, USA the Didemnum vexillum mats were limited to gravelly areas and unable to colonize the sand ridges that bounded the site, which have a mobile surface that is moved daily by the strong tidal currents (Valentine et al., 2007b). In addition, evidence found the species can also not survive being buried or smothered by coarse or fine grained sediment. Furthermore, in Holyhead marina, Didemnum vexillum colonies were contained in the harbour and established on artificial pontoons, and they were not present on the natural seabed under the pontoon, which is composed of silty mud or on deeper sections of mooring chains that are immersed in mud at low spring tides (Griffiths et al., 2009). 

However, some studies on Georges Bank, USA and Sandwich, Massachusetts observed colonies were able to survive partial covering by sand (Bullard et al., 2007; Valentine et al., 2007a). Gittenberger et al. (2015) reported that Didemnum vexillum was able to overgrow sandy bottom (cited Gittenberger, 2007). In the Netherlands the coastal zone is composed of mud and sand, with only shells as hard substrata. Didemnum sp. remained rare until 1996 when populations quickly expanded and it became a dominant invasive species because of an increase in available hard substrata for colonization after a cold winter between 1995 and 1996 caused a decrease in the abundance of many marine animals (Gittenberger, 2007). Thus, Didemnum vexillum was able to colonize and establish in mud and sand habitats where hard substrata were present.  

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

Didemnum vexillum has been recorded from less than 1 m to at least 81 m deep (Bullard et al., 2007; Tagliapietra et al., 2012; Tillin et al., 2020). It is abundant across various shore heights, thriving in both nearshore and offshore sites, particularly in subtidal areas. For example, colonies of Didemnum vexillum were dominant at depths between 45 to 60 m, occupying 50 to 90% of available space in two gravelly areas (more than 230 km²) composed of immobile pebble and cobble pavement on Georges Bank fishing ground, USA (Bullard et al., 2007; Valentine et al., 2007b; Lengyel et al., 2009). In addition, patchy mats have been observed covering approximately 1 to 1.5 km² of the pebble cobble seabed, which is interspersed with large boulders and 30 m deep in Long Island Sound, USA (Mercer et al., 2009). In an offshore scallop dredge survey, Didemnum sp. was found attached to cobbles and boulders at 10 to 34 m (Vercaemer et al., 2015). 

Sensitivity assessment. This biotope is likely to be unsuitable for the colonization of Didemnum vexillum due to the lack of gravel, shells, or any other hard substrata used for larval settlement. Despite the sheltered to extremely sheltered conditions of the habitat that would otherwise be suitable for Didemnum, the mobility of the sediment is unsuitable and makes it unlikely for Didemnum to become established. Therefore, resistance is assessed as ‘High’, albeit with low confidence due to no direct evidence of colonization in this biotope. Hence, resilience is assessed as ‘High’ and sensitivity is assessed as ‘Not sensitive’.

The Pacific oyster, Magallana gigas

Evidence

Intertidal sediments may be colonized by a number of invasive non-indigenous species. Invasive species that alter the character of the biotope or that predate on characterizing species are most likely to result in significant impacts. Intertidal flats may be colonized by the invasive non-indigenous species Crepidula fornicata and the Pacific oyster Magallana gigas. The two species have not only attained considerable biomasses from Scandinavian to Mediterranean countries but have also generated ecological consequences such as alterations of benthic habitats and communities and food chain changes (OSPAR, 2009b).

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

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

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

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

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

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

Although shorelines comprised of mainly mud were suggested to be unsuitable for spat settlement (Spencer et al., 1994), the presence of smaller hard substrata, such as shells or pebbles, can enable larvae to settle (Tillin et al., 2020). For example, in the River Teign estuary, Pacific oyster settlement was observed on shell-covered ground mainly attached to mussel shells, and occasionally attached to cockles, stones and common periwinkle (Littorina littorea) shells on a mud flat in the estuarine intertidal zone otherwise mainly comprised of sand and mud (Spencer et al., 1994). In addition, the Blue Lagoon on the north shore of Poole Harbour had the highest abundance of oysters on mud mixed with shingle and shell (Mckinstry & Jensen, 2013). Outside of the Blue Lagoon, oysters were also recorded on mixed substrata composed of mud, gravel, and shell (McKinstry & Jensen, 2013). In the Wadden Sea, the distribution of Magallana gigas on soft sediment shores can overlap with native bivalve species such as Cerastoderma edule, Macoma balthica and Scrobicularia plana (Troost, 2010; Herbert et al., 2012, 2016). However, these native species are likely to occur at higher shore elevations compared to the lower shore habitats preferred by the Pacific oyster (Troost, 2010; Herbert et al., 2012, 2016). For example, in the Wadden Sea greater densities of Cerastoderma edule and Macoma balthica were found above the level of Magallana gigas reef development (Herbert et al., 2012). Tillin et al. (2020) concluded that while successful invasions occurred on mudflats, Magallana gigas prefers mixed substrata. Fine mud sediments without hard substrata (such as small stones, gravel, and shell) are unlikely to be suitable (Tillin et al., 2020). The speed of Magallana gigas reef formation on soft substrata seems to be dependent on the amount of hard substrata present, developing quicker once there is a sufficient amount (Troost, 2010). Bergstrom et al. (2021) reported that the presence of Magallana gigas was partially dependent on increasing gravel content up to 15% but remained stable with increasing percentages (measured up to 80%).

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

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

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

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

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

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

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 above evidence suggests that this biotope is unsuitable for the colonization of Magallana gigas due to a lack of gravel, shells, or any other hard substrata required for successful settlement and establishment (Kochmann, 2012; Kochmann et al., 2013; Mckinstry & Jensen, 2013; Herbert et al., 2016; Tillin et al., 2020). However, some examples of the biotope, have low percentages of hard substrata (such as boulders, cobbles and shell), which could allow the settlement and establishment of Magallana gigas. The distribution of Magallana gigas can overlap with Macoma balthica and other native bivalve species (Troost, 2010; Herbert et al., 2012, 2016). However, the mid-shore and upper-shore extent of this biotope is probably not suitable for colonization of the Pacific oyster, which is found predominantly at the Mean Low Water levels to shallow subtidal (Troost, 2010; Herbert et al., 2012, 2016). But, the oyster could colonize the lower shore extent or examples of the biotope. Therefore, resistance to colonization by Magallana gigas is assessed as 'Medium' as a 'worst-case' scenario. Resilience is assessed as 'Very low' as the Magallana gigas population would need to be removed for recovery to occur. Hence, sensitivity is assessed as 'Medium'. The confidence in the assessment is 'Low' because the sensitivity of this biotope to Magallana gigas is potentially site-specific and there is a risk of its introduction by artificial means. 

Wireweed, Sargassum muticum

Evidence

Wireweed, Sargassum muticum is a circumglobal invasive species (Engelen et al., 2015). Sargassum muticum is known to grow in the shallow subtidal around the UK, usually in areas sheltered from wave action. Its distribution is limited by the availability of hard substratum (e.g. stones >10 cm) and light (Staeher et al., 2000; Strong & Dring 2011; Engelen et al., 2015).  It is most abundant between 1 and 3 m below mean water (Engelen et al., 2015).

Sensitivity assessment. This biotope is likely to be unsuitable for the colonization of Sargassum muticum due to the lack of hard substrata used for larval settlement. Despite the sheltered to extremely sheltered examples of this biotope that would otherwise be suitable for Sargassum, this biotope may be too shallow for colonization. Therefore, resistance is assessed as ‘High’, albeit with low confidence due to no direct evidence of colonization in this biotope. Hence, resilience is assessed as ‘High’ and sensitivity is assessed as ‘Not sensitive’.

Wakame, Undaria pinnatifida

Evidence

Wakame, Undaria pinnatifida is a large brown seaweed and an Invasive Non-Indigenous Species (INIS) that could out-compete native UK kelp species (see Farrell & Fletcher, 2006; Thompson & Schiel, 2012; Brodie et al., 2014; Hieser et al., 2014; Arnold et al., 2016; Epstein & Smale, 2017; Epstein & Smale, 2018; Kraan, 2017; Epstein et al., 2019a,b; Tidbury, 2020).  Undaria pinnatifida was first recorded in the UK in the Hamble Estuary in 1994 (Macleod et al., 2016).  It has since proliferated along UK coastlines. One year after its discovery at the Queen Anne Battery marina, Plymouth, it had become a major fouling plant on pontoons (Minchin & Nunn, 2014).  Although initially restricted to artificial habitats, such as marinas and ports, it is now widespread in natural habitats in several areas, including Plymouth Sound.

Undaria pinnatifida seems to settle better on artificial substrata (e.g. floats, marinas or piers) than on natural rocky shores among local kelps (Vaz-Pinto et al., 2014). It is found predominantly in low intertidal to shallow subtidal habitats (Epstein et al., 2019b) and is significantly more abundant on artificial substrata compared to natural rocky substrata (Heiser et al., 2014; Epstein & Smale, 2018).  James (2017) suggested that Undaria pinnatifida could out-compete native species on artificial substrata (such as marinas and wharf structures). 

Undaria pinnatifida prefers sites sheltered with low wave exposure and weak tidal streams (Heiser et al., 2014; Epstein & Smale, 2018).  In natural habitats, Undaria pinnatifida was not recorded if the wave fetch was greater than 642 km but increased in abundance and cover in very sheltered sites (Epstein & Smale, 2018).

Sensitivity assessment. This biotope is likely to be unsuitable for the colonization of Undaria pinnatifida due to the lack of hard substrata used for larval settlement. Despite the low shore and sheltered to extremely sheltered examples of this biotope that would otherwise be suitable, the mobility of the sediment is unsuitable and makes it unlikely for Undaria to become established. Therefore, resistance is assessed as ‘High’, albeit with low confidence due to no direct evidence of colonization in this biotope. Hence, resilience is assessed as ‘High’ and sensitivity is assessed as ‘Not sensitive’.

Other INIS

Evidence

The Manila clam (Tapes philippinarium), which was introduced to Poole Harbour for aquaculture in 1998, has become a naturalised population on the intertidal mudflats (occurring at densities of 60 clams/m² in some locations within the harbour (Jensen et al. 2004, cited in Caldow et al. 2007). Densities of Cerastoderma edule and Abra tenuis increased following the introduction of the Manila clam although the abundance of Scrobicularia plana and Macoma balthica declined (Caldow et al., 2005), although the decline of these species may have been caused by tri-butyl tin pollution (Langston et al., 2003) and may have facilitated the naturalization of the Manila clam.

The predatory veined whelk (Rapana venosa) and Hemigrapsus takinei are not established in the UK (although Hemigrapsus takinei has been recorded at two locations) could become significant predators of Cerastoderma edule and other species associated with the biotope in the future.

Evidence from the Balthica sea has shown that Macoma balthica and Hediste diversicolor can coexist in soft-sediment communities with invasive clam, Rangia cuneata (Möller & Kotta, 2017; Karlson et al., 2024; Dziaduch et al., 2025). Macoma balthica is forced to switch from suspension feeding to deposit feeding in the presence of Rangia cuneata (Möller & Kotta, 2017).

Sensitivity assessment. The above evidence suggests that the characteristic species in the biotope have not yet been adversely affected by the INIS listed. Therefore, 'Insufficient evidence' is recorded until further evidence becomes available.